# Chapter 7. Integrated Health Effects of Long-Term PM Exposure

7.1. Introduction

This chapter reviews, summarizes, and integrates the evidence on relationships between health effects and long-term exposures to various size fractions and sources of PM. Cardiopulmonary health effects of long-term exposure to PM have been examined in an extensive body of epidemiologic and toxicological studies. Both ep idemiologic and toxicological studies provide a basis for examining reproductive and developmental and cancer health outcomes with regard to long-term exposure to PM. In addition, there is a large body of epidemiologic literature evaluating the relationship between mortality and long-term exposure to PM. Conclusions from the 2004 PM AQCD are summarized briefly at the beginning of each section, and the evaluation of evidence from recent studies builds upon what was available during the previous review. For each health outcome (e.g., respiratory infections, lung function), results are summarized for studies from the specific scientific discipline, i.e., epidemiologic and toxicological studies. The major sections (i.e., cardiovascular , respiratory, reproductive/developmental, cancer) conclude with summaries of the evidence for the various health outcomes within that category and integration of the findings that lead to concl usions regarding causality based upon the framework described in Chapter 1. Determina tion of causality is made for the overall health effect category, such as cardiovascular effects, with coherence a nd plausibility being based upon the evidence from across disciplines and also across the suite of rela ted health outcomes including cause-specific mortality. Section 7.6 provides detailed discussions on the epidemiologic literature for long-term exposure to PM and mortality. In each summary section ( 7.2.11

, 7.3.9

, 7.4.3

, 7.5.4 , and 7.6.5 ), the evidence is briefly reviewed and independent conclusions drawn for relationships with PM 2.5 , PM 10-2.5 , and UF particles (UFPs).



# 7.2. Cardiovascular and Systemic Effects

Studies examining associations between long-term exposure to ambient PM (over months to years) and CVD morbidity had not been conducted a nd thus were not included in the 1996 or 2004 PM Air Quality Criteria Documents (U.S. EPA, 1996, 079380 ; U.S. EPA, 2004, 056905 ). A number of studies were included in the 2004 PM AQCD that evaluated the effect of long-term PM 2.5 exposure on cardiovascular mortality and found consist ent associations. No toxicological studies examined chronic atherosclerotic effects of PM exposure in animal models. However, a subchronic study that evaluated atherosclerosis progression in hyperlipidemic rabbits was discussed and this study provided the foundation for the subsequent work that has been conducted in this area (Suwa et al., 2002, 028588 ). No previous toxicological studies evaluated effects of subchronic or chronic PM exposure on diabetes measures, or HR or HRV change s, nor were there animal toxicological studies included in the 2004 PM AQCD that evaluated systemic inflammatory or blood coagulation markers following subchronic or chronic PM exposure. Several new epidemiologic studies have ex amined the long-term PM-CVD association among U.S. and European populations. The studie s investigate the association of both PM 2.5 and PM 10 exposures with a variety of clinical and subclinical CVD outcomes. Epidemiologic and toxicological studies have provided evidence of the a dverse effects of long-term exposure to PM 2.5 on

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 (ISA) and the Integrated Risk Information System (IRIS)

.

December 2009 7-1cardiovascular outcomes, including atherosclerosis, clinical and subclinical markers of cardiovascular morbidity, and cardiovascular mortality. The evidence of these effects from long-term exposure to PM 10-2.5 is weaker.



## 7.2.1. Atherosclerosis

Atherosclerosis is a progressive disease that contributes to several adverse outcomes, including acute coronary syndromes such as myocar dial infarction, sudden cardiac death, stroke and vascular aneurysms. It is multifaceted, beginning wi th an early injury or inflammation that promotes the extravasation of inflammatory cells. Under cond itions of oxidative or nitrosative stress and high lipid or cholesterol concentrations, the vesse l wall undergoes a chronic remodeling that is characterized by the presence of foam cells, migrated and differentiated smooth muscle cells, and ultimately a fibrous cap. The advanced lesion that develops from this process can occlude perfusion to distal tissue, causing ischemia, and erode, degrade, or even rupture, revealing coagulant initiators (tissue factor) that promote thrombosis, stenosis, an d infarction or stroke. Several detailed reviews of atherosclerosis pathology have beenpublished elsewhere (Ross, 1999, 156926 ; Stocker and Keaney, 2004, 157013 ).



## 7.2. 1.1. Epi demiologic Studies

Measures of Atherosclerosis

Although no study has examined the association between long-term PM exposure and longitudinal change in subclinical markers of athe rosclerosis, several cross sectional studies have been conducted. Markers of atherosclerosis used in these studies include coronary artery calcium (CAC), carotid intima-media thickness (CIMT), an kle-brachial index (ABI), and abdominal aortic calcium (AAC). These measures are descried briefly below. CAC represents the accumulation of calcium in coronary artery macrophages and represents an advanced stage of atherosclerosis. As such CAC is a measure of atherosclerosis assessed by noncontrast, cardiac-gated electron beam computed tomography (EBCT) or multidetector computed tomography (MDCT) of the coronary arteries in the heart (Greenland and Kizilbash, 2005, 156496 ; Hoffmann et al., 2005, 156556 ; Mollet et al., 2005, 155988 ). The prevalence of CAC is strongly related to age. Few people have detectable CAC in their second decade of life but the prevalence of CAC rises to approximately 100% by age 80 (Ardehali et al., 2007, 155662 ). Previous studies suggest that while the absence of CAC does not rule out atherosclerosis, it does imply a very low likelihood of significant arterial obstruction (Achenbach and Daniel, 2001, 156189 ; Arad et al., 1996, 155661 ; Shaw et al., 2003, 156083 ; Shemesh et al., 1996, 156085 ). Conversely, the presence of CAC confirms the existence of atherosclerotic plaque and the amount of calcification varies directly with the likelihood of obstructive disease (Ardehali et al., 2007, 155662 ). CAC is a quantified using the Agatston method (Agatston et al., 1990, 156197 ). Its repeatability depends on the laboratory and the method of calculation (O'Rourke et al., 2000, 192159 ). Agatston scores are frequently used to classify individuals into one of five groups (zero; mild; moderate; severe; extensive) or according to age- and sex-specific pe rcentiles of the CAC distribution (Erbel et al., 2007, 155768 ). CIMT is a measure of atherosclerosis assesse d by high-resolution, B-mode ultrasonography of the carotid arteries in the neck, the walls of which have inner (intimal), middle (medial) and outer (adventitial) layers (Craven et al., 1990, 155740 ; O'Leary et al., 1999, 156826 ; Wendelhag et al., 1993, 157136 ). CIMT estimates the distance in mm or μm between the innermost (blood-intima) and outermost (media-adventitia) interfaces, often by averaging over three arterial segments in the common carotid, carotid bulb, and internal carotid artery (Amato et al., 2007, 155656 ). CIMT has been associated with atherosclerosis risk factors (Heiss et al., 1991, 156535 ; O'Leary et al., 1992, 156825 ; Salonen and Salonen, 1991, 156938 ), prevalent coronary heart disease (Chambless et al., 1997, 156329 ; Geroulakos et al., 1994, 155788 ), and incident coronary and cerebral events (O'Leary et al., 1999, 156826 ; van der Meer et al., 2004, 156129 ). Several studies have indicated that CIMT measurements are accurate (Girerd et al., 1994, 156474 ; Pignoli et al., 1986, 156026 ; Wendelhag et

December 2009 7-2al., 1991, 157135 ) and reproducible (Montauban et al., 1999, 156777 ; Smilde et al., 1997, 156988 ; Willekes et al., 1999, 157147 ), especially for the common carotid artery (Montauban et al., 1999, 156777 ). ABI, which is also known as the ankle-arm or re sting (blood) pressure i ndex, is a measure of lower extremity arterial occlusive disease commonly caused by advanced atherosclerosis (Weitz et al., 1996, 156150 ). It is assessed by continuous wave Doppler and manual or automated oscillometric sphygmomanometry, the latter havi ng been shown to have higher repeatability and validity (Weitz et al., 1996, 156150 ). ABI is defined as the unitless ratio of ankle to brachial systolic blood pressures measured in mmHg. As ankle pressure is normally equal to or slightly higher than arm pressure (resulting in an ABI ≥ 1.0), epidemiologic studies typically define the normal ABI range as 0.90 to 1.50 (Resnick et al., 2004, 156048 ). Low ABI has been associated with all-cause and CVD mortality (Newman et al., 1993, 156805 ; Vogt et al., 1993, 157100 ), as well as myocardial infarction and stroke (Karthikeyan and Lip, 2007, 156626 ). AAC is a measure of atherosclerosis assesse d by non-contrast, EBCT or MDCT of the abdominal aorta. It is scored much like CAC (Agatston et al., 1990, 156197 ), but the age-specific prevalence and extent of AAC is greater, partic ularly among women and at ages >50 yr. Although AAC has not been studied as extensively as CA C, it is associated with carotid and coronary atherosclerosis as well as cardiovascular morbidity and mortality (Allison et al., 2004, 156210 ; Allison et al., 2006, 155653 ; Hollander et al., 2003, 156562 ; Khoury et al., 1997, 156636 ; Oei et al., 2002, 156820 ; Walsh et al., 2002, 157103 ; Wilson et al., 2001, 156159 ; Witteman et al., 1986, 156161 ) and measurements are sufficiently reproducib le to allow serial investigations over time (Budoff et al., 2005, 192105 ).

Study Findings

Diez Roux et al. (2008, 156401 ) conducted cross-sectional analyses of the association of three of these subclinical markers of atherosclerosis (CAC, CIMT and ABI), collected from 2000 to 2003 during baseline examinations of participants enro lled in the Multi-Ethnic Study of Atherosclerosis (MESA), with long-term exposure to PM 2.5 and PM 10 . The study population included 5,172 ethnically diverse people (53% female) residing in Baltimore, MD; Chicago, IL; Forsyth County, NC; Los Angeles, CA; New York, NY; and St. Paul, MN ranging in age from 45 to 84 yr old. Authors used spatio-temporal modeling of pollutant concentrations, weather and demographic data to impute 20-yr avg exposures to PM 2.5 and PM 10 . They reported small increases in CIMT of 1% (95% CI: 0-1.4) and 0.5% (95% CI: 0-1), which correspond to absolute changes of 8 (95% CI: 0-12) and 7 (95% CI: 0-14) μm, per 10 μ g/m 3 increase in 20-yr avg PM 10 and PM 2.5 concentration, respectively. Evidence of age-, gender-, lipid- and smoking-related susceptibility was lacking. They also reported weak, non-significant increases in the relative prevalence of CAC of 1% (95% CI: -2 to 4) and 0.5% (95% CI: -2 to 3) per 10 μ g/m 3 increase in PM 10 and PM 2.5 , respectively. Among the subset of 2,586 participants with EBCT-identified calcification, similarly weak associations were observed. There was little evidence of modifi cation of the CAC associations by demographic, socioeconomic or clinical characteristics. Finally, the authors report no differences in mean ABI with
PM 10 or PM 2.5 concentrations. The null findings for AB I exhibited little heterogeneity among participant subgroups and were similarly null when ABI was modeled as a dichotomous outcome using a cutpoint of 0.9 units. MESA investigators also examined the chronic PM 2.5 -AAC association in a residentially stable subset of 1,147 participants (mean age = 66 yr; 50% female) randomly selected from all MESA centers, except Baltimore, MD for enrollment in its Aortic Calcium Ancillary Study(Allen et al., 2009, 156209 ). The authors used kriging and inverse residence-to-monitor distance-weighted averaging of EPA AQS data to estimate 2-yr mean exposures to PM 2.5 . In cross-sectional analyses, the authors found a 6% (95% CI: -4 to 16) excess risk of a non-zero Agatston score and an 8% (95% CI: -30 to 46) increase in AAC, i.e., approximately 50 (95% CI: -251 to 385) Agatston units, per 10 μ g/m 3 increase in PM 2.5 concentration. These associations were stronger among users than nonusers of lipid lowering drugs. Kunzli et al. (2005, 087387 ) used baseline data collected between 1998-2003 from two randomized placebo-controlled clinical trials, the Vitamin E Atherosclerosis Progression Study (VEAPS) and the B-Vitamin Atherosclerosis Interv ention Trial (BVAIT), for their ancillary crosssectional analyses of the effect of long-term PM 2.5 exposure on CIMT. The study population included 798 residents of the greater Los Angeles, CA area who were more than 40 yr old at baseline and 44% were female. The authors used universal kriging of PM 2.5 data from 23 state and local monitors

December 2009 7-3operating in 2000 to estimate 1-yr avg exposure to PM 2.5 at each participant’s geocoded U.S. Postal Service ZIP code. They found a 4.2% (95% CI: -0.2 to 8.9) or approximately 32 (95% CI: -2 to 68) μ m increase in CIMT per 10 μ g/m 3 increase in PM 2.5 concentration. In contrast to findings from the relatively large, ethnically diverse, ye t geographically overlapping MESA ancillary study described above, PM-related increases in CIMT we re two- to three-fold larger among older and female participants taking lipid lowering drugs in this study. PM-related increases in CIMT were also higher in never smokers when compared with current or former smokers. Hoffmann et al. (2007, 091163 ) conducted a cross-sectional analysis of data collected at baseline (2000-2003) for 4,494 residents of Essen, Mülheim and Bochum, Germany enrolled in the Heinz Nixdorf Recall Study from 2000 to 2003. The age of participants ranged from 45-74 yr and 51% were female. In this cross-sectional study th e authors used dispersion and chemistry transport modeling of emissions, climate and topography data to estimate 1-yr avg exposure to PM 2.5 in 2002 (the midpoint of the baseline exam.) They reported an imprecise 43% (95% CI: -15 to 115) or 102 (95% CI: -77 to 273) Agatston unit increase in CAC per 10 μ g/m 3 increase in PM 2.5 . Differences in strength of association between subgroups defined by de mographic and clinical characteristics were small. The authors reported a more consistent a ssociation of CAC with traffic exposure (distance from a major roadway) than with PM 2.5 in this study. In a subsequent analysis of these data, Hoffmann et al. (2009, 190376 ) examined the PM-ABI association in this population. In this cross-sec tional study, no changes in ABI were observed in association with PM 2.5 concentration nor was evidence of effect modification by demographic and clinical characteristics apparent. As in the previous study (Hoffmann et al., 2007, 091163 ), residing near a major roadway was a stronger predictor of atherosclerotic changes. Absolute changes in ABI of -0.024 (95% CI: -0.047 to -0.001) were associated with living within 50 m of a major roadway compared to living more than 200 m away. Each of the studies described above relied on cross-sectional analyses examining differences in long-term average PM 2.5 concentrations across space (as well as time to the extent baseline examinations were conducted over time). Such asso ciations may reflect the effect of compositional differences in PM 2.5 as well as the effect of higher PM 2.5 concentrations. Most associations of PM 2.5 with CAC (Diez et al., 2008, 156401 ; Hoffmann et al., 2007, 091163 ), CIMT (Diez et al., 2008, 156401 ; Kunzli et al., 2005, 087387 ), ABI (Diez et al., 2008, 156401 ; Hoffmann et al., 2009, 190376 ) and AAC (Allen et al., 2009, 156209 ) reviewed in this section were weak and/or imprecise. However, several factors including exposure measur ement error, variation in baseline measures atherosclerosis, as well as limited power may co ntribute to the insensitivity of these cross-sectional studies to detect small differences in CAC, CIMT, ABI and AAC. The study by Hoffmann et al. (2007, 091163 ), which reported large, imprecise and non- significant increases in CAC in association with PM 2.5 , is not distinguished from the other studies reviewed by a superior study design or larger sample size. The several fold difference in th e magnitude of CIMT associations reported by Kunzli et al. (2005, 087387 ) and Diez Roux et al. (2008, 156401 ) may be related to differences between the study populations. The ambient PM concentratio ns from these studies are characterized in Table 7-1

.



## 7.2. 1.2. Tox icological Studies

In the only study of this kind described in the 2004 PM AQCD, Suwa et al. (2002, 028588 ) demonstrated more advanced atherosclerotic le sions based on phenotype and volume fraction in the left main and right coronary arteries of rabbits exposed to PM 10 (5 mg/kg, 2 times/wk×4 wk). Although this study was conducted using IT exposure methodology at a relatively high dose, it provided the first experimental evidence that PM exposure may result in progression of atherosclerosis. Recent toxicological studies conducted using inhalation exposures have replicated these findings at relevant concentrations and are discussed below.



### CAPs

New studies have demonstrated increased athe rosclerotic plaque area in aortas of ApoE -/- mice exposed to PM 2.5 CAPs for 4-6 mo (6 h/day×5 days/wk). Average CAPs concentrations ranged from 85 to 138 μ g/m 3 and all of the studies were conducted in Tuxedo or Manhattan, NY. Chen and Nadziejko (2005, 087219 ) reported that the percentage of aortic intimal surface covered by atherosclerotic lesions in ApoE -/- mice was increased. In male ApoE -/- /LDLR -/- mice, both lesion area

December 2009 7-4and cellularity in the aortic root were enhanced by Tuxedo, NY CAPs exposure, although there was no change in lipid content. Genetic prof iles within plaques recovered from ApoE -/- mice included many of the molecular pathways known to contribute to atherosclerosis, including inflammation (Floyd et al., 2009, 190350 ). Sun (2005, 087952 ) similarly demonstrated an enhancement of atherosclerosis in ApoE -/- mice exposed Tuxedo, NY CAPs. Pla que area in the aortic arch and abdominal aorta was significantly increased in the PM-exposed, high fat-chow group compared to air-exposed, high fat-chow group. Macrophage infilt ration in the abdominal aorta was also observed in the groups exposed to CAPs. A study conducted in Manhattan for 4 mo (May- September 2007) showed that PM 2.5 CAPs exposure increased atherosclerotic pl aque area and led to higher levels of macrophage infiltration, collagen deposition, a nd lipid composition in thoracic aortas of ApoE -/- mice (Ying et al., 2009, 190111 ), which is consistent with the previous two studies described that were conducted in Tuxedo, NY. Alteration of vasomotor function has been observed in aortic rings of ApoE –/– mice on a high fat diet with long-term exposure to CAPs (Sun et al., 2005, 087952 ; Ying et al., 2009, 190111 ). Sun (2005, 087952 ) reported that. PM 2.5 -exposed animals exhibited increased vasoconstrictor responsiveness to serotonin and PE. Increased ROS a nd elevated iNOS protein expression in aortic sections of CAPs-exposed mice may have resulted alterations in the NO pathway and generation of peroxynitrite that could have affected vascular reactivity. In contrast, Ying, et al. (2009, 190111 ) demonstrated decreased maximum constriction indu ced by PE following Manhattan CAPs exposure. Pretreatment with the soluble guanylate cyclase (sGC) inhibitor ODQ attenuated the response, indicating that CAPs exposure resulted in abnorma l NO/sGC signaling. Expression of iNOS mRNA and protein was increased in aortas of CAPs-e xposed mice, further supporting a role for NO production. In conjunction with increased NO, ao rtic superoxide production was demonstrated that appeared to be partially driven by increased NADPH oxidase activity. The difference in vasoconstrictor responses between these two studies may be attributable to varying durations (6 versus 4 mo, respectively) or CAPs compositions. Sun (2005, 087952 ) and Ying et al. (2009, 190111 ) reported similar relaxation responses to ACh for air- and CAPs-exposed mice. Howeve r, Manhattan CAPs-exposed mice had a markedly decreased response to A23187, indicating that NO release occurred via Ca 2+ -dependent mechanisms (Ying et al., 2009, 190111 ). Abnormal eNOS function is likely responsible for the decreased relaxation response, as activation of eNOS (but not iNOS) is Ca 2+ -dependent. A recent study (Sun et al., 2008, 157033 ) that was part of the research described above (Sun et al., 2005, 087952 ) investigated tissue factor (TF) expression in aortas, which is a major regulator of hemostasis and thrombosis following vascular injury or plaque erosion. In PM 2.5 -exposed ApoE -/- mice on a high-fat diet, TF was significantly elevated in the plaques of aortic sections compared to air-exposed mice on the high-fat diet. TF expressi on was generally detected in (1) the extracellular matrix surrounding macrophages and foam cell-rich areas; and (2) around smooth muscle cells. One new study of CAPs PM 2.5 or UFPs derived from traffic was conducted. Araujo et al. (2008, 156222 ) compared the relative impact of UF (0.01-0.18 μ m) and fine (0.01-2.5 μ m) PM inhalation on aortic lesion development in ApoE -/- mice following a 40-day exposure (5 h/day×3 days/wk for 75 total h). Animals were on a normal chow diet and exposed to CAPs in a mobile inhalation laboratory parked 300 m from a freeway in downtown Los Angeles. Exposure concentrations were ~440 μ g/m 3 for PM 2.5 and ~110 μ g/m 3 for UFPs, and the number concentrations were roughly equivalent (4.56×10 5 and 5.59×10 5 particles/cm 3 for PM 2.5 and UFPs, respectively). Significant increases in plaque size (estimated by lesions at the aortic root) were reported for mice exposed to UFPs only. The lesions were largely co mprised of macrophages with intracellular lipid accumulation. Increased total cholesterol measured at the end of the exposure protocol was observed only in the PM 2.5 group. HDL isolated from the UF PM-exposed mice demonstrated decreased antiinflammatory protective capacity against LDL-induced monocyte chemotactic activity in an in vitro assay. The livers from the UFP-exposed mice demonstrated significant increases in lipid peroxidation and several stress-related gene pr oducts (catalase, glutathione S-transferase Y

a , NADPH-quinone oxidoreductase1, su peroxide dismutase 2). Thus, UF Ps in these exposures had a substantially greater impact on the systemic response than did PM 2.5

.

December 2009 7-5

### Ambient Air

A study employing young BALB/c mice examined the effects of a 4-month exposure (24 h/day×7 days/wk) to ambient air on arte rial histopathology (Lemos et al., 2006, 088594 ). Outdoor exposure chambers were located in downtown Sao Paulo, Brazil next to streets of high traffic density. In the control chamber, PM 10 and NO 2 were filtered with 50% and 75% efficiency, respectively. The average pollutant concentra tions were 2.06 ppm for CO (8-h mean), 104.75 μ g/m 3

for NO 2 (24-h mean), 11.07 μ g/m 3 for SO 2 (24-h mean), and 35.52 μ g/m 3 for PM 10 (24-h mean) at a monitoring site within 100 m of the inhalation chambers. The pulmonary and coronary arteries demonstrated significant decreases in L/W ratio for animals exposed to the entire ambient mixture compared to controls, indicating thicker walls in these vessels. There was no difference reported for the L/W ratio in renal arteries. Morphologic exam ination suggested that the increases in L/W ratio were due to muscular hypertrophy rather than fibr osis. The results of this study indicate vascular remodeling of the pulmonary and coronary arteries, as opposed to changes in tone. To examine the role of systemic inflammation an d recruitment of monocytes into plaque tissue as a possible pathway for accelerated atherosclerosis, Yatera et al. (2008, 157162 ) exposed female Watanabe heritable hyperlipidemic rabbits (42 week old) to Ottawa PM 10 (EHC-93) via IT instillation (5 mg/rabbit; approximately 1.56 mg/kg) twice a week for 4 wk. Transfusion of whole blood harvested to from exposed and non-exposed animals to donor rabbits supplied labeled monocytes for assessment of monocyte recruitment from the blood to the aortic wall. The fraction of aortic surface and volume of aortic wall taken up by atherosclerotic plaque was increased and the number of labeled monocytes in the atherosclerotic plaques was elevated in rabbits exposed to PM 10

. In addition, labeled monocytes were attached onto the endothelium overlying atherosclerotic plaques and the number that migrated into the smooth musc le underneath plaques in aortic vessel walls was greater with PM 10 exposure compared to control. These responses were not observed in normal vessel walls. ICAM-1 and VCAM-1 expression was elevated in atherosclerotic lesions, likely indicating enhanced monocyte adhesion to endothel ium and migration into plaques. Monocytes in plaque tissue stained with immunogold demonstrate d foam cell characteristics, which were more numerous in the rabbits exposed to PM 10

.



### Gasoline Exhaust

Lund and colleagues (2007, 125741 ) used whole emissions from gasoline exhaust to investigate changes in the transcriptional regulati on of several gene products with known roles in both the chronic promotion and acute degradation/destabilization of atheromatous plaques. These 50-day exposures (6 h/day×7 days/wk) employed ApoE -/- mice on high-fat chow and the concentrations of the high exposure group were 61 μ g/m 3 for PM, 19 ppm for NO X , 80 ppm for CO, and 12.0 ppm for total hydrocarbons. The aver age particle number median diameter was approximately 15 nm (McDonald et al., 2007, 156746 ). Dilutions of gasoline engine emissions induced a concentration-dependent increase in tr anscription of matrix metalloproteinase (MMP) isoform 9, ET-1, and HO-1 in aortas; MMP-3 and -9 mRNA levels were only increased in animals in the highest exposure group. Strong increases in oxidative stress markers (nitrotyrosine and TBARS) in the aortas were also observed. However, using a hi gh-efficiency particle trap, they established that most of the effects were caused by the gaseous portion of the emissions and not the particles. This study did not directly address lesion area.



## 7.2.2. Venous Thromboembolism

One epidemiologic study examined the relationship between long term PM 10 concentration, venous thromboembolism, and laboratory measures of hemostasis (prothrombin and activated partial thomboplastin times [PT; PTT]). PT and PTT measure the extrinsic and intrinsic blood coagulation pathways, the former activated in response to blood vessel injury, the latter, key to subsequent amplification of the coagulation cascade and propagation of thrombus (Mackman et al., 2007, 156723 ). Decreases in PT and PTT are consistent wi th a hypercoagulable, prothrombotic state.

December 2009 7-6

## 7.2. 2.1. Epi demiologic Studies

Baccarelli et al. (2008, 157984 ) studied 2,081 residents (56% female) of the Lombardy region of Italy whose ages ranged from 18 to 84 yr old. In this case-control study of 871 patients with ultrasonographically or venographically diagnosed lo wer extremity deep vein thrombosis (DVT) and 1,210 of their healthy friends or relatives (1995-2005), the authors used arithmetic averaging of PM 10 data available at 53 monitors in nine geographic areas to estimate 1-yr avg residence-specific exposures. They found -0.06 (95% CI: -0.11 to 0) and -0.12 (95% CI: -0.23 to 0) decreases in standardized correlation coefficients for PT as we ll as 0.01 (95% CI: -0.03 to 0.04) and -0.09 (95% CI: -0.19 to 0.01) decreases in standardized correlation coefficients for PTT among cases and controls, respectively, per 10 μ g/m 3 increase in PM 10 . Patients with DVT who were taking heparin or coumarin anticoagulants were not asked to stop taking them before measurement of PT and aPTT. Of additional note, PT was neither adjusted for differences in reagents used to determine it nor conventionally reported as the International Normalized Ratio (INR). The ambient PM concentrations from this study are characterized in Table 7-1

.



## 7.2.3. Metabolic Syndromes

7.2. 3.1. Epi demiologic Studies

Chen and Schwartz (2008, 190106 ) studied 2,978 residentially stable participants in 33 U.S. communities (age range = 20-89 yr; 49% female) who were examined during phase 1 of the National Health and Nutrition Examination Survey III (1989-1991). In this cross-sectional study, the authors used inverse-distance weighted averaging of U. S. EPA AQS monitored data from participant and adjacent counties of residence to estimate 1-yr avg exposures to PM 10 . They found that after adjustment, residents of communities with lower PM 10 concentrations had fewer white blood cells than residents of communities with higher PM 10 concentrations. This difference increased with increasing number of metabolic abnormalities (insulin resistance; hypertension; hypertriglyceridemia; low high-density lipoprotein cholesterol; abdominal obesity) reported by the participant. This observed difference across i ndividuals with different degrees of metabolic abnormalities supports the concept that the presen ce of a metabolic syndrome may impart greater susceptibility to PM-associated long-term CVD effects.



## 7.2. 3.2. Tox icological Studies

Diabetics as a potentially susceptible subpopulation have only recently been evaluated. A toxicological study of a diet-induced obesity mouse model (C57BL/6 fed high-fat chow for 10 wk) examined the effects of a 128-day PM 2.5 CAPs exposure (mean mass concentration 72.7 μ g/m 3 ; Tuxedo, NY) on insulin resistance, adipose infla mmation, and visceral adiposity (Sun et al., 2009, 190487 ). Elevated fasting glucose and insulin levels were observed in CAPs-exposed mice compared to air-exposed during the glucose tolerance test. Ao rtic rings of mice exposed to CAPs demonstrated decreased peak relaxation to ACh or insulin, whic h was associated with reduced NO bioavailability. Additionally, insulin signaling was impaired in aortic tissue via lowered endothelial Akt phosphorylation. Increases in adipokines and systemic inflammatory markers (i.e., TNF- α , IL-6, E-selectin, ICAM-1, PAI-1, resistin, leptin) were reported for CAPs-exposed mice. CAPs resulted in increased visceral and mesenteric fat mass, as well as greater adipose tissue macrophages in epididymal fat pads and larger adipocyte size compared to mice in the filtered air group. The results of this study demonstrate that PM 2.5 exposure can exaggerate insulin resistance, visceral adiposity, and inflammation in mice fed high-fat chow.

December 2009 7-7

## 7.2.4. Systemic Inflammation, Immune Fu nction, and Blood Coagulation

7.2. 4.1. Ep idemiologic Studies

As discussed in Section 7.2.3.1 , Chen and Schwartz (2008, 190106 ) conducted a crosssectional study in 33 U.S. communities and used in verse-distance weighted averaging of U.S. EPA AQS monitored data from participant and adjacentcounties of residence to estimate 1-yr avg exposures to PM 10 (median concentration within quart iles = 23.1, 31.2, 38.8 and 53.7 μg/m 3 ). They found that after adjustment, residents of communitie s in quartile 1 had 138 (95% CI: 2-273) fewer white blood cells (×10 6 /L) than residents of communities in quar tiles 2-4. This difference increased with increasing number of metabolic abnormalities. Forbes et al. (2009, 190351 ) studied approximately 25,000 adults (age ≥ 16 yr; 53% female) who were representatively sampled from 720 English postcode sectors and participated in the Health Survey for England (1994, 1998 and 2003). In this fixed-effects meta-analysi s of year-specific crosssectional findings, the authors used dispersion mode ling of emissions and weather data to estimate 2-yr avg exposures to PM 10 at participant postcode sector centroids (median in 1994, 1998 and 2003 = 19.5, 17.9 and 16.2 μg/m 3 , respectively). They found little evidence of a PM 10 -inflammatory marker association, i.e., only a -0.08% (95% CI: -0 .25 to 0.10) decrease in fibrinogen concentration and a 0.14% (95% CI: -1.00 to 1.30) increase in CRP concentration per 1 μ g/m 3 increase in PM 10

. Calderon-Garciduenas et al. (2007, 091252 ) compared residentially stable, non-smoking healthy children (age range: 6-13 yr) living and attending school between 2003-2004 in Mexico City (historically high PM; altitude 2,250 m) and Polotitl án (historically low PM; altitude 2,380 m). In this ecologic study, residents of Mexico City (n = 59; 93% female) had fewer white blood cells and neutrophils (×10 9 /L) than residents of Polotitlán (n = 22; 69% female): unadjusted mean 6.2 (95% CI: 5.7-6.6) versus 6.9 (95% CI: 6.3-7.5) and 2.9 (95% CI: 2.3-3.5) versus 3.8 (95% CI: 3.2-4.4), respectively. Calderon-Garciduenas et al. (2009, 192107 ) subsequently compared 37 unadjusted mean measures of immune function and inflammation am ong an expanded number of these participants. They found that under a two-sided type I error rate ( α ) = 0.05, 16 (43%) of the measures were significantly different in residents of southwest Mexico City (n = 66; 48% female) than those in Polotitlán (n = 93; 57% female). However, only 8 measures were significantly different after Bon Ferroni-correction ( α = 0.05 / 37 = 0.001) and even fewer would be after adjustment for reported correlation between the measures of immune function and inflammation, e.g., CRP and lipopolysaccharide binding protein (Pearson’s r = 0.71). Two cross-sectional analyses of PM 10 concentration and markers of immune function or inflammation have been conducted with signifi cant changes observed in the NHANES population (stronger effects among those with metabolic disorders) (Chen and Schwartz, 2008, 190106 ) but not in a relative large survey of adults, which was conducted in England (Forbes et al., 2009, 190351 ). Ecological analyses comparing children in high versus low pollution regions in Mexico show differences in unadjusted blood markers that may be related to PM concentration or other unmeasured risk factors that di ffers across the communities studie d (Calderon-Garciduenas et al., 2007, 091252 ; Calderón-Garcidueñas et al., 2009, 192107 ).



## 7.2. 4.2. Tox icological Studies

In addition to the PM 2.5 study mentioned previously that show ed increased TF expression (an important initiator of thrombosis) in aortas of ApoE -/- mice following subchronic CAPs exposure (Sun et al., 2008, 157033 ), three recent studies examined hema tology and clotting parameters in rats and mice exposed to DE, gasoline exhaust, or hardwo od smoke for 1 week or 6 mo (Reed et al., 2004, 055625 ; Reed et al., 2006, 156043 ; Reed et al., 2008, 156903 ). In all studies, male and female F344 rats were exposed to the mixtures by whole- body inhalation for 6 h/day, 7 day/wk. Respiratory effects for these studies are presented in Section 7.3.3

.

December 2009 7-8

### Diesel Exhaust

The target PM concentrations in the DE study was 30, 100, 300, and 1,000 μ g/m 3 and the MMAD was 0.10-0.15 μ m (Reed et al., 2004, 055625 ). Male and female rats exposed to DE at the highest concentration (NO concentration 45.3 ppm; NO 2 concentration 4.0 ppm; CO concentration 29.8 ppm; SO 2 concentration 365 ppb) for 6 mo demonstrated decreased serum Factor VII, but no change in plasma fibrinogen or thrombin an ti-thrombin complex (TAT) (Reed et al., 2004, 055625 ). White blood cells were decreased only in female rats in the highest exposure group. Another DE study of shorter duration (4 wk, 4 h/day, 5day/wk; PM mass concentration 507 or 2201 μ g/m 3 , CO 1.3 and 4.8 ppm, NO <2.5 and 5.9 ppm, NO 2 <0.25 and 1.2 ppm, SO 2 0.2 and 0.3 ppm for low and high PM exposures, respectively) did not demonstra te changes in hematologic parameters or those related to coagulation (i.e., PT, PPT, plasma fibrinogen, D-dimer) or inflammation (i.e., CRP) in SH or WKY rats (Gottipolu et al., 2009, 190360 ). Together, these findings do not support a DE-related stimulation of blood coagulation following 1 or 6 mo of exposure.



### Hardwood Smoke

The target PM concentrations in the hardwood smoke study was 30, 100, 300, and 1,000 μ g/m 3

and the MMAD was 0.25-0.36 μ m (Reed et al., 2006, 156043 ). In male rats exposed to hardwood smoke, the mid-low group (PM concentration 113 μ g/m 3 ; NO, NO 2 , SO 2 concentrations 0 ppm; CO concentration 1,832.3 ppm) had the greatest responses in hematology parameters, including increased hematocrit, hemoglobin, lymphocytes, a nd decreased segmented neutrophils (Reed et al., 2006, 156043 ). Platelets were elevated in male and fema le rats after 1 week of exposure, but this response returned to control values following the 6-month exposure. No changes were observed for any coagulation markers at 6 mo.



### Gasoline Exhaust

PM mass in the gasoline exhaust study ranged from 6.6 to 59.1 μ g/m 3 , with the corresponding number concentration between 2.6×10 4 and 5.0×10 5 particles/cm 3 ; the dilutions for the gasoline exhaust were 1:10, 1:15 or 1:90 and filtered PM at the 1:10 dilution (Reed et al., 2008, 156903 ). Similar to the responses observed with hardwood smoke, male and female rats in the mid- and highgasoline exhaust exposure groups (NO concentrations 11.9 and 18.4 ppm; NO 2 concentrations 0.5 and 0.9 ppm; CO concentration 73.2 and 107.3 ppm; SO 2 concentration 0.38 and 0.62 ppm, respectively) demonstrated elevated hematocrit and hemoglobin; RBC count was also elevated in these groups (Reed et al., 2008, 156903 ). The only response that appeared somewhat dependent on the presence of particles was incr eased RBC in female rats at 6 mo , although the authors attributed the observed increases to the high concentration of CO. Collectively, these studies do not indicate robust systemic inflammation or coagulation responses in F344 rats following 6-month exposures to diesel, hardwood smoke, or gasoline exhaust. The limited effects that were observed could possibly be due to the varying gas concentrations in the exposure mixtures.



## 7.2.5. Renal and Vascular Function

Two recent epidemiologic studies have te sted associations between PM exposure and indicators of renal and vascular function (urinary albumin to creatinine ratio [UACR] and blood pressure). UACR is a measure of urinary albumi n excretion (National Kidney Foundation, 2008, 156796 ). When calculated as the ratio of albumin to creatinine concentrations in untimed (“spot”) urine samples, UACR approximates 24-h urinary al bumin excretion and can be used to identify albuminuria, a marker of generalized vasc ular endothelial damage (Xu et al., 2008, 157157 ). Values
≥ 30 mg/g (3.5 mg/mmol) and ≥ 300 mg/g (34 mg/mmol) usually define micro- and macroalbuminuria, both of which are associated wi th increases in CVD incidence and mortality (Bigazzi et al., 1998, 156272 ; Deckert et al., 1996, 156389 ; Dinneen and Gerstein, 1997, 156403 ; Gerstein et al., 2001, 156466 ; Mogensen, 1984, 156769 ). Several researchers have called the

December 2009 7-9dichotomization of albuminuria into question, obse rving that there is no threshold below which risk of cardiovascular and end-stage kidney dis ease disappears (Forman and Brenner, 2006, 156439 ; Knight and Curhan, 2003, 179900 ; Ruggenenti and Remuzzi, 2006, 156933 ). Systolic, diastolic, pulse, and mean arterial blood pressures (SBP; DBP; PP; MAP) in mmHg have also been used as measures of cardiova scular disease. Franklin et al. (1997, 156446 ) suggested that SBP and PP were the only two measures predicti ve of carotid stenosis in a multivariable analysis considering all 4 measures, whereas Khattar et al. (2001, 155896 ) suggested that their prognostic significance in hypertensive populations may differ by age, with SBP and PP being most predictive among those ≥ 60 yr and DBP among those <60 yr old (Khattar et al., 2001, 155896 ).



## 7.2. 5.1. Epi demiologic Studies

O’Neill et al. (2007, 156006 ) examined the association of UACR with PM 2.5 and PM 10 among members of the MESA population described previously (Diez et al., 2008, 156401 ). For this study of UACR, which included cross-sectional and longitudina l analyses, the study population was restricted to a subset of 3,901 participants (mean age = 63 yr ; 52% female) with complete covariate, outcome and exposure data at their first through third ex ams (2000-2004). In cross-sectional analyses, the authors found that after adjustment for de mographic and clinical characteristics, 10 μ g/m 3 increases in 20-yr imputed exposures to PM 2.5 and PM 10 were associated with negligible 0.002 (95% CI: -0.048 to 0.052) and -0.002 (95% CI: -0.038 to 0.035) mean differences in baseline log UACR, respectively. Similarly, small statistically non- significant decreases in the prevalence of microalbuminuria (defined in this setting as ≥ 25 mg/g) provided little evidence of an effect on renal function. These largely null cross-sectional findings mirrored those based on the study’s shorter-term (30- and 60-day) PM 2.5 and PM 10 exposures. Moreover, longitudinal analyses revealed only a weak association between 3-yr change in log UACR and 20-yr PM 10 exposure. Evidence of effect modification by demographic and geographic character istics was not apparent in either the crosssectional or longitudinal analyses. Auchincloss et al. (2008, 156234 ) focused on automated, oscillometric, sphygmomanometric measures of blood pressures in mmHg ( SBP; DBP; PP; MAP). Like O’Neill (2007, 156006 ), Diez et al. (2008, 156401 ) and Allen et al. (2007, 156006 ), Auchincloss et al. (2008, 156234 ) based their examination on the previously described MESA population. The authors included 5,112 study participants (age range = 45-84 yr; 52% female) who were free of clinically manifested CVD at their baseline exam in one of six primarily urban U.S. locations (2000-2002). In this cross-sectional study, they used arithmetic averaging of EPA AQS PM 2.5 data available at the monitor nearest to each participant’s geocoded U.S. Postal Service ZIP code centroid to estimate 30- and 60-day avg exposures to PM 2.5 . They found small nonsignificant increases of 1.5 (95% CI: -0.2 to 3.2), 0.2 (95% CI: -0.7 to 1.0), 1.3 (95% CI: 0.1 to 2.6), and 0.6 (95% CI: -0.4 to 1.7) mmHg increases in SBP, DBP, PP and MAP, respectively, per 10 μg/m 3 increase in 30-day avg PM 2.5 exposure, Associations were slightly weaker for 60-day avg PM 2.5 exposure and among participants without hypertension, during cooler weather, in the presence of low NO 2 , residing >300 m from a highway, or surrounded by lower road density. Finally, the Calderon-Garciduenas et al. (2007, 091252 ) ecologic study introduced in Section 7.2.3.1 also found that children residing in Mexico City had higher mean pulmonary artery pressure as assessed by Doppler echocardiography and fasting plasma endothelin-1 (ET-1) than residents in Polotitlán: unadjusted mean 17.5 (95% CI: 15.7-19.4) versus 14.6 (95% CI: 13.8-15.4) mmHg and 2.23 (95% CI: 1.93-2.53) versus 1.23 (95% CI: 1.11-1.35) pg/mL, respectively. Within Mexico City, ET-1 was higher in residents of the Northeast (historically higher PM 2.5 ) than those of the Southwest (historically lower PM 2.5 ). The MESA analyses of UACR (O'Neill et al., 2007, 156006 ) and the ecologic study of children living in a highly polluted area of Mexico (Calderon-Garciduenas et al., 2007, 091252 ) provide little evidence that long-term exposure to PM 2.5 had an effect on renal and vascular function, respectively. Auchincloss et al. (2008, 156234 ) reports small nonsignificant associations of blood pressure with 30- and 60-day avg PM 2.5 concentrations. PM concentrations from the analyses are characterized in Table 7-1

.

December 2009 7-10

### Table 7-1. Characterization of am bient PM concentrations from st udies of subclinical measures of cardiovascular diseases and long-term exposure.

Study Location Mean Concentration ( μ g/m 3 ) Upper Percentile Concentrations ( μ g/m 3 )

PM 10

Diez Roux et al. (2008, 156401 ) MESA: 6 Cities U.S. 20 yr imputed mean: 34 NR

O’Neill et al. (2007, 156006 ) MESA: 6 Cities U.S. Long-Term Exposure:

1982-2002: 34.7 1982-1987: 40.5 1988-1992: 38 1993-1997: 30.6 1998-2002: 29.7 Previous Month: 27.5 NR

Baccarelli et al. (2008, 157984 ) Lombardy Region Italy NR NR

Rosenlund et al. (2006, 089796 ) Stockholm, Sweden 30-y avg PM 10 (traffic) Cases: 2.6 Controls: 2.4 5th-95th %: 0.5-6 0.6-5.9

Chen and Swartz (2008, 190106 ) US Population (NHANES) Annual avg: 36.8 NR

Forbes et al. (2009, 190351 ) British Population 1994: 19.5 (median)

1998: 17.9 (median)

2003: 16.2 (median) 1994, Min-Max: 12.5-36.1

1998, Min-Max: 12.6-27.0

2003, Min-Max: 11.0-22.7

PM 2.5

Hoffmann et al.(2007, 091163 ) HNRS, 3 Cities Germany Annual avg: 22.8 NR

Allen et al. (2009, 156209 ) MESA: 5 Cities Annual avg: 15.8 Min-Max: 10.6-24.7

Kunzli et al. (2005, 087387 ) VEAPS BVAIT Annual avg: 20.3 Min-Max: 5.2-26.9

Auchincloss et al. (2008, 156234 ) MESA: 6 Cities Prior 30 days: 16.8 Prior 60 days: 16.7 NR

O’Neill (2007, 156006 ) MESA: 6 Cities U.S. Previous Month: 16.5 NR

Diez Roux et al. (2008, 156401 ) MESA: 6 Cities U.S. 20-y imputed mean: 21.7 NR

Hoffmann et al. (2009, 190376 ) HNRS: 3 Cities Germany Annual avg: 22.8 Min-max: 19.8-26.8

Calderon-Garciduenas et al. (2009, 192107 ) Southwest Mexico (high pollution)

Potitlan (low pollution) Annual avg: 25

Annual avg: <15 NR

NR

Calderon-Garciduenas et al. (2007, 091252 ) Southwest Mexico (high pollution)

Potitlan (low pollution) NR

NR NR

NR

MESA: Multi-Ethnic Study of Atherosclerosis HNRS: Heinz Nixdorf Recall Study VEAPS: Vitamin E Atherosclerosis Progression Study

BVAIT: B-Vitamin Atherosclerosis Intervention Trial



## 7.2. 5.2. Tox icological Studies

In a PM 2.5 CAPs study of 10 wk (6 h/day×5 days/wk) in Tuxedo, NY (mean mass concentration 79.1 μ g/m 3 ), there was no difference in mean arterial pressure (MAP) in SD rats between groups (Sun et al., 2008, 157032 ). When angiotensin II (Ang II) was infused during the last week of exposure to induce systemic hypertension, the MAP slope was consistently greater in the CAPs-exposed rats compared to the filtered air gro up. Furthermore, thoracic aortic rings were more responsive to phenylephrine-induced constriction and less responsive to ACh-induced relaxation in the PM+Ang II vessels. In contrast to the latter findings, the relaxation response was exaggerated in the PM+Ang II aortic segments with a Rho-kinase (ROCK) inhibitor. Superoxide production in aortic rings increased in the PM+Ang II group compared to the filtered air group and the addition of

December 2009 7-11NAD(P)H oxidase inhibitor (apocynin) or a NOS inhibitor (L-NAME) attenuated the superoxide generation. The levels of tetrahydrobiopterin (BH 4 ) were decreased in mesenteric vasculature and the heart by 46% and 41% in the PM+Ang II group compar ed to controls, respectively; furthermore, levels of BH 4 in the liver were similarly reduced, which is consistent with a systemic effect of CAPs. Together, these findings indicate that CAPs potentiate Ang II-induced hypertension and alter vascular reactivity, perhaps through activated NADPH oxidase and eNOS uncoupling that result in oxidative stress generation and triggeri ng of the Rho/ROCK signaling pathway.



## 7.2.6. Autonomic Function

7.2. 6.1. Tox icological Studies

Hwang et al. (2005, 087957 ) and Chen and Hwang (2005, 087218 ) used radiotelemetry to examine the chronic changes in HR and HRV resulting from the same CAPs exposures described previously (Chen and Nadziejko, 2005, 087219 ). The overall average CAPs exposure concentration was 133 μ g/m 3 and results indicate differing responses to CAPs between ApoE -/- mice and their genetic background strain, C57BL/6J mice (Hwang et al., 2005, 087957 ). Using the time period of 1:30-4:30 a.m., C57BL/6J mice showed a HR increas e only over the last month of exposure. In contrast, ApoE -/- mice had chronic decreases of 33.8 beat/min for HR. Changes in HRV (SDNN and rMSSD) were somewhat more complicated, with biphasic responses in ApoE -/- mice over the 5-month period (initial increase over first 6 wk, decrease over next 12 wk, and slight upward turn for remainder of the study)(Chen and Hwang, 2005, 087218 ). Increasing linear trends were observed in C57BL/6J mice for SDNN and rMSSD. The average CAPs concentration for the HRV study was 110 μ g/m 3 . However, only three C57BL/6J mice in the exposure group were included in the analysis compared to ten ApoE -/- animals, thus making it difficult to interpret the C57BL/6J mice responses (Chen and Hwang, 2005, 087218 ; Hwang et al., 2005, 087957 ).



## 7.2.7. Cardiac changes

7.2. 7.1. Tox icological studies

Two recent toxicological studies have evaluated the effects of PM on cardiac effects including pathology and gene expression. Cardiac mitochondr ial function has also been evaluated following PM exposure in rats.



### Diesel Exhaust

A recent study of DE exposure (PM mass concentration 507 or 2,201 μ g/m 3 , CO 1.3 or 4.8 ppm, NO <2.5 or 5.9 ppm, NO 2 <0.25 or 1.2 ppm, SO 2 0.2 or 0.3 ppm for low and high PM exposures, respectively; geometric median number diameter 85 nm) indicated a hypertensive-like cardiac gene expression in WKY rats that mimicked baseline patterns in air-exposed SH rats (Gottipolu et al., 2009, 190360 ). Exposure to the high concentration of DE for 4 wk (4 h/day, 5 day/wk) led to downregulation of genes involved in stress, antioxidant compensatory response, growth and extracellular matrix regulation, membrane transport of molecules, mitochondrial function, thrombosis regulation, and immune function. No genes were affected by DE in SH rats. A dose-dependent inhibition of mitochondrial aconita se activity in both rat strains was observed, indicating a DE effect on oxidative stress. It shoul d be noted that while DE-related cardiovascular effects were found in WKY rats only, pulmonary inflammation and injury were observed in both strains (Sections 7.3.3.2 and 7.3.5.1 ).

December 2009 7-12

### Model Particles

Wallenborn et al. (2008, 191171 ) examined the subchronic (5 h/day, 3 day/wk, 16 wk) pulmonary, cardiac, and systemic effects of nose-only exposure to particulate ZnSO 4 (9, 35, or 120 μ g/m 3 ) in WKY rats. Particle size was reported to be 31-44 nm measured as number median diameter. Although changes in pulmonary inflammation or injury and cardiac pathology were not observed, effects on cardiac mitochondrial protein and enzyme levels were noted (i.e., increased ferritin levels, decrease in succinate dehydrogenase activity), possibly indicating a small degree of mitochondrial dysfunction. Glutathione peroxidase, an antioxidant enzyme, was also decreased in the cardiac cytosol. Gene expression analysis identif ied alterations in cardiac genes involved in cell signaling events, ion channels regulation, and coagulation in animals exposed to the highest ZnSO 4 concentration only. This study demonstrates a possible direct effect of ZnSO 4 on extrapulmonary systems, as suggested by the lack of pulmonary effects (Section 7.3.3.2 ).



## 7.2.8. Left Ventricular Mass and Function

Van Hee et al. (2009, 192110 ) studied 3,827 participants (age range = 45-84 yr; 53% female) who underwent magnetic resonance imaging (MRI) of the heart at the baseline examination of the MESA cohort (2000-2002). This cross-sectional study focused on two MRI-based outcome measures: left ventricular mass index (LVMI, g/m 2 ) and ejection fraction (EF, %), the former estimated using the DuBois formula for body surface area , the latter as the ratio of stroke volume to end diastolic volume. The study also estimated annual mean exposures to PM 2.5 at participants’ geocoded residential addresses in 2000 using ordinary kriging of U.S. EPA AQS concentration data. In fully adjusted models, it found 3.8 (95% CI: -6.1 to 13.7) g/m 2 and -3.0% (-8.0 to 2.0) differences in LVMI and EF per 10 μg/m 3 increment in PM 2.5 . The findings were small and imprecise, albeit suggestive of a slight, PM-associated increase in the mass and decrease in the function of the left ventricle. The effect of living within 50 m of a major roadway on LVMI was greater than the effect of PM 2.5 (i.e., 1.4 g/m 2 [95% CI: 0.3-2.5] per 10 μg/m 3 .)



## 7.2.9. Clinical Outcomes in Epidemiologic Studies

Several epidemiologic studies of U.S. and Eu ropean populations have examined associations between long-term PM exposures and clini cal CVD events (Baccarelli et al., 2008, 157984 ; Hoffmann et al., 2006, 091162 ; Hoffmann et al., 2009, 190376 ; Maheswaran et al., 2005, 088683 ; Maheswaran et al., 2005, 090769 ; Miller et al., 2007, 090130 ; Rosenlund et al., 2006, 089796 ; Solomon et al., 2003, 156994 ; Zanobetti and Schwartz, 2007, 091247 ). Results from these studies are summarized in Figure 7-1 . The ambient PM concentrations from these studies are characterized in Table 7-2

.



### Coronary Heart Disease

Epidemiologic studies examining the association of coronary heart disease (CHD) with longterm PM exposure are discussed below (Hoffmann et al., 2006, 091162 ; Maheswaran et al., 2005, 090769 ; Miller et al., 2007, 090130 ; Puett et al., 2008, 156891 ; Rosenlund et al., 2006, 089796 ; Rosenlund et al., 2009, 190309 ; Zanobetti and Schwartz, 2007, 091247 ). Cases of CHD were variably defined in these studies to include history of angina pectoris, MI, coronary artery revascularization (bypass graft; angioplasty; stent; atherectomy), and congestive heart failure (CHF). Results pertaining to death from CHD are described in Section 7.6

. Miller et al. (2007, 090130 ) studied incident, validated MI, revascularization, and CHD death, both separately and collectively, among 58,610 post-menopausal female residents of 36 U.S. metropolitan areas (age range = 50-79 yr) enrolled in the Women’s Health Initiative Observational Study (WHI OS, 1994-1998). In this prospective cohort study of participants free of CVD at baseline (median duration of follow-up = 6 yr), the author s used arithmetic averaging of year 2000 EPA AQS
PM 2.5 data available at the monitor nearest to each participant’s geocoded U.S. Postal Service fivedigit ZIP code centroid to estimate 1-yr avg e xposures. They found 6% (95% CI: -15 to 34), 20% (95% CI: 0-43) and 21% (95% CI: 4-42) increases in the overall risk of MI, revascularization, and

December 2009 7-13their combination with CHD death per 10 μ g/m 3 increase in PM 2.5 , respectively. Hazards were higher within than between cities and in the obese. For the combined CVD outcome (MI, revascularization, stroke, CHD death, cerebrovascular disease), author s reported a 24% (95% CI: 9-41) increase in risk that was higher among participants at higher than lower quintiles of body mass index, waist-to-hip ratio, and waist circumference. The PM 2.5 -CVD association was stronger among non-diabetic than diabetic participants.



### Table 7-2. Characterization of am bient PM concentrations from st udies of clinical cardiovascular diseases and long-term exposure.

Study Location Mean Annual Concentration ( μ g/m 3 ) Upper Percentile Concentrations ( μ g/m 3 )

PM 10

Puett et al. (2008, 156891 ) 13 U.S. States 21.6

Zanobetti and Schwartz (2007, 091247 ) 21 U.S. Cities 28.8 Overall range NR

Rosenlund et al. (2006, 089796 ) Stockholm, Sweden 30 y avg PM 10 (traffic)

Cases: 2.6

Controls: 2.4 5th-95th Percentile

0.5-6.0

0.6-5.9

Rosenlund et al. (2009, 190309 ) Stockholm, Sweden 5-yr avg PM 10 from traffic:

Cases: 2.4 (median)

Controls: 2.2 (median)

Maheswaran et al. (2005, 090769 ) Sheffield, U.K. Range of means in each quintile: 16-23.3 NR

Baccarelli et al. (2008, 157984 ) Lombardia Region, Italy NR NR

PM 2.5

Miller et al. (2007, 090130 ) WHI: 36 Metropolitan areas Citywide avg (yr 2000): 13.5 Min-max: 4-19.3

Hoffmann et al. (2006, 091162 ) HNRS: 2 Cities Germany 23.3 NR

Hoffman et al. (2009, 190376 ) HRNS: 2 Cities German 22.8 NR

WHI: Womens Health Initiative HNRS: Hans Nixdorf Recall Study

Puett et al. (2008, 156891 ) studied incident, validated CHD, CHD death, and non-fatal MI among 66,250 female residents (mean age = 62 y r) of metropolitan statistical areas in thirteen northeastern U.S. states who were enrolled in the Nurses’ Health Study (NHS, 1992-2002). In this prospective cohort study of women without a history of non-fatal MI at baseline (maximum duration of follow-up = 4 yr), the authors used two-stage, spatially smoothed, land use regression to estimate residence-specific, 1-yr ma PM 10 exposures from U.S. EPA AQS and emissions, IMPROVE, and Harvard University monitor data. They found a 10% (9 5% CI: -6 to 29) increase in risk of first CHD event per 10 μ /m 3 increase in 1-yr avg PM 10 exposure, while the association with MI was close to the null value. The association with fatal CHD event of 30% (95% CI: 0-71) was stronger. Furthermore, associations with CHD death were higher in the obese and in the never smokers. Rosenlund et al. (2006, 089796 ) studied 2,938 residents of Stockholm County, Sweden (age range = 45-70 yr; 34% female). In this case-control study of 1,085 patients with their first, validated non-fatal MI and an age-, gender- and catchment-str atified random sample of 1,853 controls without MI (1992-1994), the authors used street canyon-adjusted dispersion modeling of emissions data to estimate 30-yr avg exposure to PM 10 (median = 2.4 μ g/m 3 ). They found that the OR for prevalent MI per 10 μ g/m 3 increase in PM 10 was 0.85 (95% CI: 0.50-1.42). The OR for fatal MI was elevated, but not statistically significant. In a more recent study, Rosenlund et al. (2009, 190309 ) evaluated 554,340 residents (age range = 15-79 yr; 49% female) of Stockholm County, Sweden (1984-1996). In this population-based, case-control study of 43,275 cases of incident, valid ated MI, the authors used dispersion modeling of traffic emissions and land use data to estimate 5-yr avg exposure to PM 10 . They found that after

December 2009 7-14adjustment for demographic, temporal, and soci oeconomic characteristics, the OR for MI per
5 μ g/m 3 increase in PM 10 was 1.04 (95% CI: 1.00-1.09). ORs were higher after restriction to fatal cases, in- or out-of-hospital deaths, and participan ts who did not move between population censuses. Authors state that control for confounding was supe rior in their previous study (Rosenlund et al., 2006, 089796 ) although the size of the population was larger in this recent study (Rosenlund et al., 2009, 190309 ). Zanobetti and Schwartz (2007, 091247 ) studied ICD-coded recurrent MI (ICD 9 410) and post-infarction CHF (ICD 9 428) among 196,131 Medicare recipients (age ≥ 65 yr; 50% female) discharged alive following MI hospitalization in 21 cities from 12 U.S. states (1985-1999). In this ecologic, open cohort study of re-hospitalization among MI survivors (mean duration of followup = 3.6 and 3.7 yr for MI and CHF, respectively) , the authors used arithmetic averaging of EPA AQS PM 10 data available in the county of hospitalizati on to estimate 1-yr avg exposures. They found17% (95% CI: 5-31) and 11% (95% CI: 3-21) increases in the risk of recurrent MI and postinfarction CHF, respectively, per 10 μ g/m 3 increase in PM 10 exposure. Hazards were somewhat higher among persons aged >75 yr. Hoffmann et al. (2006, 091162 ) studied self-reported CHD (MI or revascularization) among 3,399 residents of Essen and Mülheim, Germany (a ge range = 45-75 yr; 51% female) at the baseline exam of the Heinz Nixdorf Recall Study (2000-2003) introduced previously. In this cross-sectional ancillary study, the authors used dispersion modeling of emissions, climate and topography data to estimate 1-yr avg exposure to PM 2.5 (mean = 23.3 μ g/m 3 ). They found little evidence of an association between PM 2.5 and CHD in these data. After adjustment for geographic, demographic and clinical characteristics, the OR for prevalent CHD per 10 μ g/m 3 increase in exposure was 0.55 (95% CI: 0.14-2.11).

Study Avg Time Endpoint E ffect Estimate (95% CI)

Prospective Cohort Studies

Incident MI PM 2.5 Revascularization Stroke Cerebrovascular Disease Miller et al.(2007, 090130 )

58,610 post-menopausal women enrolled in WHI, incident, validated cases, 36 US cities 1 yr

All CVD

Puett et al. (2008, 156891 )

75,809 women in the Nurses Health Study, validated cases, 13 metro areas, NE states 1 yr 1st CHD Event PM 10

Case Control Studies Rosenlund et al. (2009, 190309 )

24,347 cases, N=276,926 randomly selected population based controls, Stockholm, Sweden 5 yr Fatal/Non-Fatal MI PM 10

Rosenlund et al. (2006, 089796 )

2,246 cases, N=3,206 randomly selected population controls, Stockholm, Sweden 30 yr Fatal/Non-Fatal MI

Baccarelli et al. (2008, 157984 )

871 cases, N=1210 healthy friend controls, Lombardy, Italy 1 yr DVT

Other Study Designs Hoffman et al. (2006, 091162 ) 1 yr Self-reported CHD PM 2.5 Hoffman et al. (2009, 190376 )

Prevalent cases at baseline in 3,399 residents, cross-sectional, 2 German cities 1 yr Self-reported PVD

Recurrent MI Hospitalization PM 10 Zanobetti & Schwartz (2007, 091247 )

196,131 MI survivors (ICD9 410), readmitted for MI or CHF, ecologic, open cohort, 21 US cities 1 yr Post-MI CHF Hospitalization

CHD Hospitalization Maheswaran et al. (2005, 090769 ; 2005, 088683 )

Hospital admissions among 199,682 residents, ecologic design, Sheffield, UK 5 yr Stroke Hospitalization

Relative Risk Estimate



### Figure 7-1. Risk estimates for the associations of clinical outcomes with long-term exposure to ambient PM 2.5 

### and PM 10


### .

December 2009 7-15In the study of 1,030 census enumeration district s in Sheffield, U.K. described previously, Maheswaran et al. (2005, 090769 ) studied 11,407 ICD-10-coded emergency hospitalizations for CHD (ICD10 I20-25) among 199,682 residents (age ≥ 45 yr; 45% female). In this ecologic study, the authors used dispersion modeling of emissions and cl imate data to estimate 5-yr avg exposure to
PM 10 . They found that after adjusting for smoking pr evalence, controlling for socioeconomic factors, and smoothing, the age- and gender-standardized rate ratios for CHD admission were 1.01 (95% CI: 0.92-1.11), 1.04 (95% CI: 0.93-1.15), 0.97 (95% CI : 0.87-1.08), and 1.07 (95% CI: 0.95-1.20) across
PM 10 quintiles. The linear trend was somewha t stronger for CHD mortality (Section 7.3 ). The study of post-menopausal women enrolled in the WHI OS by Miller et al. (2007, 090130 ) was the only U.S. study to examine the effect of PM 2.5 rather than PM 10 . This study, which provides strong evidence of an association, was distinguishe d by its prospective cohort design, validation of incident cases and large population. Puett et al. (2008, 156891 ), the other U.S. study with comparable design features, provides evidence of an association of incident CHD with long- term
PM 10 exposure. Findings from Swedish case control studies of incident validated cases of MI were not consistent. A cross-sectional study of self-reported CHD did not provide evidence of an association with PM 2.5, while findings from two ecologic studies of PM 10 indicated positive associations of CHD hospitalizations with PM 10 (Maheswaran et al., 2005, 088683 ; Zanobetti and Schwartz, 2007, 091247 ).



### Stroke

Miller et al. (2007, 090130 ) found 28% (95% CI: 2-61) and 35% (95% CI: 8-68) increases in the overall risk of validated stroke and cereb rovascular disease, respectively, per 10 μ g/m 3 increase in 1-yr avg PM 2.5 exposure. Risks were higher within than between cities. In the study of 1030 Census of enumeration districts in Sheffield, U.K. described previously, Maheswaran et al. (2005, 088683 ) studied 5,122 ICD-10-coded emergency hospital admissions for stroke (I60-69) among 199,682 residents (age ≥ 45 yr; 45% female) of 1,030 census enumeration districts in Sheffield, U.K. (1994-1999). In this ecologic study, the authors used dispersion modeling of emissions and climate data to estimate 5-yr avg exposure to PM 10 . They found that the age- and gender-standardized rate ratios for stroke admission were 1.05 (95% CI: 0.94 -1.17), 1.07 (95% CI: 0.95-1.20), 1.06 (95% CI: 0.94-1.20), and 1.15 (95% CI: 1.01-1.31) across PM 10 quintiles. Linear trend was somewhat stronger for stroke mortality (Section 7.6 ). These studies examining the long-term PM-stroke relationship provide evidence of association. Maheswaran et al. (2005, 088683 ) examined emergency room hospital admissions in Sheffield, U.K. using an ecologic design while results reported by Miller et al. (2007, 090130 ) are based on the prospective cohort study of the WHI OS population (both introduced previously).



### Peripheral Arterial Disease

The German Heinz Nixdorf Recall cross-sectional study described in Section 7.2.1.1 (Hoffmann et al., 2009, 190376 ) also evaluated the association between 1-yr avg exposure to PM 2.5 and peripheral arterial disease (self-reported history of a surgical or procedural intervention or an ABI <0.9 in one or both legs). The authors found no evidence of an increase in risk. The OR for peripheral arterial disease was 0. 87 (95% CI: 0.57-1.34) per 3.9 μ g/m 3 increase in PM 2.5 . However, evidence of an association with traffic exposure was present in these data. ORs of 1.77 (95% CI: 1.01-3.10), 1.02 (95% CI: 0.58-1.80), and 1.07 (95% CI: 0.68-1.68) for residing ≤ 50, 50-100, and 100-200 m of a major road (reference category: >200 m), respectively were observed. ORs were higher among participants with CAC scores ≤ 75th percentile, women, and smokers.



### Deep Vein Thrombosis

The Italian case-control study (introduced in Section 7.2.1.2 ) also examined the chronic PM 10 - DVT association (Baccarelli et al., 2008, 157984 ). The authors found a 70% (95% CI: 30-223) increase in the odds of DVT per 10 μ g/m 3 increase in 1-yr avg PM 10 exposure. This finding was consistent with the decreases in PT and PTT also obse rved among controls in this context as well as

December 2009 7-16the 47% (95% CI: 11-96) increase in the odds of DVT per inter-decile range (242 m) increase in the residence-to-major-roadway distan ce observed among a su bset of cases and controls (Baccarelli et al., 2009, 188183 ). The PM 10 -DVT and distance-DVT associations were both weaker among women and among users of oral contraceptives or hormone therapy.



## 7.2.10. Cardiovascular Mortality

New epidemiologic evidence reports a consisten t association between long-term exposure to
PM 2.5 and increased risk of cardiovascular mortality. There is little evidence for the long-term effects of PM 10-2.5 on cardiovascular mortality. This section fo cuses on cardiovascular mortality outcomes in response to long-term exposure to PM. The studies that investigate long-term exposure and mortality due to any specific or all (nonaccidental) causes are evaluated in Section 7.6 . A summary of the mean PM concentrations reported for the studies characterized in this section is presented in Table 7-8 , and the effect estimates are presented in Figure 7-7 and Figure 7-8

. A number of large, U.S. cohort studies have found consistent associations between long-term exposure to PM 2.5 and cardiovascular mortality. The American Cancer Society (ACS) (Pope et al. (2004, 055880 ) reported positive associations with deaths from specific cardiovascular diseases, particularly ischemic heart disease, and a gro up of cardiac conditions including dysrhythmia, heart failure and cardiac arrest (RR for cardiovascular mortality = 1.12 [95% CI: 1.08-1.15] per 10 μg/m 3

PM 2.5 ). In an additional reanalysis that extended th e follow-up period for the ACS cohort to 18 yr (1982-2000) (Krewski et al., 2009, 191193 ), investigators found effect estimates that were similar, though generally higher, than those reported in previous ACS analyses. A follow-up to the Harvard Six Cities study (Laden et al., 2006, 087605 ) used updated air pollution and mortality data and found positive associations between long-term exposure to PM 2.5 and mortality. Of special note is a statistically significant reduction in mortality risk reported with reduced long-term fine particle concentrations. This reduced mortality risk was observed for deaths due to cardiovascular and respiratory causes, but not for lung cancer deaths. The WHI cohort study (Miller et al., 2007, 090130 ) (described previously) found that each 10 μg/m 3 increase of PM 2.5 was associated with a 76% increase in the risk of death from cardiovascular disease (hazard ratio, 1.76 [95% CI : 1.25-2.47]). The WHI study not only confirms the ACS and Six City Study associations with cardiovascular mortality in yet another well characterized cohort with detailed individual-level in formation, it also has been able to consider the individual medical records of the thousands of WHI subjects over the period of the study. This has allowed the researchers to examine not only mortality, but also related morbidity in the form of heartproblems (cardiovascular events) experienced by the subjects during the study. These morbidity coassociations with PM 2.5 in the same population lend even greater support to the biological plausibility of the air pollution-mortality associations found in this study. In an analysis for the Seventh-Day Adven tist cohort in California (AHSMOG), a positive association with coronary heart disease mortality was reported among females (92 deaths; RR = 1.42 [95% CI: 1.06-1.90] per 10 μg/m 3 PM 2.5 ), but not among males (53 deaths; RR = 0.90 [95% CI: 0.76-1.05] per 10 μg/m 3 PM 2.5 ) (Chen et al., 2005, 087942 ). Associations were strongest in the subset of postmenopausal women (80 deaths; RR = 1.49 [95% CI: 1.17-1.89] per 10 μg/m 3 PM 2.5 ). The authors speculated that females may be more sensitive to air po llution-related effects, based on differences between males and females in dosimetry and exposure. As was found with PM 2.5 , a positive association with coronary heart disease mortality was reported for PM 10-2.5 and PM 10 among females (RR = 1.38 [95% CI: 0.97-1.95] per 10 μg/m 3 PM 10-2.5 ; RR = 1.22 [95% CI: 1.01-1.47] per 10 μg/m 3 PM 10 ), but not for males (RR = 0.92 [95% CI: 0.66-1.29] per 10 μg/m 3 PM 10-2.5 ; RR = 0.94 [95% CI: 0.82-1.08] per 10 μg/m 3 PM 10 ); associations were strongest in the subset of postmenopausal women (80 deaths) (Chen et al., 2005, 087942 ). Two additional studies explored the effects of PM 10 on cardiovascular mortality. The Nurses’ Health Study (Puett et al., 2008, 156891 ) is an ongoing prospective cohort study examining the relation of chronic PM 10 exposures with all-cause mortality and incident and fatal coronary heart disease consisting of 66,250 female nurses in MSAs in the northeastern region of the U.S. The association with fatal CHD occurred with the gr eatest magnitude when compared with other specified causes of death (hazard ratio 1.42 [95% CI: 1.11-1.81]). The North Rhine-Westphalia State Environment Agency (LUA NRW) initiated a cohortof approximately 4,80 0 women, and assessed whether long-term exposure to air pollution origin ating from motorized traffic and industrial sources was associated with total and cause-specific mortality (Gehring et al., 2006, 089797 ). They found

December 2009 7-17that cardiopulmonary mortality was associated with PM 10 (RR = 1.52 [95% CI: 1.09-2.15] per 10 μg/m 3 PM 10 ). In summary, the 2004 PM AQCD concluded that there was strong evidence that long-term exposure to PM 2.5 was associated with increased cardiopul monary mortality. Recent studies investigating cardiovascular mortality provide some of the strongest evidence for a cardiovascular effect of PM. A number of large cohort studies have been conducted throughout the U.S. and reported consistent increases in cardiovascular mortality related to PM 2.5 concentrations. The results of two of these studies have been replicated in independent reanalyses. These effects are coherent with short-term epidemiologic studies of CVD morbidity and mortality and with long-term epidemiologic studies of CVD morbidity. In a ddition, biological plausibi lity and coherence are provided by toxicological studies demonstrating short-term cardiovascular effects as well as
PM 2.5 -related plaque progression in chronically exposed mice.



## 7.2.11. Summary and Causal Determinations

7.2. 11.1. PM 2 .5 Epidemiologic studies examining associations between long-term exposure to ambient PM (over months to years) and CVD morbidity had not been conducted and thus were not included in the 1996 or 2004 PM AQCDs (U.S. EPA, 1996, 079380 ; U.S. EPA, 2006, 157071 ). A number of studies were included in the 2004 AQCD that evaluated the effect of long-term PM 2.5 exposure on cardiovascular mortality and found strong and consis tent associations. No toxicological studies had evaluated the effects of subchronic or chronic PM exposure on CVD effects in the 2004 PM AQCD. Recently, epidemiologic and toxicological studies ha ve provided evidence of the adverse effects of long-term exposure to PM 2.5 on cardiovascular outcomes and endpoints, including atherosclerosis and clinical and subclinical markers of cardiovascular morbidity. The strongest evidence for a CVD health effect related to long-term PM 2.5 exposure comes from epidemiologic studies of cardiovascular mortality. A number of large, multicity U.S. studies (the ACS, Six Cities Study, WHI, and AHSMOG) pr ovide consistent evidence of an effect between long-term exposure to PM 2.5 and cardiovascular mortality (Section 7.2.10 ). These studies were conducted in urban areas across the U.S. where mean concentrations ranged from 10.2-29.0 μg/m 3 ( Table 7-8 ). An epidemiologic study investigating the relationship between PM 2.5 and clinical CVD morbidity among post-menopausal women (Miller et al., 2007, 090130 ) provides evidence of an effect that is coherent with the cardiovascular mortality studies. This large, prospective cohort study of incident, validated cases found large increases inthe adjusted risk of MI, revascularization, and stroke using a 1-yr avg PM 2.5 concentration (mean = 13.5 μ g/m 3 ). A cross-sectional analyses of selfreported prevalence of CHD and peripheral arterial disease found no such increase in the odds of CVD morbidity (Hoffmann et al., 2006, 091162 ); the inconsistency of these findings with Miller et al. (2007, 090130 ) may be explained by differences in study design or location. The effect of long-term PM 2.5 exposure on pre-clinical measures of atherosclerosis (CIMT, CAC, AAC or ABI) has been studied in several populations using a cross-sectional study design. The magnitude of the PM 2.5 effects and their consistency across different measures of atherosclerosis in these studies varies widely, and they may be limited in their ability to discer n small changes in these measures. Kunzli et al. (2005, 087387 ) observed a non-significant 4.2% increase in CIMT associated with long-term PM 2.5 exposure among participants of a clini cal trial in greater Los Angeles, which was several fold higher than the 0.5% increase observed by Diez-Roux et al. (2008, 156401 ) in their analyses of MESA baseline data. The associations in MESA of CAC and ABI with long-term PM 2.5 exposure were largely null (Diez et al., 2008, 156401 ), while an increase in AAC with long-term
PM 2.5 exposure was reported (Chang et al., 2008, 180393 ). By contrast, a 43% increase in CAC was associated with long-term PM 2.5 exposure in a German study, but no similar association with ABI was observed (Hoffmann et al., 2009, 190376 ). Although the number of studies examining these relationships is limited, effect modification by use of lipid lowering drugs and smoking status was reported in more than one study of long-term PM 2.5 and PM 10 exposure. Evidence of enhanced atherosclerosis develo pment was demonstrated in new toxicological studies that report increased plaque and lesion areas, lipid deposition, and TF in aortas of ApoE -/- mice exposed to CAPs (Section 7.2.1.2 ). In addition, alterations in vasoreactivity were observed,

December 2009 7-18suggesting an impaired NO pathway. Additional toxicological studies of PM 10 are consistent with these results. Further support is provided by a study that reported decreased L/W ratio in the pulmonary and coronary arteries of mice exposed to ambient air. However, PM 2.5 CAPs derived from traffic in Los Angeles did not affect plaque size (Araujo et al., 2008, 156222 ). Collectively, these toxicological studies provide biological plausibili ty for the associations reported in epidemiologic studies. There is limited evidence for the effects of PM 2.5 on renal or vascular function. Cross-sectional and longitudinal epidemiologic analyses of PM 2.5 and UACR revealed no evidence of an effect (O'Neill et al., 2007, 156006 ), while small non-statistically significant increases in BP with 30- and 60-day avg PM 2.5 concentrations were reported (Auchincloss et al., 2008, 156234 ). A toxicological study did not show changes in MAP with CAPs, but indicated a CAPs-related potentiation of experimentally-induced hypertension (Sun et al., 2008, 157032 ). In addition, CAPs has induced changes in insulin resistance, visceral adiposity, and inflammation in a diet-induced obesity mouse model (Sun et al., 2009, 190487 ), indicating that diabetics may be a potentially susceptible population to PM exposure. In summary, a number of large U.S. cohor t studies report associations of long-term PM 2.5 concentration with cardiovascular mortality. These studies provide the strongest evidence for an effect of long-term PM 2.5 exposure on CVD effects. Addi tional evidence comes from a methodologically rigorous epidemiology study thatdemonstrates coherent associations between long-term PM 2.5 exposure and CVD morbidity among post-menopausal women. Toxicological studies demonstrate that this effect is biologically pl ausible and the effect is coherent with studies of short-term PM 2.5 exposure and CVD morbidity and mortalit y, and with long-term exposure to PM 2.5 and CVD mortality. Associations between PM 2.5 and subclinical measures of atherosclerosis are inconsistent, but cross-sectional studies may be lim ited in their ability to discern small changes in these measures. In addition, potential modification of the PM 2.5 -CVD association by smoking status and the use of lipid lowering drugs has been demonstrated in epidemiologic studies that used individual-level data. Toxicological studies provide evidence for accelerated development of atherosclerosis in ApoE -/- mice exposed to CAPs and show effects on coagulation factors, experimentally-induced hypertension, and vascular reactivity. Available studies of clinical cardiovascular disease outcomes report inconsistent results. Based on the above findings, the epidemiologic and toxicological evidence is 

### sufficient to infer a causal relationship between long-term PM 2.5 

### exposures and cardiovascular effects.



## 7.2. 11.2. PM 1 0-2.5 One epidemiologic study evaluated the relationship between long-term exposure to PM 10-2.5 and cardiovascular mortality and found a positive a ssociation with coronary heart disease mortality among females, but not for males; associations were strongest in the subset of post-menopausal women (Chen et al., 2005, 087942 ). No toxicological studies of long-term exposure to ambient

PM 10-2.5 and cardiovascular effects have be en conducted to date. Evidence is 

### inadequate to infer the presence or absence of a causal relationship.



## 7.2. 11.3. UFP s

A few toxicological studies of long-term exposur e to UFPs have been conducted. Increased plaque size was reported in mice exposed to UF CAPs derived from traffic (Araujo et al., 2008, 156222 ). Studies of diesel and gasoline exhaust repo rted relatively few changes in hematologic or coagulation parameters (Section 7.2.4.2 ) and one DE study demonstrated altered cardiac gene expression in normotensive rats that reflected th e development of hyperte nsion (Gottipolu et al., 2009, 190360 ). Whole and filtered gasoline exhaust indu ced increases in gene products involved in atheromatous plaque formation and/or degradation, but these effects were largely due to the gaseous emissions (Lund et al., 2007, 125741 ). Evidence from these studies alone is 

### inadequate to infer the presence or absence of a causal relationship

.

December 2009 7-19

# 7.3. Respiratory Effects

Several cohort studies reviewed in the 2004 PM AQCD provided evidence for relationships between long-term PM exposure and effects on the respiratory system, though it did not rule out the possibility that the observed respiratory effects ma y have been confounded by other pollutants. In 12 southern California communities in the Children’s Health Study (CHS), Gauderman et al. (2000, 012531 ; 2002, 026013 ) found that decreases in lung function growth among schoolchildren were associated with long-term exposure to PM. Declines in pulmonary function were reported with all three major PM size classes – PM 10 , PM 10-2.5 and PM 2.5 – though the three PM measures were highly correlated. In another analysis of data from the CHS cohort, McConnell et al. (1999, 007028 ), reported an increased risk of bronchitis sympto ms in children living in communities with higher
PM 10 and PM 2.5 concentrations. These results were found to be consistent with results of crosssectional analyses of the 24-city study by Dockery et al. (1996, 046219 ) and Raizenne et al. (1996, 077268 ), that were assessed in the 1996 PM AQCD. These studies reported associations between increased bronchitis rates and decreased peak flow with fine particle sulfate and fine particle acidity. However, the high correlation of PM 10 , acid vapor and NO 2 precluded clear attribution of the bronchitis effects reported by McConnell et al. (1999, 007028 ) to PM alone. In a prospective cohort study among a subset of children in the CHS (n = 110) who moved to other locations during the study period, Avol et al. (2001, 020552 ) reported that those subjects who moved to areas of lower
PM 10 showed increased growth in lung function compared with subjects who moved to communities with higher PM 10 concentrations. Finally, the 2004 PM AQCD concluded that there was strong epidemiologic evidence for associations between long-term exposures to PM 2.5 and cardiopulmonary mortality, though the respiratory effects were not separated from the cardiovascular effects in this conclusion. The 2004 PM AQCD (U.S. EPA, 2004, 056905 ) concluded that the evidence for an association between long-term exposure to PM and respiratory effects may be confounded by other pollutants. Gauderman et al. (2002, 026013 ) reported declines for FEV 1 and McConnell et al. (1999, 007028 ) reported increased ORs for bronchitic symptoms in asthmatics for PM 10 and PM 2.5 . Recent epidemiologic literature includes results from several prospective cohort studies, which found consistent, positive associations between long-term exposure to PM and respiratory morbidity. Associations were reported with PM 2.5 and PM 10 , and the studies showing associations only with
PM 10 were conducted in locations where the PM consist ed predominantly of fine particles, providing support for associations with long-term exposure to fi ne particles. These results are summarized below; further details of these studies are summarized in Annex E. Very few subchronic and chronic toxicological studies investigating respiratory effects were available in the 2004 PM AQCD. However, the 2002 EPA Health Assessment Document for DE reported that chronic exposure to DE was associated with histopathology including alveolar histiocytosis, aggregation of alveolar macrophages, tissue inflammation, increased polymorphonuclear leukocytes, hyperplasia of br onchiolar and Type 2 epithelial cells, thickened alveolar septa, edema, fibrosis, emphysema and lesions of the trachea and bronchi. Since then a number of animal toxicological studies have be en conducted involving inhalation exposure to CAPs, urban air, DE, gasoline exhaust, and wood smoke . These subchronic and chronic studies provide evidence of altered pulmonary function, inflamma tion, histopathological changes and oxidative and allergic responses following PM 2.5 exposures. These results are summarized below; further details of these studies are summarized in Annex D.



## 7.3.1. Respiratory Symptoms and Disease Incidence

7.3. 1.1. Epi demiologic Studies

New longitudinal cohort studies provide the best evidence to evaluate the relationship between long-term exposure to ambient PM and increased incidence of respiratory symptoms or disease. A summary of the mean PM concentrations reported for the long-term exposure studies characterized in this section is presented in Table 7-3

.

December 2009 7-20Bayer-Oglesby et al. (2005, 086245 ) examined the decline of ambient pollution levels and improved respiratory health demonstrated by a reduction in respiratory symptoms and diseases in school children (n = 9,591) in Switzerland. Reduced air pollution exposure resulted in improved respiratory health of children. Further, the av erage reduction of symptom prevalence was more pronounced in areas with stronger reduction of air pollution levels. The average decline of PM 10 between 1993 and 2000 across the nine study regions was 9.8 μg/m 3 (29%). Declining levels of PM 10 were associated with declining prevalence of chronic cough, bronchitis, common cold, nocturnal dry cough, and conjunctivitis symptoms, but no signifi cant associations were reported for wheezing, sneezing, asthma, and hay fever, as shown in Figure 7-2 . In Figure 7-2 , Panel (B) illustrates that on an aggregate level across regions, the mean change in adjusted prevalence of chronic cough is associated with the mean change in PM 10 levels (r = 0.78; p = 0.02). Similar associations were seen for nocturnal dry cough and conjunctivitis symptoms and PM 10 levels. Röösli et al. (2000, 010296 ; 2001, 108738 ; 2005, 156923 ) have demonstrated that PM 10 levels are homogeneously distributed within regions of Basel, Switzerland and are not substan tially affected by local traffic, justifying the single-monitor approach for assignment of PM 10 exposures. Based on parallel measurements of
PM 2.5 and PM 10 at seven sites in Switzerland, PM 2.5 and PM 10 at all sites are generally highly correlated (r 2 ranging from 0.85 to 0.98) (Gehrig and Buchmann, 2003, 139678 ), indicating that
PM 10 consists predominantly of fine particles in these locations. Schindler et al. (2009, 191950 ) reported that sustained reduction in ambient PM 10 concentrations can lead to decreases in respiratory symptoms among Swiss adults in the SAPALDIA study. They compared baseline data in 1991 to a follow-up interview in 2002 after a substantial decline in PM 10 concentrations served as a natural experiment. Each subject was assigned modelbased estimates of PM 10 concentrations averaged over the 12 mo preceding each health assessment with mean decline in PM 10 levels of 6.2 μg/m 3 (SD = 3.9 μg/m 3 ). When the authors tested the joint hypothesis of no association between the PM 10 difference and symptom incidence or persistence, positive results were obtained for regular cough, chronic cough or phlegm and wheezing but not regular phlegm or wheezing without a cold. Pierse et al. (2006, 088757 ) studied the association between primary PM 10 (particles directly emitted from local sources/traffic) and the preval ence and incidence of respiratory symptoms in a randomly sampled cohort of 4,400 children (aged 1-5 yr) in Leicestershire, England surveyed in 1998 and again in 2001. Annual exposure to primary PM 10 was calculated for the home address using the Airviro statistical dispersion model. After adjusting for confounders, mean annual exposure to locally generated PM 10 was associated with an increased prevalence of cough without a cold in both the 1998 (OR 1.21 [95% CI: 1.07-1.38], n = 2,164) and 2001 surveys (OR 1.56 [95% CI: 1.32-1.84], n = 1,756). Nordling et al. (2008, 097998 ) examined the relationship between estimated PM exposure levels and respiratory health effects in a Swedish birth cohort of preschool children (n = 4,089). The spatial distributions of PM from traffic in the st udy area were estimated w ith emission databases and statistical dispersion modeling. Children were examinedat 2 mo and 1, 2, and 4 yr of age. Using GIS methods, the average contribution of traffic-generated PM 10 above regional background to the children’s residential outdoor air pollution levels was determined. To evaluate the exposure assessment, the authors compared the estimated levels of traffic-generated PM 10 with PM 2.5 measurements from 42 locations (Hoek et al., 2002, 042364 ) and reported modeled traffic-generated
PM 10 correlated reasonably well with measured PM 2.5 (r = 0.61). Persistent wheezing (cumulative incidence up to age 4 yr) was associated with exposure to traffic-generated PM 10 (OR 2.28 [95% CI: 0.84-6.24] per 10 μ g/m 3 increase) while transient and late onset wheezing was not associated. This study demonstrates that respiratory effects may be present in preschool children.

December 2009 7-21

### Table 7-3. Characteriz ation of ambient PM concentrations from studies of respiratory symptoms/disease and long-term exposures.

Study Location Mean Annual Concentration (μg/m 3 ) Upper Percentile Concentrations (μg/m 3 )

PM 2.5

Annesi-Maesano et al. (2007, 093180 ) 6 French Cities Range of means across sites: 8.7-23.0

Avg of means across sites: 15.5

Brauer et al. (2007, 090691 ) The Netherlands 16.9 75th: 18.1

90th: 19.0

Max: 25.2

Goss et al. (2004, 055624 ) U.S. 13.7 75th: 15.9

Islam et al. (2007, 090697 ) 12 CHS/CA communities Max: 29.5

Janssen et al. (2003, 133555 ) The Netherlands 20.5 75th: 22.1

Max: 24.4

Kim et al. (2004, 087383 ) San Francisco, CA Range of means across sites: 11-15

Avg of means across sites: 12

McConnell et al. (2003, 049490 ) 12 CHS/CA communities 13.8 Max: 28.5

Morgenstern et al. (2008, 156782 ) Munich, Germany 11.1

PM 10

Bayer-Oglesby et al. (2005, 086245 ) Nine study regions in Switzerland Max: 46

Kunzli et al. (2009, 191949 ) Switzerland 21.5

Nordling et al. (2008, 097998 ) Sweden 4*

Schindler et al. (2009, 191950 ) Switzerland **

McConnell et al. (2003, 049490 ) 12 CHS/CA communities 30.8 Max: 63.5

Pierse et al. (2006, 088757 ) Leicestershire, U.K. 1.33 75th: 1.84

*Source specific; PM 10 from traffic **Only reported change in PM concentration

December 2009 7-22Source: Bayer-Oglesby et al. (2005, 086245 )



### Figure 7-2. Adjusted ORs and 95% Cls of symptoms and respiratory diseases associated with a decline of 10 μg/m 3 

### PM 10 

### levels in Swiss Surveillance Program of Childhood Allergy and Respiratory Symptoms 1 

### . Inset: Mean change in adjusted prevalence (1998-2001 to 1992-1993) versus mean change in regional annual averages of PM 10 

### (1997-2000 to 1993) for chro nic cough, across nine SCARPOL regions (An: Anières. Be: Bern. Bi: Biel . Ge: Geneva. La:, Langnau. Lu: Lugano. Mo: Montana. Pa: Payerne. Zh: Zürich).

McConnell et al. (2003, 049490 ) conducted a prospective study examining the association between air pollution and bronchitic symptoms in 475 school children with asthma in 12 Southern California communities as part of the CHS fr om 1996 to 1999. They investigated both the differences between- communities with 4-yr avg an d within-communities yearly variation in PM (i.e., PM 10 , PM 2.5 , PM 10-2.5 , EC, and OC). Based on a 10 μ g/m 3 change in PM 2.5 , within-communities effects were larger (OR 1.90 [95% CI: 1.10-2.70]) than those for between-communities (OR 1.30 [95% CI: 1.10-1.50]). The OR for the 10 μ g/m 3 range in 4-yr avg PM 2.5 concentrations across the 12 communities was 1.29 (95% CI: 1.06-1.58). Similar results were reported for PM 10 and PM 10-2.5 but the effect estimates were smaller in magnitude a nd generally not statistically significant. Withincommunity associations were not confounded by a ny time-fixed personal covariates. In two-

1 Adjusted for age, sex, nationality, parental education, number of siblings; farming status, lo w birth weight, breastfeeding, c hild who

smokes, family history of asthma, bronchitis, and/or atopy, mo ther who smokes, indoor humidity, mode of heating and cooking, carpeting, pets allowed in bedroom, removal of carpet and/or pets for health reasons, person who completed questionnaire, month when questionnaire was completed, number of days with the maximum temp erature <0°C, and belief of mother that there is an associatio n between environmental exposures and children’s respiratory health

December 2009 7-23pollutant models, the within-community effect estimates for PM 2.5 and OC were significant in the presence of several other pollutants. While the within-community single-pollutant effect of PM 2.5 (ß = 0.085/ μ g/m 3 ) was only modestly attenuated after adjust ing for some pollutants, it was markedly reduced after adjusting for NO 2 or OC. The between-community effect estimates generally were not significant in the presence of other pollutants in copollutant models. In the CHS, Islam et al. (2007, 090697 ) examined the hypothesis that ambient air pollution attenuates the reduced risk for childhood asthma that is associated with higher lung function (n = 2,057). At each age a distribution of pul monary functions exists. Haland et al. (2006, 156511 ) found evidence that children with high lung function have a reduced risk for asthma. Islam et al. (2007, 090697 ) used the CHS data to study how the association of asthma incidence with lung function is modified by long-term PM exposure. The incidence rate (IR) of newly diagnosed asthma increased from 9.5/1,000 person-years for children with percent-predicted FEF 25-75 values ≥ 120% to 20.4/1,000 person-years for children with FEF 25-75 value ≤ 100%. Over the 10th-90th percentile range for FEF 25-75 (57.1%), the hazard ratio of new onset asthma was 0.50 (95% CI: 0.35-0.71). The IR of asthma for FEF 25-75 ≥ 120% in the “high” PM 2.5 (13.7-29.5 μg/m 3 ) communities was 15.9/1,000 person-years compared to 6.4/1,000 person-years in “low” PM 2.5 (5.7-8.5 μg/m 3 ) communities. Loss of protection by high lung function against new onset asthma in the “high” PM 2.5 communities was observed for all the lung function measures. Figure 7-3 shows the effect of PM 2.5 on the association of lung function with asthma. Of all the pollutants examined (NO 2 , PM 10 , PM 2.5 , acid vapor, O 3 , EC, and OC), PM 2.5 appeared to have the strongest modify ing effect on the association between lung function with asthma as it had the highest R 2 value (0.42). Over the 10th-90th percentile range of
FEF 25-75 , the hazard ratio of new onset asthma was 0.34 (95% CI: 0.21-0.56) in a community with low PM 2.5 (<13.7 μ g/m 3 ) and 0.76 (95% CI: 0.45-1.26) in a community with high PM 2.5 ( ≥ 13.7 μ g/m 3 ). The data do not indicate that PM exposure increased rates of incident asthma among children with poor lung function at study entry because rates among those with poor lung function were similar in both low and high pollution communities.

Source: Reprinted with Permission of BMJ Publishing Group Ltd & British Thoracic Societyfrom Islam et al. (2007, 090697 )



### Figure 7-3. Effect of PM 2.5 

### on the association of lung f unction with asthma. Communityspecific hazard ratio of newly diagnosed asthma over 10-90th percentile range (57.1%) of FEF 25-75% 

### by level of ambient PM 2.5 

### ( μ g/m 3 

### ). The 12 CHS communities are shown.

In a prospective birth cohort study (n = 4,000) in The Netherlands, Brauer et al. (2007, 090691 ) assessed the development of asthma, allergic symptoms, and respiratory infection during the first 4 yr of life in relation to long-term PM 2.5 concentration at the home address with a validated model using GIS. PM 2.5 was associated with doctor-diagnosed asthma (OR = 1.32 [95% CI:

December 2009 7-241.04-1.69]) for a cumulative lifetime indicator. Th ese findings extend observations made at 2 yr of age in the same cohort (Brauer et al., 2002, 035192 ) providing greater confidence in the association. No associations were observed for bronchitis. Kunzli et al. (2009, 191949 ) used the SAPALDIA cohort study discussed previously in this section to evaluate the relationship between the 11-yr change (1991-2002) in traffic-related PM 10 and asthma incidence-adult onset asthma. In a cohort of 2,725 never-smokers without asthma at baseline (age: 18-60 yr in 1991), subjects reporting doctor-diagnosed asthma at follow-up were considered incident cases. Modeled traffic-related PM 10 levels were used. Cox proportional hazard models for time to asthma onset were used with adjustments for cofounders. The study findings suggest that PM contributes to asthma development and that reduc tions in PM decrease asthma risk. A strong feature of SAPALDIA is the ability to assign space, tim e, and source-specific pollution to each subject. Further, Kunzli et al. (2008, 129258 ) discusses the impact of attributable health risk models for exposures that are assumed to cause both chronic di sease and its exacerbations. The added impact of causing disease increases the risk compared to only exacerbations. A matched case-control study of infant bronchi olitis (ICD 9 code 466.1) hospitalization and two measures of long-term exposure – the month prior to hospitalization (subchronic) and the lifetime average (chronic) – to PM 2.5 and gaseous air pollutants in the South Coast Air Basin of southern California was conducted by Karr et al. (2007, 090719 ) among 18,595 infants born between 1995-2000. For each case, 10 controls matched on date were randomly selected from birth records. Exposure was based on PM 2.5 measurements collected every third day. The mean distance between the subjects’ residential ZIP code and the assigned monitor was generally 4-6 mi with a maximum distance of 30 mi. For 10 μ g/m 3 increases in both sub-chronic and chronic PM 2.5 exposure, an adjusted OR of 1.09 (95% CI: 1.04-1.14) was obs erved. In multipollutant model analyses, the association with PM 2.5 was robust to the inclusion of gaseous pollutants. Also, in a cohort of children in Germany, Morgenstern et al. (2008, 156782 ) modeled PM 2.5 data at birth addresses found statistically significant effects for asthmatic bronchitis, hay fever, and allergic sensitization to pollen. Goss et al. (2004, 055624 ) conducted a national study examining the relationship between air pollutants and health effects in a cohort of cystic fibrosis (CF) patients (n = 11,484) over the age of 6 yr (mean age = 18.4, SD = 10) enrolled in the Cystic Fibrosis Foundation National Patient Registry in 1999 and 2000. Exposure was assessed by linking air pollution values from the closest population monitor from the Air Quality System (AQS) with the centroid of the patient’s home ZIP code that was within 30 mi. The mean distance from the patient’s ZIP code to monitors for PM 2.5 and PM 10 was 10.8 mi (SD 7.8) and 11.5 mi (SD 7.9), respectively. PM 2.5 and PM 10 24-h avg were collected every 1 to 12 days. CF diagnosis involves genetic screening panels and a common severe mutation used is the loss of phenylalanine at the 508th position. Genotyping was available in 74% of the population and of those genotyped, 66% carried one or more delta F508 deletions. After adjusting for confounders, a 10 μ g/m 3 increase in PM 2.5 or PM 10 was associated with a 21% (95% CI: 7-33) or 8% (95% CI: 2-15) increase in the odds of two or more exacerbations, respectively. The exacerbations were defined as a CF-related pulmonary condition requiring admission to the hospital or use of home intravenous antibiotics. The estimate for the associations between pulmonary exacerbations and
PM 2.5 and PM 10 were attenuated when the models were adjusted for lung function. Brown et al. (2001, 012307 ) found that particle deposition was increased in CF and that particle distribution in the lungs was enhanced in poorly ventilated trach eobronchial regions in CF patients. Such focal deposition may partially explain the association of PM and CF exacerbation. Annesi-Maesano (2007, 093180 ) relate individual data on asthma and allergy from 5,338 school children (10.4 ± 0.7 yr) attending 108 randomly chosen schools in 6 French cities to the concentration of PM 2.5 monitored in school yards. Atopic asthma was related to PM 2.5 (OR 1.43 [95% CI: 1.07-1.91]) when high PM 2.5 concentrations (20.7 μ g/m 3 ) were compared to low PM 2.5 concentrations (8.7 μ g/m 3 ). The report is consistent with the results in an earlier paper (PenardMorand et al., 2005, 087951 ) in the same sample of children that related the findings to PM 10

. Kim et al. (2004, 087383 ) conducted a school-based cross-sectional study in the San Francisco metropolitan area in 2001 comprised of 10 ne ighborhoods to examine the relationship between traffic-related pollutants and current bronchitic symptoms and asthma obtained by parental questionnaire (n = 1,109). They related traffic-re lated pollutants (PM) and bronchitic and asthma symptoms in the past 12 mo. No multipollutant models were evaluated because of the high interpollutant correlations. PM 2.5 levels ranged across the school sites from 11 to 15 μg/m 3

. Schikowski et al. (2005, 088637 ) examined the relationship between both long-term air pollution exposure and living close to busy roads a nd COPD in the Rhine-Ruhr Basin of Germany

December 2009 7-25from 1985 to 1994 using consecutive cross-sectional studies. Seven monitoring stations that were <8 km to a woman’s home address provided TSP data that PM 10 was estimated from using a conversion factor (obtained from parallel measurement of TSP and PM 10 conducted at 7 sites in the Ruhr area). Distance to a major road was determined using GIS. The results of the study suggest that long-term exposure to air pollution from PM 10 and living near a major road might increase the risk of developing COPD and can have a detrimental effe ct on lung function. All ORs for 5-yr exposures were stronger than those for 1-yr exposures. In summary, the 2004 PM AQCD evaluated the available studies which primarily related effects to bronchitic symptoms in school-age children. New studies are using several different methods to include individual estimates of exposure to ambient PM that may reduce the impact of exposure error. The strength and consistency of the outcomes is enhanced by results being reported by several different researchers in different countries using different designs. Most recent studies have focused on children, but a few studies have also reported associations in adults. The CHS (McConnell et al., 2003, 049490 ) provides evidence in a prospective longitudinal cohort study that relates PM 2.5 and bronchitic symptoms and reports larger associations for withincommunity effects that are less subject to confounding than between-community effects. Several new studies report similar findings with long-term exposure to PM 10 in areas where fine particles are the predominant fraction of PM 10 . In England, in a cohort of 4,400 children (aged 1-5 yr), an association is seen with an increased prevalence of cough without a cold. Further evidence includes a reduction of respiratory symptoms corresponding to decreasing PM levels in “natural experiments” in both a cohort of Swiss school children (Bayer-Oglesby et al., 2005, 086245 ) and adults (Schindler et al., 2009, 191950 ). In a separate analysis of th e CHS, Islam et al. (2007, 090697 ) showed that PM 2.5 had the strongest modifying effect on the association between lung function with asthma such that loss of protection by high lung function against new onset asthma in high PM 2.5 communities was observedfor all the lung function measures from 10 to 18 yr of age. This relates new onset asthma to longterm PM exposure. In the Netherlands, Brauer et al. (2007, 090691 ) augments the literature with data examining the first 4 yr of life in a birth cohort showing an association with doctor-diagnosed asthma. Further, in an adult cohort in the SALPALDIA study, Kunzli et al. (2009, 191949 ) relate PM to asthma incidence.



## 7.3.2. Pulmonary Function

Several cohort studies reviewed in the 2004 PM AQCD provided evidence for relationships between long-term PM exposure and effects on the respiratory system. In 12 southern California communities in the Children’s Health Study (CHS), Gauderman et al. (2000, 012531 ; 2002, 026013 ) found that decreases in lung function growth among school children were associated with long-term exposure to PM. Declines in pulmonary function were reported with all three major PM size classes
– PM 10 , PM 10-2.5 and PM 2.5 – though the three PM measures were highly correlated. These results were found to be consistent with results of cross- sectional analyses of Raizenne et al. (1996, 077268 ), that was assessed in the 1996 PM AQCD. That study reported associations between decreased peak flow with fine particle sulfate a nd fine particle acidity. Finally, in a prospective cohort study among a subset of children in the CHS (n = 110) who moved to other locations during the study period, Avol et al. (2001, 020552 ) reported that those subjects who moved to areas of lower
PM 10 showed increased growth in lung function compared with subjects who moved to communities with higher PM 10 concentrations who showed decrease growth in lung function.



## 7.3. 2.1. Epi demiologic Studies

New longitudinal cohort studies have evaluate d the relationship between long-term exposure to PM and changes in measures of pulmonary function (FVC, FEV 1 , and measures of expiratory flow). Cross-sectional studies also offer supportive in formation (Annex E) and may provide insights derived from within community analysis. Lung function increases continue through early adulthood with growth and development, then declines with aging (Stanojevic et al., 2008, 157007 ; Thurlbeck, 1982, 093260 ; Zeman and Bennett, 2006, 157178 ). A summary of the mean PM concentrations reported for the long-term exposure studies characterized in this section is presented in Table 7-4

.

December 2009 7-26

### Table 7-4. Characteriz ation of ambient PM concentrations from studies of FEV 1 

### and long-term
exposures.

Study Location Mean Annual Concentration (μg/m 3) Upper Percentile Concentrations (μg/m 3)

PM 2.5
Gauderman et al. (2002, 026013 ) 12 CHS/CA communities 5-30

Gauderman et al. (2004, 056569 ) 12 CHS/CA communities 6-27

Goss et al. (2004, 055624 ) U.S. 13.7 75th: 15.9

Gotschi et al. (2008, 180364 ) 21 European cities Range of means across sites: 3.7-44.7

Avg of mean across sites: 16.8

PM 10

Downs et al. (2007, 092853 ) 8 cities in Switzerland Range of means across sites: 9-46

Avg of mean across sites: 21.6

Gauderman et al. (2002, 026013 ) 12 CHS/CA communities Range of means across sites: 13-78

Avg of mean across sites: NR

Gauderman et al. (2004, 056569 ) 12 CHS/CA communities Range of means across sites: 18-68

Avg of mean across sites: NR

Nordling et al. (2008, 097998 ) Sweden Modeled exposure

Avol et al. (2001, 020552 ) Southern CA/CHS Range of means across sites: 15.0-66.2

Rojas-Martinez et al. (2007, 091064 ) Mexico City, Mexico 75.6 75th: 92.2

90th: 112.7

The CHS prospectively examined the relationship between air pollutants and lung function (FVC, FEV 1 , MMEF) in a cohort (n = 1,759) of children between the ages of 10 and 18 yr, a period of rapid lung development (Gauderman et al., 2004, 056569 ). Air pollution monitoring stations provided data in each of the 12 study co mmunities from 1994-2000. The results for O 3, PM 10, NO 2, PM 2.5 , acid vapor, and EC are depicted in Figure 7-4 . In general, copollutant models for any pair of pollutants did not provide a substantially better fit to the data than the corresponding single-pollutant models due to the strong correlation between most pollutants. The pollution-related deficits in the average growth in lung function over the 8-yr period resulted in clinically important deficits in attained lung function at the age of 18. Downs et al. (2007, 092853 ) prospectively examined 9,651 randomly selected adults (18-60 yr of age) in eight cities in Switzerland (see alsoAckermann-Liebrich et al., 1997, 077537 ) to ascertain the relationship between reduced exposure to PM 10 and age-related decline in lung function (FVC,
FEV 1 , and FEF 25-50 ). An evaluated statistical dispersion model (Liu et al., 2007, 093093 ) provided spatially resolved concentrations of PM 10 that enabled assignment to residential addresses for the participant examinations in 1991 and 2002 that yielded a median decline of 5.3 μg/m 3 (IQR 4.1-7.5). Decreasing PM 10 concentrations attenuated the decline in l ung function. Effects were greater in tests reflecting small airway function. No other polluta nt relationships were evaluated, though a related study indicated that levels of NO 2 also declined over the same period (Ackermann-Liebrich et al., 2005, 087826 ). Generalized cross-validation essentially chose a linear fit for the concentrationresponse curve for age-related decline in lung function. These data show that improvement in air quality may slow the annual rate of decline in lung function in adulthood indicating positive consequences for public health. Further evidence on improvement in respiratory health with reduction in air pollution levels is provided from studies conducted in East Germany related to dramatic emissions reductions after the reunification in 1990 (Fryer and Collins, 2003, 156454 ; Heinrich et al., 2002, 034825 ; Sugiri et al., 2006, 088760 ). This type of “natural experiment” provides additional support for epidemiologic findings that relatively low levels of airborne particle s have respiratory effects.

December 2009 7-27Source: Adapted from Gauderman et al. (2004, 056569 ) Copyright © 2004 Massachusetts Medical Society. All rights reserved.



### Figure 7-4. Proportion of 18-yr olds with an FEV 1 

### below 80% of the predicted value plotted against the average levels of pollutants from 1994 through 2000 in the 12 southern California communities of the Children’s Health Study. AL = Alpine; AT = Atascadero; LA = Lake Arrowhead; LB = Long Beach; LE = Lake Elsinore; LM = Lompoc; LN = Lancaster; ML = Mira Loma; RV = Riverside; SD = San Dimas; SM = Santa Maria; UP = Upland.

In a prospective cohort study consisting of school -age children (n = 3,170) who were 8 yr of age at the beginning of the study, had not been di agnosed with asthma, and were located in Mexico City, Rojas-Martinez et al. (2007, 091064 ) evaluated the association between long-term exposure to
PM 10 , O 3 and NO 2 and lung function growth every 6 mo 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. The multipollutant model effect of PM 10 over the age of 8-10 yr of life in this cohort on FVC,
FEV 1 , and FEF 25-75 showed an association. Single pollutant models showed an association between ambient pollutants (O 3 , PM 10 and NO 2 ) and deficits in lung function growth. The association between PM 10 and FEF 25-75 was not statistically significant. While the estimates from copollutant models were not substantially different than single pollutant models, independent effects for pollutants could not be estimated accurately becaus e the traffic-related pollutants were correlated.

December 2009 7-28Although no PM 2.5 data were presented in this study, in a separate study Chow et al. (2002) report that during the winter of 1997 approximately 50% of PM 10 was in the PM 2.5 fraction in Mexico City. Gotschi et al. (2008, 156485 ) examined the relationship between air pollution and lung function in adults in the European Commun ity Respiratory Health Survey (ECRHS). FEV 1 and FVC were assessed at baseline and after 9 yr of follo w-up from 21 European centers (followed-up sample n = 5,610). No statistically significant associations were found between city-specific annual mean
PM 2.5 and average lung function levels which is in contrast to the results seen by AckermannLiebrich (1997, 077537 ) (SAPALDIA) and Schikowski et al. (2005, 088637 ) (SALIA) which compared across far more homogenous populations than for the population assessed in the ECRHS. Misclassification and confounding may partially explain the discrepancy in findings. In a birth cohort (n = 2,170) in Oslo, Norway, Oftedal et al. (2008, 093202 ) examined effects of exposure to PM 2.5 and PM 10 on lung function (FVC, FEV 1 , FEF 50% ). Spirometry was performed in 2,307 children aged 9-10 yr in 2001-2002. Reside ntial air pollution levels over the time period 1992-2002 were calculated using EPISODE dispersion models to provide three time scales of exposure: (1) first year of life; (2) lifetime exposure; and (3) just before the lung function test. Only single pollutant models were evaluated because air pollutants were highly correlated (r = 0.83-0.95). PM exposure was associated with changes in adjusted peak respiratory flow, especially in girls. No effect was found for forced volumes. Adjusting fo r contextual socioeconomic factors diminished associations. Results for PM 10 were similar to those for PM 2.5

. In an exploratory study, Mortimer et al. (2008, 187280 ) examined the association of prenatal and lifetime exposure to air pollutants using geocoded monthly average PM 10 levels with pulmonary function in a San Joaquin Valley, California cohort of 232 children (ages 6-11 yr) with asthma. First and second trimester PM 10 exposures (based on monthly average concentrations) had a negative effect on pulmonary function and may relate to pr enatal exposures affecting the lungs as they begin to develop at 6-wk gestation. Dales et al. (2008, 156378 ) in a cross-sectional prevalence study examined the relationship of pulmonary function and PM measures, other pollutant s, and indicators of motor vehicle emissions in Windsor, Ontario, in a cohort of 2,402 school children. PM 2.5 and PM 10 concentrations were estimated for each child’s residence at the postal code level. Each 10 μ g/m 3 increase in PM 2.5 was associated with a 7.0% decrease in FVC expressed in a percentage of predicted. In Leicester, England, investigators examined the carbon content of airway macrophages in induced sputum in 64 of 114 healthy 8-15 year-old children (Grigg et al., 2008, 156499 ; Kulkarni et al., 2006, 089257 ). The carbon content of airway macrophages (Finch et al., 2002, 054603 ; Strom et al., 1990, 157020 ) was used as a marker of individual exposure to PM 10 . Near each child’s home, exposure to PM 10 was estimated using a statistical dispersion model (Pierse et al., 2006, 088757 ). The authors reported a dose-dependent inverse a ssociation between the carbon content of airway macrophages and lung function in children and found no evidence that reduced lung function itself causes an increase in carbon content. Consistent results were obtained for both FVC and FEF 25-75

. Caution should be used when interpreting these results as the accuracy of the estimates on individual
PM 10 exposures were not validated; there is potential for confounding by ethnic origin; and there is concern that the magnitude of the changes in pulmo nary function associated with increased particle area appear large (Boushey et al., 2008, 192162 ). Nordling et al. (2008, 097998 ) discussed above in the respiratory symptoms section, also reported that lower PEF at age 4 was assoc iated with exposure to traffic-related PM 10 (-8.93 L/min [95% CI: -17.78 to -0.088]). Goss et al. (2004, 055624 ), discussed in Section 7.3.1.1 , found strong inverse relationships between FEV 1 and PM 2.5 concentrations in both cross-sectional and longitudinal analyses. In summary, recent studies have greatly ex panded the evidence available for the 2004 PM AQCD. The earlier CHS studies followed young chil dren for 2-4 yr. New analyses have been conducted that include longer follow-up periods of th is cohort through 18 yr of age (considered early adulthood for lung development (Stanojevic et al., 2008, 157007 ) and provide evidence that effects from exposure to PM 2.5 persist into early adulthood. Longitudinal studies follow effects over time and are considered to provide the best evidence as opposed to studies across communities as in cross-sectional studies. The longitudinal cohort studies in the 2004 PM AQCD provided data for children in one location in one study and new long itudinal studies have been conducted in other locations. Gauderman et al. (2004, 056569 ) reported that PM 2.5 exposure was associated with clinically and statistically significant deficits in FEV 1 attained at the age of 18 yr. Clinical significance was

December 2009 7-29defined as a FEV 1 below 80% of the predicted value, a criterion commonly used in clinical settings to identify persons at increased risk for adverse respi ratory conditions. This clinical aspect is an important enhancement over the earlier results reported in the 2004 PM AQCD. Further, the association reported in this study that evaluated the 8-yr time period into early adulthood not only provided evidence for the persistence of the effect, but in addition the strength and robustness of the outcomes were more positive, larger, and more certain than previous CHS studies of shorter followup. Supporting this result are new longitudinal cohort studies conducted by other researchers in other locations with different methods. Though these studies report results for PM 10 , available data discussed above indicate that the majority of PM 10 is composed of PM 2.5 in these areas. New studies provide positive results from Mexico City, Sweden, an d a national cystic fibrosis cohort in the U.S. One study reported null results in a European cohort described as having potential misclassification and confounding concerns as well as lacking a homogenous population potentially rendering the outcome as non-informative. A natural experiment in Switzerland, where PM levels had decreased, reported that improvement in air quality may slow the annual rate of decline in lung function in adulthood, indicating positive consequences for public health. These natural experiments are considered especially supportive. The relationship between long-term PM exposure and decreased lung function is thus seen during lung growth and lung development in school-age children into adulthood. At adult ages studies continue to show a relationship between decreased lung function and long-term PM exposure. Some newer studies attempting to study the relationship of long-term PM exposure from birth through preschool are reporting a relationship. Thus, the impact of long-term PM exposure is seen over the time period of lung function growth and development and the decline of lung function with aging. Overall, effect estimates from these studies are negative (i.e., indicating decreasing lung function) and the pattern of effects are si milar between the studies for FVC and FEV 1 . Thus, the data are consistent and coherent across several designs, locations, and researchers. With cautions noted, the results relating carbon content of airway macrophages to decreased measures of pulmonary function add plausibility to the epidemiologic findings. Some new studies are using individual estimates of exposure to ambient PM to reduce the impact of exposure error (Downs et al., 2007, 092853 ; Jerrett et al., 2005, 087381 ). As was found in the 2004 PM AQCD, the studies report associations with PM 2.5 and PM 10 , while most did not evaluate PM 10-2.5 . Associations have been reported with fine particle components, particularly EC and OC. Source apportionment me thods generally have not been used in these longterm exposure studies. However, numerous studies have evaluated exposures to PM related to traffic or motor vehicle sources. For example, Meng et al. (2007, 093275 ) investigated the associations between traffic and outdoor pollution levels and po orly controlled asthma among adults who were respondents to the California Health Interview Survey and found associations for traffic density and
PM 10 , but not PM 2.5

.



## 7.3. 2.2. Tox icological Studies

Urban Air

One new study evaluated the effects of chronic exposure to ambient levels of urban particles on lung development in the mouse (Mauad et al., 2008, 156743 ). Both functional and anatomical indices of lung development were measured. Ma le and female BALB/c mice were continuously exposed to ambient or filtered Sao Paolo air for 8 mo. Concentrations in the “polluted chamber” versus “clean chamber” were 16.8 versus 2.9 μg/m 3 PM 2.5 . Thus PM levels were reduced by filtration but not entirely eliminated. Ambient concentrations of CO, NO 2 and SO 2 were 1.7 ppm, 89.4 μg/m 3 and 8.1 μg/m 3 , respectively. Concentrations of gase ous pollutants were assumed to be similar to ambient levels in both chambers. Afte r 4 mo, the animals were mated and the offspring were divided into 4 groups to provide for a pr enatal exposure group, a postnatal exposure group, a pre and postnatal exposure group and a control group. Animals were sacrificed at 15 and 90 days of age for histological analysis of lungs. Pulmona ry pressure-volume measurements were also conducted in the 90-day-old offspring. Statistically significant reductions in inspiratory and

December 2009 7-30expiratory volumes were found in the group receiving both prenatal and postnatal exposure, but not in the groups receiving only prenatal exposure or on ly postnatal exposure, compared with controls. These changes in pulmonary function correlated with anatomical changes which are discussed in Section 7.3.5.1

.



### Diesel Exhaust

Li et al. (2007, 155929 ) exposed BALB/c and C56BL/6 mice to clean air or to low-dose DE (at a PM concentration of 100 μg/m 3 ) for 7 h/day and 5 days/week for 1, 4 and 8 wk. Average gas concentrations were reported to be 3.5 ppm CO, 2.2 ppm NO 2 , and less than 0.01 ppm SO 2 . Airway hyperresponsiveness (AHR) was evaluated by whole- body plethysmography at Day 0 and after 1, 4 and 8 wk of exposure. Short-term exposure responses are discussed in Secti on 6.3.2.3, 6.3.3.3 and 6.3.4.2. The increased sensitivity of airways to methacholine (measured as Penh) seen in C57BL/6 but not BALB/c mice at 1 week was also seen at 4 wk but not at 8 wk. This study suggests that adaptation occurs during prolonged DE exposure. Influx of inflammatory cells, markers of oxidative stress and effects of antioxidant intervention were also evaluated (Sections 7.3.3.2 and 7.3.4.1 ). Although no attempt was made in this study to determine the effects of gaseous components of DE on the measured responses, concentrations of gase s were very low suggesting that PM may have been responsible for the observed effects. In many animal studies changes in vent ilatory patterns are assessed using whole-body plethysmography, for which measurements are reported as enhanced pause (Penh). Some investigators report increased Penh as an indicator of AHR, but these are inconsistently correlated and many investigators consider Penh solely an indicat or of altered ventilatory timing in the absence of other measurements to confirm AHR. Therefore use of the terms AHR or airway responsiveness has been limited to instances in which the te rminology has been similarly applied by the study investigators. Gottipolu et al. (2009, 190360 ) exposed WKY and SH rats to filtered air or DE (particulate concentration 500 and 2,000 μg/m 3 ) for 4 h/day and 5 days/wk over a 4-wk period. Concentrations of gases were 1.3 and 4.8 ppm CO, NO <2.5 and 5.9 ppm NO, <0.25 and 1.2 ppm NO 2 , 0.2 and 0.3 ppm SO 2 for low and high PM exposures, respectively. Particle size, measured as geometric median number and volume diameters, was 85 and 220 nm, respectively. No DE-related effects were found for breathing parameters measured by whol e-body plethysmography. Other pulmonary effects are described in Sections 7.3.3.2 and 7.3.5.1

.



### Woodsmoke

One study evaluated the effects of subchronic woodsmoke exposure on pulmonary function in Brown Norway rats. Rats were exposed 3 h/day and 5 days/week for 4 and 12 wk to air or to concentrated wood smoke from the pinyon pine which is native to the U.S. Southwest (Tesfaigzi et al., 2002, 025575 ). PM concentrations in the woodsmoke were 1,000 and 10,000 μg/m 3 . The particles in this woodsmoke had a bimodal size distribution with the smaller size fraction (74%) characterized by a MMAD of 0.405 μm and the larger size fraction (26%) characterized by a MMAD of 6.7-11.7 μm. Many of these larger partic les would not be inhalable by the rat since 8 μm MMAD particles are about 50% inhalable (Ménache et al., 1995, 006533 ). Concentrations of gases were reported to be 15-106.4 ppm CO, 2.2-18.9 ppm NO, 2.4-19.7 ppm NO X and 3.5-13.8 ppm total hydrocarbon in these exposures. Respiratory function measured by whole-body plethysmography demonstrated a statistically significant increase in total pulmonary resistance in rats exposed to 1000 μg/m 3 woodsmoke. Additional effects were found at 10,000 μg/m 3 . Inflammatory and histopathological responses were also evaluated (Sections 7.3.3.2 and 7.3.5.1 ).

December 2009 7-31

## 7.3.3. Pulmonary Inflammation

7.3. 3.1. Epi demiologic Studies

One epidemiologic study examined the relati onship of airway inflammation (eNO) and PM measures, other pollutants, and indicators of motor vehicle emissions in Windsor, Ontario (Dales et al., 2008, 156378 ). This cohort of 2,402 school children estimated PM 2.5 and PM 10-2.5 for each child’s residence at the postal code level with an ev aluated statistical model (Wheeler et al., 2006, 103905 ). Each 10 μ g/m 3 increase in 1-yr PM 2.5 was associated with a 39% increase in eNO (p = 0.058). Associations between eNO and PM 10-2.5 were positive but not statistically significant.



## 7.3. 3.2. Tox icological Studies

CAPs Studies

A set of subchronic studies involved exposure of normal (C57BL1/6) mice, ApoE -/- and the double-knockout ApoE -/- /LDLR -/- mice to Tuxedo, NY CAPs for 5-6 month (March, April or May through September 2003 (Lippmann et al., 2005, 087452 ). The average PM 2.5 exposure concentration was 110 μ g/m 3 . Animals were fed a normal chow diet during the CAPs exposure period. No pulmonary inflammation was observed in response to CAPs exposure as measured by BALF cell counts and histology. The lack of a persistent pulmonary response may have been due to adaptation of the lung following repeated exposures. In fact, a parallel study examined CAPs-related gene expression in the double-knockout animals a nd found upregulation of numerous genes in lung tissue (Gunnison and Chen, 2005, 087956 ). An in vitro study conducted simultaneously found daily variations in CAPs-mediated NF- κ B activation in cultured human bronchial epithelial cells, suggesting that transcription factor-mediated gene upregulation could occur in response to CAPs (Maciejczyk and Chen, 2005, 087456 ). It should be noted that significant cardiovascular effects were observed in these subchronic studies which are discussed in Section 7.2.1.2

. Araujo et al. (2008, 156222 ) compared the relative impact of UF (0.01-0.18 μ m) versus fine (0.01-2.5 μ m) PM inhalation in ApoE -/- mice following a 40 day exposure (5 h/day×3 days/wk for 75 total hours). Animals were fed a normal chow diet and exposed to PM from November 3 -December 12, 2005 in a mobile inhalation laboratory that was parked 300 m from the 110 Freeway in downtown Los Angeles. Particles were concentrated to ~440 μ g/m 3 for PM 2.5 exposures and ~110 μ g/m 3 for the UF exposures, representing a roughly 15-fold increase in concentration from ambient levels; the number concentration of PM in the fine and UF chambers were roughly equivalent (4.56×10 5 and 5.59×10 5 particles/cm 3 , respectively). Over 50% of the UFPs were comprised of OC compared to only 25% for PM 2.5 . No major increase in BALF inflammatory cells was found in response to PM. However UFP exposure resulted in significant cardiovascular and systemic effects (Section 7.2.1.2 ).



### Diesel Exhaust

Gottipolu et al. (2009, 190360 ) exposed WKY and SH rats to filtered air or DE for 4 wk as described in Section 7.3.2.2 . Previous studies from this laboratory have shown enhanced effects ofPM in SH compared with WKY rats. Although the main focus of this recent study was on DEinduced mitochondrial oxidative stress and hypertensive gene expression in the heart (Section 7.2.7.1 ), some pulmonary effects were also found. Subchronic exposure to DE resulted in a dose-dependent increase in BALF neutrophils in both rat strains although levels of measured cytokines were not altered. Histological analysis of lung tissue from rats exposed to the higher concentration of DE demonstrated accumulation of particle-laden macrophages as well as focal alveolar hyperplasia and inflammation. Effect on indices of injury are discussed in Section 7.3.5.1

. Ishihara and Kagawa (2003, 096404 ) exposed Wistar rats to filtered air and DE containing 200, 1,000 and 3,000 μg/m 3 PM for 16 h/day and 6 days/wk for 6, 12, 18 or 24 mo. The mass median particle diameter was reported to be between 0.3 and 0.5 μm. Concentrations of gases ranged from

December 2009 7-322.93-35.67 ppm NO X , 0.23-4.57 ppm SO 2 , 1.8-21.9 ppm CO in the DE exposures. Statistically significant increases in total numbers of inflammatory cells and neutrophils in BALF were observed beginning at 6-12 mo of exposure to DE containing 1,000 and 3,000 μg/m 3 PM. When rats were exposed to DE containing 1,000 μg/m 3 PM, which was filtered to remove PM, the inflammatory cell response was significantly diminished. These results implicate the PM fraction of DE as a key determinant of the inflammation. The PM fraction wa s also found to mediate the increase in protein levels, the decrease in PGE 2 levels and alterations in mucus and surfactant components observed in BALF (Section 7.3.5.1 ). Li et al. (2007, 155929 ) exposed BALB/c and C56BL/6 mice to low dose DE as described in Section 7.3.2.2 . for 1, 4 and 8 wk. Increases in num bers of BALF macrophages and total inflammatory cells were observed in BALB/c mice at 8 wk but not 4 wk of DE exposure. Persistent increases in numbers of BALF neutrophils and lym phocytes were observed in both strains at 4 and 8 wk of DE exposure. Corresponding increases in BALF cytokines differed between the two strains. These results should be interpreted with caution sin ce comparisons were made with Day 0 controls rather than age-matched controls. No histopathologi cal changes in the lungs were seen at any time point after DE exposure. This study demonstrated differences in pulmonary responses to low dose DE between two mouse strains. AHR, pulmonary inflammation, markers of oxidative stress and effects of antioxidant intervention were also evaluated (Sections 7.3.2.2 and 7.3.4.1 ). Although no attempt was made in this study to determine the effects of gaseous components of DE on the measured responses, concentrations of gases were very low suggesting that PM may have been responsible for the observed effects. In a study by Hiramatsu et al. (2003, 155846 ), BALB/c and C57BL/ 6 mice were exposed to DE (PM concentrations 100 and 3,000 μg/m 3 ) for 1 or 3 mo. Concentrations of gases were reported to be 3.5-9.5 ppm CO, 2.2-14.8 ppm NO X , and less than 0.01 ppm SO 2 . Modest increases in BALF neutrophils and lymphocytes were observed in response to DE in both mouse strains at 1 and 3 mo. Histological analysis demonstrated diesel exposure pa rticle-laden alveolar macrophages in alveoli and peribronchial tissues at both time points. Br onchus-associated lymphoid tissue developed after 3-month exposure to the higher concentration of DE in both mouse strains. Mac-1 positive cells (a marker of phagocytic activation of alveolar macr ophages) were also increased in BALF of BALB/c mice exposed to the higher concentration of DE for 1 and 3 mo. Increased expression of several cytokines and decreased expression of iNOS mR NA was observed in DE-exposed mice at 1 and 3 mo. NF- κ B activation was also noted following 1-mont h exposure to the lower concentration of DE. No attempt was made in this study to determine the responses to gaseous components of the DE. In a study by Reed et al. (2004, 055625 ), healthy Fisher 344 rats and A/J mice were exposed to DE (PM concentration = 30, 100, 300 and 1,000 μg/m 3 ) by whole body inhalation for 6 h/day, 7 days/wk for either 1 week or 6 mo. Concentrati ons of gases were reported to be 2.0-45.3 ppm NO, 0.2-4.0 ppm NO 2 , 1.5-29.8 ppm CO and 8-365 ppb SO 2 . Short-term responses are discussed in Section 6.3.3.3 and 6.3.7.2, and sub-chronic systemic effects are presented in Section 7.2.4.1 . Six months of exposure resulted in no measurable effects on pulmonary inflammation. However numerous black particles were observed within al veolar macrophages after 6 mo of exposure. Seagrave et al. (2005, 088000 ) evaluated pulmonary responses in male and female CDF (F-344)/CrlBR rats exposed 6 h/day for 6 mo to fi ltered air or DE at concentrations ranging from 30-1000 μg/m 3 PM. Concentrations of gases were reported for the highest exposure as 45.3 ppm NO, 4.0 ppm NO 2 , 29.8 ppm CO and 2.2 ppm total vapor hydrocarbon. No changes in BALF cells were noted. A small decrease in TNF- α was seen in BALF of female rats exposed to the highest concentration of DE for 6 mo. Pulmonary injury also was evaluated (Section 7.3.5.1 ). Thus changes in BALF markers were modest and gender-specific.



### Woodsmoke

Seagrave et al. (2005, 088000 ) also evaluated pulmonary responses in male and female CDF (F344)/CrlBR rats exposed 6 h/day for 6 mo to filtered air or hardwood smoke concentrations ranging from 30-1,000 μg/m 3 PM. Concentrations of gases were reported for the highest exposure as 3.0 ppm CO and 3.1 ppm total vapor hydrocarbon . A small increase in BALF neutrophils was observed in male rats exposed to the lowest concen tration of hardwood smoke. Female rats exhibited a decrease in BALF macrophage inflammatory prot ein-2 (MIP-2) at the highest concentration of hardwood smoke. Pulmonary injury also was evaluated (Section 7.3.5.1 ). In general, responses to

December 2009 7-33hardwood smoke were more remarkable than responses to DE seen in a parallel study. However these gender-specific responses were modest and difficult to interpret. In a study by Reed et al. (2006, 156043 ), Fisher 344 rats, SHR rats, A/J mice and C57BL/6 mice were exposed to clean air or hard wood smoke (PM concentrations 30, 100, 300 and 1,000 μg/m 3 ) by whole body inhalation for 6 h/day, 7 days/wk for either 1 week or 6 mo. Concentrations of gases ranged from 229.0-14887.6 mg/m 3 for CO, 54.9-139.3 μg/m 3 for ammonia, and 177.6- 3455.0 μg/m 3 nonmethane VOC in these exposures. Short-term responses are discussed in Section 6.3.7.2 and sub-chronic effects are presented in Section 7.2.4.1 . Histological analysis of lung tissue showed minimal increases in alveolarmacrophages. The effects of hardwood smoke on bacterial clearance are discussed below (Section 7.3.7.2 ). Another study evaluated the effects of subchron ic woodsmoke exposure in Brown Norway rats and is described in detail in Section 7.3.2.2 (Tesfaigzi et al., 2002, 025575 ). Numbers of alveolar macrophages in BALF were significantly increased in rats exposed to 1,000 μg/m 3 woodsmoke for 12 wk, but no changes were seen in numbers of ot her inflammatory cells. A large percent of BALF macrophages contained carbonaceous material. Hist ological analysis of lung tissue showed minimal to mild inflammation in the epiglottis of the larynx in rats exposed to both concentrations of woodsmoke. Ramos et al. (2009, 190116 ) examined the effects of subchronic woodsmoke exposure on the development of emphysema in guinea pigs. Inflammation is thought to be involved in the pathogenesis of this form of COPD. Statistically si gnificant increases in total numbers of BALF cells were observed in guinea pigs exposed to smoke for 1-7 mo, with numbers of macrophages increased at 1-4 mo and numbers of neutrophils increase d at 4-7 mo. At 4 mo, alveolar mononuclear phagocytic and lymphocytic peribronchiolar inflammation were observed by histological analysis of lung tissue. This study is discussed in depth in Section 7.2.5.1

.



### Model Particles

Wallenborn et al. (2008, 191171 ) examined the pulmonary, cardiac and systemic effects of subchronic exposure to particulate ZnSO 4 . WKY rats were exposed nose-only to 10, 30, or 100 μ g/m 3 UFP of ZnSO 4 for 5 h/day and 3 day/wk over a 16-wk period. Particle size was reported to be 31-44 nm measured as number median di ameter. No changes in pulmonary inflammation or injury were observed although cardiac effects were noted (Section 7.2.7.1 ). This study possibly demonstrates a direct effect of ZnSO 4 on extrapulmonary systems, as suggested by the lack of pulmonary effects.



## 7.3.4. Pulmonary Oxidative Response

7.3. 4.1. Tox icological Studies

Urban Air

One new study evaluated the effects of subchronic exposure to ambient levels of urban particles on the development of emphysema in papain-treated mice (Lopes et al., 2009, 190430 ). Since oxidative stress is thought to contribute to the development of emphysema, 8-isoprostane levels were measured in lung tissue from the four groups of mice used in this study. A statistically significant increase in 8-isoprostane, a marker of oxidative stress, was observed in lungs from mice treated with papain and exposed to ambient air co mpared with the other groups of mice. This study is described in greater depth in Section 7.3.5.1

.

December 2009 7-34

### Diesel Exhaust

Li et al. (2007, 155929 ) exposed mice to low dose DE for 1, 4 and 8 wk as described in Section 7.3.2.2 . Markers of oxidative stress and effects of antioxidant intervention were evaluated in this model. While HO-1 mRNA and protein were increased in lung tissues of both mouse strains after 1 week of DE exposure (Section 6.3.4.2), at 8 wk of DE exposure, HO-1 protein levels remained high in C57BL/6 mice but returned to control values in BALB/c mice. This study demonstrates differences in pulmonary responses to low dose DE between two mouse strains. Furthermore, this study suggests that adaptati on occurs in BALB/c mice during prolonged DE exposure since the increase in HO-1 protein seen in both strains at 1 week of exposure was only seen in C57BL/6 mice at 8 wk. AHR (Section 7.3.2.2 ) and pulmonary inflammation (Section 7.3.3.2 ) were also evaluated. Although no attempt was made in this study to determine the effects of gaseous components of DE on the measured responses, concentrations of gases were very low. This suggests that PM may have been responsible for the observed effects.



## 7.3.5. Pulmonary Injury

7.3. 5.1. Tox icological Studies

Urban Air

One new study evaluated the effects of chronic exposure to ambient levels of urban particles on lung development in the mouse (Mauad et al., 2008, 156743 ). Both functional and anatomical indices of lung development were measured in mice exposed prenatally and/or postnatally as described in Section 7.3.2.2 . Animals were sacrificed at 15 a nd 90 days of age for histological analysis of lungs. Histological analysis demonstrated the presence of mild foci of macrophages containing black dots of carbon pigment in the prenatal and postnatal exposure group at 90 days. In addition, the alveolar spaces of 15-day old mice in the prenatal and postnatal exposure group were enlarged compared with controls. Morphometric analysis demonstrated statistically significant decreases in surface to volume ratio at 15 and 90 days in the prenatal and postnatal exposure group compared with controls. Since alveolarization is normally complete by 15 days of age, these results suggest incomplete alveolarization in the 15-day-old group and an enlargement of air spaces in the 90-day-old group. These anatomical changes correlated with decrements in pulmonary function which are discussed in Section 7.3.2.2

. Prolonged exposure to low levels of ambient air pollution beginning in early life has been linked to secretory changes in the nasal cavity of mice, specifically increased production of acidic mucosubstances (Pires-Neto et al., 2006, 096734 ). Six-day-old Swiss mice were continuously chamber exposed to ambient or filtered São Paulo air for 5 mo. Concentrations in the “polluted chamber” versus “clean chamber” were (in μg/m 3 ) 59.52 versus 37.08 for NO 2 , 12.52 versus 0 for BC, and 46.49 versus 18.62 for PM 2.5 . Thus, pollutant levels were reduced by filtration but not entirely eliminated. Compared to filtered air, ex posure to ambient air resulted in increased total mucus and acidic mucus in the epithelium lining the nasal septum, but no statistically significant differences in other parameters (amount of neut ral mucus, volume proportions of neutral mucus, total mucus, or nonsecretory epithelium, epithelial thickness, or ratio between neutral and acidic mucus). The physicochemical properties of mucus glycoproteins are critical to the protective function of the airway mucus layer. Acidified mu cus is more viscous, and is associated with a decrease in mucociliary transport. Thus acidic mucosubstances may represent impaired defense mechanisms in the respiratory tract. One new study evaluated the effects of subchronic exposure to ambient levels of urban particles on the development of emphysema in papain-treated mice (Lopes et al., 2009, 190430 ). Emphysema is a form of COPD caused by the destru ction of extracellular matrix in the alveolar region of the lung which results in airspace enlarg ement, airflow limitation and a reduction of the gas-exchange area of the lung. Inflammation, oxida tive stress, protease imbalance and apoptosis are thought to contribute to the development of emphysema. In this study, male BALB/c mice were

December 2009 7-35continuously exposed to ambient or filtered Sa o Paulo air for 2 mo. Concentrations of PM 2.5 in the “polluted chamber” versus “clean chamber” were 33.86 ± 2.09 versus 2.68 ± 0.38 μg/m 3 . Thus filtration reduced PM levels considerabl y. Ambient concentrations of CO and SO 2 were 1.7 ppm and 16.2 μg/m 3 respectively. No significant difference was observed in the concentrations of NO 2 in the “polluted chamber” versus “clean chamber” (60-80 μg/m 3 ). Half of the mice were pre-treated with papain by intranasal instillation in order to i nduce emphysema. Morphometric analysis of lung tissue demonstrated a statistically significant increase in mean linear intercept, a measure of airspace enlargement, in papain-treated mice compared with saline-treated controls exposed to filtered air. While exposure to ambient air failed to increase mean linear intercept values in saline-treated mice, mean linear intercept values were significantly increased in papain-treated mice exposed to ambient air compared with papain-treated mice exposed to filtered air. A similar pattern of responses was observed for the volume proportion of collagen and elastin fibers in alveolar tissue, which are markers of alveolar wall remodeling. Lung immunohi stochemical analysis dem onstrated an effect of papain, but not ambient air, on macrophage cell density and matrix metalloproteinase 12-positive cell density. No differences in caspase-3 positive cells, a marker of apoptosis, were observed between the four groups of mice. Oxidative stress was evaluated in this model as described in Section 7.3.4.1 . Taken together, results of this study dem onstrate that urban levels of PM, mainly from traffic sources, worsen protease-induced emphysema in an animal model. Pulmonary vascular remodeling, measured by a decrease in the lumen to wall ratio, was observed in mice exposed to ambient São Paulo air for 4 mo (Lemos et al., 2006, 088594 ). This study is described in greater detail in Section 7.2.1.2

. Kato and Kagawa (2003, 089563 ) exposed Wistar rats to roadside air contaminated mainly with automobile emissions (55.7-65.2 ppb NO 2 and 63-65 μg/m 3 suspended PM [SPM]) and examined the effects on respiratory tissue after 24, 48, or 60 wk of exposure. The surface of the lungs was light gray in color after all durations of exposure, and BC particle deposits accumulated with prolonged exposure. These characteristics were not evident in filtered air-exposed control animals, although filtered air contained low levels of air pollutants ( ≤ 6.2 ppb NO 2 and 15 μg/m 3 SPM). The most common change observed using transmission electron microscopy was the presence of particle laden (anthracotic) al veolar macrophages, or anthracosis, in a wide range of pulmonary tissues, including the submucosa, tracheal- and bronchiole-associated lymph nodes, alveolar wall and space, pleura, and perivascular connective tissue. These changes were evident after 24 wk and increased with duration of exposure. Other change s included increases in the number of mucus granules in goblet cells, mast cell infiltrati on (but no degranulation) after 24 wk, increased lysosomes in ciliated cells, some altered morphology of Clara cells, and hypertrophy of the alveolar walls after 48 wk. No goblet cell proliferation was observed, but slight, variable acidification of mucus granules appeared after 24 and 48 wk and disappeared after 60 wk. Anthracotic macrophages were seen in contact with plasma cells and lymphocytes in the lymphoid tissue, suggesting immune cell interaction in the immediate vicinity of par ticles. Even after 60 wk, no lymph node anthracosis was observed in the filtered air group. In a post-mortem study of lung tissues from 20 fema le lifelong residents of Mexico City, a high PM locale, histology demonstrated significantly greater amounts of fibrous tissue and muscle in the airway walls compared to subject s from Vancouver (Churg et al., 2003, 087899 ), a city with relatively low PM levels. Electron microscopy sho wed carbonaceous aggregates of UFPs, which the authors conclude penetrate into and are retained in the walls of small airways. The study shows an association between retained particles and airway remodeling in the form of excess muscle and fibrotic walls. The subjects were deemed suitable for examination based on never-smoker status, no use of biomass fuels for cooking, no known occupational particle/dust exposure, death by cause other than respiratory disease, and extended reside nce in each locale (lifelong for Mexico City and >20 yr for Vancouver). However, subjects from the two locales were not matched with respect to ethnicity, sex (20 females from Mexico City ver sus 13 females and 7 males from Vancouver), or mean age at death (66 ± 9 versus 76 ± 11), and other possibly influential factors such as exercise or diet were not considered.



### Diesel Exhaust

Gottipolu et al. (2009, 190360 ) exposed WKY and SH rats to filtered air or DE as described in Section 7.3.2.2 . Previous studies from this laboratory have shown enhanced effects of PM in SH

December 2009 7-36compared with WKY rats. Although the main focus of this recent study was on DE-induced mitochondrial oxidative stress and hypertensive gene expression in the heart (Section 7.2.7.1 ), some pulmonary effects were found. Inflammato ry effects are described in Section 7.3.3.2 . GGT activity in BALF was increased in both strains in response to the higher concentration of DE. No DE-related changes were observed in BALF protein or albumin. Histological analysis of lung tissue from rats exposed to the higher concentration of DE demonstrated accumulation of particle-laden macrophages as well as focal alveolar hyperplasi a and inflammation. No effects on indices of pulmonary function were observed (Section 7.3.2.2 .) Ishihara and Kagawa (2003, 096404 ) exposed rats to DE for up to 24 mo as described in Section 7.3.3.2 . A statistically significant increase in BALF protein was observed at 12 mo of exposure to DE containing 1,000 μg/m 3 PM. This response was attenuated when the DE was filtered to remove PM. Pulmonary inflammation wa s noted and is described in Section 7.3.3.2

. Seagrave et al. (2005, 088000 ) evaluated pulmonary responses in rats exposed to DE for up to 6 mo as described in Section 7.3.3.2 . A small increase in LDH was seen in BALF of female rats exposed to the highest concentration of DE for 6 mo. Pulmonary inflammation was also evaluated (Section 7.3.3.2 ). The changes in BALF markers in this study were modest and gender-specific.



### Gasoline Exhaust

Reed et al. (2008, 156903 ) examined a variety of health effects following subchronic inhalation exposure to gasoline engine exhaust. Male and female CDF (F344) /CrlBR rats, SHR rats and male C57BL/6 mice were exposed for 6 h/day and 7 days/wk for a period of 3 days-6 mo. The dilutions for the gasoline exhaust were 1:10, 1:15 and 1:90; filtered PM was at the 1:10 dilution. PM mass ranged from 6.6 to 59.1 μ g/m 3 , with the corresponding number concentration between 2.6×10 4

and 5.0×10 5 particles/cm 3 . Concentrations of gases ranged from 12.8-107.3 ppm CO, 2.0-17.9 ppm NO, 0.1-0.9 ppm NO 2 , 0.09-0.62 ppm SO 2 and 0.38-3.37 ppm NH 3 . Other effects are described in Sections 7.2.4.1 and 7.3.6.1 . No pulmonary inflammation or histopathological changes were noted in the F344 rats and A/J mice, except for a time-dependent increase in the number of macrophages containing PM. However statistically significant increases of 47% and 29% in BALF LDH were observed in female and male F344 rats, respectiv ely, after 6 mo of exposure to the highest concentration of engine exhaust. This response wa s absent when gasoline exhaust was filtered, implicating PM as a key determinant of this response. In addition, exposure to the highest concentration of gasoline exhaust resulted in stat istically significant decreases in hydrogen peroxide and superoxide production in unstimulated and stim ulated BALF macrophages. Hypermethylation of lung DNA was observed in male F344 rats following 6 mo of exposure to gasoline exhaust containing 30 μ g/m 3 PM. This response was PM-dependent since it was absent in mice exposed to filtered gasoline exhaust. The signifi cance of this epigenetic change in terms of respiratory health effects is not known. However, altered patterns of DNA methylation can affect gene expression and are sometimes associated with altered immune responses and/or the development of cancer.



### Woodsmoke

Seagrave et al. (2005, 088000 ) also evaluated pulmonary responses in rats exposed to hardwood smoke for 6 mo as described in Section 7.3.3.2 . Increases in BALF LDH and protein were seen in male but not female rats. Female rats exhi bited a decrease in BALF glutathione at the highest concentration of hardwood smoke. Decreases in BALF alkaline phosphatase were found in both males and females exposed to 1,000 μg/m 3 hardwood smoke. Male rats exposed to 100 and 300 μg/m 3 hardwood smoke exhibited a decrease in BALF β -glucuronidase activity. Pulmonary inflammation was also evaluated (Section 7.3.3.2 ). These changes in BALF markers in this study were modest and gender-specific. Another study evaluated the effects of subchron ic woodsmoke exposure in Brown Norway rats as described in Section 7.3.2.2 . (Tesfaigzi et al., 2002, 025575 ). Exposure to 1,000 μg/m 3 woodsmoke for12 wk resulted in a statistically significant increase in Alcian Blue- (AB) and Periodic Acid Schiff- (PAS) positive airway epithelial cells compared to controls, indicating an increase in mucous secretory cells containing neutra l and acid mucus, respectively. More significant histopathological responses were found following exposure to 10,000 μg/m 3 of DE. Pulmonary

December 2009 7-37function and inflammation were evaluated also but are not discussed here due to the extremely high exposure level (Sections 7.3.2.2 . and 7.3.3.2 ). Ramos et al. (2009, 190116 ) examined the effects of subchronic woodsmoke exposure on the development of emphysema in guinea pigs. In pa rticular, the involvement of macrophages and macrophage-derived MMP in woodsmoke-related responses was investigated. Guinea pigs were exposed to ambient air or to whole smoke from pine wood for 3 h/day and 5 days/wk over a 7-month period. PM 10 and PM 2.5 concentrations in the exposure chambers were reported to be 502 ± 34 and 363 ± 23 μg/m 3 , respectively, while the concentration of CO was less than 80 ppm. COHb levels were reported to be 6% in controls and 15-20% in smoke-exposed guinea pigs. Statistically significant decreases in body weight were observed in guinea pigs exposed to smoke for 4 or more months compared with controls. Statistically signi ficant increases in total numbers of BALF cells were observed in guinea pigs exposed to smoke for 1-7 mo, with numbers of macrophages increased at 1-4 month and numbers of neutrophils increas ed at 4-7 mo. At 4 mo, alveolar mononuclear phagocytic and lymphocytic peribronchiolar infla mmation, as well as bronchiolar epithelial and smooth muscle hyperplasia, were observed by hi stological analysis of lung tissue. Emphysematous lesions, smooth muscle hyperplasia and pulmonary arterial hypertension were noted at 7 mo. Morphometric analysis of lung tissue demonstrated statistically significant increases in mean linear intercept values, a measure of airspace enlargement, in guinea pigs at 6 and 7 mo of exposure. Statistically significant increases in elastolytic activity was observed in BALF macrophages and lung tissue homogenates at 1-7 mo of exposure. Lung collagenolytic activity was also increased at 4-7 mo of exposure and corresponded in time with the presence of active forms of MMP-2 and MMP-9 in lung tissue homogenates and BALF. Furthermore, MMP-1 and MMP-9 immunoreactivity was detected in macrophages, epithelial and inters titial cells in smoke-exposed animals at 7 mo. Increased levels of MMP-2 and MMP-9 mRNA were also found in smoke-exposed guinea pigs after 3-7 mo. Apoptosis was found in BALF macrophages (TUNEL assay) from guinea pigs exposed to smoke for 3-7 mo and in alveolar epithelial cells (caspase-3 immunoreactivity) after 7 mo. Taken together, these results provide evidence that subchronic exposure to woodsmoke leads to the development of emphysematous lesions accompanied by the accumulation of alveolar macrophages, increased levels and activation of MMPs, connec tive tissue remodeling and apoptosis. However, the high levels of CO and COHb reported in this study make it difficult to conclude that woodsmoke PM alone is responsible for these dramatic effects.



## 7.3.6. Allergic Responses

7.3. 6.1. Epi demiologic Studies

A number of epidemiologic studies have foun d associations between PM and allergic (or atopic) indicators. Allergy is a major driver of asth ma, which has been associated with PM in studies discussed in previous sections. In a study by Annesi-Maesano (2007, 093180 ) (described in Section 7.3.1.1 ) atopic asthma was related to PM 2.5 (OR 1.43 [95% CI: 1.07-1.91]) and positive skin prick test to common allergens was also increased with higher PM levels. This report is consistent with the results from an earlier study (Penard-Morand et al., 2005, 087951 ) in the same sample of children that associated allergic rhinitis and atopic dermatitis with PM 10 . Also, Morgenstern et al. (2008, 156782 ) found statistically significant effects for as thmatic bronchitis, hay fever, and allergic sensitization to pollen in a cohort of ch ildren in Germany examining modeled PM 2.5 data at birth addresses. Distance to a main road had a dose-re sponse relationship with sensitization to outdoor allergens. Nordling et al. (2008, 097998 ) (discussed above in Section 7.3.2.1 ) reported a positive association of PM 10 exposure during the first year of life with allergenic sensitization (IgE antibodies) to inhaled allergens, especially pollen. In a study by Brauer et al. (2007, 090691 ) (discussed above in Section 7.3.1.1 ) an interquartile range increase in PM 2.5 was associated with an increased risk of sensitization to food allergens (OR 1.75 [95% CI 1.23-2.47]). A significant association was found for sensitization to any allergen, but none was found for sensitization to specific indoor or outdoor aeroallergens or atopic dermatitis (eczema). In a study by Janssen et al. (2003, 133555 ), PM 2.5 was associated with allergic indicators su ch as hay fever (ever), skin prick test reactivity to outdoor allergens, current itchy rash, and conjunctivitis in Dutch children. These same outcomes were also associated with proximity of th e school to truck traffic but not car traffic,

December 2009 7-38suggesting a role for diesel-related pollution. Consistent with the aforementioned Dutch study by Brauer et al. (2007, 090691 ), PM 2.5 was not associated with eczema. Mortimer et al. (2008, 187280 ) examined the association between prenatal and early-life exposures to air pollutants with allergic sensitizati on in a cohort of 170 children with asthma, ages 6-11 yr, living in central California. Sensitization to at least one allergen was associated with higher levels of PM 10 and CO during the entire pregnancy and 2nd trimester and higher PM 10 during the first 2 yr of life. Sensitization to at least one i ndoor allergen was associated with higher exposures to
PM 10 and CO in during the entire pregnancy and during the 2nd trimester. However, no significant associations remained for PM 10 after adjustment for copollutants, effect modifiers, or potential cofounders in addition to year of birth. The aut hors advise that the large number of comparisons may be of concern and this study should be viewed as an exploratory, hypothesis-generating undertaking. In examining the National Health Interview Surv ey for the years 1997-2006, Bhattacharyya et al. (2009, 180154 ) found relationships between air quality and the prevalence of hay fever and sinusitis. However, the air quality data were not clearly defined and as such caution is required in interpretation of these results. In contrast, Bayer-Oglesby et al. (2005, 086245 ) found no significant association between declining levels of PM 10 and hay fever in Switzerland. In a study by Oftedal et al. (2007, 191948 ) conducted in Oslo, Norway, early-life exposure to PM 10 or PM 2.5 was generally not associated with sensitization to allergens in 9- to 10-yr-old children; lifetime exposures to PM 10 and PM 2.5 were associated with dust mite allergy, but the association was diminished by adjustment for socioeconomic factors . In Norway, wood burni ng in the wintertime is thought to account for about half of the PM 2.5 levels. Although associations between PM and reactivity to specific allergens have been reported in long-term studies, there is a consistent lack of correlation between PM and total IgE levels, indicating a selective enhancement of allergic responses.



## 7.3. 6.2. Tox icological Studies

Diesel Exhaust

Exposure to relatively low doses of DE has been shown to exacerbate asthmatic responses in ovalbumin (OVA) sensitized and challenged BALB/c mice (Matsumoto et al., 2006, 098017 ). Mice were intraperitoneally sensitized and intranasa lly challenged 1 day prior to inhalation exposure to DE (PM concentration 100 μg/m 3 ; CO, 3.5 ppm; NO 2 , 2.2 ppm; SO 2 <0.01 ppm) for 1 day or 1, 4, or 8 wk (7/h/day, 5 days/wk, endpoints 12 h post DE exposure). Results from the 1- and 4-wk exposures are described in Section 6.3.6.3. It should be noted that control mice were left in a clean room as opposed to undergoing chamber exposure to filtered air. The significant increases in AHR and airway sensitivity observed following shorter exposure periods did not persist at 8 wk. BALF cytokines were altered by DE exposure with only RANTES significantly elevated after 8 wk. DE had no effect on OVA challenge-induced peribronchial inflammatory or mucin positive cells. These results suggest that adaptive processes may have occurred during prolonged exposure to DE.



### Gasoline Exhaust

In a study by Reed et al. (2008, 156903 ), BALB/c mice were exposed to whole gasoline exhaust diluted 1:10 (H), 1:15 (M), or 1:90 (L), filtered exhaust at the 1:10 (HF), or clean air for 6 h/day (atmospheric characterization described inSection 6.3.6.3). GEE exposure from conception through 4 wk of age induced slight but non-significant increases in OVA-specific IgG1 in offspring but had no significant effect on airway reactivity, BALF cytokine or cell concentrations, although there were non-significant increases in lung neut rophils and eosinophils. Significant increases in total serum IgE were observed, but this effect persisted after filtration of particles and was thus attributed to gas phase components.

December 2009 7-39

### Woodsmoke

In a study by Tesfaigzi et al. (2005, 156116 ), Brown Norway rats were sensitized and challenged with OVA. Rats were exposed for 70 days to filtered air or to 1,000 μg/m 3 hardwood smoke. Particles were characterized by a MMAD of 0.36 μm. Concentrations of gases were reported to be 13.0 ppm CO and 3.1 ppm total vapor hydrocarbon with negligible NO X . Respiratory function was measured in anesthetized animals by w hole-body plethysmography and demonstrated a significant increase in functional residual capacity as well as a significant increase in dynamic lung compliance in hardwood smoke-exposed animals co mpared to controls. No change in total pulmonary resistance or airway responsiveness to methacholine was observed. BALF inflammatory cells were not increased, although histological analysis demonstrated focal inflammation including granulomatous lesion and eosinophilic infiltrations in hardwood smoke-exposed rats. Alterations of several cytokines in BALF and plasma were noted . Changes in airway epithelial mucus cells and intraepithelial stored mucosubstances were modest and did not achieve statistical significance. Results of this study demonstrate that subchronic exposure to hardwood smoke had minimal effects on pulmonary responses in a rat model of allergen sensitization and challenge.



## 7.3.7. Host Defense

7.3. 7.1. Epi demiologic Studies

Epidemiologic studies of respiratory infections i ndicate an association with PM. This is more evident when considering short-term exposures (Chapt er 6), but studies of long-term exposures have observed associations with general respiratory symptoms often caused by infection, such as bronchitis. In a birth cohort study of approximate ly 4,000 Dutch children, Brauer et al. (2007, 090691 )(described in Section 7.3.1.1 ) found significant positive associations for PM 2.5 with ear/nose/throat infections and doctor-diagnosed flu/serious cold in the first 4 yr of life. These results are consistent with an earlier study by Brauer et al. (2006, 090757 ), which found that an increase of 10 μg/m 3 PM 2.5 was associated with increased risk for ear infections in the Netherlands [OR 1.50 (95% CI, 1.00-2.22)]. A Swiss study by Bayer-Oglesby et al. (2005, 086245 ), discussed in Section 7.3.1.1 above, demonstrated that declining levels of PM 10 were associated with declining prevalence of common cold and conjunctivitis. Beca use traffic-related pollutants such as UFPs are high near major roadways and then decay exponentially over a short distance, Williams, et al. (2009, 191945 ) assessed exposure according to residential pr oximity to major roads in a Seattle area study of postmenopausal women. Proximity to major roads was associated with a 21% decrease in natural killer cell function, which is an important defense against viral infection and tumors. This finding was limited to women who reported exercising near traffic; other markers of inflammation and lymphocyte proliferation did not consistently differ according to proximity to major roads. In the Puget Sound region of Washington, Karr et al. (2009, 191946 ) reported that there may be a modest increased risk of bronchiolitis related to PM 2.5 exposure for infants born just before the peak respiratory syncytial virus (RSV) season. Risk estim ates were stronger when restricted to cases specifically attributed to RSV and for infants residing closer to highways. Emerging evidence suggests that respiratory infections, particularly infection by viruses such as RSV, can cause asthma or trigger asthma attacks.



## 7.3. 7.2. Tox icological Studies

Diesel Exhaust

DE may affect systemic immunity. The prolif erative response of A/J mouse spleen cells following stimulation with T cell mitogens was suppressed by 6 mo of daily exposure to DE at concentrations at or above 300 μg/m 3 PM (Burchiel et al., 2004, 055557 ). B cell proliferation was increased at 300 μg/m 3 but unaffected at higher concentrations (up to 1,000 μg/m 3 ). Concentrations of gases and were reported in the parallel study by Reed et al. (2004, 055625 ), described in

December 2009 7-40Section 7.3.3.2 . The Reed study reported a decrease in spleen weight in male mice (27% reduction in the 300 μg/m 3 exposure group). The immunosuppressive effect s of DE were not due to PAHs or benzo(a)pyrene (BaP)-quinones (BPQs) since there were little, if any, of these compounds present in the chamber atmosphere. It should be noted that sentinel animals were negative for mouse parvovirus at the start of the study, but seroconverted by the end of the study, indicating possible infection. Parvovirus can interfere with the modul ation of lymphocyte mitogenic responses (Baker, 1998, 156245 ). A 6-month exposure (6h/day, 7d/wk) to 30, 100, 300 or 1,000 μg/m 3 of PM in DE did not significantly affect bacterial clearance in C57BL/6 mice infected with Pseudomonas aeruginosa

, although all levels reduced bacterial clearance when the exposure only lasted a week (Harrod et al., 2005, 088144 ). Characterization of the exposure atmosphere was given by Reed et al. (2004, 055625 ) (Section 7.3.3.2 .).



### Gasoline Exhaust

In a study by Reed et al. (2008, 156903 ) (described in Section 6.3.7.2) long-term exposure to fresh gasoline exhaust (6h/day, 7d/wk for 6 mo) did not affect clearance of P. aeruginosa from the lungs of C57BL/6 mice.



### Hardwood Smoke

One study demonstrated immunosuppressive effects of hardwood smoke (Burchiel et al., 2005, 088090 ). Exposure to hardwood smoke increased proliferation of T cells from A/J mice exposed daily to 100 μg/m 3 PM for 6 mo, but produced a concentration-dependent suppression of proliferation at PM concentrations >300 μg/m 3 . No effects on B cell proliferation were observed. Concentrations of NO and NO 2 were not detectable or <40 ppb for all exposure levels. CO was reported to be 2, 4, and 13 ppm for the 100, 300 and 1,000 μg/m 3 PM concentrations, respectively. Exposure atmospheres contained significant levels of naphthalene and methylated napthalenes, fluorene, phenanthrene, and anthracene, as well as low concentrations of several metals (K, Ca, and Fe) (Burchiel et al., 2005, 088090 ). It should be noted that serologic analysis of study sentinel animals indicated infection with parvovirus , which can interfere with the modulation of lymphocyte mitogenic responses (Baker, 1998, 156245 ). In another study by Reed et al. (2006, 156043 ) C57BL/6 mice were exposed to 30-1,000 μg/m 3 hardwood smoke by whole-body inhalation for 6 mo prior to instillation of P. aeruginosa . Exposure characterizations are described in Section 7.3.3.2

. Although there was a trend toward increased clear ance with increasing exposure concentrations, there was no statistically significant effect of hardwood smoke exposure on bacterial clearance.



## 7.3.8. Respiratory Mortality

Two large U.S. cohort studies examined the effect of long-term exposure to PM 2.5 on respiratory mortality with mixed results. In the ACS study, Pope et al. (2004, 055880 ) reported positive associations with deaths from specific cardiovascular diseases, but no PM 2.5 associations were found with respiratory mortality. A follow-up to the Harvard Six Cities study (Laden et al., 2006, 087605 ) used updated air pollution and mortality da ta and found positive associations between long-term exposure to PM 2.5 and mortality. Of special note is a statistically significant reduction in mortality risk reported with reduced long-term fine particle concentrations observed for deaths due to cardiovascular and respiratory causes, but not fo r lung cancer deaths. There is some evidence for an association between PM 2.5 and respiratory mortality among post-neonatal infants (ages 1 month-1 year) (Section 7.4.1 ). In summary, when deaths due to respi ratory causes are separated from allcause (nonaccidental) and cardiopulmonary deaths, th ere is limited and inconsistent evidence for an effect of PM 2.5 on respiratory mortality, with one large cohort study finding a reduction in deaths due to respiratory causes associated with reduced PM 2.5 concentrations, and another large cohort study finding no PM 2.5 associations with respiratory mortality.

December 2009 7-41

## 7.3.9. Summary and Causal Determinations

7.3. 9.1. PM 2 .5 The epidemiologic studies reviewed in the 2004 PM AQCD suggested relationships between long-term PM 10 and PM 2.5 (or PM 2.1 ) exposures and increased incidence of respiratory symptoms and disease. One of these studies indicated associations with bronchitis in the 24-city cohort (Dockery et al., 1996, 046219 ). They also suggested relationships between long-term exposure to PM 2.5 and pulmonary function decrements in the CHS (Gauderman et al., 2000, 012531 ; Gauderman et al., 2002, 026013 ). These findings added to the database of the earlier 22-city study of PM 2.1 (Raizenne et al., 1996, 077268 ) that found an association between exposure to ambient particle strong acidity and impairment of lung function in children. No long-term exposure toxicological studies were reported in the 2004 PM AQCD. Recent studies have greatly expanded the evidence available since the 2004 PM AQCD. New analyses have been conducted that include longerfollow-up periods of the CHS cohort through 18 yr of age and provide evidence that effects from exposure to PM 2.5 persist into early adulthood. Gauderman et al. (2004, 056569 ) reported that PM 2.5 exposure was associated with clinically and statistically significant deficits in FEV 1 attained at the age of 18 yr. In addition, the strength and robustness of the outcomes were larger in magnitud e, and more precise than previous CHS studies with shorter follow-up periods. Supporting this re sult are new longitudinal cohort studies conducted by other researchers in other locations with different methods. These studies report results for PM 10 that is dominated by PM 2.5 . New studies provide positive a ssociations from Mexico City, Sweden, and a national cystic fibrosis cohort in the U.S. A natural experiment in Switzerland, where PM levels had decreased, reported that improvement in air quality may slow the annual rate of decline in lung function in adulthood, indicating positive cons equences for public health. Thus, the data are consistent and coherent across several study designs, locations and researchers. As was found in the 2004 PM AQCD, the studies report associations with PM 2.5 and PM 10 , while most did not evaluate
PM 10-2.5

. Associations have been reported with fine par ticle components, particularly EC and OC. Source apportionment methods generally have not b een used in these long-term exposure studies. Coherence and biological plausibility for th e observed associations with lung function decrements is provided by toxicological studies (Section 7.3.2.2 ). A recent study demonstrated that pre- and postnatal exposure to ambient levels of urban particles affected mouse lung development, as measured by anatomical and functional indices (Mauad et al., 2008, 156743 ). Another study suggested that the developing lung may be susceptible to PM since acute exposure to UF iron-soot decreased cell proliferation in the proximal alve olar region of neonatal rats (Pinkerton et al., 2004, 087465 ) (Section 6.3.5.3). Impaired lung developmen t is a viable mechanism by which PM may reduce lung function growth in children. Other an imal toxicological studies have demonstrated alterations in pulmonary function following exposure to DE and wood smoke (Section 7.3.2.2 ). An expanded body of epidemiologic evidence for the effect of PM 2.5 on respiratory symptoms and asthma incidence now includes prospective cohort studies conducted by different researchers in different locations, both within and outside the U.S. with different methods. The CHS provides evidence in a prospective longitudinal cohort study that relates PM 2.5 and bronchitic symptoms and reports larger associations for within-community ef fects that are less subject to confounding than between-community effects (McConnell et al., 2003, 049490 ). Several new studies report similar findings with long-term exposure to PM 10 in areas where fine particles predominate. In England, an association was seen with an increased prevalen ce of cough without a cold. Further evidence includes a reduction of respiratory symptoms corresponding to decreasing PM levels in natural experiments in cohorts of Swiss school children (Bayer-Oglesby et al., 2005, 086245 ) and adults (Schindler et al., 2009, 191950 ). New studies examined the relationship between long-term PM 2.5 exposure and asthma incidence. PM 2.5 had the strongest modifying effect on the association between lung function with asthma in an analysis of the CHS (Islam et al., 2007, 090697 ). The loss of protection by high lung function against new onset asthma in high PM 2.5 communities was observed for all the lung function measures. In the Netherlands, an association with doctor-diagnosed asthma was found in a birth cohort examining the first 4 yr of life (Brauer et al., 2007, 090691 ) Further, findings from an adult cohort suggest that traffic-related PM 10 contributes to asthma development and that reductions in PM decrease asthma risk (Kunzli et al., 2009, 191949 ).

December 2009 7-42A large proportion of asthma is driven by alle rgy, and the majority of recent epidemiologic studies examining allergic (or atopic) indicators found positive associations with PM 2.5 or PM 10 (Section 7.3.6.1 ) . Limited evidence for PM-mediated allergic responses is provided by toxicological studies of DE and woodsmoke, while effects of gasoline exhaust were attributed to gaseous components (Section 7.3.6.2 ). Long-term PM 2.5 exposure is associated with pulmonary inflammation and oxidative responses. An epidemiologic study found a relationship between PM 2.5 and increased inflammatory marker eNO among school children (Dales et al., 2008, 156378 ). Toxicological studies of pulmonary inflammation have demonstrated mixed results, with subchronic DE exposures generating increases and CAPs and wood smoke inducing little or no response (Section 7.3.3.2 ). The pulmonary inflammation observed with DE was attributable to the particle fraction. Toxicological studies also reported evidence of oxidative responses (Section 7.3.4.1 ). Adaptation to prolonged DE was observed for some oxidative responses in addition to some allergic and pulmonary function responses (Section 7.3.2.2 and 7.3.6.2 ). Additional support for the relationship between long-term PM 2.5 exposures and respiratory outcomes is provided by pulmonary injury responses observed in toxicological studies (Section 7.3.5.1 ). Markers of pulmonary injury were increased in rats exposed to DE and gasoline exhaust; and these changes were attributable to PM. Fu rther, lung DNA methylation was observed in the gasoline exhaust study. Histopathological changes have also been reported following exposure to heavily-trafficked urban air and woodsmoke. Findings include nasal and airway mucous cell hyperplasia accompanied by alterations in mucus production which can lead to a loss of mucus– mediated protective functions; exacerbation of protease-induced emphysema; and mast cell infiltration and hypertrophy of alveolar walls. Th ese results provide biological plausibility for adverse respiratory outcomes following long-term PM exposure. Limited information is available on host defense responses (Section 7.3.7 ) and respiratory mortality (Section 7.3.8 ) resulting from PM 2.5 exposure. Several recent epidemiologic studies suggest a relationship between long-term exposure to PM 2.5 or PM 10 and infection in children and infants (Section 7.3.7.1 ). A few toxicological studies suggest that DE exposure affects systemic immunity, and although impaired bacterial clearance is associated with short-term exposures to DE, neither DE or gasoline exhaust seems to have this effect after longer exposures (Section 7.3.7.2 ). In summary, the strongest evidence for a relationship between long-term exposure to PM 2.5 and respiratory morbidity is provided by epidem iologic studies demonstrating associations with decrements in lung function growth in children and with respiratory symptoms and disease incidence in adults. Mean PM 2.5 concentrations in these study locations ranged from 13.8 to 30 μ g/m 3 during the study periods. These studies provide evidence for associations in areas where PM is predominantly fine particles. A major challenge to in terpreting the results of these studies is that the PM size fractions and concentrations of other ai r pollutants are often correlated; however, the consistency of findings across different locations supports an independent effect of PM 2.5 . Recent toxicological studies provide support for the associations with PM 2.5 and decreases in lung function growth in children. Pre- and postnatal exposure to ambient levels of urban particles was found to affect mouse lung development, which provides biological plausibility for the epidemiologic findings. Recent subchronic and chronic toxicologi cal studies also demonstrate altered pulmonary function, mild inflammation, oxidative responses, histopathological changes including mucus cell hyperplasia and enhanced allergic responses in re sponse to CAPs, DE, urban air and woodsmoke and provide further coherence and biological plausi bility. Exacerbation of emphysematous lesions was noted in one study involving exposure to urban air in a heavily-trafficked area. 

### Collectively, the evidence is sufficient to conclude that the relationship between long-term PM 2.5 

### exposure and respiratory effects is likely to be causal.



## 7.3. 9.2. PM 1 0-2.5 The 2004 PM AQCD did not report long-term exposure studies for PM 10-2.5 . The only recent study to evaluate long-term exposure to PM 10-2.5 found positive, but not statistically significant associations with eNO (Dales et al., 2008, 156378 ). The evidence is 

### inadequate to determine if a causal relationship exists between long-term PM 10-2.5 

### exposures and respiratory effects

.

December 2009 7-43

## 7.3. 9.3. UFP s

The 2004 PM AQCD did not report long-term exposure studies for UFPs. The current evidence for long-term UFP effects is limited to toxicological studie s. Generally, subchronic exposure to DE induced pulmonary inflammation, which was in contrast to UF CAPs exposure (Section 7.3.3.2 ) It appeared that the PM fraction was responsible for the inflammatory response with DE exposure. Long-term exposure to DE also resulted in oxidative and allergic responses, although lung injury was not remarkable (Sections 7.3.4.1 and 7.3.6.2 ). The evidence is 

### inadequate to determine if a causal relationship exists between long-term UFP exposures and respiratory effects

.



# 7.4. Reproductive, Developmental, Prenatal and Neonatal Outcomes

7.4.1. Epidemiologic Studies

This section evaluates and summarizes the scientific evidence on PM and developmental and pregnancy outcomes and infant mortality. Infants and fetal development processes may be particularly vulnerable to PM exposure, and although the physical mechanisms are not fully understood, several hypotheses have been proposed involving direct effects on fetal health, altered placenta function, or indirect effects on the mother’s health (Bracken et al., 2003, 156288 ; Clifton et al., 2001, 156360 ; Maisonet et al., 2004, 156725 ; Schatz et al., 1990, 156073 ; Sram et al., 2005, 087442 ). Study of these outcomes can be difficult gi ven the need for detailed data and potential residential movement of mothers during pregna ncy. Two recent articles have reviewed methodological issues relating to the study of outdoor air pollution and adverse birth outcomes (Ritz and Wilhelm, 2008, 156914 ; Slama et al., 2008, 156985 ). Some of the key challenges to interpretation of these study results include the diff iculty in assessing exposure as most studies use existing monitoring networks to estimate individual exposure to ambient PM; the inability to control for potential confounders such as other risk factors that affect birth outcomes (e.g., smoking); evaluating the exposure window (e.g., trimester) of importance; and limited evidence on the physiological mechanism of these effects (Ritz and Wilhelm, 2008, 156914 ; Slama et al., 2008, 156985 ). Another uncertainty is whether PM effects differ by the child’s sex. A review of preterm birth and low birth weight studies found limited indication that effects may differ by gender, however sample size was limited (Ghosh et al., 2007, 091233 ). Previous summaries of the association between PM concentrations and pregnancy outcomes and infant mortality were presented in previous PM AQCDs. The 1996 PM AQCD concluded that although few studies had been conducted on the link between PM and infant mortality, the research “suggested an association,” particular ly for post-neonates (U.S. EPA, 1996, 079380 ). In the 2004 PM AQCD, additional evidence was available on PM’s effect on fetal and early postnatal development and mortality (U.S. EPA, 2004, 056905 ) and although some studies indicated a relationship between PM and pregnancy outcomes, others did not. Studies identifying associations found that exposure to
PM 10 early during pregnancy (first month of pregnancy) orlate in the pregnancy (6 wk prior to birth) were linked with higher risk of preterm birth, includi ng models adjusted for other pollutants, and that
PM 2.5 during the first month of pregnancy was associ ated with intrauterine growth restriction. However, other work did not identify relationships between PM 10 exposure and low birth weight. The state of the science at that time, as indi cated in the 2004 PM AQCD, was that the research provided mixed results based on studies from multip le countries, and that additional research was required to better understand the impact of PM on pregnancy outcomes and infant mortality. Considering evidence from recent studies discussed below, along with previous AQCD conclusions, epidemiologic studies consistently report associations between PM 10 and PM 2.5 exposure and low birth weight and infant mortality, especially dur ing the post-neonatal period. Animal toxicological evidence supports these associations with PM 2.5 , but provides little mechanistic information or biological plausibility. Information on the ambient concentrations of PM 10 and PM 2.5 in these study locations can be found in Table 7-5

.

December 2009 7-44

## 7.4. 1.1. LowBirth Weight

A large number of studies have investigated exposure to ambient PM and low birth weight at term, including a U.S. national study, as well as two studies in the northeast U.S., and four in California. Parker and Woodruff (2008, 156846 ) linked U.S. birth records for singletons delivered at 40-wk gestation in 2001-2003 during the months of March, June, September and December to quarterly estimates of PM exposure by county of residence and month of birth. They found an association between PM 10-2.5 and birthweight (-13 g [95% CI: -18.3 to -7.6]) per 10 μ g/m 3 increase), but no such association for PM 2.5

. Maisonet et al. (2001, 016624 ) analyzed 89,557 births (1994-96 ) in six northeastern cities (Boston and Springfield MA; Hartford CT; Philadelphia and Pittsburgh PA; and Washington DC). Each city had three PM 10 monitors measuring every sixth day. Results from multiple monitors were averaged in each city. Exposure was determined for each trimester of pregnancy and categorized by quartiles (<25, 25-30, 31-35, 36-43 μ g/m 3 ) and 95th percentile (>43 μ g/m 3 ). There was no increased risk for low birth weight at term associated with PM 10 exposure during any trimester of pregnancy. When birth weight was considered as a continuous outcome, exposure to PM 10 was not associated with a reduction in mean birth weight. In contrast, Bell et al. (2007, 093256 ) reported positive associations for both PM 2.5 and PM 10 with birth weight in a study of births (n = 358, 504) in Connecticut and Massachusetts (1999-2002). Birth data indicated county, not street address or ZIP code, so women were assigned exposure based on county residence at delivery. The difference in birth weight per 10 μ g/m 3 associated with PM 2.5 was -66.8 (95% CI: -77.7 to -55.9) g. For PM 10 it was -11.1 (95% CI: -15.0 to -7.2) g. The increased risk for low birth weight was OR = 1.054 (95% CI: 1.022-1.087) for PM 2.5 and OR = 1.027 (95% CI: 0.991-1.064) for PM 10 , based on average exposure during pregnancy. Reductions in birth weight were also associated with third trimester exposure to PM 10 and second and third trimester exposure
to PM 2.5 . Comparing this study to Maisonet et al. (2001, 016624 ), a larger sample size was able to detect a small increase in risk. In addition, birth weight was reduced more by exposure to PM 2.5 than by exposure to PM 10 . Measured PM 2.5 concentrations were not available in the earlier study. The Children’s Health Study is a population based cohort of children living in 12 southern California communities, selected on the basis of diff ering levels of air pollution (Salam et al., 2005, 087885 ), as previously discussed in Section 7.3 . The children in grades 4, 7 and 10 were recruited through schools. A subset of this cohort (n = 6,259) were born in California from 1975-1987. Of these, birth certificates were located for 4,842, including 3,901 infants born at term and 72 cases of low birth weight at term. Using the mother’s ZIP code at the time of birth, exposure was determined by inverse distance weighting of up to three PM 10 monitors within 50 km of the ZIP code centroid. If there was a PM 10 monitor within 5 km of the ZIP code centroid (40% of data), exposure from that monitor was used. Exposure was calculated for the entire pregnancy, and for each trimester of pregnancy. A 10 μ g/m 3 increase in PM 10 during the third trimester reduced mean birth weight -10.9 g (95% CI: -21.1 to -0.6) in single pollutant models , but became non-significant in copollutant models controlling for the effects of O 3 . Increased risks of low birth weight (<2,500 g) were not statistically significant (OR = 1.3 [95% CI: 0.9-1.9]). A strength of this study was the cohort data available included information on SES and smoking during pr egnancy. A limitation is the assignment of exposure based on monitoring stations up to 50 km distant; this may have introduced substantial exposure misclassification obscuring some associations.

December 2009 7-45

### Table 7-5. Characterization of ambient PM concentrations from studies of reproductive, developmental, prenatal and neonata l outcomes and l ong-term exposure.

Study Location Mean Annual Concentration (μg/m 3 ) Upper Percentile Concentrations (μg/m 3 )

PM 2.5

Basu et al. (2004, 087896 ) CA Range of means across sites: 14.5-18.2

Avg of means across sites: 16.2 Max: 26.3-34.1

Bell et al. (2007, 091059 ) CT & MA 22.3

Brauer et al. (2008, 156292 ) Vancouver, Canada 5.3 Max: 37.0

Huynh et al. (2006, 091240 ) CA Range of means across trimesters: 17.5-18.8

Avg of means across trimesters: 18.2

Jalaludin et al. (2007, 156601 ) Sydney, Australia 9.0

Liu (2007, 090429 ) Multicity, Canada 12.2 75th: 15

Loomis et al. (1999, 087288 ) Mexico City 27.4 Max: 85

Mannes et al. (2005, 087895 ) Sydney, Australia 9.4 75th: 11.2; Max: 82.1

Parker et al. (2005, 087462 ) CA 15.4

Ritz et al. (2007, 096146 ) Los Angeles, CA 20.0

Wilhelm and Ritz (2005, 088668 ) Los Angeles, CA 21.0 Max: 38.9-48.5

Woodruff et al. (2006, 088758 ) CA 19.2 a 75th: 22.7

Woodruff et al. (2008, 098386 ) U.S. Range of means across effects: 14.5-14.9 a

Avg of means across effects: 14.8 a 75th: 18.5-18.7

PM 10-2.5

Parker et al. (2008, 156013 ) U.S. 13.2 75th: 17.5

PM 10

Bell et al. (2007, 093256 ) CT & MA 22.3

Brauer et al. (2008, 156292 ) Vancouver, Canada 12.7 Max: 35.4

Chen et al. (2002, 024945 )` Washoe County, NV 31.53 75th: 39.35; Max: 157.32

Gilboa et al. (2005, 087892 ) TX 23.8 a 75th: 29

Ha et al. (2003, 042552 ) Seoul, South Korea 69.2 75th: 87.7; Max: 245.4

Hansen et al. (2006, 089818 ) Brisbane, Australia 19.6 Max: 171.7

Hansen et al. (2007, 090703 ) Brisbane, Australia 19.6 75th: 22.7; Max: 171.7

Jalaludin et al. (2007, 156601 ) Sydney, Australia 16.3

Kim et al. (2007, 156642 ) Seoul, Korea Range of means across time: 88.7-89.7

Avg of means across time: 89.2

Lee et al. (2003, 043202 ) Seoul, Korea 71.1 75th: 89.3; Max: 236.9

Leem et al. (2006, 089828 ) Incheon, Korea 53.8 a 75th: 64.6; Max: 106.39

Lipfert et al. (2000, 004103 ) U.S. 33.1 Max: 59

Maisonet et al. (2001, 016624 ) NE U.S. 31.0 a 75th: 36.1; Max: 46.5

Mannes et al. (2005, 087895 ) Sydney, Australia 16.8 75th: 19.9; Max: 104.0

Pereira et al. (1998, 007264 ) Sao Paulo, Brazil 65.04 Max: 192.8

Ritz et al. (2000, 012068 ) CA 49.3 Max: 178.8

Ritz et al. (2006, 089819 ) CA 46.3 Max: 83.5

December 2009 7-46Study Location Mean Annual Concentration (μg/m 3 ) Upper Percentile Concentrations (μg/m 3 )

Rogers and Dunlop (2006, 091232 ) GA 3.75 75th: 15.07

Romieu et al. (2004, 093074 ) Ciudad Juarez, Mexico 33.0-45.9

Sagiv et al. (2005, 087468 ) PA Range of means across time: 25.3-27.1

Avg of means across time: 26.2 Max: 68.9-156.3

Salam et al. (2005, 087885 ) CA Range of means across trimesters: 45.4-46.6

Avg of means across trimesters: 45.8

Suh et al. (2008, 192077 ) Seoul, Korea Range of means across trimesters: 54.6-61.1

Avg of means across trimesters: 58.27 75th: 62.8-67.8

Max: 85.1-107.36

Tsai et al. (2006, 090709 ) Kaohsiung, Taiwan 81.5 75th: 111.5; Max: 232.0

Wilhelm and Ritz (2005, 088668 ) Los Angeles, CA 38.1 Max: 74.6-103.7

Woodruff et al. (2008, 098386 ) U.S. Range of means across effects: 28.6-29.8 a

Avg of means across effects: 29.1 a 75th: 33.8-36.5

Yang et al. (2006, 090760 ) Taipei, Taiwan 53.2 75th: 64.9; Max: 234.9

a Median concentration

Parker et al. (2005, 087462 ) examined births in California within 5 miles of a monitoring station (n = 18,247). Only infants born at 40 wk ge station were included. Thus all infants were the same gestational age, and had been exposed in the same year. Exposure to PM 2.5 in quartiles (<11.9, 11.9-13.9, 14.0-18.4, >18.4) was associated with decrements in birth weight. Infants exposed to >13.9 μ g/m 3 experienced reductions in birth weight (thirdquartile -13.7 g (95% CI: -34.2 to 6.9), fourth quartile -36.1 g (95% CI: - 55.8 to -16.5). These are larger re ductions than have been seen in some other studies. However, this study reduced misclassification by including only women living within 5 miles of a monitoring station, and on ly included births at 40 wk gestation. Reducing misclassification should lead to a stronger a ssociation, if the association is causal. The effects of spatial variation in exposure were also investigated by Wilhelm and Ritz (2005, 088668 ). Their study included all women living in ZIP codes where 60% of the ZIP code was within two miles of a monitoring station in the Southern California Basin, and women with known addresses in Los Angeles County within 4 miles of a monitoring station. Exposure to average PM 10 in the third trimester was analyzed for incr eased risk of low birth weight at term ( ≥ 37-wk gestation). Analysis at the ZIP code level did not detect increased risk (per 10 μ g/m 3 PM 10, OR = 1.03 [95% CI: 0.97-1.09]). However the analysis based on geocoded addresses indicated that increasing exposure to
PM 10 was associated with increased risk of low birth weight for women living within 1 mile of the station where PM 10 was measured. For these women (n = 247 cases, 10,981 non-cases), each 10 μ g/m 3 increase in PM 10 was associated with a 22% increase in risk of term low birth weight (OR = 1.22 [95% CI: 1.05-1.41]). In the categorical analysis, exposure to PM 10 >44.4 μ g/m 3 was associated with a 48% increase in risk (OR = 1.48 [95% CI: 1.00-2.19]). Increased risk of low birth weight also was associated with exposure to CO in single pollutant models. However, when multipollutant models were considered, the effect s of CO were attenuated but the effects of PM 10 increased. Controlling for CO, NO 2 , and O 3 , each 10 μ g/m 3 increase in exposure to PM 10 increased risk of low birth weight 36% (OR = 1.36 [95% CI: 1.12-1.65]). Spatial variation in PM 2.5 exposure was investigated by Basu et al. (2004, 087896 ). They included only mothers who lived within 5 miles of a PM 2.5 monitor and within a California county with at least 1 monitor. To minimize potential c onfounding, they included only white (n = 8,597) or Hispanic (n = 8,114) women, who were married, be tween 20 and 30 yr of age, completed at least high school and were having their first child. Consistently, PM 2.5 exposure measured by the county monitor was more strongly associated with reductions in birth weight than exposure measured by the neighborhood monitor. The results were replicated in both the white and the Hispanic samples. Reductions in birth weight ranged from 15.2 to 43.5 g per 10 μ g/m 3 increase in PM

. 2.5 In the remaining U.S. study, Chen et al. (2002, 024945 ) analyzed 33,859 birth certificates of residents of Washoe County in northern Nevada (1991-1999). There were four sites monitoring PM 10 during the study period, it appears (not stated) that exposure was averaged over the county. A

December 2009 7-4710 μ g/m 3 increase in exposure to PM 10 during the third trimester of pregnancy was associated with an 11 g reduction in birth weight (95% CI: -2.3 to -19.8). Effects on risk of low birth weight were not statistically significant. For exposure in the third trimester of 19.77 to 44.74 μ g/m 3 compared to <19.74 μ g/m 3 the odds ratio for low birth weight was 1.05 (95% CI: 0.81-1.36). Comparing exposure >44.74 to the same reference category, th e odds ratio was 1.10 (95% CI: 0.71-1.71). Misclassification of exposure may have occu rred when exposure was averaged over a large geographic area (16,968 km 2 ). Recent international studies investigating effect s of particles on low birth weight include one in Munich (Slama et al., 2007, 093216 ), two in Canada (Brauer et al., 2008, 156292 ; Dugandzic et al., 2006, 088681 ), two in Australia (Hansen et al., 2007, 090703 ; Mannes et al., 2005, 087895 ), two in Taiwan (Lin et al., 2004, 089827 ; Yang et al., 2003, 087886 ) one in Korea (Ha et al., 2003, 042552 ) and two in Sao Paulo, Brazil (Gouveia et al., 2004, 055613 ; Medeiros and Gouveia, 2005, 089824 ). The majority of these studies found that PM concentrations were associated with low birth weight, though two studies (Hansen et al., 2007, 090703 ; Lin et al., 2004, 089827 ) found no associations. The effect estimates were similar in magnitude to those reported in the U.S. studies.



### Considerations in Interpreting Results of Low Birth Weight Studies

Studies included subjects at distances from monitoring stations varying from as close as 1 mile or 2 km, to as far as 50 km or the size of the county. Studies that only included subjects living within a short distance (1 mile, 2 km) of the monitoring station (thus likely reducing exposure measurement error) were more likely to find that PM exposure was associated with increased risk of low birth weight. However, Basu et al. (2004, 087896 ) reported a stronger association between PM 2.5 exposure and birth weight when exposure was estimated based on the county monitor, rather than the monitor within 5 miles of the residence. They suggest th at county level exposure may be more representative of where women spend their time, including not only home, but also other time spent away from home. Other pollutants also appeared to influence th e risk associated with particle exposure. In one study, exposure to PM 10 in a single pollutant model reduced birth weight by 11 g, but became nonsignificant in copollutant models with O 3 (Salam et al., 2005, 087885 ). In another study the risk associated with PM 10 exposure increased from 22% to 36% when other pollutants were included in the model (Wilhelm and Ritz, 2005, 088668 ). All but one study in the U.S. found some association between particle exposure and reduced birth weight (Maisonet et al., 2001, 016624 ). The results of international studies were inconsistent. This mi ght be related to the chemical composition of particles in the U.S., or to differences in the po llutant mixture. Studies with null results must be interpreted with caution when the comparison grou ps have significant exposure. This was certainly the situation in studies in Taiwan and Korea (Lee et al., 2003, 043202 ; Lin et al., 2004, 089827 ; Yang et al., 2003, 087886 ). Differences in geographical locati ons, study samples and linkage decisions may contribute to the diverse findings in the literature on the association between PM and birthweight, even within the U. S. (Parker and Woodruff, 2008, 156846 ).



## 7.4. 1.2. Pre term Birth

A potential association of exposure to airborne particles and preterm birth has been investigated in numerous epidemiologic studies, including some conducted in the U.S. and others in foreign countries. Three U.S. studies have been carried out by the same group of investigators in California. A natural experiment occurred when an open-hear th steel mill in Utah Valley was closed from August 1986 through September 1987. Parker et al. (2008, 156013 ) compared birth outcomes for Utah mothers within and outside of the Utah Valle y, before, during, and af ter the mill closure. They report that mothers who were pregnant around the time of the closure of the mill were less likely to deliver prematurely than mothers who were pregnant before or after. The strongest effect estimates were observed for exposure during the second trim ester (14% decrease in risk of preterm birth during mill closure). Preterm birth outside of the Ut ah Valley did not change during the time of the mill closure. In 2000, Ritz et al. (2000, 012068 ) published the first study investigating the association of preterm birth with PM in the U.S. The study population was women living in the southern California Basin. There were eight monitoring stations measuring PM 10 every 6th day during the study period.

December 2009 7-48Birth certificates (1989-1993) were analyzed for women living in ZIP codes within 2 miles of a monitoring station. Women with multiple gestations, chronic disease prior to pregnancy and women who delivered by cesarean section were excluded resulting in a study population of 48,904 women. The risk of preterm birth increased by 4% (RR = 1.04 [95% CI: 1.02-1.6]) per 10 μ g/m 3 increase in
PM 10 averaged in the 6 wk before birth. Exposure to PM 10 in the first month of pregnancy resulted in a 3% increase in risk (RR = 1.03 [95% CI: 1.01-1.05 ]). These results were robust in multipollutant models. Wilhelm and Ritz (2005, 088668 ) reinvestigated this association among women in the same area in 2005, when air pollution had declined from a mean level near 50 μ g/m 3 to a mean level near 40 μ g/m 3 . Birth certificate data from 1994-2000 was anal yzed for women living in ZIP codes within 2 miles of a monitoring station, or with addre sses within 5 miles of a monitoring station. No significant effects of exposure to PM 10 were reported. Exposure to PM 2.5 6 wk before birth resulted in an increase in preterm birth (RR = 1.19 [95% CI: 1.02-1.40]) for the highest quartile of exposure (PM 2.5 >24.3 μ g/m 3 ). Using a continuous measure of PM 2.5 , there was a 10% increase in risk for each 10 μ g/m 3 increase in PM 2.5 (RR = 1.10 [95% CI: 1.00-1.21]). There have been two major criticisms of air pollu tion studies using birth certificate data. First, that birth certificates only indicate the addre ss at birth and the exposure of women who moved during pregnancy may be misclassified; second, th at information about some important confounders may not be available (e.g., smoking). To obtain more precise information about these variables, Ritz et al. (2007, 096146 ) conducted a case-control study nested within a cohort of birth certificates (Jan 2003-Dec 2003) in Los Angeles County. Births to wo men residing in ZIP codes (n = 24) close to monitoring stations or major population centers or roadways (n = 87) were eligible (n = 58,316 births). All cases of low birth weight or preterm birth and an equal number of randomly sampled controls in the 24 ZIP codes close to monitors were selected. In the other 87 ZIP codes, 30% of cases and an equal number of controls were randomly sampled. Of 6,374 women selected for the case control study, 2,543 (40%) were interviewed. The association of preterm birth with exposure to
PM 2.5 differed between women responding to the survey and women who did not respond. Among responders, exposure to each 10 μ g/m 3 increase in PM 2.5 concentration in the first trimester increased risk to preterm birth by 23% (RR = 1.23 [95% CI: 1.02-1.48]). There was no increase in risk among non-responders (RR = 0.95 [95% CI: 0.82-1.10]), or inthe entire birth cohort (RR = 1.00 [95% CI: 0.94-1.07]). An additional case control study of preterm birth and PM 2.5 exposure (Huynh et al., 2006, 091240 ) used California birth certificate data. Singleton preterm infants (24-36-wk gestation) born in California (1999-2000) whose mother s lived within 5 miles of a PM 2.5 monitor were eligible. Each of these 10,673 preterm infants were matched to three te rm (39- to 44-wk gestation) controls (having a last menstrual period within 2 wk of the case infant), resulting in a study population of 42,692. Controlling for maternal race/ethni city, education, marital status, parity and CO exposure, exposure
to PM 2.5 >17.7 μ g/m 3 increased the risk of preterm birth by 14% (OR = 1.14 [95% CI: 1.07-1.23]). Averaging PM 2.5 exposure over the first month of pregnancy, the last 2 wk before birth, or the entire pregnancy did not substantially change the risk estimate. Two additional studies of preterm birth and exposure to particulate air pollution have been conducted in the U.S. Each has used a unique methodology. Sagiv et al. (2005, 087468 ) used time series to analyze births in four Pennsylvania co unties between January 1997 and December 2001. In this analysis, exposure to PM 10 is compared to the rate of preterm births each day. Both acute exposure (on the day of birth) and longer term ex posure (average exposure for the preceding 6 wk) were considered in the analysis. An advantage of this analysis is that days, rather than individuals are compared, so confounding by individual risk factors is minimized. For exposure averaged over the 6 wk prior to birth, there was an increase in risk (RR = 1.07 [95% CI: 0.98-1.18]), which persisted for acute exposure with a 2-day lag (RR = 1.10 [9 5% CI: 1.00-1.21]) and 5-day lag (RR = 1.07 95% CI: 0.98-1.18]). Rogers and Dunlop (2006, 091232 ) examined exposure to particles and risk of delivery of an infant weighing less than 1,500 g (all of which were preterm) from 24 counties in Georgia. The study included 69 preterm, small for gestational age (SGA) infants, 59 preterm appropriate for gestational age (AGA) infants and 197 term AGA controls . Exposure was estimated using an environmental transport model that considered PM 10 emissions from 32 geographica lly located industrial point sources, meteorological factors, and geographi c location of the birth home. Exposure was categorized by quartiles. Comparing women who de livered a preterm AGA infant to those who

December 2009 7-49delivered a term AGA infant, exposure to PM 10 >15.07 μ g/m 3 tripled the risk (OR = 3.68 [95% CI: 1.44-9.44]). Brauer et al. (2008, 156292 ) evaluated the impacts of PM 2.5 on preterm birth using spatiotemporal exposure metrics in Vancouver, Canada . The authors found similar results when they used a land-use regression model or inverse distan ce weighting as the exposure metric. For preterm births <37 wk, they reported an OR of 1.06 (95% CI : 1.01-1.11), and for preterm births <35 wk the OR increased to 1.12 (95% CI: 1.02-1.24). There were no consistent trends for early or late gestational period to be more str ongly associated with preterm births. Suh et al. (2008, 192077 ) conducted a study to determine if the effects of exposure to PM 10 during pregnancy on preterm delivery are modified by maternal polymorphisms in metabolic genes. They analyzed the effects of the gene-environment interaction between the GSTM1, GSTT1, CYP1a1-T6235C and -1462V polymorphisms and exposure to PM 10 during pregnancy on preterm birth in a case-control study in Seoul, Korea. PM 10 concentration ≥ 75th percentile alone was significant in the third trimester of pregnancy (OR = 2.33 [95% CI: 1.33-4.80]), but not in the first or second trimester. The risk of preterm delivery conferred by the GSTM1 null genotype was increased, and the highest risk was found during the third trimester of pregnancy (OR = 2.58 [95% CI: 1.34-4.97]). There were no statistical associations with the GSTT1 or CYP1A1 genotypes. When the gene-environment interaction was analyzed, the ri sk for preterm birth was substantially higher for women who carried the GSTM1 null genotype and were exposed to high levels of PM 10 ( ≥ 75th percentile) than for those who carried the GSTM1 positive genotype but were only exposed to low levels of PM 10 (<75th percentile) during the third trim ester of pregnancy (OR = 6.22, 95% CI: 2.14-18.08). In Incheon, Korea, Leem et al. (2006, 089828 ) estimated PM 10 exposure spatially as well as temporally. Exposure was based on 26 monitors a nd kriging was used to determine exposure for 120 dongs (administrative districts, mean area 7.82 km 2 , median area 1.42 km 3 ). The sample included 52,113 births, from 2001-2002. PM 10 was very weakly correlated with other pollutants. Exposure was compared in quartiles for the fi rst and third trimester of pregnancy. In the first trimester, relative risks for the second, third and fourth quartiles were RR = 1.14 (95% CI: 0.97-1.34), RR = 1.07 (95% CI: 0.94-1.37), and RR = 1.24 (95% CI: 1.09-1.41), respectively. Exposure to PM 10 in quartile one (reference group) was 26.9-45.9 μ g/m 3 ; fourth quartile exposure equaled 64.6-106.4 μ g/m 3 . The p-value for trend was 0.02. Exposure in the third tr imester was not related to preterm birth, however no information was provided to determine how ex posure in the third trimester was adjusted for women who delivered preterm. Two studies investigating risks of preterm birth related to particle exposure have been reported from Australia. In Brisbane, Hansen et al. (2006, 089818 ) studied 28,200 births (2000-2003) in an area of low PM 10 concentrations. Exposure to an interquartile range increase in PM 10 exposure in the first trimester resulted in a 15% increased risk of preterm birth (OR = 1.15 [95% CI: 1.06-1.25]). This result was strongly influenced by the effect of PM 10 exposure in the first month of pregnancy (OR = 1.19 [95% CI: 1.13-1.26]). PM 10 was correlated with O 3 (r = 0.77) in this study and O 3 also increased risk in the first trimester. No effects were associated with exposure to PM 10 in the third trimester. In Sydney, associations between exposure to pa rticles and preterm birth varied by season. Jalaludin et al. (2007, 156601 ) obtained information on all births in metropolitan Sydney (1998-2000). Exposure to PM 2.5 in the 3 mo preceding birth was associated with an increased risk of preterm birth (OR = 1.11 [95% CI: 1.04-1.19]). Additional effects were dependent on season of conception. Both PM 10 (OR = 1.3 [95% CI: 1.2-1.5]) and PM 2.5 (OR = 1.4 [95% CI: 1.3-1.6]) were associated with increased risk for conceptions in the winter. Conceptions in summer were associated with reductions in risk (PM 10 OR = 0.91 [95% CI: 0.88-0.93]) (PM 2.5 OR = 0.87 [95% CI: 0.84-0.92]). Due to both positive and negative fi ndings, the authors recommend caution in interpreting their results.



### Considerations in Analyzing Environmental Exposures and Preterm Birth

A major issue in studying environmental exposures and preterm birth is selecting the relevant exposure period, since the biological mechanisms lead ing to preterm birth and the critical periods of vulnerability are poorly understood (Bobak, 2000, 011448 ). Exposures proximate to the birth may be most relevant if exposure causes an acute effect . However, exposure occurring in early gestation

December 2009 7-50might affect placentation, with results observable later in pregnancy, or cumulative exposure during pregnancy may be the most important determinant. Th e studies reviewed have dealt with this issue in different ways. Many have considered several exposure metrics based on different periods of

exposure. Often the time periods used are the first month (or first trimester) of pregnancy and the last month (or 6 wk) prior to delivery. Using a time interval prior to delivery introduces an additional problem since cases and controls are not in the same stage of development when they are compared. For example, a preterm infant delivered at 36 wk is a 32-week fetus 4 wk prior to birth, while an infant born at term (40 wk) is a 36-week fetus 4 wk prior to birth. Only one study (Huynh et al., 2006, 091240 ) adjusted for this in the design. Many of these studies compare exposure in quarti les, using the lowest quartile as the reference (or control) group. No studies use a truly unexposed control group. If exposure in the lowest quartile confers risk, than it may be difficult to demonstra te additional risk associated with a higher quartile. Thus negative studies must be interpreted with caution. Preterm birth occurs both naturally ( idiopathic preterm ), and as a result of medical intervention ( iatrogenic preterm ). Ritz et al. (2000, 012068 ; 2007, 096146 ) excluded all births by Cesarean section, to limit their studies to idiopathic preterm. No other studies attempted to distinguish the type of preterm birth, although PM exposure maybe associated with only one type. This is a source of potential effect misclassification.



## 7.4. 1.3. Gro wth Restriction

Low birth weight has often been used as an out come measure because it is easily available and accurately recorded on birth certificates. However, low birth weight may result from either short gestation, or inadequate growth in utero. Most of the studies investigating air pollution exposure and low birth weight, limited their analysis to term infa nts to focus on inadequate growth. A number of studies were identified that specifically addressed growth restriction in utero by identifying infants who failed to meet specific growth standards. Usually these infants had birth weights less than the 10th percentile for gestational age, using an external standard. Many of these studies have been previously discussed, since they also examined ot her reproductive outcomes (low birth weight or preterm delivery). Three studies in the U.S. examined intraute rine growth. A recent study (Rich et al., 2009, 180122 ) investigated very small for gestational age (def ined as a fetal growth ratio <0.75), small for gestational age (defined as ≥ 75 and <85) and “reference” births ( ≥ 85) to women residing in New Jersey and mean air pollutant concentrations durin g the first, second and third trimesters. They reported an increased risk of SGA associated with first and third trimester PM 2.5 concentrations (1.116 [95% CI: 1.012, 1.232], and 1.106 [1.008-1.212], per 10 μg/m 3 PM 2.5 , respectively). Parker et al. (2005, 087462 ) reported a positive association between exposure to PM 2.5 . Since this study only included singleton live births at 40-wk gestation, birth weights less than 2,872 g for girls and 2,986 g for boys were designated SGA, based on births in Calif ornia. Infants exposed to the highest quartile
PM 2.5 (>18.4 μ g/m 3 ) compared to the lowest quartile PM 2.5 (<11.9 μ g/m 3 ) were 23% more likely to be small for gestational age (OR = 1.23 [95% CI: 1.03-1.50]). Very similar results were found for exposure in each of the three trimesters respect ively (OR = 1.26 [95% CI: 1.04-1.51], OR = 1.24 [95% CI: 1.04-1.49], OR = 1.21 [95% CI: 1.02-1. 43]). These results controlled for exposure to CO, which did not increase risk for SGA. In contrast, Salam et al. (2005, 087885 ) found no association between exposure to PM 10 and intrauterine growth retardation (IUGR) in the Cali fornia Children’s Health Study. IUGR was defined as less than the 15th percentile of predicted birth weight based on gestational age and sex in term infants. Apparently no external standard was used since 15% of infants in the study were designated as IUGR. An IQR increase in PM 10 exposure was not significantly associated with IUGR for the whole pregnancy (OR = 1.1 [95% CI: 0.9-1.3]) or for any specific trimester. Differences between this study and the study by Parker et al. (2005, 087462 ) include measurement of PM 10 versus PM 2.5 , a less stringent definition of IUGR, and exposures determined by monitors located much farther away from the subjects’ residences (up to 50 km ver sus within 5 mi). All of these factors could lead to misclassification. Two studies investigating particle exposure and SGA were conducted in Australia, with differing results (Hansen et al., 2007, 090703 ; Mannes et al., 2005, 087895 ). Mannes et al. (2005, 087895 ) defined SGA as birth weight less than two standard deviations below the national mean

December 2009 7-51birth weight for gestational age. In this study ther e was a statistically significant effect of exposure to both PM 10 (OR = 1.10 [95% CI: 1.00-1.48], per 10 μ g/m 3 increase) and PM 2.5 (OR = 1.34 [95% CI: 1.10-1.63], per 10 μ g/m 3 increase) for exposure during the second trimester. When analysis was restricted to births within 5 km of the monitoring station, the association for PM 10 became slightly stronger (OR = 1.22 [95% CI: 1.10-1.34]). Exposure during other trimesters of pregnancy was not associated with IUGR. In Brisbane, Hansen et al. (2007, 090703 ) examined head circumference (HC), crown heel length (CHL) and risk of SGA, defined as less than th e tenth percentile of weight for gestational age and gender based on an Australian national standard. There was no consistent relationship between
PM 10 exposure and SGA, HC or CHL in any trimester of pregnancy. PM 10 exposure was determined by averaging values from the five monitoring stations. Due to the sample size and limited number of monitoring stations, it was not possible to analyze the data for women living within 5 km of a monitoring station, as was done in Sydney. In Canada, Liu et al. (2007, 090429 ) investigated the effect of PM 2.5 exposure on fetal growth in three cities, Calgary, Edmonton and Montreal. IUGR was defined as birth weight below the tenth percentile, by sex and gestational week (37-42) fo r all singleton live births in Canada between 1986 and 2000. Models were adjusted for maternal age, parity, infant sex, season of birth, city of residence, and year of birth. A 10 μ g/m 3 increase in PM 2.5 was associated with an increased risk for IUGR (OR = 1.07 [95% CI: 1.03-1.10]) in the first tr imester, and similar risks were associated with exposure in the second or third trimesters. The effect of PM 2.5 was reduced in multipollutant models including CO and NO 2. Brauer et al. (2008, 156292 ) observed consistent increased risks of SGA for PM 2.5 , PM 10 , NO 2 , NO and CO in Vancouver, Canada (20% increase in risk in PM 2.5 and PM 10 per 10 μ g/m 3 increase). The effects were similar for exposure estimate s based on nearest monitor, inverse distance weighting, and land-use regression modeling. ORs for early or late pregnancy exposure windows were remarkably similar to those fo r the full duration of pregnancy.



## 7.4. 1.4. Bir th Defects

Four recent studies examined PM and birth defects. The Seoul, Korea study discussed above also considered congenital anomalies, defined as a defect in the child’s body structure (Kim et al., 2007, 156642 ). PM 10 levels were associated with higher risk of birth defects for the second trimester, with a 16% (95% CI: 0-34) increase in risk per 10 μg/m 3 in PM 10

. Two U.S. studies examined air pollution and risk of birth defects. Data were collected from the California Birth Defects Monitoring Program for four counties in Southern California (Los Angeles, Riverside, San Bernardino, and Orange) for the period 1987-1993, although each county included a subset of this period (Ritz et al., 2002, 023227 ). Cases (i.e., infants with birth defects) were identified as live birth infants and fetal deaths from 20-wk gestation to 1 yr post-birth, with isolated, multiple, syndrome, or chromosomal cardiac or orofaci al cleft defects. Cases were restricted to those with registry data for gestational age and reside nce ZIP code, and those with residences <10 miles from an air pollution monitor. Six types of categor ies were included: aortic defects; atrium and atrium septum defects; endocrinal and mitral valu e defects; pulmonary artery and valve defects; conotruncal defects; and ventricular septal de fects not part of the conotruncal category. PM 10 measurements were available every 6 days. While re sults indicated increased risk of birth defects for higher levels of CO or O 3 , the authors determined that results for PM 10 were inconclusive, finding no consistent trend of effect after adjustment for CO and O 3

. The other U.S. study examined birth defects through a case-control design in seven Texas counties for the period 1997-2000 (Gilboa et al., 2005, 087892 ). Births were excluded for parents <18 yr and several non-air pollution risk factors known to be associated with birth defects (e.g., maternal diabetes, holoprosencephaly in addition to oral clef). Comparison of the highest ( ≥ 29.0 μg/m 3 ) and lowest (<19.521 μg/m 3 ) quartiles of PM 10 for exposure defined as the third to eighth week of pregnancy generated an OR of 2.27 (95% CI: 1.43-3.60) for risk of isolated atrial septal defects and 1.26 (95% CI: 1.03-1.55) for individual atrial septal defects. Including other pollutants (CO, NO 2 , O 3 , SO 2 ) in the model did not greatly al ter results; numerical results for copollutant analysis were not provided. Strong evidence was not observed for a relationship between
PM 10 and the other birth defect categories. Review articles have concluded that the scientific literature is not sufficient to conclude a relationshi p between air pollution and birth defects (Sram et al., 2005, 087442 ).

December 2009 7-52A recent study of oral clefts conducted in Taiwan found no association between this birth defect and concentrations of PM 10 during the first or second gestational month (Hwang and Jaakkola, 2008, 193794 ). This population-based case-control study in cluded 653 cases and a random sample of 6,530 controls born in Taiwan between 2001 and 2003.



## 7.4. 1.5. Inf ant Mortality

Many studies have identified strong associations be tween exposure to particles and increased risk of mortality in adults or the general populat ion, including for short- and long-term exposure (Sections 6.5 and 7.6 ). Less evidence is available for the potential impact on infant mortality, although studies have been conducted in several co untries. The results of these infant mortality studies are presented here with the other reproductive and developmental outcomes because it is likely that in vitro exposures contribute to th is outcome. Both long-term and short-term exposure studies of infant mortality are included in this s ection. Results on PM and infant mortality includes a range of findings, with some studies finding a ssociations and many statistically non-significant or null effects. Yet, more consistency is observed when results are divided into the type of health outcome based on the age of infant and cause of death. An important question regarding the associati on between PM and infant mortality is the critical window of exposure during development for which infants are susceptible. Several age intervals have been explored: neonatal (<1 mo); infants (<1 yr); and postneonatal (1 mo-1 yr). Within these various age categories, multiple causes of deaths have been investigated, particularly total deaths and respiratory-related deaths. The studie s reflect a variety of study designs, particle size ranges, exposure periods, regions, and adjustment for confounders.



### Stillbirth

Only one study of stillbirths and PM was iden tified. A prospective cohort of pregnant women in Seoul, Korea from 2001 to 2004 was exam ined with respect to exposure to PM 10 (Kim et al., 2007, 156642 ). Gestational age was estimated by the last menstrual period or by ultrasound. Whereas many of the previously discussed studies of PM and pregnancy outcomes were based on national registries, this study examined medical re cords and gathered individual information through interviews on socioeconomic condition, medical hist ory, pregnancy complications, smoking, secondhand smoke exposure, and alcohol use. Mother’s exposure to PM 10 was based on residence for each month of pregnancy, each trimester defined as a three month period, and the 6 wk prior to death. Exposure was assigned by the nearest monitor. A 10 μg/m 3 increase in PM 10 in the third trimester was associated with an 8% (95% CI: 2-14) increase in risk of stillbirth. In São Paulo, Brazil, Poisson regression of stillbirth counts for the period 1991-1992 found that a 10 μg/m 3 increase in PM 10 was associated with a 0.8% increase in stillbirth rates (Pereira et al., 1998, 007264 ). When other pollutants (NO 2 , SO 2 , CO, O 3 ) were included simultaneously in the model, the association did not remain. Stillbirths we re defined as fetal loss at >28 wk of pregnancy age, weight >1,000 g, or length of fetus >35 cm.



### Neonatal Mortality and Neonatal Respiratory Mortality, <1 Month

Studies on PM and neonatal mortality (<1 month) included a time-series analysis of PM 10 for 4 yr of data (1998-2000) for São Paulo, Brazil (Lin et al., 2004, 095787 ). The analysis used daily counts of deaths from government registries and ad justed for temporal tr end, day of the week, weather, and holidays. Findings indicated that a 10 μg/m 3 increase in PM 10 was associated with a 1.71% (95% CI: 0.31-3.32) increase in risk of neonatal death. A case-crossover study of 11 yr (1989-2000) in S outhern California did not find an association between PM 10 and neonatal deaths (Ritz et al., 2006, 089819 ). Quantitative results were not provided. The authors considered adjustment for s eason, county, parity, gender, prenatal care, and maternal age, education, and race/ethnicity. These results add to previous work on PM and neonatal death, including studies identifying higher risk of neonatal mortality with higher TSP in the Czech Republic in an ecological analysis (Bobak and Leon, 1992, 044415 ) and case-crossover study (Bobak and Leon, 1999, 007678 ), and a

December 2009 7-53Poisson model study in Kagoshima City, Japan (Shinkura et al., 1999, 090050 ). An ecological study evaluated U.S. PM 10 data for the year 1990 using long-term pollution levels in 180 U.S. counties (Lipfert et al., 2000, 004103 ). Analysis considered birth weight, se x, month of birth, location by state and county, prenatal care, and mother’s race, ag e, educational level, marital status, and smoking status. County-level variables were included for so cioeconomic status, altitude, and climate. Results indicate a 13.1% increase in neonatal mortality (95% CI: 4.4-22.6) per 10 μg/m 3 PM 10 for non-low birth weight infants. Statistically significant a ssociations were also observed considering all infants or low birth weight infants. However, higher levels of SO 2 were associated with lower risk of infant mortality. When sulfate and an estimate of non-sul fate particles were included in the regression simultaneously, associations were observed with no n-sulfate particles and an inverse relationship with sulfate particles. Respiratory neonatal mort ality was not associated with higher TSP in the Czech Republic case-control study (Bobak and Leon, 1999, 007678 ).



### Infant Mortality and Infant Respiratory Mortality, <1Year

A literature search did not reveal new studies on PM and infant mortality (<1 year) since the previous PM AQCD. Previously conducted studies include a case-control study that reported associations between infant mortality and TSP levels over the period between birth and death for infants in the Czech Republic (Bobak and Leon, 1999, 007678 ). An ecological study evaluated U.S.
PM 10 data for the year 1990 using long-term pollution levels in 180 U.S. counties (Lipfert et al., 2000, 004103 ). The authors found a 9.64% (95% CI: 4.60-14.9) increase in risk of infant mortality for non-low birth weight infants per 10 μg/m 3 increase in PM 10 , a 13.4% (95% CI: -10.3 to 43.5) increase in non-low birth weight respiratory-di sease related deaths (ICD 9 460-519) and a 19.5% (95% CI: 0.07-42.8) increase in all non-low birth weight respiratory-related infant deaths (ICD 9 460-519, 769, 770).



### Postneonatal Mortality and Postneonatal Respiratory Mortality, 1 Month–1 Year

Several studies have been conducted on PM and postneonatal mortality since the previous PM AQCD, including three from the U.S., one from Mexico, and three from Asia. Two case-control studies examined the risk of PM to postneonatal death in California. Research focused on Southern California for the period 1989-2000 linked birth and death certificates and considered PM 10 2 mo prior to death with adjustment for prenatal care, gender, parity, county, season, and mother’s age, race/ethnicity, and education (Ritz et al., 2006, 089819 ). As previously noted, this study did not find an association between PM 10 and neonatal mortality (<1 month), however an association was observed for post-neonatal mortality, with a 10 μg/m 3 increase in PM 10 associated with a 4% (95% CI: 1-6) increase in risk. The exposure period of 2 wk before death was also considered, producing effect estimates of 5% (95% CI: 1-10) for the same PM 10 increment. Even larger effect estimates were observed for those who died at ages 4-12 mo. When CO, NO 2 , and O 3 were simultaneously included with PM 10 in the model, the central estimate reduced to 2% for the 2-wk exposure period and 4% for the 2-mo exposure period, and both estim ates lost statistical si gnificance. The other casecontrol study of California considered PM 2.5 from 1999 to 2000 for infants born to mothers within five miles of a PM 2.5 monitoring station (Woodruff et al., 2006, 088758 ). Infants who died during the postneonatal period were matched to infants with date of birth within 2 wk and birth weight category. Exposure was estimated from the time of birth to death. Models considered parity and maternal race, education, age, and martial status. A 10 μg/m 3 increase in PM 2.5 was associated with a 7% (95% CI: -7 to 24) increase in postneonatal death County-level PM 10 and PM 2.5 for the first 2 mo of life for births in urban U.S. counties ( ≥ 250,000 residents) from 1999 to 2002 were evaluated in relation to postneonatal mortality with GEE models (Woodruff et al., 2008, 098386 ). Births were restricted to singleton births with gestational age ≤ 44 wk, same county of residence at birt h and death, and non-missing data on birth order, birth weight, and maternal race, education, and martial status. Higher levels of either PM metric were associated with highe r risk of postneonatal mortality, with 4% (95% CI: -1 to 10) increase in mortality risk per 10 μg/m 3 in PM 10 and 4% (95% CI: -2 to 11) increase in mortality risk for the same increment of PM 2.5 . This work builds on a previous study of 86 U.S. urban areas from

December 2009 7-541989 to 1991, finding a 4% (95% CI: 2-7) increase in postneonatal mortality per 10 μg/m 3 countylevel PM 10 over the first 2 mo of life (Woodruff et al., 1997, 084271 ). In Ciudad Juarez, Mexico, a case-crossover approach was applied to data from 1997 to 2001 based on death certificates and the cumulative PM 10 for the day of death and previous two days (Romieu et al., 2004, 093074 ). A case-crossover study of Kaohsiung, Taiwan from 1994 to 2000 compared the average of PM 10 on the day of death and two previous days to PM 10 in control periods a week before and week after death (Tsai et al., 2006, 090709 ). A similar approach was also applied to 1994-2000 data from Taipei, Taiwan, also using case-crossover methods for the lag 0-2 PM 10 with referent periods the week before and after death (Yang et al., 2006, 090760 ). In these case-crossover studies, season was addressed through matching in the study design. A 10 μg/m 3 increase in PM 10 was associated with a 2.0% (95% CI: -2.8 to 7.0) increase in the Mexico study, a 0.59 (95% CI: -15.0 to 18.8) increase in postneonatal death in the Kaohsiung study, and a 1.02% (95% CI: -13.2 to 17.6) increase in the Taipei study. A study in Seoul, South Korea from 1995 to 1999 used time-series approaches adjusted for temporal trend and weat her, based on national death registries excluding accidental deaths (Ha et al., 2003, 042552 ). A 10 μg/m 3 increase in PM 10 was associated with a 3.14% (95% CI: 2.16-4.14) increase in risk of death for postneonates. A subset of the studies examining postneonatal mortality also considered the subset of postneonatal deaths from respiratory causes. These include the time-series study in South Korea, finding a 17.8% (95% CI: 14.4-21.2) increase in respiratory-mortality per 10 μg/m 3 increase in PM 10 (Ha et al., 2003, 042552 ) and the case-crossover study in Mexico, for which the same increment in
PM 10 was associated with a 1.5% (95% CI: -14.1 to 13.0) decrease in risk (Romieu et al., 2004, 093074 ). Both California case-control studies identified associations, with a 5% (95% CI: 1-10) increase in risk in Southern California (Ritz et al., 2006, 089819 ) and 57.4% (95% CI: 7.0-132) increase in California per 10 μg/m 3 PM 10 (Woodruff et al., 2006, 088758 ). The U.S. study found this increment in PM 10 to be linked with a 16% (95% CI: 6.0-28. 0) increase in respiratory postneonatal mortality, although effect estimates for PM 2.5 were not statistically significant (Woodruff et al., 2008, 098386 ). Earlier studies on respiratory-related postne onatal mortality include the study of 86 U.S. urban areas, finding statistically significant effects (Woodruff et al., 1997, 084271 ).



### Sudden Infant Death Syndrome

Three studies examining the relationship between PM and sudden infant death syndrome (SIDS) have been published from 2002 onward. These studies examined infant mortality and were thereby discussed in this section previously. A case-control study over a 12-year period (1989 to 2000) matched 10 controls to deaths (cases) in Southern California (Ritz et al., 2006, 089819 ). A 10 μg/m 3 increase in PM 10 the 2 mo prior to death was associated with a 3% (95% CI: -1 to 8) increase in SIDS. Adjusted for other pollutants (CO, NO 2 , and O 3 ), the effect estimate reduced to 1% (95% CI: -5 to 7). A case-control study, also based in California, found an OR of 1.008 (95% CI: 1.006-1.012) per 10 μg/m 3 increase in PM 2.5 , considering a SIDS definition of ICD 10 R95 (Woodruff et al., 2006, 088758 ). Due to changes in SIDS diagnosis, another SIDS definition was explored for ICD 10 R99 in addition to ICD 10 R95. Under this SIDS definiti on, the effect estimate changed to 1.03 (95% CI: 0.79-1.35). The authors also examined whether the relationship between PM 2.5 and SIDS differed by season, finding no significant difference. PM 10 and PM 10-2.5 were not associated with risk of SIDS; numerical results were not provided for these PM metrics. The third recentstudy of PM and SIDS examined U.S. urban counties from 1999 to 2002 (Woodruff et al., 2008, 098386 ). Statistically nonsignificant relationships were observed between SIDS and PM 10 or PM 2.5 in the first 2 mo of life. These studies add to earlier work, such as a U.S. study that found higher risk of SIDS with higher annual PM 2.5 levels, including in a separate analysis of normal birth weight infants (Lipfert et al., 2000, 004103 ), and a U.S. study identifying a 12% (95% CI: 7-17) increase in SIDS risk per 10 μg/m 3 in PM 10 for the first 2 mo of life for normal weight births (Woodruff et al., 1997, 084271 ). A study based on Taiwan found higher SIDS risk wi th lower visibility (Knöbel et al., 1995, 155905 ), whereas a 12-city Canadian time-series study identified no significant associations (Dales et al., 2004, 087342 ). Deaths by SIDS were identified by different meth ods in the studies, partly due to transition from ICD 9 to ICD 10 codes, but also due to diff erent choices within the research design. Two studies examined multiple appro aches (ICD 10 R95, ICD 10 R95 and R99) (Woodruff et al., 2006,

December 2009 7-55088758 ; Woodruff et al., 2008, 098386 ), and other studies investigated ICD 9 798.0 and ICD 10 R95 (Ritz and Wilhelm, 2008, 156914 ), ICD 9 798.0 (Woodruff et al., 1997, 084271 ), ICD 9 798.0 and 799.0 (Knöbel et al., 1995, 155905 ), as well as a sudden unexplained death of infant <1 year for which an autopsy did not identify a specifi c cause of death (Dales et al., 2004, 087342 ). These variations in the definition of health outcomes ad d to differences in populations and study designs. Although some findings indicate a potential effect of PM on risk of SIDS, with the strongest evidence perhaps from the case-control study in California (Woodruff et al., 2006, 088758 ), others do not find an effect or observe an uncertain a ssociation. For the relationship between PM and SIDS, a 2004 review article concluded consistent evidence exists compared to evidence for other infant mortality effects (Glinianaia et al., 2004, 087898 ), whereas other reviews found weaker or insufficient evidence (Heinrich and Slama, 2007, 156534 ). Another review concluded that the scientific literature on air pollution and SIDS suggests an effect, but that further research is needed to draw a conclusion (Tong and Colditz, 2004, 087883 ).



### Considerations for Comparisons across Studies

Comparison of results across studies can be ch allenging due to several issues, including differences in methodologies, populations and st udy areas, pollution levels, and the exposure timeframes used. Given the large variation in study designs, the methods to address potential confounders vary. For example, weather and season were addressed in the case-control studies by matching, in the time-series study through non-linearfunctions of temperature and temporal trend, and in the ecological study through county-level variables. All studies included consideration of seasonality and weather. Researchers used differ ent definitions of respiratory-related deaths, including ICD 9 460-519 (Bobak and Leon, 1999, 007678 ; Lipfert et al., 2000, 004103 ); ICD 9 460-519, 769-770 (Lipfert et al., 2000, 004103 ); ICD 9 460-519, 769, 770.4, 770.7, 770.8, 770.9, and ICD 10 J00-J98, P22.0, P22.9, P27.1, P27.9, P28.0, P28.4, P2 8.5, and P28.9 (Ritz et al., 2006, 089819 ); and ICD 9 460-519 and ICD 10 J00-J99 for any cause on death certificate (Romieu et al., 2004, 093074 ); ICD 10 J00-99 and P27.1 excluding J69.0 (Woodruff et al., 2006, 088758 ; Woodruff et al., 2008, 098386 ); and ICD 9 460-519 (Woodruff et al., 1997, 084271 ). Socioeconomic conditions were included at the individual level, typically maternal education, in many studies (e.g., Bobak and Leon, 1999, 007678 ; Ritz and Wilhelm, 2008, 156914 ; Ritz et al., 2006, 089819 ; Woodruff et al., 1997, 084271 ; Woodruff et al., 2006, 088758 ) and at the communitylevel in others (e.g., Bobak and Leon, 1992, 044415 ; Penna and Duchiade, 1991, 073325 ) or for both individual and community-level da ta (e.g., Lipfert et al., 2000, 004103 ). The time-series approach is unlikely to be confounded by socioeconomic and other variables that do not exhibit day-to-day variation. Similarly, case-crossover methods use eachcase as his/her own control, thereby negating the need for individual-level confounders such as socioeconomic status (e.g., Romieu et al., 2004, 093074 ; Tsai et al., 2006, 090709 ; Yang et al., 2006, 090760 ). All studies published after 2001 incorporated individual-level socioeconomic data or were of case-crossover or time-series design. One study specifically examined whether socioeconomic status modified the PM and mortality relationship, dividing subjects into three socioeconomic strata based on the ZIP code of residence at death (Romieu et al., 2004, 093074 ). This work, based in Mexico, found that at lower socioeconomic levels the association between PM 10 and postneonatal mortality increased. Although the overall association showed higher risk of death with higher PM 10 with statistical uncertainty, for the lowest socio-economic group, a 10 μg/m 3 increment in cumulative PM 10 over the 2 days before death was associated with a 60% (95% CI: 3-149) increase in postneonatal death. A trend of higher effect for lower socio-economic condition is observed in all 3 lag structures. Studies differ in terms of the time frame of pregnancy that was used to estimate exposure. Exposure to PM for infant mortality (<1 yr) was es timated as the levels between birth and death (Bobak and Leon, 1999, 007678 ), annual community levels (Lipfert et al., 2000, 004103 ; Penna and Duchiade, 1991, 073325 ) and the 3-5 days prior to death (Loomis et al., 1999, 087288 ). For neonatal deaths, exposure timeframes considered were the ti me between birth and death (Bobak and Leon, 1992, 044415 ; Bobak and Leon, 1999, 007678 ), annual levels (Bobak and Leon, 1999, 007678 ; Lipfert et al., 2000, 004103 ), monthly levels (Shinkura et al., 1999, 090050 ), the same day concentrations (Lin et al., 2004, 095787 ), and the 2 mo or 2 wk prior to death (Ritz et al., 2006, 089819 ). Postneonatal mortality was associated with PM concentrations based on annual levels (Bobak and Leon, 1992, 044415 ; Lipfert et al., 2000, 004103 ), between birth and death (Bobak and

December 2009 7-56Leon, 1999, 007678 ; Woodruff et al., 2006, 088758 ), 2 mo before death (Ritz et al., 2006, 089819 ), the first 2 mo of life (Woodruff et al., 1997, 084271 ; Woodruff et al., 2006, 088758 ), the day of death (Ha et al., 2003, 042552 ), and the average of the same day as death and previous 2 days (Romieu et al., 2004, 093074 ; Tsai et al., 2006, 090709 ; Yang et al., 2006, 090760 ). Thus, no consistent window of exposure was identified across the studies.
PM 10 concentrations were highest in South Korea (69.2 μg/m 3 ) (Ha et al., 2003, 042552 ) and Taiwan (81.45 μg/m 3 ) (Tsai et al., 2006, 090709 ), and lowest in the U.S. (29.1 μg/m 3 ) (Woodruff et al., 2008, 098386 ) and Japan (21.6 μg/m 3 ) (Shinkura et al., 1999, 090050 ). All studies used community-level exposure information based on ambient monitors, as opposed to exposure measured at the individual level (e.g., subject’s home) or personal monitoring. Given similar sources for multiple pollutants (e.g ., traffic), disentangling the health responses of copollutants is a challenge in the study of ambient air pollution. Several studies examined multiple pollutants, most by estimating the effect of different pollutants through several univariate models. Some studies noted the difficulty of sepa rating PM effects from those of other pollutants, but noted stronger evidence for particles than other pollutants (Bobak and Leon, 1999, 007678 ). A few studies applied copollutant models by including multiple pollutants simultaneously in the same model. Effect estimates for the relationship between PM 10 and neonatal deaths in São Paulo were reduced to a null effect when SO 2 was incorporated (Lin et al., 2004, 095787 ). Associations between
PM 10 and postneonatal mortality or respiratory postn eonatal mortality remained but lost statistical significance in a multiple pollutant model with CO, NO 2 , and O 3 (Ritz et al., 2006, 089819 ). Several review articles in recent years have ex amined whether exposure to PM affects risk of infant mortality, generally concluding that more consistent evidence has been observed for postneonatal mortality, particularly from respiratory causes (Bobak and Leon, 1999, 007678 ; Heinrich and Slama, 2007, 156534 ; Lacasaña et al., 2005, 155914 ; Sram et al., 2005, 087442 ). In one review authors identified 14 studies on infant mortality and air pollution and determined that studies on PM and infant mortality do not provi de consistent results, although more evidence waspresent for an association for some subsets of infant mortality such as postneonatal respiratoryrelated mortality (Bobak and Leon, 1999, 007678 ). The relationship between PM and postneonatal respiratory mortality was concluded to be causal in one review (Sram et al., 2005, 087442 ), and strong and consistent in another (Heinrich and Slama, 2007, 156534 ). Meta-analysis using inversevariance weighting of PM 10 studies found that a 10 μg/m 3 increase in acute PM 10 exposure was associated with 3.3% (95% CI: 2.4-4.3) increase in risk of postneonatal mortality, whereas the same increment of chronic PM 10 exposure was linked with a 4.8% (95% CI: 2.2-7.2) increase in postneonatal mortality and a 21.6% (95% CI: 10. 2-34.2) increase for respiratory postneonatal mortality (Lacasaña et al., 2005, 155914 ). Studies that examined multiple outcomes and ages of death allow a direct comparison based on the same study population and methodologies, thereby negating the concern that inconsistent results are due to underlying variation in population, approaches, etc. In this review, one study, based in Southern California identified no association for ne onatal effects (numerical results not provided) but statistically significant results for postneonatal mortality (Ritz et al., 2006, 089819 ). Figure
7-5 compares risk for the postneonatal period for respiratory and total mortality. In six of the seven studies, higher effect estimates were observed fo r respiratory-related mortality. Results from the neonatal period found higher effects for total mortal ity compared to respiratory mortality (Bobak and Leon, 1999, 007678 ) and the reverse for a study examining in fant mortality (Lipfert et al., 2000, 004103 ). Thus, there exists evidence for a stronger effect at the postneonatal period and for respiratory-related mortality, although this trend is not consistent across all studies.

December 2009 7-57

### Figure 7-5. Percent increase in postneonatal mortality per 10 μg/m 3 

### in PM 10 

### , comparing risk for total and respiratory mortality. Panel a (left) provides central estimates; panel b (right) also adds the 95% intervals. The points reflect central estimates and the lines the 95% intervals. Solid lines represent statistically significant effect estimates; dashed lines represent non-statistically significant estimates. 1



## 7.4. 1.6. Dec rements in Sperm Quality

Limited research conducted in the Czech Repub lic on the effect of ambient air pollution on sperm production has found associations between elevated air pollution and decrements in proportionately fewer motile sperm, proportionately fewer sperm with normal morphology or normal head shape, proportionately more sperm with abnormal chromatin (Selevan et al., 2000, 012578 ), and an increase in the percentage of sperm with DNA fragmentation (Rubes et al., 2005, 078091 ). These results were not specific to PM, but for exposure to a high-, medium- or low-polluted air mixture. Similarly, in Salt Lake City, Utah, PM 2.5 was associated with decreased sperm motility and morphology (Hammoud et al., 2009, 192156 ). Research in Los Angeles, California examined 5,134 semen samples from 48 donors in relation to ambien t air pollution measured 0-9, 10-14, 70-90 days before semen collection over a 2- yr period (1996-1998). Ambient O 3 during all exposure periods had a significant negative correlation with average sperm concentration, and no other pollutant measures were significantly associated with sperm quality parameters, or presente d quantitatively (Sokol et al., 2006, 098539 ).



## 7.4.2. Toxicological Studies

This section summarizes recent evidence on repr oductive health effects reported with exposure to ambient PM; no evidence was presented in this area in the 2004 PM AQCD. Studies from different toxicological rodent models allow for i nvestigation of specific mechanisms and modes of

1 Studies included are Bobak and Leon (1999, 007678 ), Ha et al. (2003, 042552 ), Ritz et al. (2006, 089819 ), Romieu et al. (2004, 093074 ),

Romieu et al. (2008, 156922 ), Woodruff et al. (1997, 084271 ), Woodruff et al. (2006, 088758 ). Findings from Bobak and Leon (1999,
007678 ) were based on TSP and were converted to PM 10 estimates assuming PM 10 /TSP = 0.8 as per summary data in the original article (Bobak and Leon, 1999, 007678 ). Findings from Woodruff et al. (1997, 084271 ) for respiratory-related mortality were based on non-low birth weight infants. Result s for Woodruff et al. (2006, 088758 ) were based on PM 2.5 and were converted to PM 10 assuming

PM 2.5 /PM 10 = 0.6.

December 2009 7-58action for reproductive changes. Emphasis is pl aced here on results from different windows of development, i.e., exposure in utero, neonatally or as an adult can affect reproductive outcomes as an adult. In addition, studies evaluating whether fertility is affected in female and/or male animals by a similar exposure, and how exposures are transmitted to the fertility of the F 1 offspring, are summarized. Hormonal changes which can lead to decreased sperm count or changes in the estrous cycle are also of interest. Studies of pregnancy losses and placental sufficiency are also reported. Most recently, the role of environmental chemicals in shifting sex ratios (also seen in epidemiologic studies) and in affecting heritable DNA changes have become outcomes of interest.



## 7.4. 2.1. Fem ale Reproductive Effects

Urban Air

Windows of exposure are important in determinin g reproductive success as an adult. Exposure as a neonate may have a drastically different impact than does a similar adult exposure. To test this, female BALB/C mice were exposed to ambient air in Sao Paulo as neonates or as adults and then were bred to non-exposed males (Mohallem et al., 2005, 088657 ). Ambient concentrations of the pollutants CO, NO 2 , PM 10 , and SO 2 were 2.2 ± 1.0 ppm, 107.8 ± 42.3 μ g/m 3 , 35.5 ± 12.8 μ g/m 3

, and 11.2 ± 5.3 μ g/m 3 , respectively. They reported decreased fertility in animals exposed as newborns, but not in adult-exposed female BALB/c mice. There were a significantly higher number of liveborn pups from dams housed in filtered chambers (PM and gaseous components removed) versus animals exposed to ambient air as ne wborns. There was also a higher incidence of implantation failures in dams reared as newbor ns in polluted chambers. Sex ratio, number of pregnancies per group, resorptions, fetal deaths, and fetal placental weights did not differ significantly by exposure group. Thus, in these studies, exposure to ambient air pollution affected future reproductive success of females if they we re exposed as neonates and not if exposed as adults.



### Diesel Exhaust

Significant work has been done in male rodent mo dels to determine the effect of PM exposure on reproductive outcomes, with fewer studies conducted using female rodents. Tsukue et al. (2004, 096643 ) exposed pregnant C57-BL mice to DE (0.1 mg/m 3 ) or to clean air (controls) for 8 h/day from GD2-13. The concentration of the gaseous materials including NO, NO X , NO 2 , CO and SO 2 are 2.2 ± 0.34 ppm, 2.5 ± 0.34 ppm, 0.0 ppm, 9. 8 ± 0.69 ppm, and <0.1 ppm (not detectable), respectively. At GD14 female fetuses were collected for analysis of mRNA for two genes involved in sexual differentiation (Ad4BP-1/SF-1 and MIS), and found no significant changes. Work by Yoshida et al. (2006, 097015 ) showed changes in these two transcripts in male ICR fetuses exposed to similar concentrations of DE, albeit with different daily durations of exposure. Further work by Yoshida et al. (2006, 097015 ) showed that of three mouse strains te sted, ICR male fetuses were the most sensitive to DE-dependent changes in these two genes. Nonetheless, strain sensitivity to DE particles may also differ by sex. Thus, it appears that female mice exposed in utero to DE show a lack of response at the mRNA level of MIS or Ad4bP-1/SF-1 , important genes in male sexual differentiation that showed DE-dependent changes in male pups from dams exposed in utero. Female fetuses have shown a decrease in BMP-15, which is related to oocyte development (Tsukue et al., 2004, 096643 ). A sensitive measure of androgenic activity in male rodents is anogenital distance (AGD), i.e., decreased AGD is seen with exposure to anti-androgenic environmental chemicals, the phthalates (Foster et al., 1980, 094701 ; Foster et al., 2001, 156442 ). To assess the role of DE exposure on reproductive success and anti-androgenic effects on offspring, Tsukue et al. (2002, 030593 ) exposed 6 week-old female C57-Bl mice to 4 mo of DE (0.3, 1.0, or 3.0 mg/m 3 ; PM MMAD of 0.4 μ m) or filtered air. DE-exposed estrous females had si gnificantly decreased uterine weight (1.0 mg/m 3 ). Some of the DE-exposed females were bred to unexposed males and DE-exposure led to increased, albeit not significantly increased, rates of pregnancy loss in mated females (up to 25%). Offspring were weighed after birth and decreases in body weig ht were observed at 6 and 8 wk (males and females, 1.0 and 3.0 mg/m 3 ) and 9 wk (females, 1.0 and 3.0 mg/m 3 ). Anogenital distance was decreased in 70-day old DE-exposed male offspring (0.3 mg/m 3 ). In female offspring at 70 days of

December 2009 7-59age, lower organ weights (adrenals, liver, and thymus) were observed (1.0 mg/m 3 ) compared to controls; thymus weight of the 0.3 mg/m 3 females was also lower at 70 days. Crown to rump length in females from dams exposed to DE (1.0 and 3.0 mg/m 3 ) was less than the control group. In conclusion, adult exposure to DE led to matern al-dependent reproductive changes that affected outcomes in offspring that manifested as decr eased pup body weight, anti-androgenic effects like decreased AGD and decreased organ weight (which may have been confounded by changes in body weight because weights were not reported as relative organ weights).



## 7.4. 2.2. Mal e Reproductive Effects

Diesel Exhaust

Studies were performed to determine PM-dependent strain sensitivity of the male reproductive tract using male steroidogenic enzymes as the mode l pathway. Three strains of pregnant mice (ICR, C57Bl/6J or ddY mice) were continuously exposed to DE at 0.1 mg/m 3 via inhalation or clean air over gestational days 2-13 (Yoshida et al., 2006, 156170 ). At GD14, dams were euthanized and fetuses were collected. Male fetuses were coll ected from each dam for mRNA analysis of genes related to male gonad development including Mulleri an inhibiting substance (MIS; crucial for sexual differentiation including Mullerian duct regression in males), steroid transgenic factor (Ad4BP/SF-1, an enzyme in the testosterone synthesis pathway), cytochrome P450 cholesterol side chain cleavage enzyme (P450scc), and other steroidogenic enzymes [17 β -hydroxysteroid dehydrogenase (HSD), cytochrome P450 17- α -hydroxylase (P450c17), and 3- β hydroxysteroid dehydrogenase (3 β HSD)]. There were significant decreases in MIS (ICR and C57BL/6 mice) and Ad4BP/SF-1 (ICR mice) compared to the control groups. The ddY strain demonstrated no changes in Ad4BP/SF-1 or MIS, which may be due to marked changes in 3 β -hD expression compared to non-DE exposed controls. From these studies, it appears that mouse strains with in utero exposure to DE show differential sensitivity in gonadal differentiation genes (mRNA) e xpression in male offspring; ICR are the most sensitive, followed by C57BL/6, with ddY mice being the least sensitive. Yoshida et al. (2006, 097015 ) also monitored changes in the ma le reproductive tract after in utero exposure to DE. Timed-pregnant ICR dams were exposed during gestation (2 days post-coitus [dpc]-16 dpc) to continuous DE (0.3, 1.0 or 3.0 mg/m 3 ) or clean air. The reproductive tracts of male offspring were monitored at 4 wk postnatally. These pups received possible continued exposure through lactation as dams were exposed to DE dur ing gestation and nursed pups. Exposure to 0.3 mg/m 3 of DE had no effect on male reproductive organ weight or serum testosterone. The intermediate concentration of 1.0 mg/m 3 induced increases in serum testosterone. Exposure to the higher concentration (1.0 and 3.0 mg/m 3 ) of DE led to significant increases in reproductive gland weight (testis, prostate, and coagulating gland). Th e organ weights are presented as absolute numbers and not adjusted for body weight, which is sometimes problematic for complete representation of hormonal changes, as body weight may confound absolute organ weight changes. Transcripts relating to male sexual differentiation (MIS and AD4BP/SF-1, 1.0 and 3.0 mg/m 3 ) were also significantly decreased. Sexual differentiation is a tightly regulated process and these changes in transcription may lead to changes that can affect genitalia development. The effects of DE exposure on male spermatoge nesis have also been demonstrated. Exposure of pregnant ICR mice to DE (2-16 dpc continuous inhalation exposure to 1.0 mg/m 3 or filtered clean air) led to impaired spermatogenesi s in offspring (Ono et al., 2007, 156007 ). Male offspring were followed at PND 8, 16, 21 (3 wk), 35 (5 wk) and 84 (12 wk). After 16 dpc, but before termination of the study, all of the animals were transferred to a regular animal care facility and received clean air exposure until the termination of the study. No cross fostering was performed in this experiment, so pups that were born to DE-exposed dams were also nursed on these dams and may have received lactational exposure to DE. The gaseous components of the diluted DE included NO, NO 2 , SO 2 , and CO 2 at concentrations of 11.75 ± 1.18, 4.62 ± 0.36 , 0.21 ± 0.01, and 4922 ± 244 ppm, respectively. Body weight was significantly depressed at PNDs 8 and 35. Accessory gland relative weight was significantly increased at PND8 and PND16 only. Serum testosterone was significantly decreased at 3 wk and was significantly increased at 12 wk. At 5 and 12 wk, daily sperm production (DSP) was significantly decreased. FSH receptor and StAR mRNA levels were significantly increased at 5 and 12 wk, respectively. Relative testis weight and rela tive epididymal weight were unchanged at all

December 2009 7-60time points. Histological changes showed sertoli cel ls with partial vacuolization and a significant increase in testicular multinucleated giant cells in the seminiferous tubules of DE-exposed animals compared to control. This study indicates that in utero exposure to DE had effects on spermatogenesis in offspring at the histological, hormonal and functional levels. In utero exposure to DE and its effect on adult body weight, sex ratio, and male reproductive gland weight was measured by Yoshida et al. (2006, 097015 ). Pregnant ICR mice were exposed by inhalation to DE (0.3, 1.0 or 3.0 mg/m 3 ) or clean air from 2 dpc to 16 dpc. Pups were allowed to nurse in clean air on exposed dams until weaning a nd at PND28, male pups were sacrificed. At this time, serum testosterone and pup reproductive gland weight was determined. Significant increases in relative reproductive organ weights were reported at 1.0 and 3.0 mg/m 3 for the seminal vesicle, testis, epididymis, coagulating gland, prostate and liver. Male pup serum testosterone was significantly increased at 1.0 mg/m 3 . Mean testosterone positively corre lated with testis weight, DSP, aromatase and steroidogenic enzyme message levels (P450cc, c17 lyase, and P450 aromatase). Sex ratio did not differ in DE-exposed animals vers us control. Male pup body weight of DE-exposed animals was significantly increased at PND28 (1.0 and 3.0 mg/m 3 ). These studies show that in utero DE-exposure led to increased serum testosterone an d increased reproductive gland weight in male offspring early in life. The effects of DE on murine adult male reproductive function were studied by exposing ICR male mice (6 wk of age) to DE (clean air control, 0.3, 1.0 or 3.0 mg/m 3 ) for 12 h/day for 6 mo with another group receiving a 1-mo recovery of clean air post-exposure (Yoshida and Takedab, 2004, 097760 ). After 6 mo of DE exposure, there was a con centration-dependent increase in degeneration of seminiferous tubules and a decrease in DSP/g of testis tissue. After 6 mo exposure to DE particles plus 1 mo of recovery in clean air, significan t decreases remained in DSP at the two highest concentrations. The effect of ingestion of depos ited PM on the fur with grooming cannot be ruled out as a possible exposure pathway in this experiment. To expand on PM-dependent changes in spermatogenesis, an eloquent DE-exposure model was designed to determine if PM or the gaseous phase of DE was responsible for changes in sperm production in rodents (Watanabe, 2005, 087985 ). Pregnant dams (F344/DuCrj rats) exposed to DE (6 h/day exposure to 0.17 or 1.71 mg/m 3 ; <90% of PM less than 0.5 μ m; NO 2 concentrations 0.10 and 0.79 ppm, respectively) or filtered air (removing PM only, low concentration filtered air and high concentration filtered air) from GD7 to parturition produced adult male offspring with a decreased number of sertoli cells and decreased DSP (PND 96) when compared to control mice exposed to clean air. The concentrations of NO 2 for the low and high filtered exposure groups were 0.1 and 0.8 ppm, respectively. Because both PM -filtered and DE-exposure groups showed the same outcomes, the effects are likely due to gaseous components of DE.



### Motorcycle Exhaust

Adult male (8-wk old) Wistar rats were expose d to motorcycle exhaust (ME) for 1 h in the morning and 1 h in the afternoon (5 day/wk) at 1:50 dilution for 4 wk (group A), 1:10 dilution for 2 wk (group B) or 4 wk (group C), or to clean air (Huang et al., 2008, 156574 ). After 4 wk of exposure, both exposed groups had significantly decreased body weight compared to the control group. All three ME exposure groups showed a decreased number of spermatids in the testis. Both 1:10 exposure groups also demonstrated decreased caudal epididymal sperm counts. Group C had significant decreased testicular weight, decreased mRNA expression for the cytochrome P450 substrate 7-ehtoxycoumarin O-de-ethylase, and increased IL-6, IL-1 β , and COX-2 mRNA levels. Decreased protein levels of the antioxidant, superoxide dismutase, and increased IL-6 protein were reported for group C when compared to control. In addition, serum testosterone was significantly decreased in group C. Co-treatment with the antioxi dant vitamin E resulted in partial attenuation of serum testosterone levels and caudal epididymal sperm counts, and returned IL-6, IL-1 β , and COX-2 ME exposure-dependent message levels to baseline. The glutathione antioxidant system and lipid peroxidation were unchanged. In conclusion, male animals exposed to ME showed significant decrements in body weight, spermatid number, and serum testosterone with an increase in inflammatory cytokines. Vitamin E co-treatme nt with ME-exposure led to an attenuation of inflammation and a partial rescue of testosterone levels and sperm numbers.

December 2009 7-61

### Summary of Toxicological Study Findings for Male Reproductive Effects

In summary, laboratory animals exposed to DE in utero or as adults manifest with abnormal effects on the male reproductive system. In uter o exposure to DE induced increased reproductive gland weight and increased serum testosterone in early life (PND28), which may lead to early puberty (albeit not measured in this study). With similar in utero DE exposures, later life outcomes include decreased DSP, aberrant sperm morphol ogy, and hormonal changes (testosterone and FSHr decrements). Chronic exposure of adult mice to DE also induced decreased DSP and seminiferous tubule degeneration. DE-dependent effects on male reproductive function have been reported in multiple animal models, with only one model separating exposure based on particulate versus gaseous components. DE and filtered air (gaseous ph ase only) exposure in utero induced sertoli cell and DSP decrements in both groups, indicating that the gaseous phase of DE was causative. Adult male rats exposed to ME manifested with decreased spermatid number, serum testosterone, and an increase in inflammatory cytokines. Significant ef fects on the male reproductive system have been demonstrated after exposure to ambient PM sources (DE or ME). Nonetheless, these models often include a complex mixture of gaseous component and PM exposure, which makes interpreting the contribution from PM alone difficult.



## 7.4. 2.3. Mul tiple Generation Effects

Urban Air

Veras et al. (2009, 190496 ) investigated pregnancy and female reproductive outcomes in BALB/c female mice exposed to ambient air or PM -filtered ambient air at one of two different time periods (before conception and during pregnancy) near an area of high traffic density in Sao Paulo, Brazil. Exposures were 27.5 and 6.5 μ g/m 3 PM 2.5 for ambient and PM-filtered air chambers, respectively, with 101 μ g/m 3 NO 2 , 1.81 μ g/m 3 CO, and 7.66 ppm SO 2 in both chambers. Two groups of 2nd generation (G2) nulliparous female mice were continuously exposed from birth. Estrous cyclicity and ovarian follicle classification were followed at PND60 (reproductive maturation) in one group. A further group was subdivided into fourgroups by exposures during pregnancy following reproductive capability and pregnancy outcomes of the G2 mice. Animals exposed to ambient air versus PM-filtered air had an extended time in estrous and thus, a reduction in the number of cycles during the study period. The number of antral follicles was significantly decreased in the ambient air versus the PM-filtered air animals. Other follicular quantification (number of small, growing or preovulatory follicles) showed no differences betw een the two chambers. There was an increase in the time necessary for mating, a decrease in the fe rtility index, and an increase in the pregnancy index in the ambient air group versus the PM-filtere d group. Specifically, in the ambient air groups, there was a significant increase in rate of the post-implantation loss in G1 and G2 groups. However, there was no statistically significant change in number of pups in the litter. Fetal weight was decreased in all treatment groups (ambient air groups G1 and G2, and PM-filtered G2) when compared to the PM-filtered G1 group or animals ra ised entirely in filtered air, showing that fetal weight was affected by both pre-gestational and gestational PM exposure. PM exposure prior to conception is associated with increased time in estrous, which in other animal models can be related to ovarian hormo ne dysfunction and ovulatory problems. These estrous alterations can contribute to fecundity issues. Th ere was no significant difference in number of preovulatory follicles in the above model, but ther e was a statistically significant decrease in the number of antral follicl es (Veras et al., 2009, 190496 ). Antral follicles are the last stage in follicle development prior to ovulation, and a decrease in an tral follicle number can be related to premature reproductive senescence, premature ovarian failure, or early menopause, which were not followed in this study. In this study (Veras et al., 2009, 190496 ), the males that were used to generate the G1 and G2 groups were also exposed to ambient air or PM-filtered ambient air, and thus the reproductive contribution of these males to the overall fertil ity and mating changes in the females cannot be totally eliminated as a possible confounder to the observed effects. Thus, these effects are hard to differentiate as male- or female-dependent and lik ely indicate a general loss of reproductive fitness. Interestingly, both pre- and gestational exposure to ambient air induced a significant loss in postDecember 2009 7-62implantation of fetuses and this may be related to placental insufficiency as has been described in other work by this lab (Veras et al., 2008, 190493 ).



## 7.4. 2.4. Rec eptor Mediated Effects

Arylhydrocarbon Receptor (AhR)

Diesel Exhaust Particles

The AhR is often activated by chemicals classified as endocrine disrupting compounds (EDCs), exogenous chemicals that behave as horm onally active agents, disrupting the physiological function of endogenous hormones. DE particles are known to activate the AhR. A recent study by Izawa et al. (2007, 190387 ) showed that certain polyphenols (quercetin from the onion) and food extracts (Ginkgo biloba extract) are able to attenuate DE particle-dependent AhR activation when measured with the Ah-Immunoassay, thus possibly at tenuating the EDC activity of DE particles.



## 7.4. 2.5. Dev elopmental Effects

Sex Ratio

Urban Air

A correlation between PM 10 exposure and a decrease in standardized sex ratios (SSRs) has been reported in humans exposed to air pollution (Lichtenfels et al., 2007, 097041 ; Wilson et al., 2000, 010288 ), with fewer numbers of male births reported. To understand this shift, two groups (control and exposed) of male Swiss mice were house d concurrently in Sao Paulo and received either ambient air exposure or filtered air (chemical and particulate filtering) from PND10 for 4 mo (Lichtenfels et al., 2007, 097041 ). Filtration efficiency for PM 2.5 , CB, and NO 2 inside the chamber was found to be 55%, 100%, and 35%, respectively. After this exposure, non-exposed females were placed in either chamber to mate. After mating, the males were sacrificed and testes collected; males exposed to ambient air showed decr eased testicular and epididymal sperm counts, decreased total number of germ cells, and decreased elongated sper matids, but no significant change in litter size. Females were housed in the chambers and sacrificed on GD19 when the number of pups born alive and the sex ratio were obtained. There was a si gnificant decrease in the SSR for pups born after living in the ambient air-exposed chamber compared to the filtered chamber. In this study, a shift in SSR has been shown for both humans and rodents exposed to air pollution, but other studies with DE exposure (Yoshida et al., 2006, 156170 ) or ambient air in Sao Paulo (Mohallem et al., 2005, 088657 ) showed no changes in rodent sex ratio. Possible exposure to PM and other components of ambient air via ingestion during grooming cannot be ruled out in this rodent model.



### Immunological Effects: Placenta

Diesel Exhaust

Placental insufficiency can lead to the loss of a pregnancy or to adverse fetal outcomes. DEexposure has been shown to induce inflammation in various models. Fujimoto et al. (2005, 096556 ) assessed cytokine/immunological changes of DE-dependent inhalation exposure on the placenta during pregnancy. Pregnant Slc:CR mice were exposed to DE (0.3, 1.0, or 3.0 mg/m 3 ; PM MMAD of 0.4 μm) or clean air from 2 to 13 dpc and dams, placenta, and pups were collected at 14 dpc. There was a significant increase in the number of absorbed placentas in DE-exposed animals (0.3

December 2009 7-63and 3.0 mg/m 3 ) with a significant decrease in the number of absorbed placentas in DE-exposed animals at the middle concentration (1.0 mg/m 3 ). Absorbed placentas from DE exposed mice had undetectable levels of CYP1A1 and twofold increases in TNF- α ; CYP1A1 placental mRNA from healthy placentas of DE-exposed mice was unc hanged versus control. IL-2, IL-5, IL-12 α , IL-12 β and GM-CSF mRNA significantly increased in placent as of DE-exposed animals (0.3 and 3.0 mg/m 3 ). Fujimoto et al. (2005, 096556 ) reported DE-induced significant increases in multiple inflammatory markers in the placenta with significant increases in the number of absorbed placentas.



### Immunological Effects: Asthma

Model Particles

In utero exposure may confer susceptibility to PM-induced asthmatic responses in offspring. Exposure of pregnant BALB/c mice to aerosolized RO FA leachate by inhalation or to DE particles intranasally increases asthma susceptibility to their offspring (Fedulov et al., 2008, 097482 ; Hamada et al., 2007, 091235 ). The offspring from dams exposed for 30 min to 50 mg/mL ROFA 1, 3, or 5 days prior to delivery responded to OVA immunization and aerosol challenge with airway hyperreactivity and increased antigen-specific Ig E and IgG1 antibodies (Hamada et al., 2007, 091235 ). Airway hyperreactivity was also observed in th e offspring of dams intranasally instilled with 50 μg of DE particles or TiO 2 , or 250 μg CB, indicating that the same effect could be demonstrated using relatively “inert ” particles (Fedulov et al., 2008, 097482 ). Pregnant mice were particularly sensitive to exposure to DE or TiO 2 particles, and genetic analysis indicated differential expression of 80 genes in response to TiO 2 in pregnant dams. Thus pregnancy and in utero exposure may enhance responses to PM, and exposure to even relatively inert particles may result in offspring predisposed to asthma.



### Placental Morphology

Urban Air

Exposure to ambient air pollution during pregnancy is associated with reduced fetal weight in both human and animal models. The effect of particulate urban air pollution on the functional morphology of the mouse placenta was explored by exposing second generation mice in one of four groups to urban Sao Paulo air (PM was 67% PM 2.5 , mainly of vehicular origin) or filtered air (Veras et al., 2008, 190493 ). Experimental design was: group F-F comprised of mice that were raised in filtered air chambers and completed pregnancy in f iltered air chambers; group F-nF raised in filtered air and pregnant in ambient air; group nF-nF raised and completed pregnancy in non-filtered air chambers; and group nF-F mice raised in ambient air and received filtered air during pregnancy.
Mean PM 2.5 concentrations in the F and nF chambers were 6.5 and 27.5 μg/m 3 , respectively. Exposure was from PND20-PND60. After this exposure, the animals were mated and then maintained in their respective chambers during pr egnancy. Pregnancy was terminated at GD8 (near term) with placentas and fetuses collected for analysis. Exposure to ambient PM pre-gestationally or gest ationally led to significantly smaller fetal weight (total litter weight). Pregestational exposure to ambient air induced significant increases in fetal capillary surface area and total mass-speci fic conductance, but this may be explained by reduced maternal/dam blood space and diameters. Gestational exposure to non-filtered air was associated with reduced volume, diameter (calib er) and surface area of maternal blood space with compensatory greater fetal capil lary surface and oxygen diffusion conduction rates. Intravascular barrier thickness, a quantitative relationship be tween trophoblast volume and the combined surfaces of maternal blood spaces and fetal capillaries, was not reduced with ambient air exposure. This study provides evidence that fetal/placental circulatory adaptation to maternal blood deficits after ambient PM exposure may not be sufficient to overcome PM-d ependent birth weight deficits in mice exposed to ambient air, with the magnitude of this effect greater in the gestationally-exposed groups.

December 2009 7-64

### Placental Weights and Birth Outcomes

Urban Air

Pregnant female Swiss mice were exposed to ambi ent air (Sao Paulo) or filtered air over various portions of gestation to determine if ther e was an association between fetal or placental weight or birth outcomes with exposure to air pollution (Rocha et al., 2008, 096685 ). The reported ambient concentrations of PM 10 (42 ± 17 μ g/m 3 ), NO 2 (97 ± 39 μ g/m 3 ), and SO 2 (9 ± 4 μ g/m 3 ) were measured 100 m away from the rodent exposure ch ambers. By using six time windows of exposure that covered 1-3 wk of gestation (the entire ge station period in a mouse), a significant decrease in near-term fetal weight (GD19) was induced by ambient air-exposure during the first week of gestation. Decreased placental weight could be induced by ambient air exposure during any of the 3 wk of gestation. This study points to possible windows of exposure that may be important in evaluating epidemiologic study results.



### Neurodevelopmental Effects

Diesel Exhaust

The diagnosis of autism is on the rise in the Western world with its etiology mostly unknown. Autism-associated cell loss is brain region-specific and hypothesized to be developmental in origin. Sugamata et al. (2006, 097166 ) exposed pregnant ICR mice to DE (0.3 mg/m 3 ) continuously from 2 dpc to 16 dpc. Pups with in utero exposure to DE were nursed in clean air chambers, but may have received gastro-intestinal exposure via lactational transfer of various components of DE. At 11 wk of age, cerebellar brain tissue was collected. Ear lier work has shown that DE particles (<0.1 μ m) have been detected in the brains (cerebral cortex and hippocampus) of newborn pups who were born to dams exposed to DE during pregnancy (Sugamata et al., 2006, 097166 ). Histological analysis of DEexposed pup cerebella revealed significant increases in caspase-3 (c-3) positive cells compared to control and significant decreases in cerebella Purkinje cell numbers in DE-exposed animals versus control. The ratio of cells positive for apoptosis (c-3 positive) showed a nearly significant sex difference with males displaying increased apoptosi s versus females (p = 0.09). In humans with autism, the cerebellum has a decreased number of Purkinje cells, which is thought to be fetal and developmental in origin; further, these authors sp eculate that humans may be more sensitive to DEdependent neuronal brain changes, as the human placenta is two-layers thick compared to the mouse placenta that is four layers thick.



### Behavioral Effects

Diesel Exhaust Particles

Body weight decrements at birth have recently been associated through the Barker hypothesis with adverse adult outcomes. Thus, many public ations have begun to focus on decreased birth weight for gestational age and associated adult changes. Hougaard et al. (2008, 156570 ) exposed 40 timed-pregnant C57BL/6 dams to DE particles re ference materials (SRM 2975) via inhalation over GD7-GD19 of pregnancy. They found significantly d ecreased pup weight at weaning, albeit not at birth. PM-dependent liver changes were monitored by following various inflammatory and genotoxicity-related mRNA transcripts and there we re no significant differences in pups at PND2. The comet assay from PND2 pup livers showed no significant differences in DNA damage between DE particle-exposed and control animals. The proho rmone, thyroxine, was unchanged in control and DE particle-exposed dams and offspring at wean ing. At 2 mo, female DE particle-exposed pups required less time than controls to locate the platform in its new location during the first trial of the spatial reversal learning task in the Morris water maze. Thus, DE particle exposure during in utero development led to behavioral changes without bod y weight at weaning or changes in inflammatory markers or thyroid hormone levels.

December 2009 7-65Diesel Exhaust

The effect of in utero DE exposure on CNS mo tor function was evaluated in male pups (ICR mice) after dams received DE exposure ( 8h/d×5d/wk) from GD2-GD17 (Yokota et al., 2009, 190518 ). The exposure atmosphere contained concentrations of 1.0 mg/m 3 for particle mass, 2.67 ppm CO, 0.23 ppm NO 2 , and <0.01 ppm SO 2 . Spontaneous motor activity was significantly decreased in pups (PND35), as was the dopamine metabolite homovanillic acid measured in the striatum and nucleus accumbens, indicating decrease d dopamine (DA) turnover. However, DA levels were unchanged in the same areas of the brain. The authors conclude that these data demonstrate that maternal exposure to DE induced hypolocomotion, si milar to earlier studies with adult and neonatal DE particle exposure (Peters et al., 2000, 001756 ), with decreased extracellular DA release.



### Lactation

Diesel Exhaust

Tozuka et al. (2004, 090864 ) monitored the transfer of PAHs to fetuses and breast milk of F344 rats exposed to DE (6h/day) for 2 wk from GD7-GD 20 (minus 4 days for the weekend when no exposure occurred) with PM 10 concentration of 1.73 mg/m 3 . At PND 14, breast milk was collected. Fifteen PAHs were monitored in the DE exposure chamber and seven were quantified in dam blood with levels of phenanthrene (Phe), anth racene (Ant) and benz[a]anthracene (BaA) in the DE group being significantly higher than the contro l group. In breast milk, Ant, fluoranthene (Flu), pyrene (Pyr), and chrysene (Chr) showed significant increases in the DE group compared to the control group. BaA tended to be about fourfold higher than the control group in breast milk, but the increase was not significant. PAHs in dam livers of DE versus control were not significantly different. The results of this study demonstrate th at PAHs derived from DE are transferred across the placenta from the DE-exposed dam to the fetus. Lacta tional transfer through the breast milk is also likely as PAHs were detected in dam breast milk, but this should be confirmed in future studies that cross foster control and exposed da ms and pups. The lipophilicity of the PAH based on its structure likely affected its uptake in the dam, as PAHs with 3 or 4 rings were found in maternal blood and PAHs with 5 or 6 rings were not detected.



### Heritable DNA Changes and Epigenetic Changes

Ambient Air

To address the role of ambient air exposure on heritable changes, Somers et al. (2004, 078098 ) exposed mice to ambient air in at a rural Canadian s ite or at an urban site near a steel mill. They showed that offspring of mice exposed to ambient air in urban regions inherited paternal-origin expanded simple tandem repeat (ESTR) mutations 1.9-2.1 times more frequently than offspring of mice exposed to HEPA filtered air or those exposed to rural ambient air. Mouse expanded simple tandem repeat (ESTR) DNA is composed of short base pair repeats which are unstable in the germline and tend to mutate by insertion or deletion of repeat units. In vivo and in situ studies have shown that murine ESTR loci are susceptible to ionizing radiation, and other environmental mutatgen-dependent germline mutations, and are thus good markers of exposure to environmental contaminants. Expanding upon the above work and to determine if PM or the gaseous phase of the urban air was responsible for heritable mutations, Yauk et al. (2008, 157164 ) exposed mature male C57Bl×CBA F1 hybrid mice to either HEPA-filtered air or to ambient air in Hamilton, Ontario, Canada for 3 or 10 wk, or 10 wk plus 6 wk of clean air exposure (16 wk). Sperm DNA was monitored for expanded simple tandem repeat (ESTR) mutations. In addition, male-germ line (spermatogonial stem cell) DNA methylation was m onitored post-exposure. This area in Hamilton is near two steel mills and a major highway. Air qua lity data provided by the Ontario Ministry of the Environment showed the highest concentrations of TSP and metals at week 4 (93.8 ± 17 and 3.6 ± 0.7 μg/m 3 , respectively) and PAH at week 3 (8.3 ± 1.7 ng/m 3 ). Mutation frequency at ESTR

December 2009 7-66Ms6-hm locus in sperm DNA from mice exposed 3 or 10 wk did not show elevated ESTR mutation frequencies, but there was a significant increase in ESTR mutation frequency at 16 wk in ambient air-exposed males versus HEPA filter-exposed animals, pointing to a PM-dependent mechanism of action. When compared to HEPA filter air-expos ed males, ambient air-exposed males manifested with hypermethylation of germ-line DNA at 10 and 16 wk. These PM-dependent epigenetic modifications (hypermethylation) were not seen in th e halploid stage (3 wk) of spermatogenesis, but were nonetheless seen in early stages of spermatogenesis (10 wk) and remained significantly elevated in mature sperm even after removal of the mouse from the environmental exposure (16 wk). Thus, these studies indicate that the ambient PM phase and not the gaseous phase is responsible for the increased frequency of heritable DNA mutations and epigenetic modifications.



## 7.4.3. Summary and Causal Determinations

7.4. 3.1. PM 2 .5 The 1996 PM AQCD concluded that while few studies had been conducted on the linkbetween PM and infant mortality, the research “suggested an association,” particularly for postneonates (U.S. EPA, 1996, 079380 ). In the 2004 PM AQCD, additional evidence was available on PM’s effect on fetal and early postnatal development and mortality and while some studies indicated a relationship between PM and pregnancy outcomes, others did not (U.S. EPA, 2004, 056905 ). Studies identifying associations found that exposure to PM 10 early during pregnancy (first month of pregnancy) or late in the pregnancy (6 wk prior to birth) were linked with higher risk of preterm birth, including models adjusted for other pollutants, and that PM 2.5 during the first month of pregnancy was associated with IUGR. However, ot her work did not identify relationships between
PM 10 exposure and low birth weight. The state of the science at that time, as indicated in the 2004 PM AQCD, was that the research provided mixed re sults based on studies from multiple countries. Building on the evidence characterized in the previous AQCDs, recent epidemiologic studies conducted in the U.S. and Europe were able to examine the effects of PM 2.5 , and all found an increased risk of low birth weight (Section 7.4.1 ). Exposure to PM 2.5 was usually associated with greater reductions in birth weight than exposure to PM 10 . All of the studies that examined the relationship between PM 2.5 and preterm birth report positive asso ciations, and most were statistically significant. The studies evaluating the association between PM 2.5 and growth restriction all found positive associations, with the strongest evid ence coming when exposure was assessed during the first or second trimester (Section 7.4.1 ). For infant mortality (<1 yr), several studies examined PM 2.5 and found positive associations (Section 7.4.1 ). Animal toxicological studies reported effect s including decreased uterine weight, limited evidence of male reproductive effects, and c onflicting reports of reproductive outcomes in male offspring, particularly in studies of DE (Section 7.4.2 ). Toxicological studies also reported effects for several development outcomes, including immunological effects (placental and related to asthma), neurodevelopmental and behavioral effects (Section 7.4.2 ). In summary evidence is accumulating from epidemiologic studies for effects on low birth weight and infant mortality, especially due to re spiratory causes during the post-neonatal period. The
mean PM 2.5 concentrations during the study periods ranged from 5.3-27.4 μg/m 3 . Exposure to PM 2.5 was usually associated with greater reducti ons in birth weight than exposure to PM 10 . Several U.S. studies of PM 10 investigating fetal growth reported 11-g decrements in birth weight associated with
PM 10 exposure. Most of these studies were conducted in California, where PM 2.5 and PM 10-2.5 contribute almost equally to the PM 10 mass concentration. So while these results can not be attributed to one size fraction or the other, the c onsistency of the results strengthens the interpretation that particle exposure may be causally related to reductions in birth weight. Similarly, animal evidence supported an association between PM 2.5 and PM 10 exposure and adverse reproductive and developmental outcomes, but provided little mechanisti c information or biological plausibility for an association between long-term PM exposure and adve rse birth outcomes, including low birth weight, or infant mortality. Epidemiologic studies do not consistently report associations between PM exposure and preterm birth, growth restriction, birth defects or decreased sperm quality. New evidence from animal toxicological studies on heritable mutations is of great interest, and warrants further investigation. Overall, the epid emiologic and toxicological evidence is 

### suggestive of a

December 2009 7-67

### causal relationship between long-term exposures to PM 2.5 

### and reproductive and developmental outcomes.



## 7.4. 3.2. PM 1 0-2.5

Evidence is 

### inadequate to determine if a causal relationship exists between longterm exposure to PM 10-2.5 

### and developmental and reproductive outcomes because studies have not been conducted in sufficient quantity or quality to draw any conclusion. A single study found an association between PM 10-2.5 and birthweight (-13 g [95% CI: -18.3 to -7.6] per 10 μ g/m 3 increase), but no such association for PM 2.5 (Parker et al., 2008, 156013 ).



## 7.4. 3.3. UFP s

The 2004 PM AQCD did not report long-term exposure studies for UFPs. No epidemiologic or animal toxicology studies have been conducted to evaluate the effects of long-term UFP exposure and reproductive and developmental effects. Ambi ent air exposures, which likely include UFPs, are reported in this ISA but there is no delineation of the separate contribution from UFPs. The evidence
is 

### inadequate to determine if a causal relationship exists between long-term UFP exposures and reproductive and developmental effects

.



# 7.5. Cancer, Mutagenicity, and Genotoxicity

Evidence from epidemiologic and animal toxi cological studies has been accumulating for more than three decades regarding the mutagenicity and carcinogenicity of PM in the ambient air. DE has been identified as one source of PM in ambi ent air, and has been extensively studied for its carcinogenic potential. In 1989, the International Agency for Research on Cancer (IARC) found that there was sufficient evidence that extracts of DE particles were carcinogenic in experimental animals and that there was limited evidence for the carc inogenic effect of DE in humans (IARC, 1989, 002958 ). This conclusion was based on studies in which organic extracts of DE particles were used to evaluate the effects of concentrates of the organic compounds associated with carbonaceous soot particles. These extracts were applied to the skin or administered by IT instillation or intrapulmonary implantation to mice, rats, or Syrian hamsters and an excess of tumors on the skin, lung or at the site of injection were observed. In 2002, the U.S. EPA reviewed over 30 epidemio logic studies that investigated the potential carcinogenicity of DE. These studies, on average, found that long-term occupational exposures to DE were associated with a 40% increase in th e relative risk of lung cancer (U.S. EPA, 2002, 042866 ). In the same report the U.S. EPA concluded that extensive studies with salmonella had unequivocally demonstrated mutagenic activity in both particulate and gaseous fractions of DE. They further concluded that DE may present a lung cancer hazard to humans (U.S. EPA, 2002, 042866 ). The particulate phase appeared to have the greatest contribution to the carcinogenic effect. Both the particle core and the associated organic compounds demonstrated carcinogenic properties, although a role for the gas-phase components of DE could not be ruled out. Almost the entire diesel particle mass is ≤ 10 μ m in diameter (PM 10 ), with approximately 94% of the mass of these particles <2.5 μ m in diameter (PM 2.5 ), including a subgroup with a large number of UFPs (U.S. EPA, 2002, 042866 ). U.S. EPA considered the weight of evidence for potential human carcinogenicity for DE to be strong, even though inferences were involved in the overall assessment, and concluded that DE is “likely to be carcinogenic to humans by inhalation” and that this hazard applies to environmental exposures (U.S. EPA, 2002, 042866 ). Two recent reviews of the mutagenicity (Claxton et al., 2004, 089008 ) and carcinogenicity (Claxton and Woodall, 2007, 180391 ) of ambient air have characterized the animal toxicological literature on ambient air pollution and cancer. The majority of these toxicological studies have been conducted using IT instillation or dermal routes of exposure. Generally, the toxicological evidence reviewed in this ISA has been limited to inhala tion studies conducted with lower concentrations of

December 2009 7-68PM (<2 mg/m 3 ), relevant to current ambient concentr ations and the current regulatory standard (Section 1.3). Because this ISA fo cuses on toxicological studies which use the inhalation route of exposure, it is possible that important evidence for the role of PM in mutagenicity, tumorigenicity, and/or carcinogenicity may be missed. In order to accurately characterize the relationship between PM and cancer and be consistent with the EPA Guidelines for Carcinogen Risk Assessment (U.S. EPA, 2005, 086237 ), these reviews (that include studies that employ IT instillation and dermal routes of exposure) are summarized briefly. Claxton et al. (2004, 089008 ) reviewed the mutagenicity of air in the Salmonella (Ames) assay, and showed that hundreds of compounds id entified in ambient air from varying chemical classes are mutagenic and that the commonly monitored PAHs could not account for the majority of mutagenicity associated with most ai rborne particles. They concluded that the smallest particles have the highest toxicity per particulate mass, with the PM 2.5 size fraction having greater mutagenic and cytotoxic potential than the PM 10 size fraction, which had a higher mutagenic potential than the TSP size fraction. One study reviewed by Claxton et al. (2004, 089008 ) found that the cytotoxic potential of PM 2.5 was higher in wintertime samples than in summertime samples. A series of studies on source apportionment for ambient particle mutagenic activity reviewed by Claxton et al. (2004, 089008 ) indicate that mobile sources (cars and diesel trucks) account for most of the mutagenic activity. Claxton and Woodall (2007, 180391 ) reviewed many studies that examined the rodent carcinogenicity of extracts of ambient PM sample s; the PM was of various size classes, often from TSP samples. Among a variety of mouse and rat strains, application methods, and samples employed, the authors found no pattern that would suggest the routine use of a particular strain or protocol would be more informative than another. The primary conclusion that comes from the analysis of rodent carcinogenicity studies is that the most polluted urban air samples tested to date are carcinogenic; the contribution of PM and different size classes of PM to the carcinogenic effects of ambient air has not been delineated. The differences in response by the various strains of inbred mice indicate that the genetic background of an individual can influence tumorigenic response. Studies examining different components of ambient PM (e.g., PAHs) confirm that ambient air contains multiple carcinogens, and that the carcinog enic potential of particles from different airsheds can be quite different. Therefore, one would expect the incidence of cancers related to ambient air exposure in different metropolitan areas to differ. Numerous epidemiologic and animal toxicological studies of ambient PM and their contributing sources have been conducted to assess the relative mutagenic or genotoxic potential. Studies previously reviewed in the 2004 PM AQCD (U.S. EPA, 2004, 056905 ) provide evidence that ambient PM as well as PM from specific combusti on sources (e.g., fossil fuels) is mutagenic in vivo and in vitro. Building on these results, data from recent epidemiologic and animal toxicological studies that evaluated the carcinogenic, mutageni c and/or genotoxic effects of PM, PM-constituents, and combustion emission source particles are reviewed in this section.



## 7.5.1. Epidemiologic Studies

The 2004 PM AQCD reported on original and follow-up analyses for three prospective cohort studies that examined the relationship between PM and lung cancer incidence and mortality. Based on these findings, as well as on the results from case-control and ecologic studies, the 2004 PM AQCD concluded that long-term PM exposure may in crease the risk of lung cancer incidence and mortality. The largest of the three prospective cohort studies included in the 2004 PM AQCD was the ACS study (Pope et al., 2002, 024689 ). This study was the follow-up to the original ACS study (Pope et al., 1995, 045159 ), and included a longer follow-up period and reported a statistically significant association between PM 2.5 exposure and lung cancer mortality. A 14- to 16-yr prospective study conducted using the Six Cities Study cohort reported a slightly elevated risk of lung cancer mortality for individuals living in the most polluted city (mean
PM 10 : 46.5 μg/m 3 ; mean PM 2.5 29.6 μg/m 3 ) as compared to the least polluted city (mean PM 10 : 18.2 μg/m 3 ; mean PM 2.5 11.0 μg/m 3 ) but the association was not statistically significant (Dockery et al., 1993, 044457 ). Re-analysis of the AHSMOG cohort, a study of non-smoking whites living in California, concluded that elevated long-term exposure to PM 10 was associated with lung cancer incidence among both men and women (Beeson et al., 1998, 048890 ). The original study had reported an excess of incident lung cancers on ly among women (Abbey et al., 1991, 042668 ). Further reanalysis

December 2009 7-69of this cohort revealed an association between PM 10 and lung cancer mortality among men but no association among women (Abbey et al., 1999, 047559 ). In addition, McDonnell et al. (2000, 010319 ) reported increases in lung cancer mo rtality with long-term exposure to PM 2.5 in the AHSMOG cohort; no association was seen for PM 10-2.5

.



## 7.5. 1.1. Lun g Cancer Mortality and Incidence

The following sections will examine extensions of the above mentioned cohort studies and new studies published since the 2004 PM AQCD. The section includes discussion of both lung cancer incidence and mortality, as well as markers of exposure/susceptibility. A summary of the mean PM concentrations reported for the new studies is presented in Table 7-6 . In addition, a summary of the associations for lung cancer mortality and incidence are presented in Table 7-7 and Figure 7-7 (Section 7.6 ) Further discussion of all-cause and cau se-specific mortality is presented in Section 7.6

.



### Table 7-6. Characterization of am bient PM concentrations from r ecent studies of cancer and longterm exposures to PM.

Study Location Pollutant Mean Annual Concentration (μg/m 3) Upper Percentile Concentrations (μg/m 3)

Brunekreef et al. (2009, 191947 ) The Netherlands PM 2.5 28.3 Max: 36.8

Bonner et al. (2005, 088993 ) Western NY State TSP 44

Jerret et al. (2005, 087600 ) Los Angeles, California PM 2.5 Max:27.1

Laden et al. (2006, 087605 ) 6 U.S. cities PM 2.5 Range of means across sites: 10.2-29.0

Avg of means across sites: 16.4

PM 2.5 15 Max: 22 Naess et al. (2007, 090736 ) Oslo, Norway PM 10 19 Max: 30

Palli et al. (2008, 156837 ) Florence, Italy PM 10 NR

PM 2.5 Max: 46-120 Pedersen et al. (2006, 156848 ) Czech Republic PM 10 Max: 120-238.6

Sorensen et al. (2005, 083053 ) Copenhagen, Denmark PM 2.5 Range of means across sites: 12.6-20.7

Avg of means across sites: 16.7 75th: 24.3-27.7

PM 10 Max: 55 Sram et al. (2007, 188457 ) Czech Republic PM 2.5 Max: 38

PM 10 Range of means across sites: 36.4-55.6

Avg of means across sites: 46.0 Sram et al. (2007, 192084 ) Czech Republic

PM 2.5 Range of means across sites: 24.8-44.4

Avg of means across sites: 34.6

Vineis et al. (2006, 192089 ) Multi-city, Europe PM 10 Range of means across sites: 19.9-73.4

Avg of means across sites: 35.4

Vinzents et al. (2005, 087482 ) Copenhagen, Denmark PM 10 Range of means across sites: 16.9-23.5

Avg of means across sites: 20.2

A subset of the ACS cohort study from 1982 to 2000 that included only residents of Los Angeles, California was used to examine the association between PM 2.5 and lung cancer mortality while adjusting for both individual and ne ighborhood covariates (Jerrett et al., 2005, 087600 ). There was a positive association between PM 2.5 and lung cancer mortality when adjusting for 44 individual covariates (RR 1.44 [95% CI: 0.98-2.11] per 10 μg/m 3 increase in PM 2.5 ). However, including all potential individual and neighborhood covariates as sociated with mortality reduced the association

December 2009 7-70(RR 1.20 [95% CI: 0.79-1.82] per 10 μg/m 3 increase in PM 2.5 ). A recent re-analysis of the full ACS cohort also demonstrated a positive association between PM 2.5 and lung cancer mortality (RR 1.11 [95% CI: 1.04-1.18]) (Krewski et al., 2009, 191193 ). The authors observed modification of this risk by educational attainment, with those completing a high school degree or less having greater risk. In addition to utilizing the ACS cohort for a natio nwide analysis, this same study conducted two regional assessments, one in the New York City area and the other in the Los Angeles area. No association was detected between PM 2.5 and lung cancer mortality in the analysis of the region included in the New York City analysis. A positive association was observed in the Los Angeles-area analysis using an unadjusted model, but this association did not persist after control for individual, ecologic, and copollutant covariates. The Six Cities Study was extended to include data from 1990-1998, a period including 1,368 deaths and 54,735 person-years (Laden et al., 2006, 087605 ). An elevated risk ratio for lung cancer mortality was reported when the entire follow-up period (1974-1998) was included in the analysis (RR 1.27 [95% CI 0.96-1.69] per 10 μg/m 3 increase in average annual PM 2.5 ). However, estimated decreases in PM 2.5 were not associated with reduced lung cancer mortality (RR 1.06 (95% CI: 0.43-2.62] for every 10 μg/m 3 reduction in PM 2.5 ). Naess et al. (2007, 090736 ) studied individuals aged 51-90 yr living in Oslo, Norway in 1992. Death certificate data were obtained for 1992-1 998 and information on PM was collected from 1992-1995. Women had a larger association of lung cancer mortality with PM 2.5 compared to men. Similar results were reported for PM 10

. Most recently, Brunekr eef et al. (2009, 191947 ) used the Netherlands cohort study (NLCS) on diet and cancer to conduct a re-analysis of the research performed by Beelen et al. (2008, 156263 ) examining the association between PM and both lung cancer mortality and incidence. After 10 yr of follow-up, there was no association between PM 2.5 and lung cancer mortality for either the analysis of the full cohort (n = 105,296) (RR 1.06 [95% CI: 0.82-1.38] per 10 μg/m 3 increase in PM 2.5 ) or the case-cohort (n = 4,075) (RR 0.87 [95% CI: 0.52-1. 47]). There was also no association with black smoke or traffic density variables, although living near a major roadway was associated with an elevated relative risk for lung cancer in the full cohort analysis (RR 1.20 [95% CI: 0.98-1.47]). The association was not present in the case-cohort analysis (RR 1.07 [95% CI: 0.70-1.64]). In addition to lung cancer mortality, Brunekreef et al. (2009, 191947 ) also examined the association with lung cancer incidence using 11.3 yr of follow-up data. In both the full cohort and the case-cohort analyses no association was reported between PM 2.5 and lung cancer incidence (full cohort: RR 0.81 [95% CI: 0.63-1.04]; case-c ohort: RR 0.67 [95% CI: 0.41-1.10] per 10 μg/m 3

increase in PM 2.5 ). The same was true for analyses of BS and traffic density variables. The association between PM and incident lung cancers was examined in the European Prospective Investigation into Cancer and Nutrition study (EPIC) (Vineis et al., 2006, 192089 ). Within this cohort, a nested case-control study, th e GenAir study, included cases of incident cancer and controls matched on age, gender, smoking status, country of recruitment, and time between recruitment and diagnosis. Only non-smokers and former smokers who had quit smoking at least 10 yr prior were included. The study included 113 cases and 312 controls. No association was seen between PM 10 and lung cancer (OR 0.91 [95% CI: 0.70-1.18] per 10 μg/m 3 ). The OR was elevated when cotinine, a marker for cigarette exposure, was included in the model but the authors state that this is probably due to small study size (OR 2.85 [95% CI: 0.97-8.33] comparing ≥ 11 μg/m 3 to <11 μg/m 3 ). Control for other potential confounders, such as BMI, education level, and intake of fruit and vegetables, did not have a large impact on the estimate.

December 2009 7-71

### Table 7-7. Associations* between ambient PM con centrations from select studies of lung cancer mortality and incidence.

Study Cohort Location Years Analysis subgroup Effect Estimate (95% CI)

MORTALITY

- PM 2.5 Dockery et al. (1993, 044457 ) ‡ Six-Cities Six cities across the U.S. 1974-1991 1.18 (0.89-1.57)

Krewski et al. (2000, 012281 ) ‡ Six-Cities-Re-analysis Six cities across the U.S. 1974-1991 1.16 (0.86-1.23)

Laden et al. (2006, 087605 ) Six-Cities Six cities across the U.S. 1974-1998 1.27 (0.96-1.69)

Beelen et al. (2008, 156263 ) NLCS Netherlands 1987-1996 Full Cohort 1.06 (0.82-1.38)

Beelen et al. (2008, 156263 ) NLCS Netherlands 1987-1996 Case Cohort 0.87 (0.52-1.47)

Brunekreef et al. (2009, 191947 ) NLCS-Re-analysis Netherlands 1987-1996 Full Cohort 1.06 (0.82-1.38)

Brunekreef et al. (2009, 191947 ) NLCS-Re-analysis Netherlands 1987-1996 Case Cohort 0.87 (0.52-1.47)

Pope et al. (1995, 045159 ) ‡ ACS U.S. 1982-1989 1.01 (0.91-1.12)

Pope et al. (2002, 024689 ) ‡ ACS U.S. 1982-2000 1.13 (1.04-1.22)

Jerret et al. (2005, 087600 ) ACS-LA Los Angeles 1982-2000 Intra-metro Los Angeles 1.44 (0.98-2.11)

Krewski et al. (2009, 191193 ) ACS-Re-analysis U.S. 1982-2000 1.11 (1.04-1.18)

Krewski et al. (2009, 191193 ) ACS-Re-analysis New York City 1982-2000 Intra-metro New York City 0.90 (0.29-2.78)

Krewski et al. (2009, 191193 ) ACS-Re-analysis Los Angeles 1982-2000 Intra-metro Los Angeles 1.31 (0.90-1.92)

McDonnell et al. (2000, 010319 ) † AHSMOG California 1973-1977 Men 1.39 (0.79-2.46)

Naess et al. (2007, 090736 ) Oslo, Norway 1992-1998 Men, 51-70 yrs 1.18 (0.93-1.52)

Naess et al. (2007, 090736 ) Oslo, Norway 1992-1998 Men, 71-90 yrs 1.18 (0.93-1.52)

Naess et al. (2007, 090736 ) Oslo, Norway 1992-1998 Women, 51-70 yrs 1.83 (1.36-2.47)

Naess et al. (2007, 090736 ) Oslo, Norway 1992-1998 Women, 71-90 yrs 1.45 (1.05-2.02)

MORTALITY

- PM 10 McDonnell et al. (2000, 010319 ) † AHSMOG California 1973-1977 Men 1.23 (0.84-1.80)

Naess et al. (2007, 090736 ) † Oslo, Norway 1992-1998 Men, 51-70 yrs 1.12 (0.95-1.33)

Naess et al. (2007, 090736 ) Oslo, Norway 1992-1998 Men, 71-90 yrs 1.14 (0.97-1.36)

Naess et al. (2007, 090736 ) † Oslo, Norway 1992-1998 Women, 51-70 yrs 1.50 (1.23-1.84)

Naess et al. (2007, 090736 ) † Oslo, Norway 1992-1998 Women, 71-90 yrs 1.29 (1.03-1.60)

INCIDENCE

- P

M

2.5 Beelen et al. (2008, 155681 ) NLCS Netherlands 1987-1996 Full Cohort 0.81 (0.63-1.04)

Beelen et al. (2008, 155681 ) NLCS Netherlands 1987-1996 Case Cohort 0.65 (0.41-1.04)

Brunekreef et al. (2009, 191947 ) NLCS-Re-analysis Netherlands 1987-1996 Full Cohort 0.81 (0.63-1.04)

Brunekreef et al. (2009, 191947 ) NLCS-Re-analysis Netherlands 1987-1996 Case Cohort 0.67 (0.41-1.10)

INCIDENCE – PM 10 Beeson et al. (1998, 048890 ) AHSMOG California 1977-1992 Men 1.99 (1.32-3.00)

Vineis et al. (2006, 192089 ) GenAir Europe 1993-1999 Case-Control 0.91 (0.70-1.18)

*per 10 μg/m 3 increase †Results from the paper were standardized to10 μg/m 3 [For McDonnell et al. (2000, 010319 ) the non-standardized results were reported based on IQR increments (24.3 μg/m 3 for PM 2.5 and 29.5 μg/m 3 for

PM 10 ). For Naess et al. (2007, 090736 ) the original hazard ratios were calculated based on quartiles of PM exposure. The results were converted to10 μg/m 3 using the mean range of the four quartiles (3.95 μg/m 3 for PM 2.5 and 5.88 μg/m 3 for PM 10 )]. ‡Study was included in the 2004 PM AQCD

December 2009 7-72

## 7.5. 1.2. Oth er Cancers

Bonner et al. (2005, 088993 ) conducted a population-based, case-control study of the association between ambient exposure to PAHs in early life and breast cancer incidence among women living in Erie and Niagara counties in th e state of New York. Cases (n = 1,166 of which 841 were post-menopausal) were women with primary br east cancer, and controls (n = 2,105 of which 1,495 were post-menopausal) were frequency matc hed to the cases by age, race, and county of residence. TSP was used as a proxy for PAH exposure. Annual average TSP concentrations (1959-1997) were obtained from the New York State Department of Environmental Conservation for Erie and Niagara Counties. Among postmenopausal wo men, exposure to high concentrations of TSP (>140 μg/m 3 ) at birth was associated with an OR of 2.42 for breast cancer (95% CI: 0.97-6.09) relative to low concentrations of TSP (<84 μg/m 3 ). ORs were elevated for pollution exposures at age of menarche (OR: 1.45 [95% CI: 0.74-2.87]) and ag e at first birth (OR: 1.33 [95% CI: 0.87-2.06]) among postmenopausal women. Among premenopausa l women, exposure to high concentrations of TSP at birth was associated with an OR for breast cancer incidence of 1.79 (95% CI: 0.62-5.10) relative to low exposure levels, exposure at age of menarche was associated with an OR of 0.66 (95% CI: 0.38-1.16), and exposure at age of first birth was associated with an OR of 0.52 (95% CI: 0.22-1.20).



## 7.5. 1.3. Mar kers of Exposure or Susceptibility

Several studies looked at markers of exposure or susceptibility as the outcome associated with short-term exposure. These studies are included here because they may be relevant to the mechanism that leads to cancer associated with long-term exposures. For example, inflammation can contribute to carcinogenesis by inducing genomic instability, wh ich can then lead to altered gene expression, enhanced proliferation, and resistance to apopt otic signals. Reactive oxygen and nitrogen species, provided by PM components or inflammation path ways, can cause molecular damage leading to cellular transformation. Elevated inflammatory cytokines, chemokines, and prostaglandins promote tumor growth and angiogenesis, which in turn promotes metastasis and malignant invasion. In particular, IL-6, IL-8, IL-1 β , COX-2, and TNF- α have been implicated in these processes (Kundu and Surh, 2008, 198840 ). Several lines of evidence support the involvement of COX-2 in the pathogenesis of lung cancer (Lee et al., 2008, 198811 ). Both short- and long-term exposure studies demonstrate relationships between various forms of PM and increased production of these inflammatory mediators, both in the lungs and ci rculation. Additionally, limited evidence suggests that exposure to PM (Chen and Schwartz, 2008, 190106 ), or traffic (Williams et al., 2009, 191945 ), or residence in a polluted airshed (Calderon-Garciduenas et al., 2007, 091252 ; CalderónGarcidueñas et al., 2009, 192107 ) are associated with decreases in the number or function of natural killer cells or other white blood cells, indicating suppression of anti-tumor defenses. A study performed in the Czech Republic compared 53 male policemen working at least 8 hours per day outdoors in urban air with age- and sex-matched controls who spent at least 90% of their day indoors (n = 52) (Sram et al., 2007, 188457 ). During the sampling period, two monitors from downtown and suburban areas detected levels of air pollutants in the following ranges: PM 10 32-55 μg/m 3 , PM 2.5 27-38 μg/m 3 , c-PAHs 18-22 ng/m 3 , and B[a]P 2.5-3.1 ng/m 3 using a VAPS monitor (measurements taken with a HiVol monitor, which has a lower flow rate, had a mean for
PM 10 of 62.6 μg/m 3 ). c-PAHs detected on personal monitors during sampling days had a mean of 12.04 ng/m 3 among the policemen and 6.17 ng/m 3 among the controls. No difference in percent of chromosomal aberrations was observed between th e policemen and control group using conventional cytogenetic analysis. However, using fluorescent in situ hybridization (FISH), a difference in chromosomal aberrations between the policemen a nd control group was reported. For example, the percentage of aberrant cells, as well as the ge nomic frequency of translocations per 100 cells, was about 1.4-fold greater in the policemen. This was largely driven by a difference in chromosomal aberrations between nonsmoking policemen and nonsmoking controls. A similar study that included only the policemen (n = 60), reported that the mean exposure to c-PAHs and B[a]P for 40-50 days before sampling was associated with chromosomal aberrations when analyzed with FISH (Sram et al., 2007, 192084 ). However, when included in a model with other covariates, the association with these variables was null. No association was present with use of conventional cytogenetic analysis. Palli et al. (2008, 156837 ) investigated the correlation between ambient PM 10 concentrations and individual levels of DNA bulky adducts. Study participants were 214 healthy adults aged

December 2009 7-7335-64 yr at enrollment who resided in the city of Florence, Italy. This study was conducted between 1993 and 1998. PM 10 exposure levels were based on daily environmental measures provided by two types of urban monitoring stations (high-traffic and low-traffic). The researchers assessed correlation between DNA bulky adducts measured in blood samples and PM 10 concentrations prior to blood sample collection. Time windows of PM 10 exposure evaluated in this study were 0-5 days, 0-10 days, 0-15 days, 0-30 days, 0-60 days, and 0-90 days prior to blood sample collection. Overall, average
PM 10 concentrations decreased during the study period, with some fluctuations. Quantitative values were not reported, but PM 10 appeared to range between approximately 30 and 100 μg/m 3 for hightraffic stations, and between approximately 20 and 50 μg/m 3 for low-traffic stations. This study found that levels of DNA bulky adducts among non-smoking workers with occupational traffic exposure were positively correlated with cumulative PM 10 levels from high-traffic stations during approximately 2 wk preceding blood sample collect ion (0-5 days: r = 0.55, p = 0.03; 0-10 days: r = 0.58, p = 0.02; 0-15 days: r = 0.56, p = 0.02) . DNA bulky adducts were not associated with PM 10 levels among Florence residents with no occ upational exposure to vehicle emissions or among smokers. DNA bulky adducts were not associated with PM 10 levels assessed by low-traffic urban monitoring stations. The association between personal exposure to water-soluble transition metals in PM 2.5 and oxidative stress-induced DNA damage was investigated among 49 students from Central Copenhagen, Denmark (Sorensen et al., 2005, 083053 ). Researchers assessed PM 2.5 exposure by personal sampling over two weekday periods twic e in one year (November 1999 and August 2000), and determined the concentration of water-soluble transition metals (V, Cr, Fe, Ni, Cu and Pt) in these samples. In addition, lymphocyte and 24-h urine samples were analyzed for DNA damage by measuring 7-hydro-8-oxo-2’-deoxyguanosine (8-oxod G). Mean concentrations and corresponding IQR of these metals differed between months of sample collection. This study found that 8-oxodG concentration in lymphocytes was significantly associated with V and Cr concentrations, with a 1.9% increase in 8-oxodG per 1 μg/L increase in V concentration and a 2.2% increase in 8-oxodG per 1 μg/L increase in Cr concentration; these associations were independent of the PM 2.5 mass concentration. The other transition metals were not significantly associated with the 8-oxodG concentration in lymphocytes, and none of the si x measured transition metals was associated with the 8-oxodG concentration in urine. Vinzents et al. (2005, 087482 ) investigated the association between UFP and PM 10 concentrations with levels of purine oxidation and strand breaks in DNA using a crossover design in Copenhagen, Denmark. Study participants were 15 healthy nonsmoking individuals with a mean age of 25 yr. UFP exposure was evaluated using number concentration obtained in the breathing zone by portable instruments in six 18-h weekday periods from March to June 2003. Ambient concentrations for PM 10 and UFP were also measured on all exposure da ys at curbside street stations and at one urban background station. Oxidative DNA damage was assessed by evaluating strand breaks and oxidized purines in mononuclear cells isolated from venous blood the morning after exposure measurement. Mean number concentration of UFPs (street station) was 30.4×10 3 UFPs/mL (standard deviation [SD]: 1.38), mean mass concentration of PM 10 at a background monitoring station was 16.9 μg/m 3 (SD: 1.53), and mean mass concentration of PM 10 at a street station was 23.5 μg/m 3 (SD: 1.48). Mean personal exposure to UFPs was 32.4×10 3 UFPs/mL (SD: 1.49) while bicycling (5 occasions), 19.6×10 3 UFPs/mL (SD: 1.78) during other out door activities (6 occasions), and 13.4×10 3 UFPs/mL (SD: 1.96) while indoors (6 occasions). The regression coefficients of the mixedeffects models looking at level of purine oxidation were estimated as 1.50×10 − 3 (95% CI: 0.59×10 − 3

to 2.42×10 − 3 ; p = 0.002) for cumulative outdoor exposure and 1.07×10 − 3 (95% CI: 0.37×10 − 3 to 1.77×10 − 3 ; p = 0.003) for cumulative indoor exposure. Neither cumulative outdoor nor cumulative indoor exposures to UFPs were associated with st rand breaks. Neither ambient air concentrations of
PM 10 nor number concentrations of UFPs at monitoring stations were significant predictors of DNA damage. Additionally, a number of studies employed ecologic study designs, comparing the prevalence of biomarkers in populations from more polluted lo cations to those in less polluted locations. In a pilot study conducted in the Czech Republic (Pedersen et al., 2006, 156848 ), children age 5-11 yr provided 5 mL blood samples and the frequency of micronuclei (MN) in peripheral blood lymphocytes was analyzed as a measure of cytoge netic effects. Significantly higher frequencies of MN were found in younger children living in Teplice (PM 2.5 concentration = 120 μg/m 3 ) than in Prachatice (PM 2.5 concentration = 46 μg/m 3 ). The levels of c-PAHs were also much higher in Teplice (nearly 30 ng/m 3 in Teplice and about 15 ng/m 3 in Prachatice). The difference in MN frequencies

December 2009 7-74observed in the children from the two locations may be attributable to differences in exposure to air pollution, but could also be due to differences in diet or other environmental exposures. This finding is noteworthy considering MN formation in peripheral blood lymphocytes is thought to be biologically relevant for carcinogenesis. Avogbe et al. (2005, 087811 ) showed a correlation between the level of oxidative DNA damage in individuals and exposure to ambient UFPs. Formamidopyrimidine DNA glycosylase sensitive sites and the presence of DNA strand br eaks were assessed in blood and urine samples obtained from healthy, non-smoking male volunteers that lived and worked in different areas of Cotonou, Benin. Exposure to benzene was assesse d by urinary excretion of S-phenylmercapturic acid. There was a high degree of correlation between exposure to benzene and UFPs and the presence of DNA strand breaks and formamidopyr imidine DNA glycosylase sensitive sites (rural subjects < suburban subjects < residents living near high traffic roads < taxi drivers). Genotyping studies showed that the magnitude of the effect s of benzene and UFPs may be modified by polymorphisms in GSTP1 and NQO1 genes. Tovalin et al. (2006, 091322 ) evaluated the association betwee n exposure to air pollutants and the level of DNA damage using the single cell ge l electrophoresis (comet) assay. Mononuclear lymphocytes from outdoor and indoor workers from tw o areas in Mexico, Mexico City (large city) and Puebla (medium size city), were evaluated. The outcomes showed that the outdoor workers in Mexico City exhibited greater DNA damage than in door workers in the same region. Similar levels of DNA damage were observed between indoor and outdoor workers in Puebla. The level of observed DNA damage was correlated with exposure to O 3 and PM 2.5

. In summary, several recent studies have report ed an association between lung cancer mortality and long-term PM 2.5 exposure. Although many of the estimates include the null in the confidence interval, overall the results have shown a positive relationship. The two recent studies that looked at lung cancer incidence did not report an association with PM 2.5 (Brunekreef et al., 2009, 191947 ) or
PM 10 (Vineis et al., 2006, 192089 ). Studies of exposure/susceptibility markers have reported inconsistent outcomes, with some markers be ing associated with PM and others not.



## 7.5.2. Toxicological Studies

Over the past 30 yr numerous mutagenicity and genotoxicity studies of ambient PM and their contributing sources have been conducted to assess the relative mutagenic or genotoxic potential. Studies previously reviewed in the 2004 PM AQCD (U.S. EPA, 2004, 056905 ) provide compelling evidence that ambient PM and PM from specifi c combustion sources (e.g., fossil fuels) are mutagenic in vivo and in vitro. Research cite d in the 2004 AQCD demonstrated mutagenic activity of ambient PM from urban centers in California, Germany and the Netherlands. These studies suggested that ubiquitous emission sources, particularly motor vehicle emissions, rather than isolated point sources were largely responsible for the mutageni c effects. In addition, the mutagenicity was dependent upon the chemical composition of the PM with unsubstituted polyaromatic compounds and semi-polar compounds being highly mutagenic. Mutagenicity was also dependent on size, with the fine fraction of urban PM having greater effects than the coarse fraction. Genotoxic activity was demonstrated for ambient PM from two high traffi c areas (one upwind and one downwind) and a rural site. In addition, the 2004 AQCD reported that exhausts from gasoline and diesel engines were mutagenic and that DE was more potent. More mu tagenicity was observed for exhaust from cold starts than starts at room temperature. Both gase ous and particulate fractions of DE were found to be mutagenic. Sequential fractionation of extracts from gasoline and DE implicated the polar fractions, especially nitrated polynuclear aromatic compounds, as contributing greatly to mutagenicity. Among some of the other mutagenically active compounds found in the gas phase of DE are ethylene, benzene, 1,3-butadiene, acrolein and several PAHs, all of which are also present in gasoline exhaust. Also cited in the 2004 AQCD were studies demons trating mutagenic effects of emissions from wood/biomass burning, which were primarily attr ibutable to the organic fraction and not the condensate. It was noted that wood smoke induced both frameshift mutations and base pair substitution but not DNA adducts. Further, emissions from coal combustion in China were found to be mutagenic, with both polar and aromatic fract ions contributing to effects. Little data were available on the mutagenicity of coal fly ash emissions from U.S. conventional combustion plants. In conclusion, these studies provide evidence that ambient PM and combustion-derived PM are mutagenic/genotoxic. The 2004 AQCD noted that there is not a simple relationship between

December 2009 7-75mutagenic potential and carcinogenic potential inanimals or humans. No studies evaluating carcinogenic effects of PM were reported in the 2004 AQCD. Building on results of earlier studies in the 2004 PM AQCD, data from newly published studies that evaluated the mutagenic, genotoxic a nd carcinogenic effects of PM, PM-constituents, and combustion emission source partic les are reviewed. Pertinent studies are described briefly in the following paragraphs. A summary table is pr ovided in Annex D, Tables D7 and D8).



## 7.5. 2.1. Mut agenesis and Genotoxicity

In Vitro Studies

In general, studies have focused on PM a nd PM extracts for mutagenicity testing using bacteria and mammalian cell lines. PM and/or PM extracts from ambient air samples, wood smoke, and coal, diesel, or gasoline combustion have all been reported to induce mutation in S

. typhimurium and in cultured human cells (Abou et al., 2007, 098819 ; Gabelová et al., 2007, 156457 ; Gabelová et al., 2007, 156458 ; Hannigan et al., 1997, 083598 ; Hornberg et al., 1998, 095741 ). In addition, effects associated with PM and PM-associated constitue nts include induction of MN formation, DNA adduct formation, SCE, DNA strand breaks, frameshi fts and inhibition of gap-junction intercellular communication (Alink et al., 1998, 087159 ; Arlt et al., 2007, 097257 ; Avogbe et al., 2005, 087811 ; Gabelová et al., 2007, 156457 ; Gabelová et al., 2007, 156458 ; Healey et al., 2006, 156532 ; Hornberg et al., 1996, 087164 ; Hornberg et al., 1998, 095741 ; Sevastyanova et al., 2007, 156969 ). Constituents adsorbed onto individual particles play a large role in the genotoxic potential of PM. Poma et al. (2006, 096903 ) showed that fine CB particles were consistently less genotoxic than similar concentrations of PM 2.5 extracts, suggesting that the adsorbed components play a role in the genotoxic potential of PM. Total PAH and carci nogenic PAH content were correlated with the genotoxic effects of PM (De Kok et al., 2005, 088656 ; Sevastyanova et al., 2007, 156969 ). Comparison of different extracts (water-soluble ve rsus organic) by Gutierrez-Castillo et al. (2006, 089030 ) indicated that water-soluble extracts were mo re genotoxic than the corresponding organic extracts. Sharma et al. (2007, 156975 ) reported that mutagenic activity of extracted PM samples collected in and around a waste incineration plant was attributed to the moderately polar and polar fractions. The polar and crude fractions were muta genic without metabolic activation, suggesting a direct mutagenic effect. No mutagenic activity was observed from any of the nonpolar samples evaluated. Arlt and colleagues (2007, 097257 ) have shown that the PM constituents 2-nitrobenzanthrone (2-NB) and 3-nitrobenzanthrone were genotoxic in a variety of bacterial and mammalian cell systems. Conflicting data have been reported on the role of metabolic enzymes in the genotoxicity of PM and their adsorbed constituents. Arlt et al. (2007, 097257 ) reported that the PM constituent 2-NB was genotoxic in bacterial and mammalian cells. However, metabolic activation with the human Nacetyltransferase 2 or sulfotransferase (SULT1A1) en zyme was needed for the effect to be observed in human cells. Erdinger et al. (2005, 156423 ) demonstrated that mutagenic activity was not affected when metabolism was induced. de Kok et al. (2005, 088656 ) evaluated the relationship between the physical, chemical, and genotoxic effects of ambient PM. TSP, PM 10 , and PM 2.5 were sampled at different locations and the extracts were assessed for mutagenicity and induction of DNA adducts in cells. Overall, induction of rat liver S9 metabolism generally reduced the mutagenic potential via the Ames assay of the particle fractions and DNA r eactivity (induction of DNA adducts) was generally higher after metabolic activat ion. Binková et al. (2003, 156274 ) found that the addition of S9 increased PM 10 -dependent DNA adduct formation.

Ambient Air

A limited number of studies evaluated the impact of the season on the genotoxic effects of ambient PM. A few studies have indicated that greater genotoxic effects were associated with samples collected during the winter months compared to those collected in the summer (Abou et al., 2007, 098819 ; Gabelová et al., 2007, 156457 ; Gabelová et al., 2007, 156458 ). In contrast, Hannigan et al. (1997, 083598 ) indicated that no seasonal variation was observed. Studies have also shown that greater genotoxic effects were associated with smaller particle size extracts (e.g., PM 2.5 >PM 10 ) and

December 2009 7-76from samples collected in urban areas or closer to higher trafficked areas (Abou et al., 2007, 098819 ; Hornberg et al., 1998, 095741 ). de Kok et al. (2005, 088656 ) found the direct mutagenicity (Ames assay) and the direct DNA reactivity (DNA adduct formation) of the PM 2.5 size fraction was significantly higher than that of the larger size fractions (TSP, PM 10 ) at most locations. DNA damage was assessed by the Comet assay in A549 cells exposed to PM collected from a high traffic area in Copenhagen, Denmark (TSP approximately 30 μg/m 3 ) and compared to the results from exposure of A549 cells to standard reference materials (SRM1650 or SRM2975) at the same concentrations (2.5-250 μg/ml) (Danielsen et al., 2008, 192092 ). All three particles induced strand breaks and oxidized purines in a dose-dependent manner and there were no obvious differences in potency. In contrast, only the ambient PM formed 8-oxodG when incubated with calf thymus DNA , which may be due to the concentration of transition metals.

Diesel and Gasoline Exhaust

Automobile DE particles (A-D E particles) was tested in S. typhimurium strains TA98, TA100, and its derivatives (e.g., TA98NR and YG1021) and found to be more mutagenic than forklift DE particles (f-DE particles, derivative SRM2975), ba sed on PM mass. A-DE particles had 227 times more PAH-type mutagenic activity and 8-45 ti mes more nitroarene-type mutagenic activity (DeMarini et al., 2004, 066329 ). Using a diesel engine without an oxidation catalytic converter (OCC), the diesel engine exhaust particle extract produced the highest number of revertant colonies in strains TA98 and TA100 with and without S9 at several tested loads when compared to extracts from low-sulfur diesel fuel (LSDF), rapeseed oil methyl ester (RME), and soybean oil methyl ester (SME). When an OCC was installed in the exhaust pipe of the engine, all extracts reduced the number of revertant colonies in both strains with and without S9 at partial loads but increased the number of revertant colonies without S9 at rated power. At idling, DE particles extracts increased the number of revertant colonies with and without S9 (Bunger et al., 2006, 156303 ). In a separate study, engine emissions (particle extracts and condensates) from rapeseed (canola) oil were found to produce greater mutagenic effects in S typhimurium strains TA98 and TA100 than DE particles (Bunger et al., 2007, 156304 ). Additionally, DE extract (DEE) from diesel fuel containing various percentages of ethanol was also observed to induc e mutational response in two Salmonella strains. Base diesel fuel DEE and DEE from fuel with 20% ethanol caused more significant DNA damage in rat fibrocytes L-929 cells than extracts containing 5, 10, or 15% ethanol (Song et al., 2007, 155306 ). DE and gasoline engine exhaust particles, as well as their semi-volatile organic compound (SVOC) extracts, induced mutations in the two S. typhimurium strains YG1024 and YG1029 in the absence and presence of S9; the PM extracts were more mutagenic than the SVOC extracts. Additionally, all extracts except the DE extract induced DNA damage and MN formation in Chinese hamster lung V79 cells (Liu et al., 2005, 097019 ). Another study demonstrated that gasoline engine exhaust significantly increased colony formation in TA98 with and without S9 (Zhang et al., 2007, 157186 ). Jacobsen et al. (2008, 156597 ) used the FE1-Muta TM Mouse lung epithelial cell line to investigate putative mechanisms of DE particle -induced mutagenicity. Mutation ion frequencies and ROS were determined after cells were incubated wi th 37.5 or 75 μg/ml DE particles (SRM1650) for 72-h (n = 8). The mutation frequency at the 75 μg/ml dose was significantly increased (1.55-fold; p<0.001) in contrast to cells treated with 37. 5 μg/ml DE particles. DE particles-induced ROS generation 1.6- to 1.9-fold in the epithelial cell cultures after 3 h of exposure compared with the 3- to 10-fold increase in ROS production previously repo rted for CB. The authors concluded that the mutagenic activity of DE particles is likely attri butable to activity from the organic fraction that both contains reactive species and can generate ROS. In human A549 and CHO-K1 cells, the organic fraction of DE particles significantly increased the amount of Comet and MN formation, respectively, in the presence and absence of SKF-525A (a CYP450 inhibitor) and S9, respe ctively (Oh and Chung, 2006, 088296 ). The organic base and neutral fractions of DE particles also significan tly induced DNA damage but only without SKF525A, and all fractions but the moderately polar fraction (phthalates and PAH oxyderivatives) induced MN formation with and without S9 (Bao et al., 2007, 097258 ). Gasoline engine exhaust significantly induced DNA damage as measured in the Comet assay and increased the frequency of MN in human A549 cells (Zhang et al., 2007, 157186 ). In human-hamster hybrid (A L ) cells, DE particles (SRM 2975) dose-dependently increased the mutation yield at the CD59 locus; this was

December 2009 7-77significantly reduced by simultaneous treatment with phagocytosis inhibitors (Bao et al., 2007, 097258 ).

Wood Smoke

The mutagenicity of wood smoke and cigarette smoke (CS) extracts was assayed in S. typhimurium strains TA98 and TA100 (Ames assay) using the pre-incubation assay with exogenous metabolic activation (rat liver S-9). Extract s of both samples (62.5 or 125 μg total PM equivalent/ml) were equally mutagenic to strain TA98 but the wood smoke extract was less mutagenic than the CS extracts in strain TA100 (Iba et al., 2006, 156582 ).



### In Vivo studies

Ambient Air

The contribution of ambient urban roadside air exposure (4, 12, 24, 48 or 60 wk) to DNA damage was examined in the lungs, nasal mucosa, a nd livers of adult male Wistar rats in Kawasaki, Japan (Sato et al., 2003, 096615 ). Messenger RNA levels of CYP450 enzymes that catalyze the transformation of PAHs to reactive metabolites were also evaluated. Concentrations of gases were reported to be 12-182 ppb NO and 0-9 ppb NO 2 in the filtered air chamber and 33-280 ppb NO and 42-81 ppb NO 2 in the experimental group chamber. Suspe nded PM concentrations were 11-19 μg/m 3 in the filtered air chamber and 42-100 μg/m 3 (average 63 μg/m 3 ) in the experimental group chamber. Body weight significantly decreased in exposed animals at 24, 48 and 60 wk. A 4-wk exposure to urban roadside air resulted in significant increa ses in multiple DNA adducts (lung, nasal, and liver DNA adducts). With longer exposures, there were significant increases in lung (48 wk), nasal (60 wk), and liver DNA adducts (60 wk). Changes were seen in CYP1A2 mRNA at 4 wk with a 2.3-fold increase in exposed animals compared to the control group with no change observed at 60 wk; CYP1A1 mRNA was unchanged. These results indicate that exposure to ambient air in this roadside area could induce DNA adduct formation, which may be important for carcinogenicity. Earlier studies (Ichinose et al., 1997, 053264 ) have shown that 8-oxodG, a DNA adduct, is elevated along with tumor formation in a dose-dependent manner in mice administered DE particles. The finding of adducts in the liver indicated that de position of PM and its associated PAHs in the lung can have indirect effects on extrapulmonary orga ns. It should be noted that PM deposition on the fur and ingestion during grooming cannot be ruled out as a possible exposure route. Another animal toxicological study employe d “non-carcinogenic” particles obtained from pooled non-cancerous lung tissue collected during surgical lung resection from three non-smoking male patients diagnosed with lung adenocarcinomas (Tokiwa et al., 2005, 191952 ). Particles were partially purified to remove organic compounds. Mor phologically the particles were similar to DE or ambient air PM and the organic extracts from th e particles were directly mutagenic in S. typhimurium tester strains TA98, YG1021 and YG1024. BALB/c and ICR mice were intratracheally instilled with particles at doses of 0.25, 0.5, 1.0, or 2.0 mg/mouse. After 24 h, 8-oxodG was measured in lung DNA and found to be increased in ICR mice in a dose-dependent manner, reaching a maximum of ~2.75 8-oxodG/10 5 dG at the 2.0 mg dose. The response was statistically significant at doses of 0.5, 1.0, and 2.0 mg. The increased 8- oxodG levels observed in vivo was reported to be likely due to hydroxyl radicals presumed to be involved in phagocytosis of non-mutagenic particles by inflammatory cells that could induce hydroxylation of guanine residue on DNA.

Diesel Exhaust

An in vivo study employed gtp delta transgenic mice carrying the lambda EG10 on each Chromosome 17 from a C57BL/6J background to investigate the effects of DE particles on mutation frequency (Hashimoto et al., 2007, 097261 ). Mice were exposed via inhalation to DE particles or via IT instillation to DE particles or DE particle extract and lambda EG10 phages were rescued; E. coli YG6020 was infected with the phage and screened for 6-thioguanine resistance. The mutagenic potency (mutation frequency per mg) caused by DE par ticle extract was twice that of DE particles, suggesting that the mutagenicity of DE particles is attributed primarily to compounds in the extract,

December 2009 7-78since ≈ 50% of the weight of DE particles was provide d by the extract. There was no difference in mutation frequency between the 1 and 3 mg/m 3 DE particle groups after 12 wk of exposure.

Wood Smoke

One recent study measured the effect of fre shly generated hardwood smoke on CYP1A1 activity based on ethoxyresorufin O-deethylase in pulmonary microsomes recovered from male Sprague-Dawley rats exposed to hardwood smoke by nose-only inhalation exposure (Iba et al., 2006, 156582 ). CYP1A1 activity in rat lung explants treate d with extracts of the total PM (TPM) from hardwood smoke samples and from freshly generated cigarette smoke (CS) was also evaluated. Unlike CS, hardwood smoke did not induce pulmonary CYP1A1 activity or mRNA (assessed by northern blot analysis) nor did extracts of hardwood smoke TPM induce CYP1A1 protein (assessed by western blot analysis) in cultured rat lung explants. The results suggest that unique constituents that are activated by CYP1A1 may be present in CS but not hardwood smoke.



## 7.5. 2.2. Car cinogenesis

Studies published prior to the 2004 AQCD that evaluated the carcinogenicity of ambient air were reviewed by Claxton and Woodall (2007, 180391 ). Five studies involved chronic inhalation exposures in rodents. No statistically significant increase in tumorigenesis was observed following chronic exposure to urban air pollution in Los Angeles (Gardner, 1966, 015129 ; Gardner et al., 1969, 015130 ; Wayne and Chambers, 1968, 038537 ). However in a study conducted in Brazil, urban air pollution was found to enhance the formation of ur ethane-induced lung tumors in mice (Cury et al., 2000, 192100 ; Reymao et al., 1997, 084653 ). Two recent studies evaluated the carcinogenic poten tial of chronic inhalation exposures to DE (Reed et al., 2004, 055625 ) and hardwood smoke (Reed et al., 2006, 156043 ). Two indicators of carcinogenic potential, formation of MN and tumorigenesis were measured in strain A/J mice, which is a mouse model that spontaneously develops lung tumors. Exposure to DE or hardwood smoke at concentrations of 1,000 μ g/m 3 and below did not cause increased formation of MN or an increased rate of lung tumors in this cancer-prone rodent model. These studies are described below.



### Diesel Exhaust

A/J mice were exposed to 30, 100, 300 and 1000 μ g/m 3 DE for 6 h/day and 7 days/wk for 6 mo (Reed et al., 2004, 055625 ). The concentration of gases in this including NO X , NO 2 , CO, SO 2 , NH 3 , methane, non-methane VOC, and FID total hy drocarbon ranged from control to high dose group values of 0 to 50.4 ± 0.6 ppm, 0.2 ± 0.2 to 6.9 ± 3.3 ppm, 0.3 ± 0.1 to 30.9 ± 4.5 ppm, not detectable to 955.2 ± 58.4 ppb, 176.5 ± 8.8 to 9.1 ± 0.2 μg/m 3 , 1406.5 ± 253.2 to 2642.1 ± 455.9 μg/m 3 , 134.0 ± 52.1 to 1578.6 ± 256.2 μg/m 3 , 0.1 ± 0.1 to 2.2 ± 0.2 pp m, respectively. Particle sizes in the four exposure groups ranged from 0.10-0.15 μ m MMAD with geometric standard deviations of 1.4-1.8. Following the 6-mo exposure and a 6-mo recovery period, mice were collected and MN formation in blood and tumor multiplicity a nd tumor incidence were measured in lungs. No increases in formation of MN or numbers of lung adenomas were observed in DE-exposed mice compared with controls.



### Wood Smoke

A/J mice were exposed to 30, 100, 300 and 1,000 μ g/m 3 hardwood smoke or to 30, 100, 300 and 1,000 μ g/m 3 DE for 6 h/day and 7 days/wk for 6 mo (Reed et al., 2006, 156043 ). Gaseous components of the hardwood smoke included CO, NH 3 , and non-methane VOC with concentrations from control levels to high dose hardwood smoke exposure ranging from 229 ± 31 to 14887.6 ± 832.3 ppm, 139.3 ± 2.3 to 54.9 ± 1.2 μg/m 3 and 177.6 ± 10.4 to 3455.0 ± 557.2 μg/m 3 , respectively. Concentrations of NO X , NO 2 and SO 2 were reported to be null. Particle sizes in the four exposure groups ranged from 0.25-0.36 μ m MMAD with geometric standard deviations of 2.0-3.3. Following the 6-mo exposure and a 6-mo recovery period, mi ce were collected and MN formation in blood and tumor multiplicity and tumor incide nce were measured in lungs. No increases in formation of MN or

December 2009 7-79numbers of lung adenomas were observed in hardwood smoke-exposed mice compared with controls. However, hardwood smoke from this study was mutagenic in the Ames reverse mutation

assay.



## 7.5. 2.3. Sum mary of Toxicological Studies

In summary, numerous new in vitro studies confirm and extend findings reported in the 2004 AQCD that ambient PM from urban sites and comb ustion-derived PM are mutagenic and genotoxic. A small number of new studies were conducted in vivo. One of these studies demonstrated increased mutagenic potency in mice exposed to DE particle s and DE particle extract. Another study found increased formation of 8-oxodG, a DNA adduct, following IT instillation of PM in mice. A chronic inhalation study of rats exposed to urban roadside air reported increased formation of DNA adducts in nose, lung, and liver and induction of CYP1A2. Inhalation exposure of rats to hardwood smoke failed to induce CYP1A1 in another study. Finall y, two chronic inhalation studies found no evidence of carcinogenic potential for DE and hardwood smoke in a cancer-prone mouse model. Collectively, these results provide some evidence, mainly from in vitro studies, to support the biological plausibility of ambient PM-lung cancer re lationships observed in epidemiology studies.



## 7.5.3. Epigenetic Studies and Othe r Heritable DNA mutations

Two epidemiologic epigenetic studies examined the effect of PM on DNA methylation. Both studies examined methylation of Alu and long inte rspersed nuclear element-1 (LINE-1) sequences, which are located in repetitive elements. In prev ious studies, methylation of these sequences has been linked to global genomic DNA methylation content (Weisenberger et al., 2005, 192101 ; Yang et al., 2004, 192102 ). The first study included men age 55 and older who were part of the Normative Aging Study in the Boston area (Baccarelli et al., 2009, 192155 ). A stationary monitoring site located 1 km from the examination site was used to estimate ambient PM 2.5 exposure for the duration of the study (1999-2007). During the study period, the median level of PM 2.5, averaged over 7-day periods, was 9.8 μg/m 3 (interquartile range 8.0-12.0 μg/m 3 ). There was no association between PM 2.5 and Alu methylation. LINE-1 methyl ation was associated with PM 2.5 measured over the 7 days before the examinations. The second study included 63 healthy men aged 27 -55 yr working at an electric furnace steel plant (Tarantini et al., 2009, 192010 ). Blood samples were taken twice, once in the morning after 2 days of not working and once in the morning after 3 full days of work. PM 10 was measured in 11 work areas and individuals completed daily logs about the amount of time spent in each area. On average, individuals had an estimated exposure of 233.4 μg/m 3 PM 10 (range 73.4-1220.2 μg/m 3 ).Short-term exposure did not alter the methylation of Alu and LINE-1. To examine effects of longterm exposure, both blood samples were considered independent of time, and Alu and LINE-1 were examined with respect to overall estimated PM 10 exposure using mixed effects models. There was a negative association between increasing levels of PM 10 exposure and Alu and LINE-1 methylation, indicating that PM 10 causes epigenetic changes to occur with long-term exposure. This study also looked at levels of iNOS gene, which is a gene suppressed by DNA methylation. iNOS expression was not associated with long term exposure to PM 10 but was affected by methylation in the short

term. Animal toxicology studies evaluating the effect of PM exposure on changes in the epigenome and other non-epigenetic heritable DNA changes ha ve only recently been conducted. After earlier work showed increased germline mutation rates in herring gulls nesting near steel mills on Lake Ontario (Yauk and Quinn, 1996, 089093 ) further work was conducted to address air-dependent contribution to germline mutations by housing male and female Swiss Webster mice in the same area and comparing mutation rates in those animals with mutation rates of animals housed in a rural setting with less air pollution (Somers et al., 2002, 078100 ). To determine if PM or the gaseous phase of the urban air was responsible for heritable mutations, Yauk et al. (2008, 157164 ) exposed mature male C57Bl×CBA F1 hybrid mice to either HE PA-filtered air or to ambient air in Hamilton, Ontario, Canada for 3 or 10 wk, or 10 wk plus 6 wk of clean air exposure (16 wk) (also discussed in Section 7.4.2.5 ). Sperm DNA was monitored for ESTR mutations, testicular sample bulky DNA adducts, and DNA single or double strand breaks. In addition, male-germ line (spermatogonial stem

December 2009 7-80cell) DNA methylation was monitored post-exposure. Th is area in Hamilton is near two steel mills and a major highway. Air composition showed mean concentrations for TSP of 93.8 ± 17 μg/m 3

, PAH of 8.3 ± 1.7 ng/m 3 , and metal of 3.6 ± 0.7 μ g/m 3 . Mutation frequency at ESTR Ms6-hm locus in sperm DNA from mice exposed 3 or 10 wk did not show elevated ESTR mutation frequencies, but there was a significant increase in ESTR mutati on frequency at 16 wk compared to HEPA-filter control animals, pointing to a PM-dependent mechan ism of action. No detectable DNA adducts were observed in testes samples at any of the time poi nts monitored. To verify inhalation exposure to particles, DNA adducts were reported in the lungs of mice exposed for 3 wk to ambient air; no other time points showed detectable DNA adduct form ation. Hypermethylation of germ-line DNA was also observed in mice exposed to ambient air for 10 and 16 wk. These PM-dependent epigenetic modifications (hypermethylation) were not seen in th e halploid stage (3 wk) of spermatogenesis, but were nonetheless seen in early stages of spermatogenesis (10 wk) and remained significantly elevated in mature sperm even after removal of the mouse from the environmental exposure (16 wk). Thus, these studies indicate that the ambient PM phase and not the gaseous phase is responsible for the increased frequency of heritable DNA mutations and epigenetic modifications. Based on the limited evidence from these epigenetics studies, long-term exposure to PM 10 may result in epigenetic changes. PM 2.5 also potentially affects some DNA methylation content. As epigenetic research progresses, future studies examining the relationship between PM and DNA methylation will be important in more thoroughly characterizing these associations. The effect of ambient PM on heritable DNA mutations and the epigenome has been well characterized in a Canadian steel mill area. Mice exposed to ambient PM plus gases developed paternally-derived heritable DNA mutations and epigenetic changes in sperm DNA that were not observed in mice exposed to ambient air that was HEPA-filtered. This is the first animal toxicology study showing heritable effects of PM exposure on DNA mutation and the epigenome. Because the epigenetics field is so new, further work in this emerging area will expand on these PM-dependent methylation changes to determine if the results can be recapitulated at other urban sites.



## 7.5.4. Summary and Causal Determinations

7.5. 4.1. PM 2 .5 The 2004 PM AQCD reported on original and follow-up analyses for three prospective cohort studies that reported positive relationships between PM 2.5 and lung cancer mortality. Several recent, well-conducted epidemiologic studies have extended the evidence for a positive association between
PM 2.5 and lung cancer mortality (Section 7.5.1.1 ). Generally, studies have not reported associations between long-term exposure to PM 2.5 or PM 10 and lung cancer incidence (Section 7.5.1.1 ). Animal toxicological studies did not focus on specific si ze fractions of PM, but rather examined ambient PM, wood smoke, and DE particles (Section 7.5.2 ). A number of recent studies indicate that ambient urban PM, emissions from wood/biomass burning, emissions from coal combustion, and gasoline and DE are mutagenic and that PAHs are genotoxic (Section 7.5.2 ). These findings are consistent with earlier studies that concluded that ambien t PM and PM from specific combustion sources are mutagenic and genotoxic and provide biological plausibility for the results observed in the epidemiologic studies. A limited number of epidemiologic and toxicological studies on the epigenome demonstrate that PM induces changes in methylation (Section 7.5.3 ), a new area of research that will likely be expanded in the futu re. However, it has yet to be determined how these alterations in the genome could influence the in itiation and promotion of cancer. Overall, the
evidence is 

### suggestive of a causal relationship between relevant PM 2.5 

### exposures and cancer, with the strongest evidence from the epidemiologic studies of lung cancer mortality. This evidence is limited by the non-specific measure of PM size fraction in some of the epidemiologic studies and most of the animal t oxicological studies, and the inconsistency in evidence with recent epidemiologic studies for an effect on cancer incidence. There is no epidemiologic evidence for cancer related to long- term exposure to PM in organs or systems other than the lung.

December 2009 7-81

## 7.5.

7.5. 4.2. PM 1 0-2.5 The 2004 PM AQCD did not report long-term exposure studies for PM 10-2.5 . No epidemiologic studies have been conducted to evaluate the effects of long-term PM 10-2.5 exposure and cancer. The
evidence is 

### inadequate to assess the association between PM 10-2.5 

### and UFP exposures and

cancer.



## 4.3. UFP s

The 2004 PM AQCD did not report long-term exposure studies for UFPs. No epidemiologic studies have been conducted to evaluate the effe cts of long-term UFP and cancer. The evidence is 

### inadequate to determine if a causal re lationship exists between long-term UFP exposures and cancer

.



# 7.6. Mortality

In the 1996 PM AQCD, results were presented for three prospective cohort studies of adult populations: the Six Cities Study (Dockery et al., 1993, 044457 ); the ACS Study (Pope et al., 1995, 045159 ); and the AHSMOG Study (Abbey et al., 1995, 000669 ). The 1996 AQCD concluded that the chronic exposure studies, taken together, sugge sted associations between increases in mortality and long-term exposure to PM 2.5 , though there was no evidence to support an association with PM 10-

2.5 (U.S. EPA, 1996, 079380 ). Discussions of mortality and long-term exposure to PM in the 2004 PM AQCD emphasized the results of four U.S. prospective cohort studie s, but the greatest weight was placed on the findings of the ACS and the Harvard Six Cities studies, which had each undergone extensive independent reanalysis, and which were based on cohorts that were broadly representative of the U.S. population. The 2004 PM AQCD concluded that the results from the Seventh-Day Adventist (AHSMOG) cohort provided some suggestive (but less conclusive) evidence for associations, while results from the Veterans Cohort provided inconsistent evidence for associations between long-term exposures to
PM 2.5 and mortality. Collectively, the 2004 PM AQCD found that these studies provided strong evidence that long-term exposure to PM 2.5 was associated with increased risk of human mortality. Effect estimates for all-cause mortality ranged from 6 to 13% increased risk per 10 μg/m 3 PM 2.5 , while effect estimates for cardiopulmonary mortality ranged from 6 to 19% per 10 μg/m 3 PM 2.5 . For lung cancer mortality, the effect estimate was a 13% increase per 10 μg/m 3 PM 2.5 , based upon the results of the extended analysis from the ACS cohort (Pope et al., 2002, 024689 ). With regard to
PM 10-2.5 , the 2004 PM AQCD reported that no association was observed between mortality and longterm exposure to PM 10-2.5 in the ACS study (Pope et al., 2002, 024689 ), while a positive but statistically non-significant association was reported in males in the AHSMOG cohort (McDonnell et al., 2000, 010319 ). Thus, the 2004 PM AQCD concluded that there was insufficient evidence for associations between long-term exposure to PM 10-2.5 and mortality. Overall, the 2004 PM AQCD concluded that there was strong epidemiologic evidence for associations between long-term exposures to PM 2.5 and excess all-cause and cardiopulmonary mortality. At the time of the 2004 PM AQCD, only a li mited number of the chronic-exposure cohort studies had considered direct measurements of consti tuents of PM, other than sulfates. With regard to source-oriented evaluations of mortality assoc iations with long-term exposure, the 2004 PM AQCD noted only the study by Hoek et al. (2002, 042364 ), in which the authors concluded that long-term exposure to traffic-related air pollution may shorten life expectancy. However, Hoek et al. (2002, 042364 ) also noted that living near a major road might include other factors that contribute to mortality associations. There was not sufficient evidence at the time of the 2004 PM AQCD to draw conclusions on effects associated with specific components or sources of PM. New epidemiologic evidence reports a consisten t association between long-term exposure to
PM 2.5 and increased risk of mortality. There is li ttle evidence for the long-term effects of PM 10-2.5 on mortality. Although this section focuses on mortality outcomes in response to long-term exposure to PM, it does not evaluate studies that examine th e association between PM and infant mortality.

December 2009 7-82These studies are evaluated in Section 7.5 because it is possible that in utero exposures contribute to infant mortality. A summary of the mean PM concentrations reported for the studies characterized in this section is presented in Table 7-8

.



### Table 7-8. Characterization of am bient PM concentrations from st udies of mortality and long-term exposures to PM.

Study Location Mean Concentration (μg/m 3) Upper Percentile Concentrations (μg/m 3)

PM 2.5

Brunekreef et al. (2009, 191947 ) The Netherlands 28 95th: 32

99th: 33

Max: 37

Chen et al. (2005, 087942 ) Multicity, CA 29.0

Eftim et al. (2008, 099104 ) U.S. 13.6-14.1 Max: 19.1-25.1

Enstrom (2005, 087356 ) CA 23.4 Max: 36.1

Goss et al. (2004, 055624 ) U.S. 13.7 75th: 15.9

Janes et al. (2007, 090927 ) U.S. 14.0

Jerrett et al. (2005, 087600 ) Los Angeles, CA Max: 27.1

Krewski et al. (2009, 191193 ) U.S. 14.02 75 th : 16.00

90th: 26.75

95th: 27.89

Max: 30.01

Laden et al. (2006, 087605 ) Multicity, U.S. 10.2-29.0

Lipfert et al. (2006, 088218 ) U.S. 14.3

Miller et al. (2007, 090130 ) U.S. 13.5 75th: 18.3

Max: 28.3

Pope et al. (2004, 055880 ) U.S. 17.1

Schwartz et al. (2008, 156963 ) Multicity, U.S. 17.5 Max: 40

Zeger et al. (2007, 157176 ) U.S. 17.0

Zeger et al. (2008, 191951 ) U.S. 13.2 75th: 14.9

PM 10-2.5
Chen et al. (2005, 087942 ) Multicity, CA 25.4

Lipfert et al. (2006, 088218 ) U.S. 16.0

PM 10

Chen et al. (2005, 087942 ) Multicity, CA 52.6

Gehring et al. (2006, 089797 ) North Rhine, Germany 43.7-48.0 Max: 52.5-56.1

Goss et al. (2004, 055624 ) U.S. 24.8 75th: 28.9

Puett et al. (2008, 156891 ) NE U.S. 21.6

Zanobetti et al. (2008, 156177 ) U.S. 29.4

December 2009 7-83

## 7.6.1. Recent Studies of Long-Term Exp osure to PM and Mortality

Studies since the last PM AQCD include results of new analyses and insights for the ACS and Harvard Six Cities studies, further analyses from the AHSMOG and Veterans study cohorts, as well as analyses of a Cystic Fibrosis cohort and subsets of the ACS from Los Angeles and New York City. In the original analyses of the Six Cities and ACS cohort studies, no associations were found between long-term exposure to PM 10-2.5 and mortality, and the extended and follow-up analyses did not evaluate associations with PM 10-2.5 . The historical and more recent results for PM 2.5 of both the ACS and the Harvard Six Cities studies are compiled in Figure 7-6 . Moreover, since the last PM AQCD, there is a major new cohort that investigates the effects of PM 2.5 on cardiovascular mortality in the literature: the WHI study (Miller et al., 2007, 090130 ). Most recently, an ecologic cohort study of the nation’s Medicare population has been completed (Eftim et al., 2008, 099104 ). These new findings further strengthen the evidence linking long-term exposure to PM 2.5 and mortality, while providing indications that the magnitude of the PM 2.5 -mortality association is larger than previously estimated ( Figure 7-7 ). Two recent reports from the AHSMOG and Veterans study cohorts have provided some limited evidence for associations between long-term exposure to PM 10-2.5 and mortality. The original analyses of the AHSMO G cohort study found positive associations between long-term concentrations of PM 10 and 15-yr mortality due to natural causes and lung cancer (Abbey et al., 1999, 047559 ). McDonnell et al. (2000, 010319 ) reanalyzed these data and concluded that previously observed association of long-term ambient PM 10 concentrations with mortality for males were best explained by a relationship of mortality with the fine fraction of PM 10 rather than the thoracic coarse fraction of PM 10 . Recent reports from the AHSMOG study cohort, as well as the Nurses’ Health Study and a cohort of women in Germany have provided some evidence for associations between long-term exposure to PM 10 and mortality among women. 

### Harvard Six Cities: A follow-up study has used updated air pollution and mortality data; an additional 1,368 deaths occurred during the follow- up period (1990-1998) versus 1,364 deaths in the original study period (1974-1989) (Laden et al., 2006, 087605 ). Statistically significant associations are reported between long-term exposure to PM 2.5 and mortality for data for the two periods (RR = 1.16 [95% CI: 1.07-1.26] per 10 μg/m 3 PM 2.5 ). Of special note is a statistically significant reduction in mortality risk reported with reduced long-term PM 2.5 concentrations (RR = 0.73 [95% CI: 0.57-0.95] per 10 μg/m 3 PM 2.5 ). This is equivalent to an RR of 1.27 for reduced mortality risks with reduced long-term PM 2.5 concentrations. This reduced mortality risk was observed for deaths due to cardiovascular and respiratory causes, but not for lung cancer deaths. The PM 2.5 concentrations for recent years were estimated from visibility data, which introduces some uncertainty in the interpretation of the results from this study. Coupled with the results of the original analysis (Dockery et al., 1993, 044457 ), this study strongly suggests that a reduction in PM 2.5 pollution yields positive health benefits. 

### ACS Extended Analyses/Reanalysis II: Two new analyses further evaluated the associations of long-term PM 2.5 exposures with risk of mortality in 50 U.S. cities reported by Pope and colleagues (2002, 024689 ), adding new details about deat hs from specific cardiovascular and respiratory causes (Krewski, 2009, 190075 ; Pope et al., 2004, 055880 ). Pope et al. (2004, 055880 ) reported positive associations with deaths from sp ecific cardiovascular diseases, particularly ischemic heart disease (IHD), and a group of car diac conditions including dysrhythmia, heart failure and cardiac arrest (RR for cardiovascular mortality = 1.12, 95% CI 1.08-1.15 per 10 μg/m 3 PM 2.5 ), but no PM associations were found with respiratory mortality. In an additional reanalysis that extended the follow-up period for the ACS cohort to 18 yr (1982-2000) (Krewski et al., 2009, 191193 ), investigators found effect estimates that were similar, though generally higher, than those reported in pr evious ACS analyses. This reanalysis also included data for seven ecologic (neighborhood-level) contextu al (i.e., not individual-level) covariates, each of which represents local factors known or suspected to influence mortality, such as poverty level, educational attainment, and unemployment. The effect estimate for all cause mortality, based on
PM 2.5 concentrations measured in 1999-2000 was 1.03 (95% CI: 1.01-1.05). The corresponding effect estimates for deaths due to IHD and lung cancer were 1.15 (95% CI: 1.04-1.18) and 1.11 (95% CI: 1.04-1.18), respectively. In earlier analyses of this cohort, investigators found that increasing education levels appeared to reduce the effect of PM 2.5 exposure on mortality. Results from this reanalysis show a similar pattern, although with somewhat less certainty, for all causes of death except IHD, for which the pattern was reve rsed. Overall, although the addition of random effects modeling and contextual covariates to the ACS model made most effect estimates higher (but

December 2009 7-84some lower), they were not statistically different from the earlier ACS effect estimates. Thus, these new analyses, with their more extensive considera tion of potentially confounding factors, confirm the published ACS PM 2.5 -mortality results to be robust. 

### California Cancer Prevention Study: In a cohort of elderly people in 11 California counties (mean age 73 yr in 1983), an association was reported for long-term PM 2.5 exposure with all-cause deaths from 1973-1982 (RR = 1.04 [95% CI: 1.01-1.07] per 10 μg/m 3 PM 2.5 ) (Enstrom, 2005, 087356 ). However, no significant associations were reported with deaths in later time periods when PM 2.5 levels had decreased in the most pollute d counties (1983-2002) (RR = 1.00 [95% CI: 0.98-1.02] per 10 μg/m 3 PM 2.5 ). The PM 2.5 data were obtained from th e EPA’s Inhalation Particle Network (collected 1979-1983), and the locations repr esented a subset of data used in the 50-city ACS study (Pope et al., 1995, 045159 ). However, the use of average values for California counties as exposure surrogates likely leads to significant exposure error, as many California counties are large and quite topographically variable.

Cohort Study Years Mean Effect Estimate (95% CI) Original Dockery et al. (1993, 044457 ) 1974-1991 18.6 All Cause Reanalysis Krewski et al. (2000, 012281 ) 1974-1991 18.6 Temporal Changes Villeneuve et al. (2002, 042576 ) 1974-1991 18.6 Extended Laden et al. (2006, 087605 ) 1974-1998 16.4 SCS

6-Cities Medicare Eftim et al. (2008, 099104 ) 2000-2002 14.1 Original Pope et al. (1995, 045159 ) 1982-1989 18.2 Reanalysis Krewski et al.(2000, 012281 ) 1982-1989 18.2 Extended Pope et al. (2002, 024689 ) 1979-1983 21.1 Extended Pope et al. (2002, 024689 ) 1999-2000 14.0 Intra-metro LA Jerrett et al. (2005, 087600 ) 1982-2000 19.0 ACS Medicare Eftim et al. (2008, 099104 ) 2000-2002 13.6 Reanalysis II Krewski et al. (2009, 190075 ) 1982-2000 14.0 Reanalysis II - LA Krewski et al. (2009, 190075 ) 1982-2000 20.5 ACS

Reanalysis II - NYC Krewski et al. (2009, 190075 ) 1982-2000 12.8 Original Dockery et al. (1993, 044457 ) 1974-1991 18.6 CPD SCS Reanalysis Krewski et al. (2000, 012281 ) 1974-1991 18.6 Original Pope et al. (1995, 045159 ) 1982-1989 18.2 Reanalysis Krewski et al. (2000, 012281 ) 1982-1989 18.2 Extended Pope et al. (2002, 024689 ) 1979-1983 21.1 Extended Pope et al. (2002, 024689 ) 1999-2000 14.0 Intra-metro LA Jerrett et al. (2005, 087600 ) 1982-2000 19.0 Reanalysis II Krewski et al. (2009, 190075 ) 1982-2000 14.0 Reanalysis II - LA Krewski et al. (2009, 190075 ) 1982-2000 20.5 ACS

Reanalysis II - NYC Krewski et al. (2009, 190075 ) 1982-2000 12.8 SCS Extended Laden et al. (2006, 087605 ) 1974-1998 16.4 CVD Reanalysis Krewski et al. (2000, 012281 ) 1982-1989 18.2 ACS Extended Pope et al. (2004, 055880 ) 1982-2000 17.1 Extended Pope et al. (2004, 055880 ) 1982-2000 17.1 IHD Intra-metro LA Jerrett et al. (2005, 087600 ) 1982-2000 19.0 Reanalysis II Krewski et al. (2009, 190075 ) 1982-2000 14.0 Reanalysis II - LA Krewski et al. (2009, 190075 ) 1982-2000 20.5 ACS

Reanalysis II - NYC Krewski et al. (2009, 190075 ) 1982-2000 12.8 Original Dockery et al. (1993, 044457 ) 1974-1991 18.6 Lung Cancer Reanalysis Krewski et al. (2000, 012281 ) 1974-1991 18.6 SCS

Extended Laden et al. (2006, 087605 ) 1974-1998 16.4 Original Pope et al. (1995, 045159 ) 1982-1989 18.2 Extended Pope et al. (2002, 024689 ) 1979-1983 21.1 Extended Pope et al. (2002, 024689 ) 1999-2000 14.0 Intra-metro LA Jerrett et al. (2005, 087600 ) 1982-2000 19.0 Reanalysis II Krewski et al. (2009, 190075 ) 1982-2000 14.0 Reanalysis II - LA Krewski et al. (2009, 190075 ) 1982-2000 20.5 ACS

Reanalysis II - NYC Krewski et al. (2009, 190075 ) 1982-2000 12.8 SCS Extended Laden et al. (2006, 087605 ) 1974-1998 16.4 Other Extended Pope et al. (2002, 024689 ) 1979-1983 21.1 Extended Pope et al. (2002, 024689 ) 1999-2000 14.0 Intra-metro LA Jerrett et al. (2005, 087600 ) 1982-2000 19.0 ACS

Reanalysis II Krewski et al. (2009, 190075 ) 1982-2000 14.0

CPD=Cardio-Pulmonary Disease CVD=Cardiovascular Disease
IHD=Ischemic Heart Disease Relative Risk Estimate



### Figure 7-6. Mortality risk estimates associated with long-term exposure to PM 2.5 

### from the Harvard Six Cities Study (SCS) and the American Cancer Society Study (ACS).

December 2009 7-85Study Cohort Subset Mean Effect Estimate (95% CI) McDonnell et al. (2000, 010319 ) AHSMOG Males 32.0 All Cause Full Cohort 28.3 Brunekreef et al. (2009, 191947 ) NLCS-AIR Case Cohort 28.3 1973-1982 23.4 1983-2002 23.4* Enstrom (2005, 087356 ) CA Cancer Prevention

1973-2002 23.4* Jerrett et al. (2005, 087600 ) ACS-LA 19.0 Krewski et al. (2009, 191193 ) ACS Reanalysis II-LA 20.5 Laden et al. (2006, 087605 ) Harvard 6-Cities 16.4 14.3 Lipfert et al. (2006, 088218 ) Veterans Cohort
14.3 ACS Sites 13.6 Eftim et al. (2008, 099104 ) Medicare Cohort 6-Cities sites 14.1 Krewski et al. (2009, 191193 ) ACS Reanalysis II 14.0 Goss et al. (2004, 055624 ) U.S. Cystic Fibrosis 13.7 65+, Eastern 14.0 65+, Central 10.7 65+, Western 13.1 65-74, Eastern 14.0 65-74, Central 10.7 65-74, Western 13.1 65+, Eastern 14.0 75-84, Central 10.7 75-84, Western 13.1 85+, Eastern 14.0 85+, Central 10.7 Zeger et al. (2008, 191951 ) MCAPS

85+, Western 13.1 Krewski et al. (2009, 191193 ) ACS Reanalysis II-NYC 12.8 Full Cohort 28.3 CV Brunekreef et al. (2009, 191947 ) NLCS-AIR Case Cohort 28.3 Pope et al. (2004, 055880 ) ACS 17.1 Laden et al. (2006, 087605 ) Harvard 6-Cities 16.4 Males, 51-70 yrs 14.3 Males, 71-90 yrs 14.3 Females, 51-70 yrs 14.3 Naess et al. (2007, 090736 ) Oslo, Norway

Females 71-90 yrs 14.3 Miller et al. (2007, 090130 ) WHI Females 13.5 Females 29.0 CHD Chen et al. (2005, 087942 ) AHSMOG Males 29.0 Jerrett et al. (2005, 087600 ) ACS-LA 19.0 IHD Krewski et al. (2009, 190075 ) ACS Reanalysis II-LA 20.5 Pope et al. (2004, 055880 ) ACS 17.1 ACS Reanalysis II 14.0 Krewski et al. (2009, 191193 ) ACS Reanalysis II-NYC 12.8 McDonnell et al. (2000, 010319 ) AHSMOG Males 32.0 CPD Jerrett et al. (2005, 087600 ) ACS-LA 19.0 ACS Reanalysis II-LA 20.5 ACS Reanalysis II 14.0 Krewski et al. (2009, 191193 )

ACS Reanalysis II-NYC 12.8 Full Cohort 28.3 Respiratory Brunekreef et al. (2009, 191947 ) NLCS-AIR Case Cohort 28.3 Laden et al. (2006, 087605 ) Harvard 6-Cities 16.4 McDonnell et al. (2000, 010319 ) AHSMOG Males 32.0 Full Cohort 28.3 Lung Cancer Brunekreef et al. (2009, 191947 ) NLCS-AIR Case Cohort 28.3 Jerrett et al. (2005, 087600 ) ACS-LA 19.0 Krewski et al. (2009, 191193 ) ACS Reanalysis II-LA 20.5 Laden et al. (2006, 087605 ) Harvard 6-Cities 16.4 Males, 51-70 yrs 14.3 Males 71-90 yrs 14.3 Females, 51-70 yrs 14.3 Naess et al. (2007, 090736 ) Oslo, Norway

Females, 71-90 yrs 14.3 ACS Reanalysis II 14.0 Krewski et al. (2009, 191193 ) ACS Reanalysis II-NYC 12.8 Full Cohort 28.3 Other Brunekreef et al. (2009, 191947 ) NLCS-AIR Case Cohort 28.3 Jerrett et al. (2005, 087600 ) ACS-LA 19.0 Laden et al. (2006, 087605 ) Harvard 6-Cities 16.4 Krewski et al. (2009, 191193 ) ACS Reanalysis II 14.0

CV=Cardiovascular; CHD= Coronary Heart Disease
IHD=Ischemic Heart Disease; CPD=Cardio-Pulmonary Disease

* PM 2.5 data from 1973-1982 applied to all subsequent time periods Relative Risk Estimate



### Figure 7-7. Mortality risk estimates, long-term exposure to PM 2.5 

### in recent cohort studies.

December 2009 7-86

### AHSMOG: In this analysis for the Seventh-Day Adventist cohort in California, a positive, statistically significant, association with coro nary heart disease mortality was reported among females (92 deaths; RR = 1.42 [95% CI: 1.06-1.90] per 10 μg/m 3 PM 2.5 ), but not among males (53 deaths; RR = 0.90 [95% CI: 0.76-1.05] per 10 μg/m 3 PM 2.5 ) (Chen et al., 2005, 087942 ). Associations were strongest in the subset of post menopausal women (80 deaths; RR = 1.49 [95% CI: 1.17-1.89] per 10 μg/m 3 PM 2.5 ). The authors speculated that fema les may be more sensitive to air pollution-related effects, based on differences between males and females in dosimetry and exposure. As was found with PM 2.5 , a positive association with coronary heart disease mortality was reported for PM 10-25 and PM 10 among females (RR = 1.38 [95% CI: 0.97-1.95] per 10 μg/m 3 PM 10-2.5 ; RR = 1.22 [95% CI: 1.01-1.47] per 10 μg/m 3 PM 10 ), but not for males (RR = 0.92 [95% CI: 0.66- 1.29] per 10 μg/m 3 PM 10-2.5 ; RR = 0.94 [95% CI: 0.82-1.08] per 10 μg/m 3 PM 10 ); associations were strongest in the subset of postmenopausal women (80 deaths) (Chen et al., 2005, 087942 ). 

### U.S. Cystic Fibrosis cohort: A positive, but not statistically significant, association was reported for PM 2.5 in this study (RR = 1.32 [95% CI: 0.91-1.93] per 10 μg/m 3 PM 2.5 ) that primarily focused on evidence of exacerbation of respiratory symptoms (Goss et al., 2004, 055624 ). No clear association was reported for PM 10 . However, only 200 deaths had occurred in the cohort of over 11,000 people (average age in cohort was 18.4 yr), so the power of this study to detect associations was relatively low. 

### Women’s Health Initiative (WHI) Study: This nationwide cohort study considered 65,893 post-menopausal women with no history of cardiovasc ular disease who lived in 36 U.S. metropolitan areas from 1994 to 1998 (Miller et al., 2007, 090130 ). The study had a median subject follow-up time of 6 years. Miller and colleagues assessed each woman’s exposure to air pollutants using the monitor located nearest to their residence. Hazardratios were estimated for the first cardiovascular event, adjusting for age, race or ethnic group, sm oking status, educational level, household income, body-mass index, and presence or absence of di abetes, hypertension, or hypercholesterolemia. Overall, this study concludes that “long-term expos ure to fine particulate air pollution is associated with the incidence of cardiovascular disease an d death among postmenopausal women.” In terms of effect size, the study found that each increase of 10 μg/m 3 of PM 2.5 was associated with a 24% increase in the risk of a cardiovascular event (hazard ratio, 1.24 [95% CI: 1.09-1.41]) and a 76% increase in the risk of death from cardiovascular di sease (hazard ratio, 1.76 [95% CI: 1.25-2.47]). While this study found results confirmatory to the ACS and Six Cities Study, it reports much larger relative risk estimates per μg/m 3 PM 2.5 . In addition, since the study included only women without pre-existing cardiovascular disease, it could potentially be a healthier cohort population than that considered by the ACS and Six Cities Study . Indeed, the WHI Study reported only 216 cardiovascular deaths in 349,643 women-yr of follow-up, or a rate of 0.075% deaths per year (Miller et al., 2007, 090130 ), while the ACS Study reported that 10 % of subjects died of cardiovascular disease over a 16-yr follow-up period, yielding a ra te of 0.625% per year, or approximately 8 times the cardiovascular mortality rate of the WHI population (Pope et al., 2004, 055880 ). Thus, PM 2.5 impacts may yield higher relative risk esti mates in the WHI population because the PM 2.5 risk is being compared to a much lower prevailing risk of cardiovascular death in this select study population. The WHI study not only confirms the ACS and Si x City Study associations with mortality in yet another well characterized cohort with detailed in dividual-level information, it also has been able to consider the individual medical records of th e thousands of WHI subjects over the period of the study. This has allowed the researchers to examine not only mortality, but also related morbidity in the form of heart problems (cardiovascular events) experienced by the subjects during the study. As reported in this paper, this examination confirmed that there is an increased risk of cardiovascular morbidity, as well (Section 7.2.9 ). These morbidity co-associations with PM 2.5 in the same population lend even greater support to the biological plausibility of the air pollution-mortality associations found in this study. 

### Medicare Cohort Studies: Using Medicare data, Eftim and co-authors (2008, 099104 ) assessed the association of PM 2.5 with mortality for the same locations included in the ACS and Six City Study. For these locations, they estimated the chronic effects of PM 2.5 on mortality for the period 2000-2002 using mortality data for cohorts of Medicare participants and average PM 2.5 levels from monitors in the same counties included in the two studies. Using aggregate counts of mortality by county for three age groups, they estimated mort ality risk associated with air pollution adjusting for age and sex and area-level covariates (educatio n, income level, poverty, and employment), and controlled for potential confounding by cigarette smoking by including standardized mortality ratios

December 2009 7-87for lung cancer and COPD. This study is, therefore, an ecological analysis, similar to past published cross-sectional analyses, in that area-level cova riates (education, income level, poverty, and employment) are employed as controlling variab les, since individual level information is not available from the Medicare database (other th an age and sex), which includes virtually all Americans aged 65 or greater. Exposures are also ecol ogical in nature, as central site data are used as indices of exposure. These results indicated that a 10 μg/m 3 increase in the yearly average PM 2.5 concentration is associated with 10.9% (95% CI : 9.0-12.8) and with 20.8% (95% CI: 14.8-27.1) increases in all-cause mortality for the ACS and Six Cities Study counties, respectively. The estimates are somewhat higher than those reported by the original investigators, and there may be several possible explanations for this apparent increa se, especially that this is an older population than the ACS cohort. Perhaps the most likely explanation is that the lack of personal confounder information (e.g., past personal smoking information) led to an insufficient control for the effects of these other variables’ effects on mortality, inflating the pollution effect es timates somewhat, similar to what has been found in the ACS analyses when only ecological-level control variables were included. The ability of the Eftim et al. (2008, 099104 ) study results to qualitatively replicate the original individual-level cohort study (e.g., ACS a nd Six Cities Study) results suggests that past ecological cross-sectional mortality study results may also provide useful insights into the nature of the association, especially when used for considera tion of time trends, or for comparisons of the relative (rather than absolute) sizes of risks between different pollutants or PM components in health effects associations. Janes et al. (2007, 090927 ) used the same nationwide Medicare mortality data to examine the association between monthly averages of PM 2.5 over the preceding 12 mo and monthly mortality rates in 113 U.S. counties from 2000 to 2002. They decomposed the association between PM 2.5 and mortality into two components: (1) the association between “national trends” in PM 2.5 and mortality; and (2) the association between “local trends,” defined as county-specific deviations from national trends. This second component is posited to provide evidence as to whether counties having steeper declines in PM 2.5 also have steeper declines in mortality relative to the national trend. They report that the exposure effect estimates are different at these two spatiotemporal scales, raising concerns about confounding bias in these analyses. The authors assert that the association between trends in
PM 2.5 and mortality at the national scale is more li kely to be confounded than is the association between trends in PM 2.5 and mortality at the local scale and, if the association at the national scale is set aside, that there is little evidence of an association between 12-month exposure to PM 2.5 and mortality in this analysis. However, in response, Pope and Burnett (2007, 090928 ) point out that such use of long-term time trends as the primary so urce of exposure variability has been avoided in most other air pollution epidemiology studies beca use of such concerns about potential confounding of such time-trend associations. By linking monitoring data to the U.S. Medicar e system by county of residence, Zeger et al. (2007, 157176 ) analyzed Medicare mortality records, comprising over 20 million enrollees in the 250 largest counties during 2000-2002. The authors es timated log-linear regression models having age-specific county level mortality rates as the outcome and, as the main predictor, the average PM 2.5 pollution level in each county during 2000. Area-lev el covariates were used to adjust for socioeconomic status and smoking. The authors reported re sults under several degrees of adjustment for spatial confounding and with stratification into eastern, central and western U.S. counties. A 10 μ g/m 3 increase in PM 2.5 was associated with a 7.6% increase in mortality (95% CI: 4.4-10.8). When adjusted for spatial confounding, the estimated log-relative risks dropped by 50%. Zeger et al. (2007, 157176 ) found a stronger association in the eastern counties than nationally, with no evidence of an association in western counties. In a subsequent report, Zeger et al. (2008, 191951 ) created a new retrospective cohort, the Medicare Cohort Air Pollution Study (MCAPS), consisting of 13.2 million persons residing in 4,568 ZIP codes in urban areas having geographic centroids within 6 miles of a PM 2.5 monitor. Using this cohort, they investigated the relationship between 6-yr avg exposure to PM 2.5 and mortality risk over the period 2000-2005. When divided by region, the associations between long-term exposure to
PM 2.5 and mortality for the eastern and central ZIP code s were qualitatively similar to those reported in the ACS and Six Cities Study, with 11.4% (95% CI: 8.8-14.1) and 20.4% (95% CI: 15.0-25.8) increases per 10 μg/m 3 increase in PM 2.5 in the eastern and central regions, respectively. The MCAPS results included evidence of differing PM 2.5 relative risks by age and geographic location, where risk declines with increasing age category until ther e is no evidence of an association among persons

December 2009 7-88≥ 85 yr of age, and there is no evidence of a positive association for the 640 urban ZIP codes in the western region of the U.S. Using hospital discharge data, Zanobetti et al. (2008, 156177 ) constructed a cohort of persons discharged with COPD using Medicare data be tween 1985 and 1999. Positive associations in the survival analyses were reported for single year and multiple-year lag exposures, with a hazard ratio for total mortality of 1.22 (95% CI: 1.17-1.27) per 10 μg/m 3 increase in PM 10 over the previous 4 years. 

### Veterans Cohort: A recent reanalysis of the Veterans cohort data focused on exposure to traffic-related air pollution (traffic density based on traffic flow rate data and road segment length) reported a stronger relationship between mortality w ith long-term exposure to traffic than with PM 2.5 mass (Lipfert et al., 2006, 088218 ). A significant association was reported between total mortality and PM 2.5 in single-pollutant models (RR = 1.12 [95% CI: 1.04-1.20] per 10 μg/m 3 PM 2.5 ). This risk estimate is larger than results reported in a previ ous study of this cohort. In multipollutant models including traffic density, the association with PM 2.5 was reduced and lost statistical significance. Traffic emissions contribute to PM 2.5 so it would be expected th at the two would be highly correlated, and, thus, these mul tipollutant model results should be interpreted with caution. In a companion study, Lipfert et al. (2006, 088218 ) used data from EPA’s fine particle speciation network, and reported findings for PM 2.5 which were similar to those reported by Lipfert et al. (2006, 088218 ). In this study (Lipfert et al., 2006, 088218 ), a significant association was reported between long-term exposure to PM 10-2.5 and total mortality in a single-pollutant model (RR = 1.07, 95% CI: 1.01-1.12 per 10 μg/m 3 PM 10-2.5 ). However, the association became negative and not statistically significant in a model that included traffic density . As it would be expected that traffic would contribute to the PM 10-2.5 concentrations, it is difficult to inte rpret the results of these multipollutant analyses. 

### Nurses’ Health Study Cohort: The Nurses’ Health Study (Puett et al., 2008, 156891 ) is an ongoing prospective cohort study examining the relation of chronic PM 10 exposures with all-cause mortality and incident and fatal CHD consisting of 66,250 female nurses in MSAs in the northeastern region of the U.S. All cause mortality was statistically significantly associated with average PM 10 exposures in the time period 3-48 mo preceding death. The association was strongest with average PM 10 exposure in the 24 mo prior to death (hazard ratio 1.16 [95% CI: 1.05-1.28]) and weakest with exposure in the month prior to deat h (hazard ratio 1.04 [95% CI: 0.98-1.11]). The association with fatal CHD occurred with the greatest magnitude with mean exposure in the 24 mo prior to death (hazard ratio 1.42 [95% CI: 1.11-1.81]). 

### Netherlands Cohort Study (NLCS): The Netherlands Cohort Study (Brunekreef et al., 2009, 191947 ) estimates the effects of traffic-related air pollution on cause specific mortality in a cohort of approximately 120,000 subjects aged 55-69 yr at enrollment. For a 10 μg/m 3 increase in
PM 2.5 concentration, the relative risk for natural-cause mortality in the full cohort was 1.06 (95% CI: 0.97-1.16), similar in magnitude to the results re ported by the ACS. In a case-cohort analysis adjusted for additional potential confounders, ther e were no associations between air pollution and mortality. 

### German Cohort: The North Rhine-Westphalia State Environment Agency (LUA NRW) initiated a cohort of approximately 4,800 women, and assessed whether long-term exposure to air pollution originating from motorized traffic and industrial sources was associated with total and cause-specific mortality (Gehring et al., 2006, 089797 ). They found that cardiopulmonary mortality was associated with PM 10 (RR = 1.52 [95% CI: 1.09-2.15] per 10 μg/m 3 PM 10 ).



## 7.6.2. Composition and Source-O riented Analyses of PM

As discussed in the 2004 PM AQCD, only a very limited number of the chronic exposure cohort studies have included direct measurements of chemical-specific PM constituents other than sulfates, or assessments of source-oriented effects, in their analyses. One exception is the Veterans Cohort Study, which looked at associations with some constituents, and traffic. 

### Veterans Cohort: Using data from EPA’s fine particle speciation network, Lipfert et al. (2006, 088756 ) reported a positive association for mortality with sulfates. Using 2002 data from the fine particle speciation network, positive associ ations were found between mortality and long-term exposures to nitrates, EC, Ni and V, as well as traffic density and peak O 3 concentrations. In

December 2009 7-89multipollutant models, associations with traffic dens ity remained significant, as did nitrates, Ni and V in some models. 

### Netherlands Cohort Study: Beelen et al. (2008, 156263 ) studied the association between long-term exposure to traffic-related air pollution and mortality in a Dutch cohort. They used data from an ongoing cohort study on diet and cancer with 120,852 subjects who were followed from 1987 to 1996. Exposure to BS, NO 2 , SO 2 , and PM 2.5 , as well as various exposure variables related to traffic, were estimated at the home address. Traf fic intensity on the nearest road was independently associated with mortality. Relative risks (CI) for a 10 μg/m 3 increase in BS concentrations (difference between 5th and 95th percentile) were 1.05 (95% CI: 1.00-1.11) for natural cause, 1.04 (95% CI: 0.95-1.13) for cardiovascular, 1.22 (9 5% CI: 0.99-1.50) for respiratory, 1.03 (95% CI: 0.88-1.20) for lung cancer, and 1.04 (95% CI: 0.97-1 .12) for mortality other than cardiovascular, respiratory, or lung cancer. Results were similar for NO 2 and PM 2.5 , but no associations were found for SO 2 . Traffic-related air pollution and several traffic exposure variables were associated with mortality in the full cohort, although the relative risks were generally small. Associations between natural-cause and respiratory mortalit y were statistically significant for NO 2 and BS. These results add to the evidence that long-term exposure to traf fic-related particulate air pollution is associated with increased mortality. Given the general dearth of published source-orient ed studies of the mortality impacts of longterm PM exposure components, and given that the recent Medicare Cohort study now indicates that such ecological cross-sectional studies can be usef ul for evaluating time trends and/or comparisons across pollution components, it may well be that examining past cross-sectional studies comparing source-oriented components of PM may be informat ive. In particular, Ozkaynak and Thurston (1987, 072960 ), utilized the chemical speciation conductedin the Inhalable Particle (IP) Network to conduct a chemical constituent and source-oriented evaluation on long-term PM exposure and mortality in the U.S. They analyzed the 1980 U.S. vital statistics and available ambient air pollution data bases for sulfates and fine, inhalable, and TS P mass. Using multiple regression analyses, they conducted a cross-sectional analysis of the associa tion between various particle measures and total mortality. Results from the various analyses indicated the importance of considering particle size, composition, and source information in modeling of particle pollution health effects. Of the independent mortality predictors considered, pa rticle exposure measures most related to the respirable fraction of the aerosols, such as fine pa rticles and sulfates, were most consistently and significantly associated with the reported SMSA-specific total annual mortality rates. On the other hand, particle mass measures that included PM 10-2.5 (e.g., total suspended particles and inhalable particles) were often found to be non-significant pr edictors of total mortality. Furthermore, based on the application of PM 2.5 source apportionment, particles from industrial sources and from coalcombustion were indicated to be more significant contributors to human mortality than fine soilderived particles.



## 7.6.3. Within-City Effects of PM Exposure

Much of the exposure gradient in the national- scale cohort studies was due to city-to-citydifferences in regional air pollution, raising the possibility that some or all of the original PMsurvival associations may have been driven instead by city-to-city differences in some unknown (non-pollution) confounder variable. This has been evaluated by three recent studies. 

### ACS, Los Angeles: To investigate this issue, two new analyses using ACS data focused on neighborhood-to-neighborhood differences in urban air pollutants, using data from 23 PM 2.5 monitoring stations in the Los Angeles area, and ap plying interpolation methods (Jerrett et al., 2005, 087600 ) or land use regression methods (Krewski et al., 2009, 191193 ) to assign exposure levels to study individuals. This resulted in both improved exposure assessment and an increased focus on local sources of PM 2.5 . Significant associations between PM 2.5 and mortality from all causes and cardiopulmonary diseases were reported with the magnitude of the relative risks being greater than those reported in previous assessments. In general, the associations for PM 2.5 and mortality using these two methods for exposure assessment were similar, though the use of land use regression resulted in somewhat smaller hazard ratios and tighter CIs (see Table 7-9 ). This indicates that city-tocity confounding was not the cause of the associations found in the earlier ACS Cohort studies. This provides evidence that reducing exposure erro r can result in stronger associations between PM 2.5 and mortality than generally observed in broader studies having less exposure detail.

December 2009 7-90

### Table 7-9. Comparison of results from ACS intra-urban analysis of Los Angeles and New York City using kriging or land use regression to estimate exposure.

Cause of

Death Los Angeles:

Hazard Ratio 1 and 95% Confidence Interval Using Kriging 2

(Jerrett et al., 2005, 087600 ) Los Angeles:

Hazard Ratio 1 and 95% Confidence Interval Using Land Use Regression 3

(Krewski et al., 2009, 191193 ) New York City:

Hazard Ratio 1 and 95% Confidence Interval Using Land Use Regression 4

(Krewski et al., 2009, 191193 )

All Cause 1.11 (0.99-1.25) 1.13 (1.01-1.25) 0.86 (0.63-1.18)

IHD 1.25 (0.99-1.59) 1.26 (1.02-1.56) 1.56 (0.87-2.88)

CPD 1.07 (0.91-1.26) 1.09 (0.94-1.26) 0.66 (0.41-1.08)

Lung Cancer 1.20 (0.79-1.82) 1.31 (0.90-1.92) 0.90 (0.29-2.78)

1 Hazard ratios presented per 10 μg/m 3 increase in PM 2.5 2 Model included parsimonious contextual covariates

3 Model included parsimonious individual level (23) and ecologic (4) covariates

4 Model included all 44 individual level and 7 ecologic covariates.



### ACS, New York: Krewski et al. (2009, 191193 ) applied the same techniques used in the land use regression analysis of Los Angeles to an investigation conducted in New York City. Annual average concentrations were calculated for each of 62 monitors from 3 yr of daily monitoring data for 1999-2001. Those data were combined with land-use data collected from traffic counting systems, roadway network maps, satellite photos of the study area, and local government planning and tax-assessment maps to assign estimated exposur es to the ACS participants. The investigators did not observe elevated effect estimates for all ca use, CPD or lung cancer deaths, but IHD did show a positive association with PM 2.5 concentration. The difference be tween the 90th and 10th percentiles of the 3-yr avg PM 2.5 concentration was 1.5 μg/m 3 and the difference between the minimum and maximum values of the 3-yr avg PM 2.5 concentration was 7.8 μg/m 3 . This narrow range in PM 2.5 exposure contrasts across the New York City me tropolitan area and may well account for the inconclusive results in this city-specific anal ysis. Relatively uniform exposures would reduce the power of the statistical models to detect patterns of mortality relative to exposure and estimate the association with precision. 

### WHI Study: This study also investigated the within- ve rsus between-city effects in its cities. As shown in Figure 7-8 , similar effects for both the within and between-city analyses demonstrate that this association is not due to some other (non-pollution) confounder differing between the various cities, strengthening confidence in the overall pollution-effect estimates.

December 2009 7-91Source: Miller et al. (2007, 090130 ) Copyright © 2007 Massachusetts Medical Society. All rights reserved.



### Figure 7-8. Plots of the relative risk of deat h from cardiovascular disease from the Women’s Health Initiative study displaying the between-city and within-city contributions to the overall association between PM 2.5 

### and cardiovascular mortality windows of exposure-effects.



## 7.6.4. Effects of Different Long-term Exposure Windows

The delay between changes in exposure and changes in health has important policy implications. Schwartz et al. (2008, 156963 ) investigated this issue using an extended follow-up of the Harvard Six Cities Study. Cox proportional hazar ds models were fit to control for smoking, body mass index, and other covariates. Penalized splines were fit in a flexible functional form to the concentration response to examine its shape, and the degrees of freedom for the curve were selected based on Akaike’s information criterion (AIC). Th e researchers also used model averaging as an alternative approach, where multiple models are fit explicitly and averaged, weighted by their probability of being correct given the data. Th e lag relationship by model was averaged across a range of unconstrained distributed la g models (i.e., same year, 1 yr prio r, 2 yr prior, etc.). Results of the lag comparison are shown in Figure 7-9 indicating that the effects of changes in exposure on mortality are seen within 2 yr. The authors al so noted that the concentration-response curve was linear, clearly continuing below the level of the current U.S. air quality standard of 15 μg/m 3

.

December 2009 7-92Source: Schwartz et al. (2008, 156963 )



### Figure 7-9. The model-averaged estimated effect of a 10-μg/m 3 

### increase in PM 2.5 

### on all-cause mortality at different lags (in years) between exposure and death. Each lag is estimated independently of the others. Al so shown are the pointwise 95% CIs for each lag, based on jackknife estimates.

Similarly, the effect of long-term exposure to PM 10 on the risk of death in a large multicity study of elderly subjects discharged alive following an admission for COPD found the effect was not limited to the exposure in each year of follow-up, and had larger cumulative effects spread over the follow-up year and three precedi ng years (Zanobe tti et al., 2008, 156177 ). Röösli et al. (2005, 156923 ) took an alternative approach to determining the window over which the mortality effects of long-term polluti on exposures occurred. They fit the model shown in Figure 7-10 using k = 0.5 based on the Utah Steel Strike (Pope, 1989, 044461 ) and the Ireland coal ban study (Clancy et al., 2002, 035270 ). They found that roughly 75% of health benefits are observed in the first 5 years, as shown in Table 7-10 . These results are consistent with the findings of Schwartz et al. (2008, 156963 ). Puett et al. (2008, 156891 ) also compared different long-term exposure lags, with exposure periods ranging from 1 month to 48 mo prior to death. They found statistically significant associations with average PM 10 exposures in the time period 3-48 mo prior to death, with the strongest associations in the 24 mo prior to death and the weakest with exposure in the 1 mo prior to death.

December 2009 7-93Source: Reprinted with Permission from Oxford University Press & the International Epidemiological Society from Röösli et al. ( 2005, 156923 )



### Figure 7-10. Time course of relative risk of death after a sudden decrease in air pollution exposure during the year 2000, assuming a steady state model (solid line) and a dynamic model (bold dashed line). The th in dashed line refe rs to the reference scenario.

Table 7-10. Distribution of the effect of a hypothetical reduction of 10 μg/m 3 

### PM 10 

### in 2000 on allcause mortality 2000-2009 in Switzerland.

Y ea r 1999 2000 2001 2002 2003 2004 2005 2006 2007 2008 2009

Proportion of total effect (%) - 39.3 23.9 14.5 8.8 5.3 3.2 2.0 1.2 0.7 0.4

Relative risk (per 10 μg/m 3 reduction in PM 10 ) 1.0 0.9775 0.9863 0.9917 0.9950 0.9969 0.9981 0.9989 0.9993 0.9996 0.9997

Relative risk and proportion of total effect in each year are shown, assuming a time constant k of 0.5 Source: Röösli et al. (2005, 156923 )

In the reanalysis of the ACS cohort, the inves tigators calculated time windows of exposure as average concentrations during successive 5-yr perio ds preceding the date of death (Krewski et al., 2009, 191193 ). The investigators considered the time window with the best-fitting model (judged by the AIC statistic) to be the period during which pollution had the strongest influence on mortality. Overall, the differences between the time peri ods were small and demonstrated no definitive patterns. High correlations between exposure levels in the three periods may have reduced the ability of this analysis to detect any differences in the relative importance of the time windows. The investigators did not analyze any time periods smaller than 5 yr, so the results are not directly comparable to those reported by Schwartz et al. (2008, 156963 ), Röösli et al. (2005, 156923 ), and Puett et al. (2008, 156891 ).

December 2009 7-94Generally, these results indicate a developi ng coherence of the air pollution mortality literature, suggesting that the health benefits from reducing air pollution do not require a long latency period and would be expected within a few years of intervention.



## 7.6.5.

7.6. 

## Summary and Causal Determinations

5.1. P M 2.5

In the 1 996 PM AQCD (U.S. EPA, 1996, 079380 ), results were presented for three prospective cohort studies of adult populations: the Six Cities Study (Dockery et al., 1993, 044457 ); the ACS Study (Pope et al., 1995, 045159 ); and the AHSMOG Study (Abbey et al., 1995, 000669 ). The 1996 AQCD concluded that the chronic exposure studies, taken together, suggested associations between increases in mortality and long-term exposure to PM 2.5 , though there was no evidence to support an association with PM 10-2.5 (U.S. EPA, 1996, 079380 ). Discussions of mortality and long-term exposure to PM in the 2004 PM AQCD emphasized the results of four U.S. prospective cohort studies, but the greatest weight was placed on th e findings of the ACS and the Harvard Six Cities studies, which had undergone extensive independent reanalysis, and which were based on cohorts that were broadly representative of the U.S. pop ulation. Collectively, the 2004 PM AQCD found that these studies provided strong evidence that long-term exposure to PM 2.5 was associated with increased risk of human mortality. The recent evidence is largely consistent with past studies, further supporting the evidence of associations between long-term PM 2.5 exposure and increased risk of human mortality (Section 7.6 ) in areas with mean concentrations from 13.2 to 29 μg/m 3 ( Figure 7-7 ). New evidence from the Six Cities cohort study shows a relatively large risk estimat e for reduced mortality risk with decreases in
PM 2.5 (Laden et al., 2006, 087605 ). The results of new analyses from the Six Cities cohort and the ACS study in Los Angeles suggest that previous and current studies may have underestimated the magnitude of the association (Jerrett et al., 2005, 087600 ). With regard to mortality by cause-ofdeath, recent ACS analyses indi cate that cardiovascular mortality primarily accounts for the total mortality association with PM 2.5 among adults, and not respiratory mortality. The recent WHI cohort study shows even higher cardiovascular risks per μ g/m 3 than found in the ACS study, but this is likely due to the fact that the study included only post-menopausal women without pre-existing cardiovascular disease (Miller et al., 2007, 090130 ). There is additional evidence for an association between PM 2.5 exposure and lung cancer mortality (Section 7.5.1.1 ). The WHI study also considered within versus between city mortality, as well as morbidity co-associations with PM 2.5 in the same population. The first showed that the results are not due to between city confounding, and the morbidity analyses show the coherence of the mortality association across health endpoints, supporting the biological plausibility of the air pollution-mortality associations found in these studies. Results from a new study examining the relationship between life expectancy and PM 2.5 and the findings from a multiyear expert judgment study that comprehensively characterizes the size and uncertainty in estimates of mortality re ductions associated with decreases in PM 2.5 in the U.S draw conclusions that are consistent with an association between long-term exposure to PM 2.5 and mortality (Pope et al., 2009, 190107 ; Roman et al., 2008, 156921 ). Pope et al. (2009, 190107 ) report that a decrease of 10 μg/m 3 in the concentration of PM 2.5 is associated with an estimated increase in mean (± SE) life expectancy of 0.61 ± 0.20 y ear. For the approximate period of 1980-2000, the average increase in life expectancy was 2.72 yr among the 211 counties in the analysis. The authors note that reduced air pollution was only one factor c ontributing to increased life expectancies, with its effects overlapping with those of other factors. Roman et al. (2008, 156921 ) applied state-of-the-art expert judgment elicitation techniques to develop probabilistic uncertainty distributions that reflect the broader array of uncertainties in the concentration-response relationship. This study foll owed best standard practices for expert elicitations based on the body of literature accumu lated over the past two decades. The resulting
PM 2.5 effect estimate distributions, elicited from 12 of the world’s leading experts on this issue, are shown in Figure 7-11 . They indicate both larger central estimates of mortality reductions for decreases in long-term PM 2.5 exposure in the U.S. (averaging almost 1% per μ g/m 3 PM 2.5 ) than reported (for example) by the ACS Study (i.e., 0.6% per μg/m 3 PM 2.5 in Pope et al. (2002, 024689 ),

December 2009 7-95and a wider distribution of uncertainty by each expert than provided by any one of the PM 2.5 epidemiologic studies. However, a composite uncertain ty range of the overall mean effect estimate (i.e., based upon all 12 experts’ estimates, but not provided in Figure 7-11 ) would be much narrower, and closer to that derived from the ACS study than indicated for any one expert shown in Figure 7-11

.

Source: Reprinted with Permission of ACS from Roman et al. (2008, 156921 )



### Figure 7-11. Experts’ mean effect estimates and uncertainty distributions for the PM 2.5 

### mortality concentration-response coefficient for a 1 μg/m 3 

### change in annual average PM 2.5


### .

Overall, recent evidence supports the strong evidence reported in the 2004 PM AQCD (U.S. EPA, 2004, 056905 ) that long-term exposure to PM 2.5 is associated with an increased risk of human mortality. When looking at the cause of d eath, the strongest evidence comes from mortality due to cardiovascular disease, with additional evidence supporting an association between PM 2.5 and lung cancer mortality ( Figure 7-7 ). Fewer studies evaluate the respiratory component of cardiopulmonary mortality, and the evidence to support an association with long-term exposure to
PM 2.5 and respiratory mortality is weak ( Figure 7-7 ). Together these findings are consistent and coherent with the evidence from epidemiologic, controlled human exposure, and animal toxicological studies for the effects of short- and lo ng-term exposure to PM on cardiovascular effects presented in Sections 6.2 and 7.2 , respectively. Evidence of short- and long-term exposure to PM 2.5 and respiratory effects (Sections 6.3 and 7.3 , respectively) and infant mortality (Section 7.4 ) are coherent with the weak respiratory mortality effects. Additionally, the evidence for short- and longterm cardiovascular and respiratory morbidity prov ides biological plausibility for mortality due to cardiovascular or respiratory disease. The most recent evidence for the association between longterm exposure to PM 2.5 and mortality is particularly strong fo r women. Collectively, the evidence is 

### sufficient to conclude that the relationship between long-term PM 2.5 

### exposures and mortality is causal.

December 2009 7-96

## 7.6.5.2.

7.6.5.3. PM 10-2.5

In the 2 004 PM AQCD, results from the ACS and Six Cities study analyses indicated that
PM 10-2.5 was not associated with mortality. Evidence is still limited to adequately characterize the association between PM 10-2.5 and PM sources and/or components. The new findings from AHSMOG and Veterans cohort studies provide limited evid ence of associations between long-term exposure to

PM 10-2.5 and mortality in areas with mean concentrations from 16 to 25 μg/m 3 . The evidence 

### for PM 10-2.5 

### is inadequate to determine if a causal relati onship exists between long-term exposures and mortality

.



## UFPs

The 2004 PM AQCD did not report long-term exposure studies for UFPs. No epidemiologic studies have been conducted to evaluate the effe cts of long-term UFP exposure and mortality. The
evidence is 

### inadequate to determine if a causal relationship exists between long-term UFP exposures and mortality

.

December 2009 7-97

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