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 epidemiologic 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 conclusions regarding causality based upon the framework 

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

such as cardiovascular effects, with coherence and plausibility being based upon the evidence from 

across disciplines and also across the suite of related 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 PM2.5, 

PM10-2.5, 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 and 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 PM2.5

exposure on cardiovascular mortality and found consistent 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 changes, 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 examined the long-term PM-CVD association among 

U.S. and European populations. The studies investigate the association of both PM2.5 and PM10

exposures with a variety of clinical and subclinical CVD outcomes. Epidemiologic and toxicological 

studies have provided evidence of the adverse effects of long-term exposure to PM2.5 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 7-1 
cardiovascular outcomes, including atherosclerosis, clinical and subclinical markers of 

cardiovascular morbidity, and cardiovascular mortality. The evidence of these effects from long-term 

exposure to PM10-2.5 is weaker. 

7.2.1. Atherosclerosis 

Atherosclerosis is a progressive disease that contributes to several adverse outcomes, 

including acute coronary syndromes such as myocardial infarction, sudden cardiac death, stroke and 

vascular aneurysms. It is multifaceted, beginning with an early injury or inflammation that promotes 

the extravasation of inflammatory cells. Under conditions of oxidative or nitrosative stress and high 

lipid or cholesterol concentrations, the vessel 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, and infarction or stroke. Several detailed reviews of 

atherosclerosis pathology have been published elsewhere (Ross, 1999, 156926; Stocker and Keaney, 

2004, 157013). 

7.2.1.1. Epidemiologic Studies 

Measures of Atherosclerosis 

Although no study has examined the association between long-term PM exposure and 

longitudinal change in subclinical markers of atherosclerosis, several cross sectional studies have 

been conducted. Markers of atherosclerosis used in these studies include coronary artery calcium 

(CAC), carotid intima-media thickness (CIMT), ankle-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 percentiles of the CAC distribution (Erbel et al., 

2007, 155768). 

CIMT is a measure of atherosclerosis assessed 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 7-2 
al., 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 resting (blood) pressure index, 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 having 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 assessed 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, particularly among women and at ages >50 yr. Although 

AAC has not been studied as extensively as CAC, 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 reproducible 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 enrolled in the Multi-Ethnic Study of Atherosclerosis 

(MESA), with long-term exposure to PM2.5 and PM10. 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 PM2.5 and PM10. 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 μg/m3 increase in 20-yr avg PM10 and PM2.5 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 μg/m3

 increase in PM10 and PM2.5, respectively. Among the 

subset of 2,586 participants with EBCT-identified calcification, similarly weak associations were 

observed. There was little evidence of modification of the CAC associations by demographic, 

socioeconomic or clinical characteristics. Finally, the authors report no differences in mean ABI with 

PM10 or PM2.5 concentrations. The null findings for ABI 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 PM2.5-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 PM2.5. 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 μg/m3

 increase in PM2.5 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 Intervention Trial (BVAIT), for their ancillary crosssectional analyses of the effect of long-term PM2.5 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 PM2.5 data from 23 state and local monitors 

December 2009 7-3 
operating in 2000 to estimate 1-yr avg exposure to PM2.5 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) μm increase in CIMT per 10 μg/m3

 increase in PM2.5 concentration. In contrast to findings from 

the relatively large, ethnically diverse, yet geographically overlapping MESA ancillary study 

described above, PM-related increases in CIMT were 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 the authors used dispersion and chemistry transport 

modeling of emissions, climate and topography data to estimate 1-yr avg exposure to PM2.5 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 μg/m3 increase in PM2.5. Differences in 

strength of association between subgroups defined by demographic and clinical characteristics were 

small. The authors reported a more consistent association of CAC with traffic exposure (distance 

from a major roadway) than with PM2.5 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-sectional study, no changes in ABI were observed in 

association with PM2.5 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 PM2.5 concentrations across space (as well as time to the extent baseline 

examinations were conducted over time). Such associations may reflect the effect of compositional 

differences in PM2.5 as well as the effect of higher PM2.5 concentrations. Most associations of PM2.5

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 measurement error, variation in baseline measures 

atherosclerosis, as well as limited power may contribute 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 PM2.5, is not distinguished from the other studies reviewed by a superior study design or larger 

sample size. The several fold difference in the 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 concentrations from these studies are characterized in Table 7-1. 

7.2.1.2. Toxicological Studies 

In the only study of this kind described in the 2004 PM AQCD, Suwa et al. (2002, 028588) 

demonstrated more advanced atherosclerotic lesions based on phenotype and volume fraction in the 

left main and right coronary arteries of rabbits exposed to PM10 (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 atherosclerotic plaque area in aortas of ApoE-/- mice 

exposed to PM2.5 CAPs for 4-6 mo (6 h/day×5 days/wk). Average CAPs concentrations ranged from 

85 to 138 μg/m3

 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 7-4 
and cellularity in the aortic root were enhanced by Tuxedo, NY CAPs exposure, although there was 

no change in lipid content. Genetic profiles 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. Plaque 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 infiltration 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 PM2.5 CAPs exposure increased atherosclerotic plaque area and led to higher levels of 

macrophage infiltration, collagen deposition, and 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. PM2.5-exposed animals exhibited increased vasoconstrictor 

responsiveness to serotonin and PE. Increased ROS and 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 induced by PE following Manhattan CAPs exposure. 

Pretreatment with the soluble guanylate cyclase (sGC) inhibitor ODQ attenuated the response, 

indicating that CAPs exposure resulted in abnormal NO/sGC signaling. Expression of iNOS mRNA 

and protein was increased in aortas of CAPs-exposed mice, further supporting a role for NO 

production. In conjunction with increased NO, aortic 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. However, Manhattan CAPs-exposed mice had a markedly 

decreased response to A23187, indicating that NO release occurred via Ca2+-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 Ca2+-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 PM2.5-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 expression 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 PM2.5 or UFPs derived from traffic was conducted. Araujo et al. 

(2008, 156222) compared the relative impact of UF (0.01-0.18 μm) and fine (0.01-2.5 μm) 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 μg/m3

 for PM2.5 and ~110 μg/m3 for UFPs, and the number concentrations 

were roughly equivalent (4.56×105

 and 5.59×105 particles/cm3 for PM2.5 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 comprised of macrophages with intracellular lipid 

accumulation. Increased total cholesterol measured at the end of the exposure protocol was observed 

only in the PM2.5 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 products (catalase, glutathione S-transferase Ya, NADPH-quinone oxidoreductase1, superoxide dismutase 2). Thus, UFPs in these exposures had a 

substantially greater impact on the systemic response than did PM2.5. 

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 arterial 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, PM10 and NO2 were filtered with 50% and 75% efficiency, 

respectively. The average pollutant concentrations were 2.06 ppm for CO (8-h mean), 104.75 μg/m3

for NO2 (24-h mean), 11.07 μg/m3

 for SO2 (24-h mean), and 35.52 μg/m3 for PM10 (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 examination suggested that the increases in L/W ratio 

were due to muscular hypertrophy rather than fibrosis. 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 and 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 PM10 (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 PM10. 

In addition, labeled monocytes were attached onto the endothelium overlying atherosclerotic plaques 

and the number that migrated into the smooth muscle underneath plaques in aortic vessel walls was 

greater with PM10 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 endothelium and migration into plaques. Monocytes in 

plaque tissue stained with immunogold demonstrated foam cell characteristics, which were more 

numerous in the rabbits exposed to PM10. 

Gasoline Exhaust 

Lund and colleagues (2007, 125741) used whole emissions from gasoline exhaust to 

investigate changes in the transcriptional regulation 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 μg/m3

 for PM, 19 ppm for NOX, 80 ppm for CO, 

and 12.0 ppm for total hydrocarbons. The average particle number median diameter was 

approximately 15 nm (McDonald et al., 2007, 156746). Dilutions of gasoline engine emissions 

induced a concentration-dependent increase in transcription 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 high-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 PM10 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 with a hypercoagulable, prothrombotic state. 

December 2009 7-6 
7.2.2.1. Epidemiologic 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 lower extremity deep vein thrombosis (DVT) and 

1,210 of their healthy friends or relatives (1995-2005), the authors used arithmetic averaging of PM10

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 well 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 μg/m3

 increase in PM10. 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. Epidemiologic 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 PM10. They found that after 

adjustment, residents of communities with lower PM10 concentrations had fewer white blood cells 

than residents of communities with higher PM10 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 individuals with different degrees of metabolic 

abnormalities supports the concept that the presence of a metabolic syndrome may impart greater 

susceptibility to PM-associated long-term CVD effects. 

7.2.3.2. Toxicological 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 PM2.5 CAPs exposure (mean mass concentration 72.7 μg/m3

; 

Tuxedo, NY) on insulin resistance, adipose inflammation, 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. Aortic rings of mice exposed to CAPs demonstrated 

decreased peak relaxation to ACh or insulin, which 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 PM2.5 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 Function, and Blood Coagulation 

7.2.4.1. Epidemiologic 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 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 PM10 (median concentration within quartiles = 23.1, 31.2, 38.8 and 53.7 μg/m3

). They 

found that after adjustment, residents of communities in quartile 1 had 138 (95% CI: 2-273) fewer 

white blood cells (×106

/L) than residents of communities in quartiles 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-analysis of year-specific crosssectional findings, the authors used dispersion modeling of emissions and weather data to estimate 

2-yr avg exposures to PM10 at participant postcode sector centroids (median in 1994, 1998 and 2003 

= 19.5, 17.9 and 16.2 μg/m3, respectively). They found little evidence of a PM10-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 μg/m3

 increase in PM10. 

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 Polotitlán (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 (×109

/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 among 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 PM10 concentration and markers of immune function or 

inflammation have been conducted with significant 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 differs across the communities studied (Calderon-Garciduenas et al., 

2007, 091252; Calderón-Garcidueñas et al., 2009, 192107). 

7.2.4.2. Toxicological Studies 

In addition to the PM2.5 study mentioned previously that showed 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 hematology and clotting parameters in rats 

and mice exposed to DE, gasoline exhaust, or hardwood 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 μg/m3 and the 

MMAD was 0.10-0.15 μm (Reed et al., 2004, 055625). Male and female rats exposed to DE at the 

highest concentration (NO concentration 45.3 ppm; NO2 concentration 4.0 ppm; CO concentration 

29.8 ppm; SO2 concentration 365 ppb) for 6 mo demonstrated decreased serum Factor VII, but no 

change in plasma fibrinogen or thrombin anti-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 μg/m3, CO 

1.3 and 4.8 ppm, NO <2.5 and 5.9 ppm, NO2 <0.25 and 1.2 ppm, SO2 0.2 and 0.3 ppm for low and 

high PM exposures, respectively) did not demonstrate 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 μg/m3

and the MMAD was 0.25-0.36 μm (Reed et al., 2006, 156043). In male rats exposed to hardwood 

smoke, the mid-low group (PM concentration 113 μg/m3; NO, NO2, SO2 concentrations 0 ppm; CO 

concentration 1,832.3 ppm) had the greatest responses in hematology parameters, including 

increased hematocrit, hemoglobin, lymphocytes, and decreased segmented neutrophils (Reed et al., 

2006, 156043). Platelets were elevated in male and female 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 μg/m3, with the corresponding 

number concentration between 2.6×104

 and 5.0×105 particles/cm3; 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; NO2 concentrations 0.5 

and 0.9 ppm; CO concentration 73.2 and 107.3 ppm; SO2 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 increased 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 tested 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 albumin 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 albumin excretion and can be used to identify 

albuminuria, a marker of generalized vascular 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 with 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 7-9 
dichotomization of albuminuria into question, observing that there is no threshold below which risk 

of cardiovascular and end-stage kidney disease 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 cardiovascular disease. Franklin et al. (1997, 156446) suggested 

that SBP and PP were the only two measures predictive 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. Epidemiologic Studies 

O’Neill et al. (2007, 156006) examined the association of UACR with PM2.5 and PM10 among 

members of the MESA population described previously (Diez et al., 2008, 156401). For this study of 

UACR, which included cross-sectional and longitudinal 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 exams (2000-2004). In cross-sectional analyses, the 

authors found that after adjustment for demographic and clinical characteristics, 10 μg/m3

 increases 

in 20-yr imputed exposures to PM2.5 and PM10 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) PM2.5 and PM10 exposures. Moreover, longitudinal analyses revealed only a weak 

association between 3-yr change in log UACR and 20-yr PM10 exposure. Evidence of effect 

modification by demographic and geographic characteristics 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 PM2.5 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 PM2.5. 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/m3

 increase in 30-day avg PM2.5 exposure, 

Associations were slightly weaker for 60-day avg PM2.5 exposure and among participants without 

hypertension, during cooler weather, in the presence of low NO2, 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 PM2.5) than those of 

the Southwest (historically lower PM2.5). 

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 PM2.5 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 PM2.5 concentrations. PM concentrations from the analyses are 

characterized in Table 7-1. 

December 2009 7-10 
Table 7-1. Characterization of ambient PM concentrations from studies of subclinical measures of 

cardiovascular diseases and long-term exposure. 

Study Location Mean Concentration (μg/m3) Upper Percentile 

Concentrations (μg/m3) 

PM10

 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 PM10 (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 

PM2.5 

 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. Toxicological Studies 

In a PM2.5 CAPs study of 10 wk (6 h/day×5 days/wk) in Tuxedo, NY (mean mass 

concentration 79.1 μg/m3

), 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 group. 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 7-11 
NAD(P)H oxidase inhibitor (apocynin) or a NOS inhibitor (L-NAME) attenuated the superoxide 

generation. The levels of tetrahydrobiopterin (BH4) were decreased in mesenteric vasculature and the 

heart by 46% and 41% in the PM+Ang II group compared to controls, respectively; furthermore, 

levels of BH4 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 triggering of the Rho/ROCK signaling pathway. 

7.2.6. Autonomic Function 

7.2.6.1. Toxicological 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 μg/m3

 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 increase 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 μg/m3

. 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. Toxicological studies 

Two recent toxicological studies have evaluated the effects of PM on cardiac effects including 

pathology and gene expression. Cardiac mitochondrial 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 μg/m3, CO 1.3 or 

4.8 ppm, NO <2.5 or 5.9 ppm, NO2 <0.25 or 1.2 ppm, SO2 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 aconitase activity in both rat strains was observed, 

indicating a DE effect on oxidative stress. It should 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 ZnSO4 (9, 35, or 

120 μg/m3

) 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 identified alterations in cardiac genes involved in cell 

signaling events, ion channels regulation, and coagulation in animals exposed to the highest ZnSO4

concentration only. This study demonstrates a possible direct effect of ZnSO4 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/m2

) 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 PM2.5 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/m2

 and -3.0% (-8.0 to 2.0) differences 

in LVMI and EF per 10 μg/m3 increment in PM2.5. 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 PM2.5 (i.e., 1.4 g/m2

 [95% CI: 0.3-2.5] per 10 μg/m3.) 

7.2.9. Clinical Outcomes in Epidemiologic Studies 

Several epidemiologic studies of U.S. and European populations have examined associations 

between long-term PM exposures and clinical 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 authors used arithmetic averaging of year 2000 EPA AQS 

PM2.5 data available at the monitor nearest to each participant’s geocoded U.S. Postal Service fivedigit ZIP code centroid to estimate 1-yr avg exposures. 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 7-13 
their combination with CHD death per 10 μg/m3

 increase in PM2.5, respectively. Hazards were higher 

within than between cities and in the obese. For the combined CVD outcome (MI, revascularization, 

stroke, CHD death, cerebrovascular disease), authors 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 PM2.5-CVD association was stronger among non-diabetic than 

diabetic participants. 

Table 7-2. Characterization of ambient PM concentrations from studies of clinical cardiovascular 

diseases and long-term exposure. 

Study Location Mean Annual 

Concentration (μg/m3) 

Upper Percentile 

Concentrations (μg/m3) 

PM10

 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 PM10 (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 PM10 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 

PM2.5 

 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 yr) 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 PM10 exposures from U.S. EPA AQS and emissions, IMPROVE, and 

Harvard University monitor data. They found a 10% (95% CI: -6 to 29) increase in risk of first CHD 

event per 10 μ/m3

 increase in 1-yr avg PM10 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-stratified 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 PM10 (median = 2.4 μg/m3

). They found that the OR for prevalent MI 

per 10 μg/m3

 increase in PM10 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, validated MI, the authors used dispersion modeling of 

traffic emissions and land use data to estimate 5-yr avg exposure to PM10. They found that after 

December 2009 7-14 
adjustment for demographic, temporal, and socioeconomic characteristics, the OR for MI per 

5 μg/m3

 increase in PM10 was 1.04 (95% CI: 1.00-1.09). ORs were higher after restriction to fatal 

cases, in- or out-of-hospital deaths, and participants who did not move between population censuses. 

Authors state that control for confounding was superior 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 PM10 data available in the county of hospitalization to estimate 1-yr avg exposures. They found 

17% (95% CI: 5-31) and 11% (95% CI: 3-21) increases in the risk of recurrent MI and postinfarction CHF, respectively, per 10 μg/m3

 increase in PM10 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 (age 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 PM2.5 (mean = 23.3 μg/m3

). They found little evidence of an 

association between PM2.5 and CHD in these data. After adjustment for geographic, demographic 

and clinical characteristics, the OR for prevalent CHD per 10 μg/m3 increase in exposure was 0.55 

(95% CI: 0.14-2.11). 

Study Avg Time Endpoint Effect Estimate (95% CI) 

Prospective Cohort Studies 

Incident MI PM2.5

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 PM10

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 PM10

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 PM2.5

Hoffman et al. (2009, 190376) 

Prevalent cases at baseline in 3,399 residents, 

cross-sectional, 2 German cities 

1 yr Self-reported PVD 

Zanobetti & Schwartz (2007, 091247) Recurrent MI Hospitalization PM10

196,131 MI survivors (ICD9 410), readmitted for 

MI or CHF, ecologic, open cohort, 21 US cities 

1 yr 

Post-MI CHF Hospitalization 

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

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 PM2.5 and PM10. 

December 2009 7-15 
In the study of 1,030 census enumeration districts 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 climate data to estimate 5-yr avg exposure to 

PM10. They found that after adjusting for smoking prevalence, 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 

PM10 quintiles. The linear trend was somewhat 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 PM2.5 rather than PM10. This study, which provides 

strong evidence of an association, was distinguished 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 

PM10 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 PM2.5, while findings from two ecologic studies of PM10 indicated positive 

associations of CHD hospitalizations with PM10 (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 cerebrovascular disease, respectively, per 10 μg/m3

 increase 

in 1-yr avg PM2.5 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 PM10. 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 PM10 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 PM2.5

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 μg/m3

 increase in PM2.5. 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 PM10-

DVT association (Baccarelli et al., 2008, 157984). The authors found a 70% (95% CI: 30-223) 

increase in the odds of DVT per 10 μg/m3

 increase in 1-yr avg PM10 exposure. This finding was 

consistent with the decreases in PT and PTT also observed among controls in this context as well as 

December 2009 7-16 
the 47% (95% CI: 11-96) increase in the odds of DVT per inter-decile range (242 m) increase in the 

residence-to-major-roadway distance observed among a subset of cases and controls (Baccarelli et 

al., 2009, 188183). The PM10-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 consistent association between long-term exposure to 

PM2.5 and increased risk of cardiovascular mortality. There is little evidence for the long-term effects 

of PM10-2.5 on cardiovascular mortality. This section focuses 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 PM2.5 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 group 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/m3

PM2.5). 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 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 PM2.5

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/m3 increase of PM2.5 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 information, 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 heart 

problems (cardiovascular events) experienced by the subjects during the study. These morbidity coassociations with PM2.5 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 Adventist 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/m3

 PM2.5), but not among males (53 deaths; RR = 0.90 [95% CI: 

0.76-1.05] per 10 μg/m3

 PM2.5) (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/m3

 PM2.5). The 

authors speculated that females may be more sensitive to air pollution-related effects, based on 

differences between males and females in dosimetry and exposure. As was found with PM2.5, a 

positive association with coronary heart disease mortality was reported for PM10-2.5 and PM10 among 

females (RR = 1.38 [95% CI: 0.97-1.95] per 10 μg/m3

 PM10-2.5; RR = 1.22 [95% CI: 1.01-1.47] per 

10 μg/m3 PM10), but not for males (RR = 0.92 [95% CI: 0.66-1.29] per 10 μg/m3 PM10-2.5; RR = 0.94 

[95% CI: 0.82-1.08] per 10 μg/m3

 PM10); associations were strongest in the subset of 

postmenopausal women (80 deaths) (Chen et al., 2005, 087942). 

Two additional studies explored the effects of PM10 on cardiovascular mortality. The Nurses’ 

Health Study (Puett et al., 2008, 156891) is an ongoing prospective cohort study examining the 

relation of chronic PM10 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 greatest 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 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 

December 2009 7-17 
that cardiopulmonary mortality was associated with PM10 (RR = 1.52 [95% CI: 1.09-2.15] per 

10 μg/m3

 PM10). 

In summary, the 2004 PM AQCD concluded that there was strong evidence that long-term 

exposure to PM2.5 was associated with increased cardiopulmonary 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 PM2.5 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 addition, biological plausibility and coherence are 

provided by toxicological studies demonstrating short-term cardiovascular effects as well as 

PM2.5-related plaque progression in chronically exposed mice. 

7.2.11.Summary and Causal Determinations 

7.2.11.1. PM2.5 

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 PM2.5 exposure on 

cardiovascular mortality and found strong and consistent 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 have provided evidence of the adverse effects of 

long-term exposure to PM2.5 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 PM2.5 exposure comes 

from epidemiologic studies of cardiovascular mortality. A number of large, multicity U.S. studies 

(the ACS, Six Cities Study, WHI, and AHSMOG) provide consistent evidence of an effect between 

long-term exposure to PM2.5 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/m3

(Table 7-8). An epidemiologic study investigating the relationship between PM2.5 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 in the adjusted risk of MI, revascularization, and 

stroke using a 1-yr avg PM2.5 concentration (mean = 13.5 μg/m3

). 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 PM2.5 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 PM2.5 effects and their consistency across different measures of atherosclerosis in 

these studies varies widely, and they may be limited in their ability to discern small changes in these 

measures. Kunzli et al. (2005, 087387) observed a non-significant 4.2% increase in CIMT associated 

with long-term PM2.5 exposure among participants of a clinical 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 PM2.5

exposure were largely null (Diez et al., 2008, 156401), while an increase in AAC with long-term 

PM2.5 exposure was reported (Chang et al., 2008, 180393). By contrast, a 43% increase in CAC was 

associated with long-term PM2.5 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 PM2.5 and PM10 exposure. 

Evidence of enhanced atherosclerosis development 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 (Section7.2.1.2). In addition, alterations in vasoreactivity were observed, 

December 2009 7-18 
suggesting an impaired NO pathway. Additional toxicological studies of PM10 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, PM2.5 CAPs derived from 

traffic in Los Angeles did not affect plaque size (Araujo et al., 2008, 156222). Collectively, these 

toxicological studies provide biological plausibility for the associations reported in epidemiologic 

studies. 

There is limited evidence for the effects of PM2.5 on renal or vascular function. Cross-sectional 

and longitudinal epidemiologic analyses of PM2.5 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 PM2.5 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. cohort studies report associations of long-term PM2.5

concentration with cardiovascular mortality. These studies provide the strongest evidence for an 

effect of long-term PM2.5 exposure on CVD effects. Additional evidence comes from a 

methodologically rigorous epidemiology study that demonstrates coherent associations between 

long-term PM2.5 exposure and CVD morbidity among post-menopausal women. Toxicological 

studies demonstrate that this effect is biologically plausible and the effect is coherent with studies of 

short-term PM2.5 exposure and CVD morbidity and mortality, and with long-term exposure to PM2.5

and CVD mortality. Associations between PM2.5 and subclinical measures of atherosclerosis are 

inconsistent, but cross-sectional studies may be limited in their ability to discern small changes in 

these measures. In addition, potential modification of the PM2.5-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 PM2.5 exposures and cardiovascular effects.

7.2.11.2. PM10-2.5

One epidemiologic study evaluated the relationship between long-term exposure to PM10-2.5

and cardiovascular mortality and found a positive association 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 

PM10-2.5 and cardiovascular effects have been conducted to date. Evidence is inadequate to infer 

the presence or absence of a causal relationship. 

7.2.11.3. UFPs 

A few toxicological studies of long-term exposure 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 reported 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 the development of hypertension (Gottipolu et al., 

2009, 190360). Whole and filtered gasoline exhaust induced 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 may 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 – PM10, PM10-2.5 and PM2.5 – 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 symptoms in children living in communities with higher 

PM10 and PM2.5 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 PM10, acid vapor and NO2 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 

PM10 showed increased growth in lung function compared with subjects who moved to communities 

with higher PM10 concentrations. Finally, the 2004 PM AQCD concluded that there was strong 

epidemiologic evidence for associations between long-term exposures to PM2.5 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 FEV1 and McConnell et al. (1999, 007028) 

reported increased ORs for bronchitic symptoms in asthmatics for PM10 and PM2.5. 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 PM2.5 and PM10, and the studies showing associations only with 

PM10 were conducted in locations where the PM consisted predominantly of fine particles, providing 

support for associations with long-term exposure to fine 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 bronchiolar 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 been 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, inflammation, histopathological changes and oxidative and 

allergic responses following PM2.5 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. Epidemiologic 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 7-20 
Bayer-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 average reduction of symptom prevalence was more 

pronounced in areas with stronger reduction of air pollution levels. The average decline of PM10

between 1993 and 2000 across the nine study regions was 9.8 μg/m3

 (29%). Declining levels of PM10

were associated with declining prevalence of chronic cough, bronchitis, common cold, nocturnal dry 

cough, and conjunctivitis symptoms, but no significant 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 PM10 levels (r = 0.78; p = 0.02). Similar associations were seen 

for nocturnal dry cough and conjunctivitis symptoms and PM10 levels. Röösli et al. (2000, 010296; 

2001, 108738; 2005, 156923) have demonstrated that PM10 levels are homogeneously distributed 

within regions of Basel, Switzerland and are not substantially affected by local traffic, justifying the 

single-monitor approach for assignment of PM10 exposures. Based on parallel measurements of 

PM2.5 and PM10 at seven sites in Switzerland, PM2.5 and PM10 at all sites are generally highly 

correlated (r2

 ranging from 0.85 to 0.98) (Gehrig and Buchmann, 2003, 139678), indicating that 

PM10 consists predominantly of fine particles in these locations. 

Schindler et al. (2009, 191950) reported that sustained reduction in ambient PM10

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 PM10 concentrations served as a natural experiment. Each subject was assigned modelbased estimates of PM10 concentrations averaged over the 12 mo preceding each health assessment 

with mean decline in PM10 levels of 6.2 μg/m3

 (SD = 3.9 μg/m3). When the authors tested the joint 

hypothesis of no association between the PM10 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 PM10 (particles directly 

emitted from local sources/traffic) and the prevalence 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 PM10 was calculated for the home address 

using the Airviro statistical dispersion model. After adjusting for confounders, mean annual exposure 

to locally generated PM10 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 study area were estimated with emission databases and 

statistical dispersion modeling. Children were examined at 2 mo and 1, 2, and 4 yr of age. Using GIS 

methods, the average contribution of traffic-generated PM10 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 PM10 with PM2.5

measurements from 42 locations (Hoek et al., 2002, 042364) and reported modeled traffic-generated 

PM10 correlated reasonably well with measured PM2.5 (r = 0.61). Persistent wheezing (cumulative 

incidence up to age 4 yr) was associated with exposure to traffic-generated PM10 (OR 2.28 [95% CI: 

0.84-6.24] per 10 μg/m3 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. Characterization of ambient PM concentrations from studies of respiratory 

symptoms/disease and long-term exposures. 

Study Location Mean Annual Concentration (μg/m3) 

Upper Percentile 

Concentrations 

(μg/m3) 

PM2.5 

 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 

PM10

 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; PM10 from traffic 

**Only reported change in PM concentration 

December 2009 7-22 
Source: 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/m3 PM10 levels in Swiss Surveillance Program of 

Childhood Allergy and Respiratory Symptoms1. Inset: Mean change in adjusted 

prevalence (1998-2001 to 1992-1993) versus mean change in regional annual 

averages of PM10 (1997-2000 to 1993) for chronic 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 from 1996 to 1999. They investigated both the 

differences between- communities with 4-yr avg and within-communities yearly variation in PM 

(i.e., PM10, PM2.5, PM10-2.5, EC, and OC). Based on a 10 μg/m3 change in PM2.5, 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 μg/m3 range in 4-yr avg PM2.5 concentrations across the 12 

communities was 1.29 (95% CI: 1.06-1.58). Similar results were reported for PM10 and PM10-2.5 but 

the effect estimates were smaller in magnitude and generally not statistically significant. Withincommunity associations were not confounded by any time-fixed personal covariates. In two-

 

1

 Adjusted for age, sex, nationality, parental education, number of siblings; farming status, low birth weight, breastfeeding, child who 

smokes, family history of asthma, bronchitis, and/or atopy, mother 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 temperature <0°C, and belief of mother that there is an association 

between environmental exposures and children’s respiratory health 

December 2009 7-23 
pollutant models, the within-community effect estimates for PM2.5 and OC were significant in the 

presence of several other pollutants. While the within-community single-pollutant effect of PM2.5

(ß = 0.085/μg/m3

) was only modestly attenuated after adjusting for some pollutants, it was markedly 

reduced after adjusting for NO2 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 pulmonary 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 FEF25-75 values ≥ 120% to 

20.4/1,000 person-years for children with FEF25-75 value ≤ 100%. Over the 10th-90th percentile 

range for FEF25-75 (57.1%), the hazard ratio of new onset asthma was 0.50 (95% CI: 0.35-0.71). The 

IR of asthma for FEF25-75 ≥ 120% in the “high” PM2.5 (13.7-29.5 μg/m3

) communities was 15.9/1,000 

person-years compared to 6.4/1,000 person-years in “low” PM2.5 (5.7-8.5 μg/m3) communities. Loss 

of protection by high lung function against new onset asthma in the “high” PM2.5 communities was 

observed for all the lung function measures. Figure 7-3 shows the effect of PM2.5 on the association 

of lung function with asthma. Of all the pollutants examined (NO2, PM10, PM2.5, acid vapor, O3, EC, 

and OC), PM2.5 appeared to have the strongest modifying effect on the association between lung 

function with asthma as it had the highest R2

 value (0.42). Over the 10th-90th percentile range of 

FEF25-75, the hazard ratio of new onset asthma was 0.34 (95% CI: 0.21-0.56) in a community with 

low PM2.5 (<13.7 μg/m3

) and 0.76 (95% CI: 0.45-1.26) in a community with high PM2.5

(≥ 13.7 μg/m3

). 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 Society from Islam et al. (2007, 090697) 

Figure 7-3. Effect of PM2.5 on the association of lung function with asthma. Communityspecific hazard ratio of newly diagnosed asthma over 10-90th percentile range 

(57.1%) of FEF25-75% by level of ambient PM2.5 (μg/m3). 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 PM2.5 concentration at the home address with a validated 

model using GIS. PM2.5 was associated with doctor-diagnosed asthma (OR = 1.32 [95% CI: 

December 2009 7-24 
1.04-1.69]) for a cumulative lifetime indicator. These 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 PM10 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 PM10 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 reductions in PM decrease asthma risk. A strong feature 

of SAPALDIA is the ability to assign space, time, 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 disease and its exacerbations. The added impact of 

causing disease increases the risk compared to only exacerbations. 

A matched case-control study of infant bronchiolitis (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 PM2.5 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 PM2.5 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 μg/m3

 increases in both sub-chronic and chronic PM2.5 exposure, an 

adjusted OR of 1.09 (95% CI: 1.04-1.14) was observed. In multipollutant model analyses, the 

association with PM2.5 was robust to the inclusion of gaseous pollutants. Also, in a cohort of children 

in Germany, Morgenstern et al. (2008, 156782) modeled PM2.5 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 PM2.5 and PM10 

was 10.8 mi (SD 7.8) and 11.5 mi (SD 7.9), respectively. PM2.5 and PM10 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 μg/m3 increase in PM2.5 or PM10 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 

PM2.5 and PM10 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 tracheobronchial 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 PM2.5 monitored in school yards. Atopic asthma was related to PM2.5 (OR 1.43 

[95% CI: 1.07-1.91]) when high PM2.5 concentrations (20.7 μg/m3

) were compared to low PM2.5

concentrations (8.7 μg/m3). 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 PM10.

Kim et al. (2004, 087383) conducted a school-based cross-sectional study in the San Francisco 

metropolitan area in 2001 comprised of 10 neighborhoods to examine the relationship between 

traffic-related pollutants and current bronchitic symptoms and asthma obtained by parental 

questionnaire (n = 1,109). They related traffic-related pollutants (PM) and bronchitic and asthma 

symptoms in the past 12 mo. No multipollutant models were evaluated because of the high 

interpollutant correlations. PM2.5 levels ranged across the school sites from 11 to 15 μg/m3

. 

Schikowski et al. (2005, 088637) examined the relationship between both long-term air 

pollution exposure and living close to busy roads and COPD in the Rhine-Ruhr Basin of Germany 

December 2009 7-25 
from 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 PM10 was estimated from using a 

conversion factor (obtained from parallel measurement of TSP and PM10 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 PM10 and living near a major road might increase the risk of 

developing COPD and can have a detrimental effect 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 PM2.5 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 PM10 in areas where fine particles are 

the predominant fraction of PM10. 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 the CHS, Islam et al. (2007, 090697) showed that PM2.5 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 PM2.5 communities was observed 

for 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 

– PM10, PM10-2.5 and PM2.5 – 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 and 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 

PM10 showed increased growth in lung function compared with subjects who moved to communities 

with higher PM10 concentrations who showed decrease growth in lung function. 

7.3.2.1. Epidemiologic Studies 

New longitudinal cohort studies have evaluated the relationship between long-term exposure 

to PM and changes in measures of pulmonary function (FVC, FEV1, and measures of expiratory 

flow). Cross-sectional studies also offer supportive information (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. Characterization of ambient PM concentrations from studies of FEV1 and long-term 

exposures. 

Study Location Mean Annual 

Concentration (μg/m3) Upper Percentile Concentrations (μg/m3)

PM2.5 

 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 

PM10

 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, FEV1, 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 communities from 1994-2000. The results for O3, PM10, NO2, 

PM2.5, 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 PM10 and age-related decline in lung function (FVC, 

FEV1, and FEF25-50). An evaluated statistical dispersion model (Liu et al., 2007, 093093) provided 

spatially resolved concentrations of PM10 that enabled assignment to residential addresses for the 

participant examinations in 1991 and 2002 that yielded a median decline of 5.3 μg/m3

 (IQR 4.1-7.5). 

Decreasing PM10 concentrations attenuated the decline in lung function. Effects were greater in tests 

reflecting small airway function. No other pollutant relationships were evaluated, though a related 

study indicated that levels of NO2 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 particles have respiratory effects. 

December 2009 7-27 
Source: 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 FEV1 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 diagnosed with asthma, and were located in Mexico 

City, Rojas-Martinez et al. (2007, 091064) evaluated the association between long-term exposure to 

PM10, O3 and NO2 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 PM10 over the age of 8-10 yr of life in this cohort on FVC, 

FEV1, and FEF25-75 showed an association. Single pollutant models showed an association between 

ambient pollutants (O3, PM10 and NO2) and deficits in lung function growth. The association 

between PM10 and FEF25-75 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 because the traffic-related pollutants were correlated. 

December 2009 7-28 
Although no PM2.5 data were presented in this study, in a separate study Chow et al. (2002) report 

that during the winter of 1997 approximately 50% of PM10 was in the PM2.5 fraction in Mexico City. 

Gotschi et al. (2008, 156485) examined the relationship between air pollution and lung 

function in adults in the European Community Respiratory Health Survey (ECRHS). FEV1 and FVC 

were assessed at baseline and after 9 yr of follow-up from 21 European centers (followed-up sample 

n = 5,610). No statistically significant associations were found between city-specific annual mean 

PM2.5 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 PM2.5 and PM10 on lung function (FVC, FEV1, FEF50%). Spirometry was performed in 

2,307 children aged 9-10 yr in 2001-2002. Residential 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 for contextual socioeconomic factors diminished 

associations. Results for PM10 were similar to those for PM2.5. 

In an exploratory study, Mortimer et al. (2008, 187280) examined the association of prenatal 

and lifetime exposure to air pollutants using geocoded monthly average PM10 levels with pulmonary 

function in a San Joaquin Valley, California cohort of 232 children (ages 6-11 yr) with asthma. First 

and second trimester PM10 exposures (based on monthly average concentrations) had a negative 

effect on pulmonary function and may relate to prenatal 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 pollutants, and indicators of motor vehicle emissions in 

Windsor, Ontario, in a cohort of 2,402 school children. PM2.5 and PM10 concentrations were 

estimated for each child’s residence at the postal code level. Each 10 μg/m3

 increase in PM2.5 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 PM10. Near each child’s home, 

exposure to PM10 was estimated using a statistical dispersion model (Pierse et al., 2006, 088757). 

The authors reported a dose-dependent inverse association 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 FEF25-75. 

Caution should be used when interpreting these results as the accuracy of the estimates on individual 

PM10 exposures were not validated; there is potential for confounding by ethnic origin; and there is 

concern that the magnitude of the changes in pulmonary 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 associated with exposure to traffic-related PM10 (-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 FEV1 and PM2.5 concentrations in both cross-sectional and 

longitudinal analyses. 

In summary, recent studies have greatly expanded the evidence available for the 2004 PM 

AQCD. The earlier CHS studies followed young children for 2-4 yr. New analyses have been 

conducted that include longer follow-up periods of this 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 PM2.5 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 longitudinal studies have been conducted in other 

locations. 

Gauderman et al. (2004, 056569) reported that PM2.5 exposure was associated with clinically 

and statistically significant deficits in FEV1 attained at the age of 18 yr. Clinical significance was 

December 2009 7-29 
defined as a FEV1 below 80% of the predicted value, a criterion commonly used in clinical settings 

to identify persons at increased risk for adverse respiratory 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 PM10, available data 

discussed above indicate that the majority of PM10 is composed of PM2.5 in these areas. New studies 

provide positive results from Mexico City, Sweden, and 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 similar between the studies for FVC and FEV1. 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 PM2.5 and PM10, 

while most did not evaluate PM10-2.5. Associations have been reported with fine particle components, 

particularly EC and OC. Source apportionment methods 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 poorly controlled asthma among adults who were 

respondents to the California Health Interview Survey and found associations for traffic density and 

PM10, but not PM2.5. 

7.3.2.2. Toxicological 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. Male 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/m3

 PM2.5. Thus PM levels were reduced by 

filtration but not entirely eliminated. Ambient concentrations of CO, NO2 and SO2 were 1.7 ppm, 

89.4 μg/m3

 and 8.1 μg/m3, respectively. Concentrations of gaseous pollutants were assumed to be 

similar to ambient levels in both chambers. After 4 mo, the animals were mated and the offspring 

were divided into 4 groups to provide for a prenatal 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. Pulmonary pressure-volume measurements were also 

conducted in the 90-day-old offspring. Statistically significant reductions in inspiratory and 

December 2009 7-30 
expiratory volumes were found in the group receiving both prenatal and postnatal exposure, but not 

in the groups receiving only prenatal exposure or only 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/m3) 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 NO2, and less than 0.01 ppm SO2. 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 Section 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 gases were very low suggesting that PM may have 

been responsible for the observed effects. 

In many animal studies changes in ventilatory 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 indicator 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 terminology 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/m3

) 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 NO2, 0.2 and 

0.3 ppm SO2 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 whole-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/m3

. 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 particles 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 NOX 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/m3

 woodsmoke. Additional effects were found at 10,000 μg/m3. 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. Epidemiologic Studies 

One epidemiologic study examined the relationship 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 PM2.5 and PM10-2.5 for each child’s 

residence at the postal code level with an evaluated statistical model (Wheeler et al., 2006, 103905). 

Each 10 μg/m3

 increase in 1-yr PM2.5 was associated with a 39% increase in eNO (p = 0.058). 

Associations between eNO and PM10-2.5 were positive but not statistically significant. 

7.3.3.2. Toxicological 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 PM2.5 exposure 

concentration was 110 μg/m3

. 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 and 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-κB 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 μm) versus fine 

(0.01-2.5 μm) 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 μg/m3

 for PM2.5 exposures and 

~110 μg/m3

 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 (4.56×105

 and 5.59×105 particles/cm3, respectively). Over 50% of the UFPs were 

comprised of OC compared to only 25% for PM2.5. 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 of 

PM 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/m3

 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 7-32 
2.93-35.67 ppm NOX, 0.23-4.57 ppm SO2, 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/m3

 PM. When rats were 

exposed to DE containing 1,000 μg/m3 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 was also found to mediate the increase in protein 

levels, the decrease in PGE2 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 numbers 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 lymphocytes 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 since comparisons were made with Day 0 controls 

rather than age-matched controls. No histopathological 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/m3

) for 1 or 3 mo. Concentrations of gases were reported 

to be 3.5-9.5 ppm CO, 2.2-14.8 ppm NOX, and less than 0.01 ppm SO2. 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 particle-laden alveolar macrophages in alveoli 

and peribronchial tissues at both time points. Bronchus-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 macrophages) 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 mRNA was observed in DE-exposed mice at 1 and 

3 mo. NF-κB activation was also noted following 1-month 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/m3

) by whole body inhalation for 6 h/day, 

7 days/wk for either 1 week or 6 mo. Concentrations of gases were reported to be 2.0-45.3 ppm NO, 

0.2-4.0 ppm NO2, 1.5-29.8 ppm CO and 8-365 ppb SO2. 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 alveolar 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 filtered air or DE at concentrations ranging from 

30-1000 μg/m3

 PM. Concentrations of gases were reported for the highest exposure as 45.3 ppm NO, 

4.0 ppm NO2, 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/m3

 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 concentration of hardwood smoke. Female rats exhibited 

a decrease in BALF macrophage inflammatory protein-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 7-33 
hardwood 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 hardwood smoke (PM concentrations 30, 100, 300 and 

1,000 μg/m3) 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/m3

 for CO, 54.9-139.3 μg/m3 for ammonia, 

and 177.6- 3455.0 μg/m3

 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 alveolar macrophages. The effects of hardwood smoke on 

bacterial clearance are discussed below (Section 7.3.7.2). 

Another study evaluated the effects of subchronic 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/m3 woodsmoke for 

12 wk, but no changes were seen in numbers of other inflammatory cells. A large percent of BALF 

macrophages contained carbonaceous material. Histological 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 significant 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 increased 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 ZnSO4. WKY rats were exposed nose-only to 10, 30, or 

100 μg/m3 UFP of ZnSO4 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 diameter. 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 ZnSO4 on extrapulmonary systems, as suggested by the lack of 

pulmonary effects. 

7.3.4. Pulmonary Oxidative Response 

7.3.4.1. Toxicological 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 compared 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 adaptation 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. Toxicological 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 and 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/m3

) 59.52 versus 37.08 for NO2, 12.52 versus 0 for 

BC, and 46.49 versus 18.62 for PM2.5. Thus, pollutant levels were reduced by filtration but not 

entirely eliminated. Compared to filtered air, exposure 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 neutral 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 mucus 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 destruction of extracellular matrix in the alveolar 

region of the lung which results in airspace enlargement, airflow limitation and a reduction of the 

gas-exchange area of the lung. Inflammation, oxidative stress, protease imbalance and apoptosis are 

thought to contribute to the development of emphysema. In this study, male BALB/c mice were 

December 2009 7-35 
continuously exposed to ambient or filtered Sao Paulo air for 2 mo. Concentrations of PM2.5 in the 

“polluted chamber” versus “clean chamber” were 33.86 ± 2.09 versus 2.68 ± 0.38 μg/m3

. Thus 

filtration reduced PM levels considerably. Ambient concentrations of CO and SO2 were 1.7 ppm and 

16.2 μg/m3 respectively. No significant difference was observed in the concentrations of NO2 in the 

“polluted chamber” versus “clean chamber” (60-80 μg/m3

). Half of the mice were pre-treated with 

papain by intranasal instillation in order to induce 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 immunohistochemical analysis demonstrated 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 demonstrate 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 NO2 and 63-65 μg/m3

 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 NO2 and 15 μg/m3

SPM). The most common change observed using transmission electron microscopy was the presence 

of particle laden (anthracotic) alveolar 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 changes included increases in the number of mucus 

granules in goblet cells, mast cell infiltration (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 particles. 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 female 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 subjects from Vancouver (Churg et al., 2003, 087899), a city with 

relatively low PM levels. Electron microscopy showed 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 residence 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 versus 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 7-36 
compared 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. Inflammatory 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 hyperplasia 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/m3 PM. This response was attenuated when the DE was filtered 

to remove PM. Pulmonary inflammation was 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 μg/m3

, with the corresponding number concentration between 2.6×104

and 5.0×105

 particles/cm3. Concentrations of gases ranged from 12.8-107.3 ppm CO, 2.0-17.9 ppm 

NO, 0.1-0.9 ppm NO2, 0.09-0.62 ppm SO2 and 0.38-3.37 ppm NH3. 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, respectively, after 6 mo of exposure to the highest 

concentration of engine exhaust. This response was 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 statistically significant decreases in hydrogen peroxide 

and superoxide production in unstimulated and stimulated BALF macrophages. Hypermethylation of 

lung DNA was observed in male F344 rats following 6 mo of exposure to gasoline exhaust 

containing 30 μg/m3

 PM. This response was PM-dependent since it was absent in mice exposed to 

filtered gasoline exhaust. The significance 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 exhibited 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/m3

 hardwood smoke. Male rats exposed to 100 and 

300 μg/m3

 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 subchronic woodsmoke exposure in Brown Norway rats 

as described in Section 7.3.2.2. (Tesfaigzi et al., 2002, 025575). Exposure to 1,000 μg/m3

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 neutral and acid mucus, respectively. More significant 

histopathological responses were found following exposure to 10,000 μg/m3

 of DE. Pulmonary 

December 2009 7-37 
function 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 particular, 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. PM10 and PM2.5 concentrations in the exposure chambers were reported to be 502 ± 34 and 

363 ± 23 μg/m3

, 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 significant 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 increased at 4-7 mo. At 4 mo, alveolar mononuclear 

phagocytic and lymphocytic peribronchiolar inflammation, as well as bronchiolar epithelial and 

smooth muscle hyperplasia, were observed by histological 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 interstitial 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, connective 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. Epidemiologic Studies 

A number of epidemiologic studies have found associations between PM and allergic (or 

atopic) indicators. Allergy is a major driver of asthma, 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 PM2.5 (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 PM10. Also, Morgenstern et al. 

(2008, 156782) found statistically significant effects for asthmatic bronchitis, hay fever, and allergic 

sensitization to pollen in a cohort of children in Germany examining modeled PM2.5 data at birth 

addresses. Distance to a main road had a dose-response relationship with sensitization to outdoor 

allergens. Nordling et al. (2008, 097998) (discussed above in Section 7.3.2.1) reported a positive 

association of PM10 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 PM2.5 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), PM2.5 was associated with allergic indicators such 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 the school to truck traffic but not car traffic, 

December 2009 7-38 
suggesting a role for diesel-related pollution. Consistent with the aforementioned Dutch study by 

Brauer et al. (2007, 090691), PM2.5 was not associated with eczema. 

Mortimer et al. (2008, 187280) examined the association between prenatal and early-life 

exposures to air pollutants with allergic sensitization 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 PM10 and CO during the entire pregnancy and 2nd trimester and higher PM10 during the 

first 2 yr of life. Sensitization to at least one indoor allergen was associated with higher exposures to 

PM10 and CO in during the entire pregnancy and during the 2nd trimester. However, no significant 

associations remained for PM10 after adjustment for copollutants, effect modifiers, or potential 

cofounders in addition to year of birth. The authors 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 Survey 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 PM10 and hay fever in Switzerland. In a study by Oftedal et 

al. (2007, 191948) conducted in Oslo, Norway, early-life exposure to PM10 or PM2.5 was generally 

not associated with sensitization to allergens in 9- to 10-yr-old children; lifetime exposures to PM10

and PM2.5 were associated with dust mite allergy, but the association was diminished by adjustment 

for socioeconomic factors . In Norway, wood burning in the wintertime is thought to account for 

about half of the PM2.5 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. Toxicological 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 intranasally challenged 1 day prior to inhalation exposure to 

DE (PM concentration 100 μg/m3; CO, 3.5 ppm; NO2, 2.2 ppm; SO2 <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 in Section 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 neutrophils 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/m3

 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 NOX. Respiratory function 

was measured in anesthetized animals by whole-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 compared 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. Epidemiologic Studies 

Epidemiologic studies of respiratory infections indicate an association with PM. This is more 

evident when considering short-term exposures (Chapter 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 approximately 4,000 Dutch children, Brauer et al. (2007, 

090691)(described in Section 7.3.1.1) found significant positive associations for PM2.5 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/m3 PM2.5 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 PM10 were associated with declining 

prevalence of common cold and conjunctivitis. Because 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 proximity 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 PM2.5 exposure for infants born just before the peak 

respiratory syncytial virus (RSV) season. Risk estimates 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. Toxicological Studies 

Diesel Exhaust 

DE may affect systemic immunity. The proliferative 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/m3

 PM (Burchiel et al., 2004, 055557). B cell proliferation was 

increased at 300 μg/m3

 but unaffected at higher concentrations (up to 1,000 μg/m3). Concentrations 

of gases and were reported in the parallel study by Reed et al. (2004, 055625), described in 

December 2009 7-40 
Section 7.3.3.2. The Reed study reported a decrease in spleen weight in male mice (27% reduction in 

the 300 μg/m3

 exposure group). The immunosuppressive effects 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 modulation of lymphocyte mitogenic responses (Baker, 

1998, 156245). A 6-month exposure (6h/day, 7d/wk) to 30, 100, 300 or 1,000 μg/m3

 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/m3 PM for 6 mo, but produced a concentration-dependent suppression of 

proliferation at PM concentrations >300 μg/m3

. No effects on B cell proliferation were observed. 

Concentrations of NO and NO2 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/m3 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/m3 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 clearance 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 PM2.5 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 PM2.5 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 data and found positive associations between 

long-term exposure to PM2.5 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 for lung cancer deaths. There is some evidence for 

an association between PM2.5 and respiratory mortality among post-neonatal infants (ages 1 month-1 

year) (Section 7.4.1). In summary, when deaths due to respiratory causes are separated from allcause (nonaccidental) and cardiopulmonary deaths, there is limited and inconsistent evidence for an 

effect of PM2.5 on respiratory mortality, with one large cohort study finding a reduction in deaths due 

to respiratory causes associated with reduced PM2.5 concentrations, and another large cohort study 

finding no PM2.5 associations with respiratory mortality. 

December 2009 7-41 
7.3.9. Summary and Causal Determinations 

7.3.9.1. PM2.5 

The epidemiologic studies reviewed in the 2004 PM AQCD suggested relationships between 

long-term PM10 and PM2.5 (or PM2.1) 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 PM2.5 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 PM2.1 (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 longer follow-up periods of the CHS cohort through 18 yr 

of age and provide evidence that effects from exposure to PM2.5 persist into early adulthood. 

Gauderman et al. (2004, 056569) reported that PM2.5 exposure was associated with clinically and 

statistically significant deficits in FEV1 attained at the age of 18 yr. In addition, the strength and 

robustness of the outcomes were larger in magnitude, and more precise than previous CHS studies 

with shorter follow-up periods. Supporting this result are new longitudinal cohort studies conducted 

by other researchers in other locations with different methods. These studies report results for PM10

that is dominated by PM2.5. New studies provide positive associations 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 consequences 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 PM2.5 and PM10, while most did not evaluate 

PM10-2.5. Associations have been reported with fine particle components, particularly EC and OC. 

Source apportionment methods generally have not been used in these long-term exposure studies. 

Coherence and biological plausibility for the observed associations with lung function 

decrements is provided by toxicological studies (Section7.3.2.2). 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 alveolar region of neonatal rats (Pinkerton et al., 2004, 

087465) (Section 6.3.5.3). Impaired lung development is a viable mechanism by which PM may 

reduce lung function growth in children. Other animal 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 PM2.5 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 PM2.5 and bronchitic symptoms and 

reports larger associations for within-community effects 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 PM10 in areas where fine particles predominate. In England, an 

association was 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 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 PM2.5 exposure and asthma 

incidence. PM2.5 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 PM2.5 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 PM10 contributes to asthma development and that reductions in PM 

decrease asthma risk (Kunzli et al., 2009, 191949). 

December 2009 7-42 
A large proportion of asthma is driven by allergy, and the majority of recent epidemiologic 

studies examining allergic (or atopic) indicators found positive associations with PM2.5 or PM10 

(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 PM2.5 exposure is associated with pulmonary inflammation and oxidative 

responses. An epidemiologic study found a relationship between PM2.5 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 PM2.5 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. Further, 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. These 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 PM2.5 exposure. Several recent epidemiologic studies 

suggest a relationship between long-term exposure to PM2.5 or PM10 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 PM2.5

and respiratory morbidity is provided by epidemiologic studies demonstrating associations with 

decrements in lung function growth in children and with respiratory symptoms and disease incidence 

in adults. Mean PM2.5 concentrations in these study locations ranged from 13.8 to 30 μg/m3

 during 

the study periods. These studies provide evidence for associations in areas where PM is 

predominantly fine particles. A major challenge to interpreting the results of these studies is that the 

PM size fractions and concentrations of other air pollutants are often correlated; however, the 

consistency of findings across different locations supports an independent effect of PM2.5. Recent 

toxicological studies provide support for the associations with PM2.5 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 toxicological studies also demonstrate altered pulmonary 

function, mild inflammation, oxidative responses, histopathological changes including mucus cell 

hyperplasia and enhanced allergic responses in response to CAPs, DE, urban air and woodsmoke and 

provide further coherence and biological plausibility. 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 PM2.5 exposure 

and respiratory effects is likely to be causal.

7.3.9.2. PM10-2.5

The 2004 PM AQCD did not report long-term exposure studies for PM10-2.5. The only recent 

study to evaluate long-term exposure to PM10-2.5 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 PM10-2.5 exposures and respiratory effects. 

December 2009 7-43 
7.3.9.3. UFPs 

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 studies. 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 given the need for detailed data and potential 

residential movement of mothers during pregnancy. 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 difficulty 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,” particularly 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 

PM10 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 

PM2.5 during the first month of pregnancy was associated with intrauterine growth restriction. 

However, other work did not identify relationships between PM10 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 results based on studies from multiple 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 PM10 and PM2.5 exposure and low 

birth weight and infant mortality, especially during the post-neonatal period. Animal toxicological 

evidence supports these associations with PM2.5, but provides little mechanistic information or 

biological plausibility. Information on the ambient concentrations of PM10 and PM2.5 in these study 

locations can be found in Table 7-5. 

December 2009 7-44 
7.4.1.1. Low Birth 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 PM10-2.5 and birthweight (-13 g [95% CI: -18.3 to -7.6]) per 10 μg/m3

 increase), 

but no such association for PM2.5. 

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 PM10 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 μg/m3

) and 95th percentile (>43μg/m3). There was no increased 

risk for low birth weight at term associated with PM10 exposure during any trimester of pregnancy. 

When birth weight was considered as a continuous outcome, exposure to PM10 was not associated 

with a reduction in mean birth weight. 

In contrast, Bell et al. (2007, 093256) reported positive associations for both PM2.5 and PM10

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 μg/m3

 associated with PM2.5

was -66.8 (95% CI: -77.7 to -55.9) g. For PM10 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 PM2.5 and OR = 1.027 (95% CI: 

0.991-1.064) for PM10, based on average exposure during pregnancy. Reductions in birth weight 

were also associated with third trimester exposure to PM10 and second and third trimester exposure 

to PM2.5. 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 PM2.5 than 

by exposure to PM10. Measured PM2.5 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 differing 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 PM10 monitors within 50 km of the ZIP code centroid. If 

there was a PM10 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 μg/m3 increase in PM10 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 O3. 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 pregnancy. 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 neonatal outcomes and long-term exposure. 

Study Location Mean Annual 

Concentration (μg/m3) 

Upper Percentile 

Concentrations (μg/m3) 

PM2.5 

 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 19.2a 75th: 22.7 

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

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

PM10-2.5

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

PM10

 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 23.8a 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 53.8a 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. 31.0a 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 7-46 
Study Location Mean Annual 

Concentration (μg/m3) 

Upper Percentile 

Concentrations (μg/m3) 

 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: 28.6-29.8a 

Avg of means across effects: 29.1a 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 gestation were included. Thus all infants were the 

same gestational age, and had been exposed in the same year. Exposure to PM2.5 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 μg/m3 experienced reductions in birth weight (third quartile -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 reductions 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 only included births at 40 wk gestation. Reducing 

misclassification should lead to a stronger association, 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 PM10

in the third trimester was analyzed for increased risk of low birth weight at term (≥ 37-wk gestation). 

Analysis at the ZIP code level did not detect increased risk (per 10 μg/m3

 PM10, OR = 1.03 [95% CI: 

0.97-1.09]). However the analysis based on geocoded addresses indicated that increasing exposure to 

PM10 was associated with increased risk of low birth weight for women living within 1 mile of the 

station where PM10 was measured. For these women (n = 247 cases, 10,981 non-cases), each 

10 μg/m3

 increase in PM10 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 PM10 >44.4 μg/m3

 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 effects of CO were attenuated but the effects of PM10

increased. Controlling for CO, NO2, and O3, each 10 μg/m3

 increase in exposure to PM10 increased 

risk of low birth weight 36% (OR = 1.36 [95% CI: 1.12-1.65]). 

Spatial variation in PM2.5 exposure was investigated by Basu et al. (2004, 087896). They 

included only mothers who lived within 5 miles of a PM2.5 monitor and within a California county 

with at least 1 monitor. To minimize potential confounding, they included only white (n = 8,597) or 

Hispanic (n = 8,114) women, who were married, between 20 and 30 yr of age, completed at least 

high school and were having their first child. Consistently, PM2.5 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 μg/m3

 increase in PM2.5. 

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 PM10

during the study period, it appears (not stated) that exposure was averaged over the county. A 

December 2009 7-47 
10 μg/m3

 increase in exposure to PM10 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 μg/m3

 compared to 

<19.74 μg/m3 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, the odds ratio was 1.10 (95% CI: 0.71-1.71). 

Misclassification of exposure may have occurred when exposure was averaged over a large 

geographic area (16,968 km2

). 

Recent international studies investigating effects 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 PM2.5 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 that 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 the risk associated with particle exposure. In one 

study, exposure to PM10 in a single pollutant model reduced birth weight by 11 g, but became nonsignificant in copollutant models with O3 (Salam et al., 2005, 087885). In another study the risk 

associated with PM10 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 might be related to the chemical composition of 

particles in the U.S., or to differences in the pollutant mixture. Studies with null results must be 

interpreted with caution when the comparison groups 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 locations, 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. Preterm 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-hearth 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 Valley, before, during, and after 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 trimester (14% decrease in risk of preterm birth 

during mill closure). Preterm birth outside of the Utah 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 PM10 every 6th day during the study period. 

December 2009 7-48 
Birth 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 μg/m3

 increase in 

PM10 averaged in the 6 wk before birth. Exposure to PM10 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 μg/m3

 to a mean level near 

40 μg/m3. Birth certificate data from 1994-2000 was analyzed for women living in ZIP codes within 

2 miles of a monitoring station, or with addresses within 5 miles of a monitoring station. No 

significant effects of exposure to PM10 were reported. Exposure to PM2.5 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 

(PM2.5 >24.3 μg/m3). Using a continuous measure of PM2.5, there was a 10% increase in risk for each 

10 μg/m3

 increase in PM2.5 (RR = 1.10 [95% CI: 1.00-1.21]). 

There have been two major criticisms of air pollution studies using birth certificate data. First, 

that birth certificates only indicate the address at birth and the exposure of women who moved 

during pregnancy may be misclassified; second, that 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 women 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 

PM2.5 differed between women responding to the survey and women who did not respond. Among 

responders, exposure to each 10 μg/m3 increase in PM2.5 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 in the entire birth cohort (RR = 1.00 [95% CI: 

0.94-1.07]). 

An additional case control study of preterm birth and PM2.5 exposure (Huynh et al., 2006, 

091240) used California birth certificate data. Singleton preterm infants (24-36-wk gestation) born in 

California (1999-2000) whose mothers lived within 5 miles of a PM2.5 monitor were eligible. Each of 

these 10,673 preterm infants were matched to three term (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/ethnicity, education, marital status, parity and CO exposure, exposure 

to PM2.5 >17.7 μg/m3

 increased the risk of preterm birth by 14% (OR = 1.14 [95% CI: 1.07-1.23]). 

Averaging PM2.5 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 counties between January 1997 and December 2001. In 

this analysis, exposure to PM10 is compared to the rate of preterm births each day. Both acute 

exposure (on the day of birth) and longer term exposure (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 [95% 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 PM10 emissions from 32 geographically located industrial point 

sources, meteorological factors, and geographic location of the birth home. Exposure was 

categorized by quartiles. Comparing women who delivered a preterm AGA infant to those who 

December 2009 7-49 
delivered a term AGA infant, exposure to PM10>15.07 μg/m3

 tripled the risk (OR = 3.68 [95% CI: 

1.44-9.44]). 

Brauer et al. (2008, 156292) evaluated the impacts of PM2.5 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 distance 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 strongly associated with preterm births. 

Suh et al. (2008, 192077) conducted a study to determine if the effects of exposure to PM10

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 PM10 during pregnancy on preterm 

birth in a case-control study in Seoul, Korea. PM10 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 risk for preterm birth was substantially higher for 

women who carried the GSTM1 null genotype and were exposed to high levels of PM10 (≥ 75th 

percentile) than for those who carried the GSTM1 positive genotype but were only exposed to low 

levels of PM10 (<75th percentile) during the third trimester of pregnancy (OR = 6.22, 95% CI: 

2.14-18.08). 

In Incheon, Korea, Leem et al. (2006, 089828) estimated PM10 exposure spatially as well as 

temporally. Exposure was based on 26 monitors and kriging was used to determine exposure for 120 

dongs (administrative districts, mean area 7.82 km2

, median area 1.42 km3). The sample included 

52,113 births, from 2001-2002. PM10 was very weakly correlated with other pollutants. Exposure 

was compared in quartiles for the first 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 PM10 in quartile one 

(reference group) was 26.9-45.9 μg/m3

; fourth quartile exposure equaled 64.6-106.4 μg/m3. The 

p-value for trend was 0.02. Exposure in the third trimester was not related to preterm birth, however 

no information was provided to determine how exposure 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 PM10 concentrations. Exposure to an interquartile range increase in PM10 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 PM10 exposure in the first month of pregnancy 

(OR = 1.19 [95% CI: 1.13-1.26]). PM10 was correlated with O3 (r = 0.77) in this study and O3 also 

increased risk in the first trimester. No effects were associated with exposure to PM10 in the third 

trimester. 

In Sydney, associations between exposure to particles and preterm birth varied by season. 

Jalaludin et al. (2007, 156601) obtained information on all births in metropolitan Sydney 

(1998-2000). Exposure to PM2.5 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 PM10 (OR = 1.3 [95% CI: 1.2-1.5]) and PM2.5 (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 (PM10 OR = 0.91 [95% CI: 0.88-0.93]) (PM2.5 OR = 0.87 [95% CI: 

0.84-0.92]). Due to both positive and negative findings, 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 leading 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 7-50 
might affect placentation, with results observable later in pregnancy, or cumulative exposure during 

pregnancy may be the most important determinant. The 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 quartiles, using the lowest quartile as the reference 

(or control) group. No studies use a truly unexposed control group. If exposure in the lowest quartile 

confers risk, than it may be difficult to demonstrate additional risk associated with a higher quartile. 

Thus negative studies must be interpreted with caution. 

Preterm birth occurs both naturally (idiopathic preterm), and as a result of medical 

intervention (iatrogenic preterm). Ritz et al. (2000, 012068; 2007, 096146) excluded all births by 

Cesarean section, to limit their studies to idiopathic preterm. No other studies attempted to 

distinguish the type of preterm birth, although PM exposure maybe associated with only one type. 

This is a source of potential effect misclassification. 

7.4.1.3. Growth Restriction 

Low birth weight has often been used as an outcome 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 infants 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 other reproductive outcomes (low birth weight or 

preterm delivery). 

Three studies in the U.S. examined intrauterine growth. A recent study (Rich et al., 2009, 

180122) investigated very small for gestational age (defined 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 during the first, second and third trimesters. They 

reported an increased risk of SGA associated with first and third trimester PM2.5 concentrations 

(1.116 [95% CI: 1.012, 1.232], and 1.106 [1.008-1.212], per 10 μg/m3

 PM2.5, respectively). Parker et 

al. (2005, 087462) reported a positive association between exposure to PM2.5. 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 California. Infants exposed to the highest quartile 

PM2.5 (>18.4 μg/m3

) compared to the lowest quartile PM2.5 (<11.9 μg/m3) 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 respectively (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 PM10 and 

intrauterine growth retardation (IUGR) in the California 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 PM10 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 PM10 versus PM2.5, 

a less stringent definition of IUGR, and exposures determined by monitors located much farther 

away from the subjects’ residences (up to 50 km versus 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 7-51 
birth weight for gestational age. In this study there was a statistically significant effect of exposure to 

both PM10 (OR = 1.10 [95% CI: 1.00-1.48], per 10 μg/m3

 increase) and PM2.5 (OR = 1.34 [95% CI: 

1.10-1.63], per 10 μg/m3

 increase) for exposure during the second trimester. When analysis was 

restricted to births within 5 km of the monitoring station, the association for PM10 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 the tenth percentile of weight for gestational age 

and gender based on an Australian national standard. There was no consistent relationship between 

PM10 exposure and SGA, HC or CHL in any trimester of pregnancy. PM10 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 PM2.5 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) for 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 μg/m3 increase in PM2.5 was associated with an increased risk for 

IUGR (OR = 1.07 [95% CI: 1.03-1.10]) in the first trimester, and similar risks were associated with 

exposure in the second or third trimesters. The effect of PM2.5 was reduced in multipollutant models 

including CO and NO2. 

Brauer et al. (2008, 156292) observed consistent increased risks of SGA for PM2.5, PM10, NO2, 

NO and CO in Vancouver, Canada (20% increase in risk in PM2.5 and PM10 per 10 μg/m3

 increase). 

The effects were similar for exposure estimates 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 for the full duration of pregnancy. 

7.4.1.4. Birth 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). PM10 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/m3 in PM10. 

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 orofacial cleft defects. Cases were restricted to those 

with registry data for gestational age and residence ZIP code, and those with residences <10 miles 

from an air pollution monitor. Six types of categories were included: aortic defects; atrium and 

atrium septum defects; endocrinal and mitral value defects; pulmonary artery and valve defects; 

conotruncal defects; and ventricular septal defects not part of the conotruncal category. PM10

measurements were available every 6 days. While results indicated increased risk of birth defects for 

higher levels of CO or O3, the authors determined that results for PM10 were inconclusive, finding no 

consistent trend of effect after adjustment for CO and O3. 

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/m3

) and lowest (<19.521 μg/m3) quartiles of PM10 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, NO2, O3, SO2) in the model did not greatly alter results; numerical results for 

copollutant analysis were not provided. Strong evidence was not observed for a relationship between 

PM10 and the other birth defect categories. Review articles have concluded that the scientific 

literature is not sufficient to conclude a relationship between air pollution and birth defects (Sram et 

al., 2005, 087442). 

December 2009 7-52 
A recent study of oral clefts conducted in Taiwan found no association between this birth 

defect and concentrations of PM10 during the first or second gestational month (Hwang and Jaakkola, 

2008, 193794). This population-based case-control study included 653 cases and a random sample of 

6,530 controls born in Taiwan between 2001 and 2003. 

7.4.1.5. Infant Mortality 

Many studies have identified strong associations between exposure to particles and increased 

risk of mortality in adults or the general population, 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 countries. 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 this outcome. Both long-term and short-term exposure 

studies of infant mortality are included in this section. Results on PM and infant mortality includes a 

range of findings, with some studies finding associations 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 association 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 studies 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 identified. A prospective cohort of pregnant women 

in Seoul, Korea from 2001 to 2004 was examined with respect to exposure to PM10 (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 records and gathered individual information through 

interviews on socioeconomic condition, medical history, pregnancy complications, smoking, secondhand smoke exposure, and alcohol use. Mother’s exposure to PM10 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/m3

 increase in PM10 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/m3

 increase in PM10 was associated with a 0.8% increase in stillbirth rates (Pereira et al., 

1998, 007264). When other pollutants (NO2, SO2, CO, O3) were included simultaneously in the 

model, the association did not remain. Stillbirths were 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 PM10 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 adjusted for temporal trend, day of the week, 

weather, and holidays. Findings indicated that a 10 μg/m3

 increase in PM10 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 Southern California did not find an association 

between PM10 and neonatal deaths (Ritz et al., 2006, 089819). Quantitative results were not 

provided. The authors considered adjustment for season, 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 7-53 
Poisson model study in Kagoshima City, Japan (Shinkura et al., 1999, 090050). An ecological study 

evaluated U.S. PM10 data for the year 1990 using long-term pollution levels in 180 U.S. counties 

(Lipfert et al., 2000, 004103). Analysis considered birth weight, sex, month of birth, location by state 

and county, prenatal care, and mother’s race, age, educational level, marital status, and smoking 

status. County-level variables were included for socioeconomic status, altitude, and climate. Results 

indicate a 13.1% increase in neonatal mortality (95% CI: 4.4-22.6) per 10 μg/m3

 PM10 for non-low 

birth weight infants. Statistically significant associations were also observed considering all infants 

or low birth weight infants. However, higher levels of SO2 were associated with lower risk of infant 

mortality. When sulfate and an estimate of non-sulfate particles were included in the regression 

simultaneously, associations were observed with non-sulfate particles and an inverse relationship 

with sulfate particles. Respiratory neonatal mortality 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. 

PM10 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/m3

 increase in PM10, a 13.4% (95% CI: -10.3 to 43.5) 

increase in non-low birth weight respiratory-disease 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 PM10 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 PM10 and neonatal mortality (<1 month), however an association was 

observed for post-neonatal mortality, with a 10 μg/m3

 increase in PM10 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 PM10 increment. Even larger effect estimates 

were observed for those who died at ages 4-12 mo. When CO, NO2, and O3 were simultaneously 

included with PM10 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 estimates lost statistical significance. The other casecontrol study of California considered PM2.5 from 1999 to 2000 for infants born to mothers within 

five miles of a PM2.5 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/m3

 increase in PM2.5 was associated with a 

7% (95% CI: -7 to 24) increase in postneonatal death 

County-level PM10 and PM2.5 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 birth 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 higher risk of postneonatal mortality, with 4% (95% CI: -1 to 10) 

increase in mortality risk per 10 μg/m3

 in PM10 and 4% (95% CI: -2 to 11) increase in mortality risk 

for the same increment of PM2.5. This work builds on a previous study of 86 U.S. urban areas from 

December 2009 7-54 
1989 to 1991, finding a 4% (95% CI: 2-7) increase in postneonatal mortality per 10 μg/m3

 countylevel PM10 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 PM10 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 PM10 on the day of death and two previous days to PM10 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 PM10 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/m3 increase in PM10

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 weather, based on national death registries excluding 

accidental deaths (Ha et al., 2003, 042552). A 10 μg/m3

 increase in PM10 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/m3

 increase in PM10

(Ha et al., 2003, 042552) and the case-crossover study in Mexico, for which the same increment in 

PM10 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/m3

 PM10 (Woodruff et al., 2006, 088758). The U.S. study found this 

increment in PM10 to be linked with a 16% (95% CI: 6.0-28.0) increase in respiratory postneonatal 

mortality, although effect estimates for PM2.5 were not statistically significant (Woodruff et al., 2008, 

098386). Earlier studies on respiratory-related postneonatal 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/m3 increase in PM10 the 2 mo prior to death was associated with a 3% (95% CI: -1 to 8) 

increase in SIDS. Adjusted for other pollutants (CO, NO2, and O3), 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/m3

 increase in PM2.5, 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 definition, the effect estimate changed to 1.03 (95% CI: 

0.79-1.35). The authors also examined whether the relationship between PM2.5 and SIDS differed by 

season, finding no significant difference. PM10 and PM10-2.5 were not associated with risk of SIDS; 

numerical results were not provided for these PM metrics. The third recent study 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 PM10 or PM2.5 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 PM2.5 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/m3

 in PM10 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 with 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 methods in the studies, partly due to transition 

from ICD 9 to ICD 10 codes, but also due to different choices within the research design. Two 

studies examined multiple approaches (ICD 10 R95, ICD 10 R95 and R99) (Woodruff et al., 2006, 

December 2009 7-55 
088758; 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 specific cause of death (Dales et al., 2004, 087342). These 

variations in the definition of health outcomes add 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 association. 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 challenging due to several issues, including 

differences in methodologies, populations and study 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-linear functions of temperature and temporal trend, 

and in the ecological study through county-level variables. All studies included consideration of 

seasonality and weather. Researchers used different 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, P28.5, 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 data (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 each case 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 PM10 and postneonatal mortality increased. Although the 

overall association showed higher risk of death with higher PM10 with statistical uncertainty, for the 

lowest socio-economic group, a 10 μg/m3 increment in cumulative PM10 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 estimated 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 time 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 7-56 
Leon, 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. 

PM10 concentrations were highest in South Korea (69.2 μg/m3

) (Ha et al., 2003, 042552) and 

Taiwan (81.45 μg/m3

) (Tsai et al., 2006, 090709), and lowest in the U.S. (29.1 μg/m3) (Woodruff et 

al., 2008, 098386) and Japan (21.6 μg/m3

) (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 separating 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 PM10 and neonatal deaths in São Paulo were 

reduced to a null effect when SO2 was incorporated (Lin et al., 2004, 095787). Associations between 

PM10 and postneonatal mortality or respiratory postneonatal mortality remained but lost statistical 

significance in a multiple pollutant model with CO, NO2, and O3 (Ritz et al., 2006, 089819). 

Several review articles in recent years have examined 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 provide consistent results, although more evidence was 

present 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 PM10 studies found that a 10 μg/m3

 increase in acute PM10 exposure was 

associated with 3.3% (95% CI: 2.4-4.3) increase in risk of postneonatal mortality, whereas the same 

increment of chronic PM10 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 neonatal effects (numerical results not provided) 

but statistically significant results for postneonatal mortality (Ritz et al., 2006, 089819). Figure 

7-5compares risk for the postneonatal period for respiratory and total mortality. In six of the seven 

studies, higher effect estimates were observed for respiratory-related mortality. Results from the 

neonatal period found higher effects for total mortality compared to respiratory mortality (Bobak and 

Leon, 1999, 007678) and the reverse for a study examining infant 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/m3

 in PM10, 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. Decrements in Sperm Quality 

Limited research conducted in the Czech Republic 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, PM2.5 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 ambient air pollution measured 0-9, 10-14, 70-90 days 

before semen collection over a 2-yr period (1996-1998). Ambient O3 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 presented quantitatively (Sokol et al., 

2006, 098539). 

7.4.2. Toxicological Studies 

This section summarizes recent evidence on reproductive 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 investigation 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 PM10 estimates assuming PM10/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. Results for Woodruff et al. (2006, 088758) were based on PM2.5 and were converted to PM10 assuming 

PM2.5/PM10 = 0.6. 

December 2009 7-58 
action for reproductive changes. Emphasis is placed 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 F1 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. Female Reproductive Effects 

Urban Air 

Windows of exposure are important in determining 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, NO2, PM10, and SO2 were 2.2 ± 1.0 ppm, 107.8 ± 42.3 μg/m3

 , 35.5 ± 12.8 μg/m3, 

and 11.2 ± 5.3 μg/m3

, 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 newborns. There was also a higher incidence of 

implantation failures in dams reared as newborns 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 were exposed as neonates and not if exposed as adults. 

Diesel Exhaust 

Significant work has been done in male rodent models 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/m3

) or to clean air (controls) for 8 h/day 

from GD2-13. The concentration of the gaseous materials including NO, NOX, NO2, CO and SO2

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 tested, 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/m3

; PM MMAD of 0.4 μm) or 

filtered air. DE-exposed estrous females had significantly decreased uterine weight (1.0 mg/m3

). 

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 weight were observed at 6 and 8 wk (males and 

females, 1.0 and 3.0 mg/m3

) and 9 wk (females, 1.0 and 3.0 mg/m3). Anogenital distance was 

decreased in 70-day old DE-exposed male offspring (0.3 mg/m3

). In female offspring at 70 days of 

December 2009 7-59 
age, lower organ weights (adrenals, liver, and thymus) were observed (1.0 mg/m3

) compared to 

controls; thymus weight of the 0.3 mg/m3

 females was also lower at 70 days. Crown to rump length 

in females from dams exposed to DE (1.0 and 3.0 mg/m3

) was less than the control group. In 

conclusion, adult exposure to DE led to maternal-dependent reproductive changes that affected 

outcomes in offspring that manifested as decreased 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. Male Reproductive Effects 

Diesel Exhaust 

Studies were performed to determine PM-dependent strain sensitivity of the male reproductive 

tract using male steroidogenic enzymes as the model pathway. Three strains of pregnant mice (ICR, 

C57Bl/6J or ddY mice) were continuously exposed to DE at 0.1 mg/m3

 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 collected from each dam for mRNA analysis of genes 

related to male gonad development including Mullerian 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 [17β-hydroxysteroid dehydrogenase (HSD), 

cytochrome P450 17-α-hydroxylase (P450c17), and 3-βhydroxysteroid dehydrogenase (3βHSD)]. 

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 3β-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) expression 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 male 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/m3

) 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 during gestation and nursed pups. Exposure to 0.3 

mg/m3

 of DE had no effect on male reproductive organ weight or serum testosterone. The 

intermediate concentration of 1.0 mg/m3

 induced increases in serum testosterone. Exposure to the 

higher concentration (1.0 and 3.0 mg/m3

) of DE led to significant increases in reproductive gland 

weight (testis, prostate, and coagulating gland). The 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/m3

) 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 spermatogenesis have also been demonstrated. Exposure 

of pregnant ICR mice to DE (2-16 dpc continuous inhalation exposure to 1.0 mg/m3

 or filtered clean 

air) led to impaired spermatogenesis 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, NO2, SO2, and 

CO2 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 relative epididymal weight were unchanged at all 

December 2009 7-60 
time points. Histological changes showed sertoli cells 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/m3

) or clean air from 2 dpc to 16 dpc. Pups were allowed to 

nurse in clean air on exposed dams until weaning and 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/m3

 for the seminal vesicle, 

testis, epididymis, coagulating gland, prostate and liver. Male pup serum testosterone was 

significantly increased at 1.0 mg/m3

. Mean testosterone positively correlated 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 versus control. Male pup body weight of DE-exposed 

animals was significantly increased at PND28 (1.0 and 3.0 mg/m3

). These studies show that in utero 

DE-exposure led to increased serum testosterone and 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/m3

) 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 concentration-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, significant decreases remained in DSP at the two highest 

concentrations. The effect of ingestion of deposited 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/m3

; <90% of PM less than 0.5 μm; NO2 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 NO2 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 exposed 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-1β, 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 antioxidant vitamin E resulted in partial attenuation of 

serum testosterone levels and caudal epididymal sperm counts, and returned IL-6, IL-1β, 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-treatment 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 utero 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 morphology, 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 phase 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 effects 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. Multiple 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 μg/m3

 PM2.5 for ambient and PM-filtered air chambers, 

respectively, with 101 μg/m3 NO2, 1.81 μg/m3 CO, and 7.66 ppm SO2 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 four groups 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 between the two chambers. There was an increase in 

the time necessary for mating, a decrease in the fertility index, and an increase in the pregnancy 

index in the ambient air group versus the PM-filtered 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 raised 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 hormone dysfunction and ovulatory problems. These estrous 

alterations can contribute to fecundity issues. There was no significant difference in number of 

preovulatory follicles in the above model, but there was a statistically significant decrease in the 

number of antral follicles (Veras et al., 2009, 190496). Antral follicles are the last stage in follicle 

development prior to ovulation, and a decrease in antral 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 fertility 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 likely indicate a general loss of reproductive fitness. 

Interestingly, both pre- and gestational exposure to ambient air induced a significant loss in postDecember 2009 7-62 
implantation 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. Receptor 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 hormonally 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 attenuating the EDC activity of DE particles. 

7.4.2.5. Developmental Effects 

Sex Ratio 

Urban Air 

A correlation between PM10 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 housed 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 PM2.5, CB, and NO2 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 decreased testicular and epididymal sperm counts, decreased total 

number of germ cells, and decreased elongated spermatids, 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 significant 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/m3

; 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 7-63 
and 3.0 mg/m3

) with a significant decrease in the number of absorbed placentas in DE-exposed 

animals at the middle concentration (1.0 mg/m3

). 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 unchanged versus control. IL-2, IL-5, IL-12α, IL-12β and 

GM-CSF mRNA significantly increased in placentas of DE-exposed animals (0.3 and 3.0 mg/m3

). 

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 ROFA 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 IgE and IgG1 antibodies (Hamada et al., 2007, 

091235). Airway hyperreactivity was also observed in the offspring of dams intranasally instilled 

with 50 μg of DE particles or TiO2, 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 TiO2 particles, and genetic analysis indicated differential 

expression of 80 genes in response to TiO2 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% PM2.5, 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 filtered 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 PM2.5 concentrations in the F and nF chambers were 6.5 and 27.5 μg/m3

, respectively. 

Exposure was from PND20-PND60. After this exposure, the animals were mated and then 

maintained in their respective chambers during pregnancy. Pregnancy was terminated at GD8 (near 

term) with placentas and fetuses collected for analysis. 

Exposure to ambient PM pre-gestationally or gestationally 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-specific 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 (caliber) and surface area of maternal blood space with 

compensatory greater fetal capillary surface and oxygen diffusion conduction rates. Intravascular 

barrier thickness, a quantitative relationship between 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-dependent 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 ambient air (Sao Paulo) or filtered air over 

various portions of gestation to determine if there 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 PM10 (42 ± 17 μg/m3), NO2 (97 ± 39 μg/m3), and SO2 (9 ± 4 μg/m3) were 

measured 100 m away from the rodent exposure chambers. By using six time windows of exposure 

that covered 1-3 wk of gestation (the entire gestation 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/m3

) 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. Earlier work has shown that DE particles (<0.1 μm) 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 apoptosis 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 speculate 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 publications 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 reference materials (SRM 2975) via inhalation over 

GD7-GD19 of pregnancy. They found significantly decreased 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 were 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 prohormone, thyroxine, was unchanged in control and 

DE particle-exposed dams and offspring at weaning. 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 body weight at weaning or changes in inflammatory 

markers or thyroid hormone levels. 

December 2009 7-65 
Diesel Exhaust 

The effect of in utero DE exposure on CNS motor 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/m3

 for particle mass, 

2.67 ppm CO, 0.23 ppm NO2, and <0.01 ppm SO2. Spontaneous motor activity was significantly 

decreased in pups (PND35), as was the dopamine metabolite homovanillic acid measured in the 

striatum and nucleus accumbens, indicating decreased 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, similar 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 PM10 concentration of 1.73 mg/m3

. 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), anthracene (Ant) and benz[a]anthracene (BaA) in the 

DE group being significantly higher than the control 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 that PAHs derived from DE are transferred across the 

placenta from the DE-exposed dam to the fetus. Lactational 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 dams 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 site 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 monitored post-exposure. This area in Hamilton is 

near two steel mills and a major highway. Air quality 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/m3

, respectively) and PAH at week 3 (8.3 ± 1.7 ng/m3). Mutation frequency at ESTR 

December 2009 7-66 
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 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-exposed 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 the 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. PM2.5 

The 1996 PM AQCD concluded that while few studies had been conducted on the link 

between 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 PM10 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 PM2.5 during the first month of 

pregnancy was associated with IUGR. However, other work did not identify relationships between 

PM10 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 results 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 PM2.5, and all found an 

increased risk of low birth weight (Section 7.4.1). Exposure to PM2.5 was usually associated with 

greater reductions in birth weight than exposure to PM10. All of the studies that examined the 

relationship between PM2.5 and preterm birth report positive associations, and most were statistically 

significant. The studies evaluating the association between PM2.5 and growth restriction all found 

positive associations, with the strongest evidence coming when exposure was assessed during the 

first or second trimester (Section 7.4.1). For infant mortality (<1 yr), several studies examined PM2.5

and found positive associations (Section 7.4.1). 

Animal toxicological studies reported effects including decreased uterine weight, limited 

evidence of male reproductive effects, and conflicting 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 respiratory causes during the post-neonatal period. The 

mean PM2.5 concentrations during the study periods ranged from 5.3-27.4 μg/m3

. Exposure to PM2.5

was usually associated with greater reductions in birth weight than exposure to PM10. Several U.S. 

studies of PM10 investigating fetal growth reported 11-g decrements in birth weight associated with 

PM10 exposure. Most of these studies were conducted in California, where PM2.5 and PM10-2.5

contribute almost equally to the PM10 mass concentration. So while these results can not be 

attributed to one size fraction or the other, the consistency 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 PM2.5 and PM10 exposure and adverse reproductive and 

developmental outcomes, but provided little mechanistic information or biological plausibility for an 

association between long-term PM exposure and adverse 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 epidemiologic and toxicological evidence is suggestive of a 

December 2009 7-67 
causal relationship between long-term exposures to PM2.5 and reproductive and 

developmental outcomes.

7.4.3.2. PM10-2.5

Evidence is inadequate to determine if a causal relationship exists between longterm exposure to PM10-2.5 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 PM10-2.5 and birthweight (-13 g [95% CI: -18.3 to -7.6] per 10 μg/m3

increase), but no such association for PM2.5 (Parker et al., 2008, 156013). 

7.4.3.3. UFPs 

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. Ambient 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 toxicological 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 ambient 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 carcinogenic 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 epidemiologic 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 the 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 μm in diameter (PM10), with approximately 94% of the mass of these particles 

<2.5 μm in diameter (PM2.5), 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 inhalation studies conducted with lower concentrations of 

December 2009 7-68 
PM (<2 mg/m3

), relevant to current ambient concentrations and the current regulatory standard 

(Section 1.3). Because this ISA focuses 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 identified 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 airborne particles. They concluded that the smallest particles have 

the highest toxicity per particulate mass, with the PM2.5 size fraction having greater mutagenic and 

cytotoxic potential than the PM10 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 PM2.5 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 samples; 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 carcinogenic 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 combustion 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, mutagenic 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 increase 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 PM2.5 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 

PM10: 46.5 μg/m3

; mean PM2.5 29.6 μg/m3) as compared to the least polluted city (mean PM10: 

18.2 μg/m3

; mean PM2.5 11.0 μg/m3) 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 PM10 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 only among women (Abbey et al., 1991, 042668). Further reanalysis 

December 2009 7-69 
of this cohort revealed an association between PM10 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 mortality with long-term exposure to PM2.5 in the 

AHSMOG cohort; no association was seen for PM10-2.5. 

7.5.1.1. Lung 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 cause-specific mortality is presented in 

Section 7.6. 

Table 7-6. Characterization of ambient PM concentrations from recent studies of cancer and longterm exposures to PM. 

Study Location Pollutant Mean Annual 

Concentration (μg/m3) Upper Percentile Concentrations (μg/m3)

Brunekreef et al. (2009, 191947) The Netherlands PM2.5 28.3 Max: 36.8 

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

Jerret et al. (2005, 087600) Los Angeles, California PM2.5 Max:27.1 

Laden et al. (2006, 087605) 6 U.S. cities PM2.5

Range of means across sites: 10.2-29.0 

Avg of means across sites: 16.4 

PM2.5 15 Max: 22 

Naess et al. (2007, 090736) Oslo, Norway 

PM10 19 Max: 30 

Palli et al. (2008, 156837) Florence, Italy PM10 NR 

PM2.5 Max: 46-120 

Pedersen et al. (2006, 156848) Czech Republic 

PM10 Max: 120-238.6 

Sorensen et al. (2005, 083053) Copenhagen, Denmark PM2.5

Range of means across sites: 12.6-20.7 

Avg of means across sites: 16.7 

75th: 24.3-27.7 

PM10 Max: 55 

Sram et al. (2007, 188457) Czech Republic 

PM2.5 Max: 38 

PM10

Range of means across sites: 36.4-55.6 

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

PM2.5

Range of means across sites: 24.8-44.4 

Avg of means across sites: 34.6 

Vineis et al. (2006, 192089) Multi-city, Europe PM10

Range of means across sites: 19.9-73.4 

Avg of means across sites: 35.4 

Vinzents et al. (2005, 087482) Copenhagen, Denmark PM10

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 PM2.5 and lung cancer mortality 

while adjusting for both individual and neighborhood covariates (Jerrett et al., 2005, 087600). There 

was a positive association between PM2.5 and lung cancer mortality when adjusting for 44 individual 

covariates (RR 1.44 [95% CI: 0.98-2.11] per 10 μg/m3

 increase in PM2.5). However, including all 

potential individual and neighborhood covariates associated with mortality reduced the association 

December 2009 7-70 
(RR 1.20 [95% CI: 0.79-1.82] per 10 μg/m3

 increase in PM2.5). A recent re-analysis of the full ACS 

cohort also demonstrated a positive association between PM2.5 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 nationwide 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 PM2.5 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/m3 increase in average annual PM2.5). However, estimated 

decreases in PM2.5 were not associated with reduced lung cancer mortality (RR 1.06 (95% CI: 

0.43-2.62] for every 10 μg/m3 reduction in PM2.5). 

Naess et al. (2007, 090736) studied individuals aged 51-90 yr living in Oslo, Norway in 1992. 

Death certificate data were obtained for 1992-1998 and information on PM was collected from 

1992-1995. Women had a larger association of lung cancer mortality with PM2.5 compared to men. 

Similar results were reported for PM10. 

Most recently, Brunekreef 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 PM2.5 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/m3 increase in PM2.5) 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 PM2.5 and lung cancer incidence (full 

cohort: RR 0.81 [95% CI: 0.63-1.04]; case-cohort: RR 0.67 [95% CI: 0.41-1.10] per 10 μg/m3

increase in PM2.5). 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, the 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 PM10 and lung cancer (OR 0.91 [95% CI: 0.70-1.18] per 10 μg/m3

). 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/m3

 to 

<11 μg/m3

). 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 concentrations from select studies of lung cancer 

mortality and incidence. 

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

MORTALITY - PM2.5

 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 - PM10

 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 - PM2.5

 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 – PM10

 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/m3 increase 

†Results from the paper were standardized to10 μg/m3

 [For McDonnell et al. (2000, 010319) the non-standardized results were reported based on IQR increments (24.3 μg/m3 for PM2.5 and 29.5 μg/m3 for 

PM10). For Naess et al. (2007, 090736) the original hazard ratios were calculated based on quartiles of PM exposure. The results were converted to10 μg/m3

 using the mean range of the four quartiles (3.95 

μg/m3

 for PM2.5 and 5.88 μg/m3 for PM10)]. 

‡Study was included in the 2004 PM AQCD 

December 2009 7-72 
7.5.1.2. Other 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 the state of New York. Cases (n = 1,166 of which 841 

were post-menopausal) were women with primary breast cancer, and controls (n = 2,105 of which 

1,495 were post-menopausal) were frequency matched 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 women, exposure to high concentrations of TSP 

(>140 μg/m3

) 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/m3

). ORs were elevated for pollution exposures at age 

of menarche (OR: 1.45 [95% CI: 0.74-2.87]) and age at first birth (OR: 1.33 [95% CI: 0.87-2.06]) 

among postmenopausal women. Among premenopausal 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. Markers 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, which can then lead to altered gene expression, 

enhanced proliferation, and resistance to apoptotic signals. Reactive oxygen and nitrogen species, 

provided by PM components or inflammation pathways, 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-1β, 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 circulation. 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: PM10

32-55 μg/m3

, PM2.5 27-38 μg/m3, c-PAHs 18-22 ng/m3, and B[a]P 2.5-3.1 ng/m3 using a VAPS 

monitor (measurements taken with a HiVol monitor, which has a lower flow rate, had a mean for 

PM10 of 62.6 μg/m3

). c-PAHs detected on personal monitors during sampling days had a mean of 

12.04 ng/m3 among the policemen and 6.17 ng/m3 among the controls. No difference in percent of 

chromosomal aberrations was observed between the policemen and control group using conventional 

cytogenetic analysis. However, using fluorescent in situ hybridization (FISH), a difference in 

chromosomal aberrations between the policemen and control group was reported. For example, the 

percentage of aberrant cells, as well as the genomic 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 PM10 concentrations 

and individual levels of DNA bulky adducts. Study participants were 214 healthy adults aged 

December 2009 7-73 
35-64 yr at enrollment who resided in the city of Florence, Italy. This study was conducted between 

1993 and 1998. PM10 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 PM10 concentrations prior to blood 

sample collection. Time windows of PM10 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 

PM10 concentrations decreased during the study period, with some fluctuations. Quantitative values 

were not reported, but PM10 appeared to range between approximately 30 and 100 μg/m3 for hightraffic stations, and between approximately 20 and 50 μg/m3 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 PM10 levels from high-traffic stations during 

approximately 2 wk preceding blood sample collection (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 PM10

levels among Florence residents with no occupational exposure to vehicle emissions or among 

smokers. DNA bulky adducts were not associated with PM10 levels assessed by low-traffic urban 

monitoring stations. 

The association between personal exposure to water-soluble transition metals in PM2.5 and 

oxidative stress-induced DNA damage was investigated among 49 students from Central 

Copenhagen, Denmark (Sorensen et al., 2005, 083053). Researchers assessed PM2.5 exposure by 

personal sampling over two weekday periods twice 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-oxodG). 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 PM2.5 mass 

concentration. The other transition metals were not significantly associated with the 8-oxodG 

concentration in lymphocytes, and none of the six measured transition metals was associated with 

the 8-oxodG concentration in urine. 

Vinzents et al. (2005, 087482) investigated the association between UFP and PM10 

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 PM10 and UFP were also measured on all exposure days 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 30.4×103 UFPs/mL (standard 

deviation [SD]: 1.38), mean mass concentration of PM10 at a background monitoring station was 

16.9 μg/m3

 (SD: 1.53), and mean mass concentration of PM10 at a street station was 23.5 μg/m3 (SD: 

1.48). Mean personal exposure to UFPs was 32.4×103 UFPs/mL (SD: 1.49) while bicycling (5 

occasions), 19.6×103 UFPs/mL (SD: 1.78) during other outdoor activities (6 occasions), and 

13.4×103 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 strand breaks. Neither ambient air concentrations of 

PM10 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 locations 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 cytogenetic effects. Significantly higher frequencies of 

MN were found in younger children living in Teplice (PM2.5 concentration = 120 μg/m3

) than in 

Prachatice (PM2.5 concentration = 46 μg/m3

). The levels of c-PAHs were also much higher in Teplice 

(nearly 30 ng/m3 in Teplice and about 15 ng/m3

 in Prachatice). The difference in MN frequencies 

December 2009 7-74 
observed 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 breaks 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 assessed 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 formamidopyrimidine DNA glycosylase sensitive sites (rural 

subjects < suburban subjects < residents living near high traffic roads < taxi drivers). Genotyping 

studies showed that the magnitude of the effects of benzene and UFPs may be modified by 

polymorphisms in GSTP1 and NQO1 genes. 

Tovalin et al. (2006, 091322) evaluated the association between exposure to air pollutants and 

the level of DNA damage using the single cell gel electrophoresis (comet) assay. Mononuclear 

lymphocytes from outdoor and indoor workers from two 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 indoor 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 O3 and PM2.5. In summary, several recent studies have reported an association between lung cancer mortality 

and long-term PM2.5 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 PM2.5 (Brunekreef et al., 2009, 191947) or 

PM10 (Vineis et al., 2006, 192089). Studies of exposure/susceptibility markers have reported 

inconsistent outcomes, with some markers being 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 specific combustion sources (e.g., fossil fuels) are 

mutagenic in vivo and in vitro. Research cited 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 mutagenic 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 traffic 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 mutagenicity was observed for exhaust from cold 

starts than starts at room temperature. Both gaseous 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 demonstrating mutagenic effects of emissions from 

wood/biomass burning, which were primarily attributable 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 fractions 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 7-75 
mutagenic potential and carcinogenic potential in animals 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 and carcinogenic effects of PM, PM-constituents, 

and combustion emission source particles are reviewed. Pertinent studies are described briefly in the 

following paragraphs. A summary table is provided in Annex D, Tables D7 and D8). 

7.5.2.1. Mutagenesis and Genotoxicity 

In Vitro Studies 

In general, studies have focused on PM and 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 constituents include induction of MN formation, DNA 

adduct formation, SCE, DNA strand breaks, frameshifts 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 PM2.5 extracts, suggesting that the adsorbed components play a role in the 

genotoxic potential of PM. Total PAH and carcinogenic 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 versus organic) by Gutierrez-Castillo et al. (2006, 

089030) indicated that water-soluble extracts were more 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 mutagenic 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) enzyme 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, PM10, and PM2.5 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 reactivity (induction of DNA adducts) was generally 

higher after metabolic activation. Binková et al. (2003, 156274) found that the addition of S9 

increased PM10-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., PM2.5>PM10) and 

December 2009 7-76 
from 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 PM2.5 size fraction was significantly higher than that of the 

larger size fractions (TSP, PM10) 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/m3

) 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-DE 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), based on PM mass. A-DE particles had 227 times 

more PAH-type mutagenic activity and 8-45 times 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 induce 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-MutaTM 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 with 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 reported for CB. The authors concluded that the 

mutagenic activity of DE particles is likely attributable 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, respectively (Oh and Chung, 2006, 088296). The organic base and 

neutral fractions of DE particles also significantly 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 (AL) cells, DE 

particles (SRM 2975) dose-dependently increased the mutation yield at the CD59 locus; this was 

December 2009 7-77 
significantly 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). Extracts 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, and 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 NO2 in the filtered air chamber and 33-280 ppb NO and 

42-81 ppb NO2 in the experimental group chamber. Suspended PM concentrations were 11-19 μg/m3

in the filtered air chamber and 42-100 μg/m3

 (average 63 μg/m3) 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 increases 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 deposition of PM and its associated PAHs in the lung 

can have indirect effects on extrapulmonary organs. 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 employed “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. Morphologically the particles were similar to DE or 

ambient air PM and the organic extracts from the 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/105

 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 particle extract was twice that of DE particles, 

suggesting that the mutagenicity of DE particles is attributed primarily to compounds in the extract, 

December 2009 7-78 
since ≈50% of the weight of DE particles was provided by the extract. There was no difference in 

mutation frequency between the 1 and 3 mg/m3

 DE particle groups after 12 wk of exposure. 

Wood Smoke 

One recent study measured the effect of freshly 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 treated 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. Carcinogenesis 

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 urethane-induced lung tumors in mice (Cury et al., 

2000, 192100; Reymao et al., 1997, 084653). 

Two recent studies evaluated the carcinogenic potential 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 μg/m3

 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 μg/m3 DE for 6 h/day and 7 days/wk for 

6 mo (Reed et al., 2004, 055625). The concentration of gases in this including NOX, NO2, CO, SO2, 

NH3, methane, non-methane VOC, and FID total hydrocarbon 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/m3

, 1406.5 ± 253.2 to 2642.1 ± 

455.9 μg/m3

, 134.0 ± 52.1 to 1578.6 ± 256.2 μg/m3, 0.1 ± 0.1 to 2.2 ± 0.2 ppm, respectively. Particle 

sizes in the four exposure groups ranged from 0.10-0.15 μm 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 and 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 μg/m3 hardwood smoke or to 30, 100, 300 

and 1,000 μg/m3 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, NH3, 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/m3

 and 177.6 ± 10.4 to 3455.0 ± 557.2 μg/m3, respectively. 

Concentrations of NOX, NO2 and SO2 were reported to be null. Particle sizes in the four exposure 

groups ranged from 0.25-0.36 μm MMAD with geometric standard deviations of 2.0-3.3. Following 

the 6-mo exposure and a 6-mo recovery period, mice were collected and MN formation in blood and 

tumor multiplicity and tumor incidence were measured in lungs. No increases in formation of MN or 

December 2009 7-79 
numbers 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. Summary 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 combustion-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 particles 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. Finally, 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 relationships observed in epidemiology studies. 

7.5.3. Epigenetic Studies and Other Heritable DNA mutations 

Two epidemiologic epigenetic studies examined the effect of PM on DNA methylation. Both 

studies examined methylation of Alu and long interspersed nuclear element-1 (LINE-1) sequences, 

which are located in repetitive elements. In previous 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 PM2.5 exposure for the duration of the study 

(1999-2007). During the study period, the median level of PM2.5, averaged over 7-day periods, was 

9.8 μg/m3

 (interquartile range 8.0-12.0 μg/m3). There was no association between PM2.5 and Alu 

methylation. LINE-1 methylation was associated with PM2.5 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. PM10 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/m3

 PM10 (range 73.4-1220.2 μg/m3). 

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 PM10 exposure using mixed effects models. There was a 

negative association between increasing levels of PM10 exposure and Alu and LINE-1 methylation, 

indicating that PM10 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 PM10 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 have 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 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) (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 7-80 
cell) DNA methylation was monitored post-exposure. This area in Hamilton is near two steel mills 

and a major highway. Air composition showed mean concentrations for TSP of 93.8 ± 17 μg/m3

, 

PAH of 8.3 ± 1.7 ng/m3

, and metal of 3.6 ± 0.7 μg/m3. 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 mutation frequency at 16 wk compared to HEPA-filter 

control animals, pointing to a PM-dependent mechanism of action. No detectable DNA adducts were 

observed in testes samples at any of the time points 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 formation. 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 the 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 PM10 may 

result in epigenetic changes. PM2.5 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. PM2.5

The 2004 PM AQCD reported on original and follow-up analyses for three prospective cohort 

studies that reported positive relationships between PM2.5 and lung cancer mortality. Several recent, 

well-conducted epidemiologic studies have extended the evidence for a positive association between 

PM2.5 and lung cancer mortality (Section 7.5.1.1). Generally, studies have not reported associations 

between long-term exposure to PM2.5 or PM10 and lung cancer incidence (Section 7.5.1.1). Animal 

toxicological studies did not focus on specific size 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 ambient 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 future. However, it has yet to be determined how these 

alterations in the genome could influence the initiation and promotion of cancer. Overall, the 

evidence is suggestive of a causal relationship between relevant PM2.5 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 toxicological 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. PM10-2.5

The 2004 PM AQCD did not report long-term exposure studies for PM10-2.5. No epidemiologic 

studies have been conducted to evaluate the effects of long-term PM10-2.5 exposure and cancer. The 

evidence is inadequate to assess the association between PM10-2.5 and UFP exposures and 

cancer. 

4.3. UFPs 

The 2004 PM AQCD did not report long-term exposure studies for UFPs. No epidemiologic 

studies have been conducted to evaluate the effects of long-term UFP and cancer. The evidence is 

inadequate to determine if a causal relationship 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, suggested associations between increases in mortality 

and long-term exposure to PM2.5, though there was no evidence to support an association with PM10-

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 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 

PM2.5 and mortality. Collectively, the 2004 PM AQCD found that these studies provided strong 

evidence that long-term exposure to PM2.5 was associated with increased risk of human mortality. 

Effect estimates for all-cause mortality ranged from 6 to 13% increased risk per 10 μg/m3

 PM2.5, 

while effect estimates for cardiopulmonary mortality ranged from 6 to 19% per 10 μg/m3 PM2.5. For 

lung cancer mortality, the effect estimate was a 13% increase per 10 μg/m3

 PM2.5, based upon the 

results of the extended analysis from the ACS cohort (Pope et al., 2002, 024689). With regard to 

PM10-2.5, the 2004 PM AQCD reported that no association was observed between mortality and longterm exposure to PM10-2.5 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 PM10-2.5 and mortality. Overall, the 2004 PM AQCD 

concluded that there was strong epidemiologic evidence for associations between long-term 

exposures to PM2.5 and excess all-cause and cardiopulmonary mortality. 

At the time of the 2004 PM AQCD, only a limited number of the chronic-exposure cohort 

studies had considered direct measurements of constituents of PM, other than sulfates. With regard 

to source-oriented evaluations of mortality associations 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 consistent association between long-term exposure to 

PM2.5 and increased risk of mortality. There is little evidence for the long-term effects of PM10-2.5 on 

mortality. Although this section focuses on mortality outcomes in response to long-term exposure to 

PM, it does not evaluate studies that examine the association between PM and infant mortality. 

December 2009 7-82 
These 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 ambient PM concentrations from studies of mortality and long-term 

exposures to PM. 

Study Location Mean Concentration 

(μg/m3) Upper Percentile Concentrations (μg/m3)

PM2.5 

 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 

75th: 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 

PM10-2.5

 Chen et al. (2005, 087942) Multicity, CA 25.4 

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

PM10

 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 Exposure 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 PM10-2.5 and mortality, and the extended and follow-up analyses did 

not evaluate associations with PM10-2.5. The historical and more recent results for PM2.5 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 PM2.5 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 PM2.5 and mortality, while 

providing indications that the magnitude of the PM2.5-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 PM10-2.5 and 

mortality. The original analyses of the AHSMOG cohort study found positive associations between 

long-term concentrations of PM10 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 PM10 concentrations with mortality for males 

were best explained by a relationship of mortality with the fine fraction of PM10 rather than the 

thoracic coarse fraction of PM10. 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 PM10 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 PM2.5 and mortality for data for the two periods 

(RR = 1.16 [95% CI: 1.07-1.26] per 10 μg/m3

 PM2.5). Of special note is a statistically significant 

reduction in mortality risk reported with reduced long-term PM2.5 concentrations (RR = 0.73 [95% 

CI: 0.57-0.95] per 10 μg/m3

 PM2.5). This is equivalent to an RR of 1.27 for reduced mortality risks 

with reduced long-term PM2.5 concentrations. This reduced mortality risk was observed for deaths 

due to cardiovascular and respiratory causes, but not for lung cancer deaths. The PM2.5

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 PM2.5

pollution yields positive health benefits. 

ACS Extended Analyses/Reanalysis II: Two new analyses further evaluated the 

associations of long-term PM2.5 exposures with risk of mortality in 50 U.S. cities reported by Pope 

and colleagues (2002, 024689), adding new details about deaths 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 specific cardiovascular diseases, particularly 

ischemic heart disease (IHD), and a group 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/m3

 PM2.5), 

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 previous ACS analyses. This reanalysis also included 

data for seven ecologic (neighborhood-level) contextual (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 

PM2.5 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 PM2.5 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 reversed. Overall, although the addition of random 

effects modeling and contextual covariates to the ACS model made most effect estimates higher (but 

December 2009 7-84 
some lower), they were not statistically different from the earlier ACS effect estimates. Thus, these 

new analyses, with their more extensive consideration of potentially confounding factors, confirm 

the published ACS PM2.5-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 PM2.5 exposure with 

all-cause deaths from 1973-1982 (RR = 1.04 [95% CI: 1.01-1.07] per 10 μg/m3

 PM2.5) (Enstrom, 

2005, 087356). However, no significant associations were reported with deaths in later time periods 

when PM2.5 levels had decreased in the most polluted counties (1983-2002) (RR = 1.00 [95% CI: 

0.98-1.02] per 10 μg/m3 PM2.5). The PM2.5 data were obtained from the EPA’s Inhalation Particle 

Network (collected 1979-1983), and the locations represented 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 

SCS Original Dockery et al. (1993, 044457) 1974-1991 18.6 CPD

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

ACS Reanalysis Krewski et al. (2000, 012281) 1982-1989 18.2 

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 PM2.5 from the 

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

December 2009 7-85 
Study Cohort Subset Mean Effect Estimate (95% CI) 

McDonnell et al. (2000, 010319) AHSMOG Males 32.0 All Cause 

Brunekreef et al. (2009, 191947) NLCS-AIR Full Cohort 28.3 

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 

Lipfert et al. (2006, 088218) Veterans Cohort 14.3 

14.3 

Eftim et al. (2008, 099104) Medicare Cohort ACS Sites 13.6 

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 

Brunekreef et al. (2009, 191947) NLCS-AIR Full Cohort 28.3 CV 

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 

Chen et al. (2005, 087942) AHSMOG Females 29.0 CHD 

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 

Krewski et al. (2009, 191193) ACS Reanalysis II 14.0 

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 

Brunekreef et al. (2009, 191947) NLCS-AIR Full Cohort 28.3 Respiratory 

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 

Krewski et al. (2009, 191193) ACS Reanalysis II 14.0 

ACS Reanalysis II-NYC 12.8 

Brunekreef et al. (2009, 191947) NLCS-AIR Full Cohort 28.3 Other 

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 

* PM2.5 data from 1973-1982 applied to all subsequent time periods Relative Risk Estimate 

Figure 7-7. Mortality risk estimates, long-term exposure to PM2.5 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 coronary heart disease mortality was reported among 

females (92 deaths; RR = 1.42 [95% CI: 1.06-1.90] per 10 μg/m3

 PM2.5), but not among males 

(53 deaths; RR = 0.90 [95% CI: 0.76-1.05] per 10 μg/m3

 PM2.5) (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/m3

 PM2.5). The authors speculated that females may be more sensitive to air 

pollution-related effects, based on differences between males and females in dosimetry and 

exposure. As was found with PM2.5, a positive association with coronary heart disease mortality was 

reported for PM10-25 and PM10 among females (RR = 1.38 [95% CI: 0.97-1.95] per 10 μg/m3 PM10-2.5;

RR = 1.22 [95% CI: 1.01-1.47] per 10 μg/m3 PM10), but not for males (RR = 0.92 [95% CI: 0.66-

1.29] per 10 μg/m3

 PM10-2.5; RR = 0.94 [95% CI: 0.82-1.08] per 10 μg/m3 PM10); 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 PM2.5 in this study (RR = 1.32 [95% CI: 0.91-1.93] per 10 μg/m3

 PM2.5) that primarily 

focused on evidence of exacerbation of respiratory symptoms (Goss et al., 2004, 055624). No clear 

association was reported for PM10. 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 cardiovascular 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. Hazard ratios were estimated for the first cardiovascular 

event, adjusting for age, race or ethnic group, smoking status, educational level, household income, 

body-mass index, and presence or absence of diabetes, hypertension, or hypercholesterolemia. 

Overall, this study concludes that “long-term exposure to fine particulate air pollution is associated 

with the incidence of cardiovascular disease and death among postmenopausal women.” In terms of 

effect size, the study found that each increase of 10 μg/m3

 of PM2.5 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 disease (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/m3

 PM2.5. 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 rate of 0.625% per year, or approximately 8 times 

the cardiovascular mortality rate of the WHI population (Pope et al., 2004, 055880). Thus, PM2.5

impacts may yield higher relative risk estimates in the WHI population because the PM2.5 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 Six City Study associations with mortality in 

yet another well characterized cohort with detailed individual-level information, 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 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 PM2.5 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 PM2.5 with mortality for the same locations included in the ACS and Six 

City Study. For these locations, they estimated the chronic effects of PM2.5 on mortality for the 

period 2000-2002 using mortality data for cohorts of Medicare participants and average PM2.5 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 mortality risk associated with air pollution adjusting 

for age and sex and area-level covariates (education, income level, poverty, and employment), and 

controlled for potential confounding by cigarette smoking by including standardized mortality ratios 

December 2009 7-87 
for lung cancer and COPD. This study is, therefore, an ecological analysis, similar to past published 

cross-sectional analyses, in that area-level covariates (education, income level, poverty, and 

employment) are employed as controlling variables, since individual level information is not 

available from the Medicare database (other than age and sex), which includes virtually all 

Americans aged 65 or greater. Exposures are also ecological in nature, as central site data are used as 

indices of exposure. These results indicated that a 10 μg/m3

 increase in the yearly average PM2.5

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 increase, 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 estimates 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 and 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 consideration 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 PM2.5 over the preceding 12 mo and monthly mortality 

rates in 113 U.S. counties from 2000 to 2002. They decomposed the association between PM2.5 and 

mortality into two components: (1) the association between “national trends” in PM2.5 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 PM2.5 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 

PM2.5 and mortality at the national scale is more likely to be confounded than is the association 

between trends in PM2.5 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 PM2.5 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 source of exposure variability has been avoided in 

most other air pollution epidemiology studies because of such concerns about potential confounding 

of such time-trend associations. 

By linking monitoring data to the U.S. Medicare 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 estimated log-linear regression models having 

age-specific county level mortality rates as the outcome and, as the main predictor, the average PM2.5

pollution level in each county during 2000. Area-level covariates were used to adjust for socioeconomic status and smoking. The authors reported results under several degrees of adjustment for 

spatial confounding and with stratification into eastern, central and western U.S. counties. A 

10 μg/m3

 increase in PM2.5 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 PM2.5 monitor. Using this 

cohort, they investigated the relationship between 6-yr avg exposure to PM2.5 and mortality risk over 

the period 2000-2005. When divided by region, the associations between long-term exposure to 

PM2.5 and mortality for the eastern and central ZIP codes 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/m3

 increase in PM2.5 in the eastern and central regions, respectively. The MCAPS 

results included evidence of differing PM2.5 relative risks by age and geographic location, where risk 

declines with increasing age category until there 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 between 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/m3

 increase in PM10 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 with long-term exposure to traffic than with PM2.5

mass (Lipfert et al., 2006, 088218). A significant association was reported between total mortality 

and PM2.5 in single-pollutant models (RR = 1.12 [95% CI: 1.04-1.20] per 10 μg/m3 PM2.5). This risk 

estimate is larger than results reported in a previous study of this cohort. In multipollutant models 

including traffic density, the association with PM2.5 was reduced and lost statistical significance. 

Traffic emissions contribute to PM2.5 so it would be expected that the two would be highly 

correlated, and, thus, these multipollutant 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 PM2.5 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 PM10-2.5 and total mortality in a single-pollutant model (RR = 1.07, 95% CI: 

1.01-1.12 per 10 μg/m3 PM10-2.5). 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 PM10-2.5 concentrations, it is difficult to interpret 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 PM10 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 PM10 exposures in the time period 3-48 mo preceding death. The association was strongest 

with average PM10 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 death (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/m3

 increase in 

PM2.5 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 reported by the ACS. In a case-cohort analysis 

adjusted for additional potential confounders, there 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 PM10 (RR = 1.52 [95% CI: 1.09-2.15] per 10 μg/m3 PM10). 

7.6.2. Composition and Source-Oriented 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 associations were found between mortality and long-term 

exposures to nitrates, EC, Ni and V, as well as traffic density and peak O3 concentrations. In 

December 2009 7-89 
multipollutant models, associations with traffic density 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, NO2, SO2, and PM2.5, as well as various exposure variables related to 

traffic, were estimated at the home address. Traffic intensity on the nearest road was independently 

associated with mortality. Relative risks (CI) for a 10 μg/m3

 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 (95% 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 NO2 and PM2.5, but no associations were found 

for SO2. 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 mortality were statistically significant for NO2 and BS. These results 

add to the evidence that long-term exposure to traffic-related particulate air pollution is associated 

with increased mortality. 

Given the general dearth of published source-oriented 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 useful 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 informative. In particular, Ozkaynak and Thurston (1987, 

072960), utilized the chemical speciation conducted in 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 TSP mass. Using multiple regression analyses, they 

conducted a cross-sectional analysis of the association 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, particle exposure measures most related to the 

respirable fraction of the aerosols, such as fine particles 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 PM10-2.5 (e.g., total suspended particles and inhalable 

particles) were often found to be non-significant predictors of total mortality. Furthermore, based on 

the application of PM2.5 source apportionment, particles from industrial sources and from coal 

combustion 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-city 

differences 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 PM2.5

monitoring stations in the Los Angeles area, and applying 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 PM2.5. Significant associations between PM2.5 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 PM2.5 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 error can result in stronger associations between PM2.5 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 Ratio1 and 95% Confidence 

Interval Using Kriging2 

(Jerrett et al., 2005, 087600) 

Los Angeles: 

Hazard Ratio1

 and 95% Confidence 

Interval Using Land Use Regression3 

(Krewski et al., 2009, 191193) 

New York City: 

Hazard Ratio1

 and 95% Confidence 

Interval Using Land Use Regression4 

(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/m3 increase in PM2.5 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 exposures to the ACS participants. The investigators 

did not observe elevated effect estimates for all cause, CPD or lung cancer deaths, but IHD did show 

a positive association with PM2.5 concentration. The difference between the 90th and 10th percentiles 

of the 3-yr avg PM2.5 concentration was 1.5 μg/m3 and the difference between the minimum and 

maximum values of the 3-yr avg PM2.5 concentration was 7.8 μg/m3

. This narrow range in PM2.5

exposure contrasts across the New York City metropolitan area and may well account for the 

inconclusive results in this city-specific analysis. 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- versus 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 7-91 
Source: Miller et al. (2007, 090130) 

Copyright © 2007 Massachusetts Medical Society. All rights reserved. 

Figure 7-8. Plots of the relative risk of death from cardiovascular disease from the Women’s 

Health Initiative study displaying the between-city and within-city contributions 

to the overall association between PM2.5 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 hazards 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). The 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. The lag relationship by model was averaged across a 

range of unconstrained distributed lag models (i.e., same year, 1 yr prior, 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 also noted that the concentration-response curve was 

linear, clearly continuing below the level of the current U.S. air quality standard of 15 μg/m3

. 

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

Figure 7-9. The model-averaged estimated effect of a 10-μg/m3 increase in PM2.5 on all-cause 

mortality at different lags (in years) between exposure and death. Each lag is 

estimated independently of the others. Also shown are the pointwise 95% CIs for 

each lag, based on jackknife estimates. 

Similarly, the effect of long-term exposure to PM10 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 preceding years (Zanobetti 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 pollution 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 PM10 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 7-93 
Source: 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 thin dashed line refers to the reference 

scenario. 

Table 7-10. Distribution of the effect of a hypothetical reduction of 10 μg/m3 PM10 in 2000 on allcause mortality 2000-2009 in Switzerland. 

Year 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/m3 reduction in PM10) 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 investigators calculated time windows of exposure as 

average concentrations during successive 5-yr periods 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 periods 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 7-94 
Generally, these results indicate a developing 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. PM2.5

In the 1996 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 PM2.5, though there was no evidence to support an 

association with PM10-2.5 (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 the 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. population. Collectively, the 2004 PM AQCD found that 

these studies provided strong evidence that long-term exposure to PM2.5 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 PM2.5 exposure and increased risk of human mortality (Section 7.6) 

in areas with mean concentrations from 13.2 to 29 μg/m3

 (Figure 7-7). New evidence from the Six 

Cities cohort study shows a relatively large risk estimate for reduced mortality risk with decreases in 

PM2.5 (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 indicate that cardiovascular mortality primarily accounts for the total 

mortality association with PM2.5 among adults, and not respiratory mortality. The recent WHI cohort 

study shows even higher cardiovascular risks per μg/m3 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 PM2.5 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 PM2.5 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 PM2.5 and 

the findings from a multiyear expert judgment study that comprehensively characterizes the size and 

uncertainty in estimates of mortality reductions associated with decreases in PM2.5 in the U.S draw 

conclusions that are consistent with an association between long-term exposure to PM2.5 and 

mortality (Pope et al., 2009, 190107; Roman et al., 2008, 156921). Pope et al. (2009, 190107) report 

that a decrease of 10 μg/m3

 in the concentration of PM2.5 is associated with an estimated increase in 

mean (± SE) life expectancy of 0.61 ± 0.20 year. 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 contributing 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 followed best standard practices for expert 

elicitations based on the body of literature accumulated over the past two decades. The resulting 

PM2.5 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 PM2.5 exposure in the U.S. (averaging almost 1% per μg/m3 PM2.5) than 

reported (for example) by the ACS Study (i.e., 0.6% per μg/m3

 PM2.5 in Pope et al. (2002, 024689), 

December 2009 7-95 
and a wider distribution of uncertainty by each expert than provided by any one of the PM2.5

epidemiologic studies. However, a composite uncertainty 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 PM2.5

mortality concentration-response coefficient for a 1 μg/m3 change in annual 

average PM2.5. 

Overall, recent evidence supports the strong evidence reported in the 2004 PM AQCD 

(U.S. EPA, 2004, 056905) that long-term exposure to PM2.5 is associated with an increased risk of 

human mortality. When looking at the cause of death, the strongest evidence comes from mortality 

due to cardiovascular disease, with additional evidence supporting an association between PM2.5 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 

PM2.5 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 long-term exposure to PM on cardiovascular effects 

presented in Sections 6.2 and 7.2, respectively. Evidence of short- and long-term exposure to PM2.5

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 provides biological plausibility for mortality due to 

cardiovascular or respiratory disease. The most recent evidence for the association between longterm exposure to PM2.5 and mortality is particularly strong for women. Collectively, the evidence is 

sufficient to conclude that the relationship between long-term PM2.5 exposures and 

mortality is causal. 

December 2009 7-96 
7.6.5.2. 

7.6.5.3. 

PM10-2.5 

In the 2004 PM AQCD, results from the ACS and Six Cities study analyses indicated that 

PM10-2.5 was not associated with mortality. Evidence is still limited to adequately characterize the 

association between PM10-2.5 and PM sources and/or components. The new findings from AHSMOG 

and Veterans cohort studies provide limited evidence of associations between long-term exposure to 

PM10-2.5 and mortality in areas with mean concentrations from 16 to 25 μg/m3

. The evidence for 

PM10-2.5 is inadequate to determine if a causal relationship 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 effects 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 
Chapter 7 References 

Abbey DE; Lebowitz MD; Mills PK; Petersen FF; Beeson WL; Burchette RJ (1995). Long-term ambient concentrations of 

particulates and oxidants and development of chronic disease in a cohort of nonsmoking California residents. 

Presented at In: Phalen, R. F.; Bates, D. V., eds. Proceedings of the colloquium on particulate air pollution and 

human mortality and morbidity; January 1994; Irvine, CA. Inhalation Toxicol. 7: 19-34., Irvine, CA. 000669

Abbey DE; Mills PK; Petersen FF; Beeson WL (1991). Long-term ambient concentrations of total suspended particulates 

and oxidants as related to incidence of chronic disease in California Seventh-Day Adventists. Environ Health 

Perspect, 94: 43-50. 042668

Abbey DE; Nishino N; McDonnell WF; Burchette RJ; Knutsen SF; Beeson WL; Yang JX (1999). Long-term inhalable 

particles and other air pollutants related to mortality in nonsmokers. Am J Respir Crit Care Med, 159: 373-382. 

047559

Abou Chakra OR; Joyeux M; Nerriere E; Strub MP; Zmirou-Navier D (2007). Genotoxicity of organic extracts of urban 

airborne particulate matter: an assessment within a personal exposure study. Chemosphere, 66: 1375-81. 098819

Achenbach S; Daniel WG (2001). Noninvasive coronary angiography--an acceptable alternative? N Engl J Med, 345: 

1909-1910. 156189

Ackermann-Liebrich U; Kuna-Dibbert B; Probst-Hensch NM; Schindler C; Dietrich DF; Stutz EZ; Bayer-Oglesby L; 

Baum F; Brandli O; Brutsche M; Downs SH; Keidel D; Gerbase MW; Imboden M; Keller R; Knopfli B; Kunzli 

Nellweger J-P; Leuenberger P; SALPADIA Team (2005). Follow-up of the Swiss Cohort study on air pollution and 

lung diseases in adults (SAPALDIA 2) 1991-2003: methods and characterization of participants. Int J Public Health, 

50: 245-263. 087826

Ackermann-Liebrich U; Leuenberger P; Schwartz J; Schindler C; Monn C; Bolognini B; Bongard JP; Brandli O; 

Domenighetti G; Elsasser S; Grize L; Karrer W; Keller R; Keller-Wossidlo H; Kunzli N; Martin BW; Medici TC; 

Pliger B; Wuthrich B; Zellweger JP; Zemp E (1997). Lung function and long term exposure to air pollutants in 

Switzerland. Am J Respir Crit Care Med, 155: 122-129. 077537

Agatston AS; Janowitz WR; Hildner FJ; Zusmer NR; Viamonte M Jr; Detrano R (1990). Quantification of coronary artery 

calcium using ultrafast computed tomography. J Am Coll Cardiol, 15: 827-832. 156197

Alink GM; Sjogren M; Bos RP; Doekes G; Kromhout H; Scheepers PTJ (1998). Effect of airborne particles from selected 

indoor and outdoor environments on gap-junctional intercellular communication. Toxicol Lett, 96/97: 209-213. 

087159

Allen RW; Criqui MH; Diez Roux AV; Allison M; Shea S; Detrano R; Sheppard L; Wong N; Hinckley Stukovsky K; 

Kaufman JD (2009). Fine particulate air pollution, proximity to traffic, and aortic atherosclerosis: The multi-ethnic 

study of atherosclerosis. Epidemiology, 20: 254-264. 156209

Allison MA; Cheung P; Criqui MH; Langer RD; Wright CM (2006). Mitral and aortic annular calcification are highly 

associated with systemic calcified atherosclerosis. Circulation, 113: 861-866. 155653

Allison MA; Criqui MH; Wright CM (2004). Patterns and risk factors for systemic calcified atherosclerosis. Arterioscler 

Thromb Vasc Biol, 24: 331-336. 156210

Amato M; Montorsi P; Ravani A; Oldani E; Galli S; Ravagnani PM; Tremoli E; Baldassarre D (2007). Carotid intimamedia thickness by B-mode ultrasound as surrogate of coronary atherosclerosis: correlation with quantitative 

coronary angiography and coronary intravascular ultrasound findings. Eur Heart J, 28: 2094-2101. 155656

Annesi-Maesano I; Moreau D; Caillaud D; Lavaud F; Le Moullec Y; Taytard A; Pauli G; Charpin D (2007). Residential 

proximity fine particles related to allergic sensitisation and asthma in primary school children. Respir Med, 101: 

1721-1729. 093180

 

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 7-98 
Arad Y; Spadaro LA; Goodman K; Lledo-Perez A; Sherman S; Lerner G; Guerci AD (1996). Predictive value of electron 

beam computed tomography of the coronary arteries. 19-month follow-up of 1173 asymptomatic subjects. 

Circulation, 93: 1951-1953. 155661

Araujo JA; Barajas B; Kleinman M; Wang X; Bennett BJ; Gong KW; Navab M; Harkema J; Sioutas C; Lusis AJ; Nel AE 

(2008). Ambient particulate pollutants in the ultrafine range promote early atherosclerosis and systemic oxidative 

stress. Circ Res, 102: 589-596. 156222

Ardehali R; Nasir K; Kolandaivelu A; Budoff MJ; Blumenthal RS (2007). Screening patients for subclinical atherosclerosis 

with non-contrast cardiac CT. Atherosclerosis, 192: 235-242. 155662

Arlt VM; Glatt H; Gamboa da Costa G; Reynisson J; Takamura-Enya T; Phillips DH (2007). Mutagenicity and DNA 

adduct formation by the urban air pollutant 2-nitrobenzanthrone. Toxicol Sci, 98: 445-57. 097257

Auchincloss AH; Roux AV; Dvonch JT; Brown PL; Barr RG; Daviglus ML; Goff DC; Kaufman JD; O'Neill MS (2008). 

Associations between Recent Exposure to Ambient Fine Particulate Matter and Blood Pressure in the Multi-Ethnic 

Study of Atherosclerosis (MESA). Environ Health Perspect, 116: 486-491. 156234

Avogbe PH; Ayi-Fanou L; Autrup H; Loft S; Fayomi B; Sanni A; Vinzents P; Moller P (2005). Ultrafine particulate matter 

and high-level benzene urban air pollution in relation to oxidative DNA damage. Carcinogenesis, 26: 613-620. 

087811

Avol EL; Gauderman WJ; Tan SM; London SJ; Peters JM (2001). Respiratory effects of relocating to areas of differing air 

pollution levels. Am J Respir Crit Care Med, 164: 2067-2072. 020552

Baccarelli A; Martinelli I; Pegoraro V; Melly S; Grillo P; Zanobetti A; Hou L; Bertazzi PA; Mannucci PM; Schwartz J 

(2009). Living near Major Traffic Roads and Risk of Deep Vein Thrombosis. Circulation, 119: 3118-3124. 188183

Baccarelli A; Martinelli I; Zanobetti A; Grillo P; Hou LF; Bertazzi PA; Mannucci PM; Schwartz J (2008). Exposure to 

Particulate Air Pollution and Risk of Deep Vein Thrombosis. Arch Intern Med, 168: 920-927. 157984

Baccarelli A; Wright RO; Bollati V; Tarantini L; Litonjua AA; Suh HH; Zanobetti A; Sparrow D; Vokonas PS; Schwartz J 

(2009). Rapid DNA Methylation Changes after Exposure to Traffic Particles . Am J Respir Crit Care Med, 179: 

572-578. 192155

Baker DG (1998). Natural pathogens of laboratory mice, rats, and rabbits and their effects on research. Clin Microbiol Rev, 

11: 231-266. 156245

Bao L; Chen S; Wu L; Hei Tom K; Wu Y; Yu Z; Xu A (2007). Mutagenicity of diesel exhaust particles mediated by cellparticle interaction in mammalian cells. Toxicol Sci, 229: 91-100. 097258

Basu R; Woodruff TJ; Parker JD; Saulnier L; Schoendorf KC (2004). Comparing exposure metrics in the relationship 

between PM25 and birth weight in California. J Expo Sci Environ Epidemiol, 14: 391-396. 087896

Bayer-Oglesby L; Grize L; Gassner M; Takken-Sahli K; Sennhauser FH; Neu U; Schindler C; Braun-Fahrlander C (2005). 

Decline of ambient air pollution levels and improved respiratory health in Swiss children. Environ Health Perspect, 

113: 1632-1637. 086245

Beelen R; Hoek G; van den Brandt PA; Goldbohm RA; Fischer P; Schouten LJ; Armstrong B; Brunekreef B (2008). Longterm exposure to traffic-related air pollution and lung cancer risk. Epidemiology, 19: 702-710. 155681

Beelen R; Hoek G; van den Brandt PA; Goldbohm RA; Fischer P; Schouten LJ; Jerrett M; Hughes E; Armstrong B; 

Brunekreef B (2008). Long-term effects of traffic-related air pollution on mortality in a Dutch cohort (NLCS-AIR 

study). Environ Health Perspect, 116: 196-202. 156263

Beeson WL; Abbey DE; Knutsen SF (1998). Long-term concentrations of ambient air pollutants and incident lung cancer 

in California adults: results from the AHSMOG study. Environ Health Perspect, 106: 813-823. 048890

Bell ML; Ebisu K; Belanger K (2007). Ambient air pollution and low birth weight in Connecticut and Massachusetts. 

Environ Health Perspect, 115: 1118-24. 091059

Bell ML; Kim JY; Dominici F (2007). Potential confounding of particulate matter on the short-term association between 

ozone and mortality in multisite time-series studies. Environ Health Perspect, 115: 1591-1595. 093256

Bhattacharyya N (2009). Air quality influences the prevalence of hay fever and sinusitis. Laryngoscope, 119: 429-433. 

180154

December 2009 7-99 
Bigazzi R; Bianchi S; Baldari D; Campese VM (1998). Microalbuminuria predicts cardiovascular events and renal 

insufficiency in patients with essential hypertension. J Hypertens, 16: 1325-1333. 156272

Binková B; Cerná M; Pastorková A; Jeli ́nek R; Beneš I; Novák J; Šrám RJ (2003). Biological activities of organic 

compounds adsorbed onto ambient air particles: comparison between the cities of Teplice and Prague during the 

summer and winter seasons 2000–2001. Mutat Res Fund Mol Mech Mutagen, 525: 43-59. 156274

Bobak M (2000). Outdoor air pollution, low birth weight, and prematurity. Environ Health Perspect, 108: 173-176. 011448

Bobak M; Leon DA (1992). Air pollution and infant mortality in the Czech Republic, 1986-1988. Lancet, 8826: 1010-1014. 

044415

Bobak M; Leon DA (1999). The effect of air pollution on infant mortality appears specific for respiratory causes in the 

postneonatal period. Epidemiology, 10: 666-670. 007678

Bonner MR; Han D; Nie J; Rogerson P; Vena JE; Muti P; Trevisan M; Edge SB; Freudenheim JL (2005). Breast cancer risk 

and exposure in early life to polycyclic aromatic hydrocarbons using total suspended particulates as a proxy 

measure. Cancer Epidemiol Biomarkers Prev, 14: 53-60. 088993

Boushey HA; Armstrong B; Brauer M; Brunekreef B; Buckpitt A; Hoidal JR; London S; Reid N; Rom WN; Russell A 

(2008). Critique, Health Review Committee. In Grigg J; Kulkarni N; Pierse N; Rushton L; O?Callaghan C; Rutman 

A (Ed.),Research report 134: Black-pigmented material in airway macrophages from healthy children: Association 

with lung function and modeled PM10 (pp. 25-33). Boston, MA: Health Effects Institute. 192162

Bracken MB; Triche EW; Belanger K; Saftlas A; Beckett WS; Leaderer BP (2003). Asthma symptoms, severity, and drug 

therapy: a prospective study of effects on 2205 pregnancies. Obstet Gynecol, 102: 739-753. 156288

Brauer M; Gehring U; Brunekreef B; De Jongste J; Gerritsen J; Rovers M; Wichmann H-E; Wijga A; Heinrich J (2006). 

Traffic-related air pollution and otitis media. Environ Health Perspect, 114: 1414-1418. 090757

Brauer M; Hoek G; Smit HA; De Jongste JC; Gerritsen J; Postma DS; Kerkhof M; Brunekreef B (2007). Air pollution and 

development of asthma, allergy and infections in a birth cohort. Eur Respir J, 29: 879-888. 090691

Brauer M; Hoek G; Van Vliet P; Meliefste K; Fischer PH; Wijga A; Koopman LP; Neijens HJ; Gerritsen J; Kerkhof M; 

Heinrich J; Bellander T; Brunekreef B (2002). Air pollution from traffic and the development of respiratory 

infections and asthmatic and allergic symptoms in children. Am J Respir Crit Care Med, 166: 1092-1098. 035192

Brauer M; Lencar C; Tamburic L; Koehoorn M; Demers P; Karr C (2008). A cohort study of traffic-related air pollution 

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

Brown JS; Zeman KL; Bennett WD (2001). Regional deposition of coarse particles and ventilation distribution in healthy 

subjects and patients with cystic fibrosis. J Aerosol Med Pulm Drug Deliv, 14: 443-454. 012307

Brunekreef B; Beelen R; Hoek G; Schouten L; Bausch-Goldbohm S; Fischer P; Armstrong B; Hughes E; Jerrett M; van 

den Brandt P (2009). Effects of long-term exposure to traffic-related air pollution on respiratory and cardiovascular 

mortality in the Netherlands: The NLCS-AIR Study. Health Effects Institute. Boston, MA. 139. 191947

Budoff MJ; Takasu J; Katz R; Mao S; Shavelle DM; O'Brien KD; Blumenthal RS; Carr JJ; Kronmal R (2005). 

Reproducibility of CT measurements of aortic valve calcification, mitral annulus calcification, and aortic wall 

calcification in the multi-ethnic study of atherosclerosis. Acad Radiol, 13: 166-172. 192105

Bunger J; Krahl J; Weigel A; Schroder O; Bruning T; Muller M; Hallier E; Westphal G (2006). Influence of fuel properties, 

nitrogen oxides, and exhaust treatment by an oxidation catalytic converter on the mutagenicity of diesel engine 

emissions. Arch Toxicol, 80: 540-546. 156303

Bunger J; Schappler-Scheele B; Hilgers R; Hallier E (2007). A 5-year follow-up study on respiratory disorders and lung 

function in workers exposed to organic dust from composting plants. Int Arch Occup Environ Health, 80: 306-312. 

156304

Burchiel SW; Lauer FT; Dunaway SL; Zawadzki J; McDonald JD; Reed MD (2005). Hardwood smoke alters murine 

splenic T cell responses to mitogens following a 6-month whole body inhalation exposure. Toxicol Appl Pharmacol, 

202: 229-236. 088090

Burchiel SW; Lauer FT; McDonald JD; Reed MD (2004). Systemic immunotoxicity in AJ mice following 6-month whole 

body inhalation exposure to diesel exhaust. Toxicol Appl Pharmacol, 196: 337-345. 055557

December 2009 7-100 
Calderón-Garcidueñas L; Macías-Parra M; Hoffmann HJ; Valencia-Salazar G; Henríquez-Roldán C; Osnaya N; Monte OC; 

Barragán-Mejía G; Villarreal-Calderon R; Romero L; Granada-Macías M; Torres-Jardón R; Medina-Cortina H; 

Maronpot RR (2009). Immunotoxicity and environment: immunodysregulation and systemic inflammation in 

children. Toxicol Pathol, 37: 161-169. 192107

Calderon-Garciduenas L; Vincent R; Mora-Tiscareno A; Franco-Lira M; Henriquez-Roldan C; Barragan-Mejia G; GarridoGarcia L; Camacho-Reyes L; Valencia-Salazar G; Paredes R; Romero L; Osnaya H; Villarreal-Calderon R; TorresJardon (2007). Elevated plasma endothelin-1 and pulmonary arterial pressure in children exposed to air pollution. 

Environ Health Perspect, 115: 1248-1253. 091252

Chambless LE; Heiss G; Folsom AR; Rosamond W; Szklo M; Sharrett AR; Clegg LX (1997). Association of coronary 

heart disease incidence with carotid arterial wall thickness and major risk factors: the Atherosclerosis Risk in 

Communities (ARIC) Study, 1987-1993. Am J Epidemiol, 146: 483-494. 156329

Chang J; Delfino RJ; Gillen D; Tjoa T; Nickerson B; Cooper D (2008). Repeated respiratory hospital encounters among 

children with asthma and residential proximity to traffic. J Occup Environ Med, 66: 90-98. 180393

Chen JC; Schwartz J (2008). Metabolic syndrome and inflammatory responses to long-term particulate air pollutants. 

Environ Health Perspect, 116: 612-617. 190106

Chen L; Yang W; Jennison BL; Goodrich A; Omaye ST (2002). Air pollution and birth weight in northern Nevada, 1991-

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

Chen LC; Hwang JS (2005). Effects of subchronic exposures to concentrated ambient particles (CAPs) in mice IV 

Characterization of acute and chronic effects of ambient air fine particulate matter exposures on heart-rate 

variability. Inhal Toxicol, 17: 209-216. 087218

Chen LC; Nadziejko C (2005). Effects of subchronic exposures to concentrated ambient particles (CAPs) in mice V CAPs 

exacerbate aortic plaque development in hyperlipidemic mice. Inhal Toxicol, 17: 217-224. 087219

Chen LH; Knutsen SF; Shavlik D; Beeson WL; Petersen F; Ghamsary M; Abbey D (2005). The association between fatal 

coronary heart disease and ambient particulate air pollution: Are females at greater risk? Environ Health Perspect, 

113: 1723-1729. 087942

Churg A; Brauer M; del Carmen Avila-Casado M; Fortoul TI; Wright JL (2003). Chronic exposure to high levels of 

particulate air pollution and small airway remodeling. Environ Health Perspect, 111: 714-718. 087899

Clancy L; Goodman P; Sinclair H; Dockery DW (2002). Effect of air pollution control on death rates in Dublin, Ireland: an 

intervention study. Lancet, 360: 1210-1214. 035270

Claxton LD; Matthews PP; Warren SH (2004). The genotoxicity of ambient outdoor air, a review: Salmonella mutagenicity. 

DNA Repair (Amst), 567: 347-399. 089008

Claxton LD; Woodall GM (2007). A review of the mutagenicity and rodent carcinogenicity of ambient air. Mutat Res Rev 

Mutat Res, 636: 36-94. 180391

Clifton VL; Giles WB; Smith R; Bisits AT; Hempenstall PA; Kessell CG; Gibson PG (2001). Alterations of placental 

vascular function in asthmatic pregnancies. Am J Respir Crit Care Med, 164: 546-554. 156360

Craven TE; Ryu JE; Espeland MA; Kahl FR; McKinney WM; Toole JF; McMahan MR; Thompson CJ; Heiss G; Crouse JR 

3rd (1990). Evaluation of the associations between carotid artery atherosclerosis and coronary artery stenosis. A 

case-control study. Circulation, 82: 1230-1242. 155740

Cury PM; Lichtenfels AJ; Reymão MS; Conceição GM; Capelozzi VL; Saldiva PH (2000). Urban levels of air pollution 

modifies the progression of urethane-induced lung tumours in mice. Pathol Res Pract, 196: 627-633. 192100

Dales R; Burnett RT; Smith-Doiron M; Stieb DM; Brook JR (2004). Air pollution and sudden infant death syndrome. 

Pediatrics, 113: 628-631. 087342

Dales R; Wheeler A; Mahmud M; Frescura AM; Smith-Doiron M; Nethery E; Liu L (2008). The Influence of Living Near 

Roadways on Spirometry and Exhaled Nitric Oxide in Elementary Schoolchildren. Environ Health Perspect, 116: 

1423-1427. 156378

Danielsen PH; Loft S; Møller P (2008). DNA damage and cytotoxicity in type II lung epithelial (A549) cell cultures after 

exposure to diesel exhaust and urban street particles. Part Fibre Toxicol, 5: 6. 192092

December 2009 7-101 
Deckert T; Yokoyama H; Mathiesen E; Ronn B; Jensen T; Feldt-Rasmussen B; Borch-Johnsen K; Jensen JS (1996). Cohort 

study of predictive value of urinary albumin excretion for atherosclerotic vascular disease in patients with insulin 

dependent diabetes. Br Med J, 312: 871-874. 156389

DeMarini DM; Brooks LR; Warren SH; Kobayashi T; Gilmour MI; Singh P (2004). Bioassay-directed fractionation and 

Salmonella mutagenicity of automobile and forklift diesel exhaust particles. Environ Health Perspect, 112: 814-819. 

066329

De Kok TM; Hogervorst JG; Briede JJ; Van Herwijnen MH; Maas LM; Moonen EJ; Driece HA; Kleinjans JC (2005). 

Genotoxicity and physicochemical characteristics of traffic-related ambient particulate matter. Environ Mol 

Mutagen, 46: 71-80. 088656

Diez Roux AV; Auchincloss AH; Franklin TG; Raghunathan T; Barr RG; Kaufman J; Astor B; Keeler J (2008). Long-term 

exposure to ambient particulate matter and prevalence of subclinical atherosclerosis in the Multi-Ethnic Study of 

Atherosclerosis. Am J Epidemiol, 167: 667-675. 156401

Dinneen SF; Gerstein HC (1997). The association of microalbuminuria and mortality in non-insulin-dependent diabetes 

mellitus. A systematic overview of the literature. Arch Intern Med, 157: 1413-1418. 156403

Dockery DW; Damokosh AI; Neas LM; Raizenne M; Spengler JD; Koutrakis P; Ware JH; Speizer FE (1996). Health 

effects of acid aerosols on North American children: respiratory symptoms and illness. Environ Health Perspect, 

104: 500-505. 046219

Dockery DW; Pope CA III; Xu X; Spengler JD; Ware JH; Fay ME; Ferris BG Jr; Speizer FE (1993). An association 

between air pollution and mortality in six US cities. N Engl J Med, 329: 1753-1759. 044457

Downs SH; Schindler C; Liu L-JS; Keidel D; Bayer-Oglesby L; Brutsche MH; Gerbase MW; Keller R; Kunzli N; 

Leuenberger P; Probst-Hensch NM; Tschopp J-M; Zellweger J-P; Rochat T; Schwartz J; Ackermann-Liebrich U; 

(2007). Reduced exposure to PM10 and attenuated age-related decline in lung function. J Expo Sci Environ 

Epidemiol, 15: 185-204. 092853

Dugandzic R; Dodds L; Stieb D; Smith-Doiron M (2006). The association between low level exposures to ambient air 

pollution and term low birth weight: A retrospective cohort study. Environ Health, 5: 3. 088681

Eftim SE; Samet JM; Janes H; McDermott A; Dominici F (2008). Fine particulate matter and mortality: a comparison of 

the six cities and American Cancer Society cohorts with a medicare cohort. Epidemiology, 19: 209-216. 099104

Enstrom JE (2005). Fine particulate air pollution and total mortality among elderly Californians, 1973-2002. Inhal Toxicol, 

17: 803-816. 087356

Erbel R; Mohlenkamp S; Kerkhoff G; Budde T; Schmermund A (2007). Non-invasive screening for coronary artery 

disease: calcium scoring. Heart, 93: 1620-1629. 155768

Erdinger L; Durr M; Hopker KA (2005). Correlations between mutagenic activity of organic extracts of airborne particulate 

matter, NOx and sulphur dioxide in southern Germany: results of a two-year study. Environ Sci Pollut Res Int, 12: 

10-20. 156423

Fedulov AV; Leme A; Yang Z; Dahl M; Lim R; Mariani TJ; Kobzik L (2008). Pulmonary exposure to particles during 

pregnancy causes increased neonatal asthma susceptibility. Am J Respir Cell Mol Biol, 38: 57-67. 097482

Finch GL; Hobbs CH; Blair LF; Barr EB; Hahn FF; Jaramillo RJ; Kubatko JE; March TH; White RK; Krone JR; Menache 

MG; Nikula KJ; Mauderly JL; Van Gerpen J; Merceica MD; Zielinska B; Stankowski L; Burling K; Howell S; 

Mauderly JL (2002). Effects of subchronic inhalation exposure of rats to emissions from a diesel engine burning 

soybean oil-derived biodiesel fuel. Inhal Toxicol, 14: 1017-1048. 054603

Floyd HS; Chen LC; Vallanat B; Dreher K (2009). Fine ambient air particulate matter exposure induces molecular 

alterations associated with vascular disease progression within plaques of atherosclerotic susceptible mice. Inhal 

Toxicol, 21: 394-403. 190350

Forbes LJ; Patel MD; Rudnicka AR; Cook DG; Bush T; Stedman JR; Whincup PH; Strachan DP; Anderson RH (2009). 

Chronic exposure to outdoor air pollution and markers of systemic inflammation. Epidemiology, 20: 245-253. 

190351

Forman JP; Brenner BM (2006). 'Hypertension' and 'microalbuminuria': the bell tolls for thee. Kidney Int, 69: 22-28. 

156439

December 2009 7-102 
Foster PMD; Mylchreest E; Gaido KW; Sar M (2001). Effects of phthalate esters on the developing reproductive tract of 

male rats. Hum Reprod Update, 7: 231-235. 156442

Foster PMD; Thomas LV; Cook MW; Gangolli SD (1980). Study of the testicular effects and changes in zinc excretion 

produced by some n-alkyl phthalates in the rat. Toxicol Appl Pharmacol, 54: 392-398. 094701

Franklin SS; Sutton-Tyrrell K; Belle SH; Weber MA; Kuller LH (1997). The importance of pulsatile components of 

hypertension in predicting carotid stenosis in older adults. J Hypertens, 15: 1143-1150. 156446

Fryer ME; Collins CD (2003). Model intercomparison for the uptake of organic chemicals by plants. Environ Sci Technol, 

37: 1617–1624. 156454

Fujimoto A; Tsukue N; Watanabe M; Sugawara I; Yanagisawa R; Takano H; Yoshida S; Takeda K (2005). Diesel exhaust 

affects immunological action in the placentas of mice. Environ Toxicol, 20: 431-440. 096556

Gabelová A; Valovicova Z; Bacova G; Labaj J; Binkova B; Topinka J; Sevastyanova O; Sram RJ; Kalina I; Habalova V; 

Popov TA; Panev T; Farmer PB (2007). Sensitivity of different endpoints for in vitro measurement of genotoxicity 

of extractable organic matter associated with ambient airborne particles (PM10). Mutat Res Fund Mol Mech 

Mutagen, 620: 103-113. 156458

Gabelová A; Valovicova Z; Labaj J; Bacova G; Binkova B; Farmer Peter B (2007). Assessment of oxidative DNA damage 

formation by organic complex mixtures from airborne particles PM(10). Mutat Res, 620: 135-144. 156457

Gardner MB (1966). Biological effects of urban air pollution III Lung tumors in mice. Arch Environ Occup Health, 12: 

305-313. 015129

Gardner MB; Loosli CG; Hanes B; Blackmore W; Teebken D (1969). Histopathologic findings in rats exposed to ambient 

and filtered air. Arch Environ Occup Health, 19: 637-647. 015130

Gauderman WJ; Avol E; Gilliland F; Vora H; Thomas D; Berhane K; McConnell R; Kuenzli N; Lurmann F; Rappaport E; 

Margolis H; Bates D; Peters J (2004). The effect of air pollution on lung development from 10 to 18 years of age. N 

Engl J Med, 351: 1057-1067. 056569

Gauderman WJ; Gilliland GF; Vora H; Avol E; Stram D; McConnell R; Thomas D; Lurmann F; Margolis HG; Rappaport 

EB; Berhane K; Peters JM (2002). Association between air pollution and lung function growth in southern 

California children: results from a second cohort. Am J Respir Crit Care Med, 166: 76-84. 026013

Gauderman WJ; McConnell R; Gilliland F; London S; Thomas D; Avol E; Vora H; Berhane K; Rappaport EB; Lurmann F; 

Margolis HG; Peters J (2000). Association between air pollution and lung function growth in southern California 

children. Am J Respir Crit Care Med, 162: 1383-1390. 012531

Gehrig R; Buchmann B (2003). Characterising seasonal variations and spatial distribution of ambient PM10 and PM25 

concentrations based on long-term Swiss monitoring data. Atmos Environ, 37: 2571-2580. 139678

Gehring U; Heinrich J; Kramer U; Grote V; Hochadel M; Sugiri D; Kraft M; Rauchfuss K; Eberwein HG; Wichmann H-E 

(2006). Long-term exposure to ambient air pollution and cardiopulmonary mortality in women. Epidemiology, 17: 

545-551. 089797

Geroulakos G; O'Gorman DJ; Kalodiki E; Sheridan DJ; Nicolaides AN (1994). The carotid intima-media thickness as a 

marker of the presence of severe symptomatic coronary artery disease. Eur Heart J, 15: 781-785. 155788

Gerstein HC; Mann JF; Yi Q; Zinman B; Dinneen SF; Hoogwerf B; Halle JP; Young J; Rashkow A; Joyce C; Nawaz S; 

Yusuf S (2001). Albuminuria and risk of cardiovascular events, death, and heart failure in diabetic and nondiabetic 

individuals. JAMA, 286: 421-426. 156466

Ghosh R; Rankin J; Pless-Mulloli T; Glinianaia S (2007). Does the effect of air pollution on pregnancy outcomes differ by 

gender? A systematic review. Environ Res, 105: 400-408. 091233

Gilboa SM; Mendola P; Olshan AF; Langlois PH; Savitz DA; Loomis D; Herring AH; Fixler DE (2005). Relation between 

ambient air quality and selected birth defects, seven county study, Texas, 1997-2000. Am J Epidemiol, 162: 238-

252. 087892

Girerd X; Mourad JJ; Acar C; Heudes D; Chiche S; Bruneval P; Mignot JP; Billaud E; Safar M; Laurent S (1994). 

Noninvasive measurement of medium-sized artery intima-media thickness in humans: in vitro validation. J Vasc 

Res, 31: 114-120. 156474

December 2009 7-103 
Glinianaia FV; Rankin J; Bell R; Pless-Mulloli T; Howel D (2004). Does particulate air pollution contribute to infant 

death? A systematic review. Environ Health Perspect, 112: 1365-1371. 087898

Goss CH; Newsom SA; Schildcrout JS; Sheppard L; Kaufman JD (2004). Effect of ambient air pollution on pulmonary 

exacerbations and lung function in cystic fibrosis. Am J Respir Crit Care Med, 169: 816-821. 055624

Gotschi T; Heinrich J; Sunyer J; Kunzli N (2008). Long-term effects of ambient air pollution on lung function: A review. 

Epidemiology, 19: 690-701. 156485

Gotschi T; Sunyer J; Chinn S; de Marco R; Forsberg B; Gauderman JW; Garcia-Esteban R; Heinrich J; Jacquemin B; Jarvis 

D; Ponzio M; Villani S; Kunzli N (2008). Air pollution and lung function in the European Community Respiratory 

Health Survey. Int J Epidemiol, 37: 1349-1358. 180364

Gottipolu RR; Wallenborn JG; Karoly ED; Schladweiler MC; Ledbetter AD; Krantz T; Linak WP; Nyska A; Johnson JA; 

Thomas R; Richards JE; Jaskot RH; Kodavanti UP (2009). One-month diesel exhaust inhalation produces 

hypertensive gene expression pattern in healthy rats. Environ Health Perspect, 117: 38-46. 190360

Gouveia N; Bremner SA; Novaes HMD (2004). Association between ambient air pollution and birth weight in Sao Paulo, 

Brazil. J Epidemiol Community Health, 58: 11-17. 055613

Greenland P; Kizilbash MA (2005). Coronary computed tomography in coronary risk assessment. J Cardpulm Rehabil, 25: 

3-10. 156496

Grigg J; Kulkarni N; Pierse N; Rushton L; O'Callaghan C; Rutman A (2008). Black-pigmented material in airway 

macrophages from healthy children: association with lung function and modeled PM10. Heath Effects Institute. 

Boston, MA. Research Report 134. 156499

Gunnison A; Chen LC (2005). Effects of subchronic exposures to concentrated ambient particles (CAPs) in mice VI Gene 

expression in heart and lung tissue. Inhal Toxicol, 17: 225-233. 087956

Gutierrez-Castillo ME; Roubicek DA; Cebrian-Garcia ME; De Vizcaya-Ruiz A; Sordo-Cedeno M; Ostrosky-Wegman P 

(2006). Effect of chemical composition on the induction of DNA damage by urban airborne particulate matter. 

Environ Mol Mutagen, 47: 199-211. 089030

Ha E-H; Lee J-T; Kim H; Hong Y-C; Lee (2003). Infant susceptibility of mortality to air pollution in Seoul, South Korea. 

Pediatrics, 111: 284-290. 042552

Haland G; Carlsen KC; Sandvik L; Devulapalli CS; Munthe-Kaas MC; Pettersen M; Carlsen KH (2006). Reduced lung 

function at birth and the risk of asthma at 10 years of age. N Engl J Med, 355: 1682-1689. 156511

Hamada K; Suzaki Y; Leme A; Ito T; Miyamoto K; Kobzik L; Kimura H (2007). Exposure of pregnant mice to an air 

pollutant aerosol increases asthma susceptibility in offspring. J Toxicol Environ Health A Curr Iss, 70: 688-695. 

091235

Hammoud A; Carrell DT; Gibson M; Sanderson M; Parker-Jones K; Matthew Peterson C (2009). Decreased sperm motility 

is associated with air pollution in Salt Lake City. Fertil Steril, TBD: TBD. 192156

Hannigan MP; Cass GR; Penman BW; Crespi CL; Lafleur AL; Busby WF Jr; Thilly WG (1997). Human cell mutagens in 

Los Angeles air. Environ Sci Technol, 31: 438-447. 083598

Hansen C; Neller A; Williams G; Simpson R (2006). Maternal exposure to low levels of ambient air pollution and preterm 

birth in Brisbane, Australia. BJOG, 113: 935-941. 089818

Hansen C; Neller A; Williams G; Simpson R (2007). Low levels of ambient air pollution during pregnancy and fetal 

growth among term neonates in Brisbane, Australia. Environ Res, 103: 383-389. 090703

Harrod KS; Jaramillo RJ; Berger JA; Gigliotti AP; Seilkop SK; Reed MD (2005). Inhaled diesel engine emissions reduce 

bacterial clearance and exacerbate lung disease to Pseudomonas aeruginosa infection in vivo. Toxicol Sci, 83: 155-

165. 088144

Hashimoto AH; Amanuma K; Hiyoshi K; Sugawara Y; Goto S; Yanagisawa R; Takano H; Masumura K-I; Nohmi T; Aoki Y 

(2007). Mutations in the lungs of gpt delta transgenic mice following inhalation of diesel exhaust. Environ Mol 

Mutagen, 48: 682-693. 097261

Healey K; Smith EC; Wild CP; Routledge MN (2006). The mutagenicity of urban particulate matter in an enzyme free 

system is associated with the generation of reactive oxygen species. Mutat Res Fund Mol Mech Mutagen, 602: 1-6. 

156532

December 2009 7-104 
Heinrich J; Hoelscher B; Frye C; Meyer I; Pitz M; Cyrys J; Wjst M; Neas L; Wichmann H-E (2002). Improved air quality 

in reunified Germany and decreases in respiratory symptoms. Epidemiology, 13: 394-401. 034825

Heinrich J; Slama R (2007). Fine particles, a major threat to children. Int J Hyg Environ Health, 210: 617-622. 156534

Heiss G; Sharrett AR; Barnes R; Chambless LE; Szklo M; Alzola C (1991). Carotid atherosclerosis measured by B-mode 

ultrasound in populations: associations with cardiovascular risk factors in the ARIC study. Am J Epidemiol, 134: 

250-256. 156535

Hiramatsu K; Azuma A; Kudoh S; Desaki M; Takizawa H; Sugawara I (2003). Inhalation of diesel exhaust for three 

months affects major cytokine expression and induces bronchus-associated lymphoid tissue formation in murine 

lungs. Exp Lung Res, 29: 607-622. 155846

Hoek G; Brunekreef B; Goldbohm S; Fischer P; Van den Brandt PA (2002). Association between mortality and indicators 

of traffic-related air pollution in the Netherlands: A cohort study. Lancet, 360: 1203-1209. 042364

Hoffmann B; Moebus S; Kroger K; Stang A; Mohlenkamp S; Dragano N; Schmermund A; Memmesheimer M; Erbel R; 

Jockel K-H (2009). Residential exposure to urban pollution, ankle-brachial index, and peripheral arterial disease. 

Epidemiology, 20: 280-288. 190376

Hoffmann B; Moebus S; Mohlenkamp S; Stang A; Lehmann N; Dragano N; Schmermund A; Memmesheimer M; Mann K; 

Erbel R; Jockel K-H; Heinz Nixdorf Recall Study Investigative Group (2007). Residential exposure to traffic is 

associated with coronary atherosclerosis. Circulation, 116: 489-496. 091163

Hoffmann B; Moebus S; Stang A; Beck E-M; Dragano N; Mohlenkamp S; Schmermund A; Memmesheimer M; Mann K; 

Erbel R; Jockel K-H; Heinz Nixdorf RECALL Study Investigative Group (2006). Residence close to high traffic 

and prevalence of coronary heart disease. Eur Heart J, 27: 2696-2702. 091162

Hoffmann MH; Shi H; Schmitz BL; Schmid FT; Lieberknecht M; Schulze R; Ludwig B; Kroschel U; Jahnke N; Haerer W; 

Brambs HJ; Aschoff AJ (2005). Noninvasive coronary angiography with multislice computed tomography. JAMA, 

293: 2471-2478. 156556

Hollander M; Hak AE; Koudstaal PJ; Bots ML; Grobbee DE; Hofman A; Witteman JC; Breteler MM (2003). Comparison 

between measures of atherosclerosis and risk of stroke: the Rotterdam Study. Stroke, 34: 2367-2372. 156562

Hornberg C; Maciuleviciute L; Seemayer NH (1996). Sister chromatid exchanges in rodent tracheal epithelium exposed in 

vitro to environmental pollutants. Toxicol Lett, 88: 45-53. 087164

Hornberg C; Maciuleviciute L; Seemayer NH; Kainka E (1998). Induction of sister chromatid exchanges (SCE) in human 

tracheal epithelial cells by the fractions PM-10 and PM-25 of airborne particulates. Toxicol Lett, 96/97: 215-220. 

095741

Hougaard KS; Jensen KA; Nordly P; Taxvig C; Vogel U; Saber AT; Wallin H (2008). Effects of prenatal exposure to diesel 

exhaust particles on postnatal development, behavior, genotoxicity and inflammation in mice. Part Fibre Toxicol, 5: 

3. 156570

Huang JY; Liao JW; Liu YC; Lu SY; Chou CP; Chan WH; Chen SU; Ueng TH (2008). Motorcycle exhaust induces 

reproductive toxicity and testicular interleukin-6 in male rats. Toxicol Sci, 103: 137-148. 156574

Huynh M; Woodruff TJ; Parker JD; Schoendorf KC (2006). Relationships between air pollution and preterm birth in 

California. Paediatr Perinat Epidemiol, 20: 454-461. 091240

Hwang BF; Jaakkola JJ (2008). Ozone and other air pollutants and the risk of oral clefts. Environ Health Perspect, 116: 

1411-1415. 193794

Hwang J-S; Nadziejko C; Chen LC (2005). Effects of subchronic exposures to concentrated ambient particles (CAPs) in 

mice: III Acute and chronic effects of CAPs on heart rate, heart-rate fluctuation, and body temperature. Inhal 

Toxicol, 17: 199-207. 087957

IARC (1989). Diesel and gasoline engine exhausts. Lyon, France: International Agency for Research on Cancer. 002958

Iba MM; Fung J; Chung L; Zhao J; Winnik B; Buckley BT; Chen LC; Zelikoff JT; Kou YR (2006). Differential inducibility 

of rat pulmonary CYP1A1 by cigarette smoke and wood smoke. Mutat Res Fund Mol Mech Mutagen, 606: 1-11. 

156582

Ichinose T; Yajima Y; Nagashima M; Takenoshita S; Nagamachi Y; Sagai M (1997). Lung carcinogenesis and formation of 

8-hydroxy-deoxyguanosine in mice by diesel exhaust particles. Carcinogenesis, 18: 185-192. 053264

December 2009 7-105 
Ishihara Y; Kagawa J (2003). Chronic diesel exhaust exposures of rats demonstrate concentration and time-dependent 

effects on pulmonary inflammation. Inhal Toxicol, 15: 473-492. 096404

Islam T; Gauderman WJ; Berhane K; McConnell R; Avol E; Peters JM; Gilliland FD (2007). The relationship between air 

pollution, lung function and asthma in adolescents. Thorax, 62: 957-963. 090697

Izawa H; Watanabe G; Taya K; Sagai M (2007). Inhibitory effects of foods and polyphenols on activation of aryl 

hydrocarbon receptor induced by diesel exhaust particles. Environ Sci J Integr Environ Res, 14: 149-156. 190387

Jacobsen NR; Mrller P; Cohn CA; Loft S; Vogel U; Wallin H (2008). Diesel exhaust particles are mutagenic in FE1-

MutaTM Mouse lung epithelial cells. Mutat Res Fund Mol Mech Mutagen, 641: 54-57. 156597

Jalaludin B; Mannes T; Morgan G; Lincoln D; Sheppeard V; Corbett S (2007). Impact of ambient air pollution on 

gestational age is modified by season in Sydney, Australia. Environ Health, 6: 16. 156601

Janes H; Dominici F; Zeger SL (2007). Trends in air pollution and mortality: an approach to the assessment of unmeasured 

confounding. Epidemiology, 18: 416-423. 090927

Janssen NAH; Brunekreef B; van Vliet P; Aarts F; Maliefste K; Harssema H; Fischer P (2003). The relationship between 

air pollution from heavy traffic and allergic sensitization, bronchial byperresponsiveness, and respiratory symptoms 

in Dutch schoolchildren. Epidemiology, 111: 1512-1518. 133555

Jerrett M; Burnett RT; Ma R; Pope CA III; Krewski D; Newbold KB; Thurston G; Shi Y; Finkelstein N; Calle EE; Thun MJ 

(2005). Spatial analysis of air pollution and mortality in Los Angeles. Epidemiology, 16: 727-736. 087600

Jerrett M; Buzzelli M; Burnett RT; DeLuca PF (2005). Particulate air pollution, social confounders, and mortality in small 

areas of an industrial city. Soc Sci Med, 60: 2845-2863. 087381

Karr C; Lumley T; Schreuder A; Davis R; Larson T; Ritz B; Kaufman J (2007). Effects of subchronic and chronic exposure 

to ambient air pollutants on infant bronchiolitis. Am J Epidemiol, 165: 553-560. 090719

Karr C; Rudra C; Miller K; Gould T; Larson T; Sathyanarayana S; Koenig J (2009). Infant exposure to fine particulate 

matter and traffic and risk of hospitalization for RSV bronchiolitis in a region with lower ambient air pollution. 

Environ Res, 109: 321-327. 191946

Karthikeyan VJ; Lip GYH (2007). Peripheral artery disease and hypertension: the relation between ankle-brachial index 

and mortality. J Hum Hypertens, 21: 762-765. 156626

Kato A; Kagawa J (2003). Morphological effects in rat lungs exposed to urban roadside air. Inhal Toxicol, 15: 799-818. 

089563

Khattar RS; Swales JD; Dore C; Senior R; Lahiri A (2001). Effect of aging on the prognostic significance of ambulatory 

systolic, diastolic, and pulse pressure in essential hypertension. Circulation, 104: 783-789. 155896

Khoury Z; Schwartz R; Gottlieb S; Chenzbraun A; Stern S; Keren A (1997). Relation of coronary artery disease to 

atherosclerotic disease in the aorta, carotid, and femoral arteries evaluated by ultrasound. Am J Cardiol, 80: 1429-

1433. 156636

Kim JJ; Smorodinsky S; Lipsett M; Singer BC; Hodgson AT; Ostro B (2004). Traffic-related air pollution near busy roads: 

the East Bay children's Respiratory Health Study. Am J Respir Crit Care Med, 170: 520-526. 087383

Kim OJ; Ha EH; Kim BM; Park HS; Jung WJ; Lee BE; Suh YJ; Kim YJ; Lee JT; Kim H; Hong YC (2007). PM10 and 

pregnancy outcomes: a hospital-based cohort study of pregnant women in Seoul. J Occup Environ Med, 49: 1394-

1402. 156642

Knight EL; Curhan GC (2003). Albuminuria: moving beyond traditional microalbuminuria cut-points. Current Opinion 

Nephrol Hypertens, 12: 283-284. 179900

Knöbel HH; Chen CJ; Liang KY (1995). Sudden infant death syndrome in relation to weather and optimetrically measured 

air pollution in Taiwan. Pediatrics, 96: 1106-1110. 155905

Krewski D (2009). Evaluating the effects of ambient air pollution on life expectancy. N Engl J Med, 360: 413-415. 190075

Krewski D; Burnett RT; Goldberg MS; Hoover K; Siemiatycki J; Jerrett M; Abrahamowicz M; White WH (2000). 

Reanalysis of the Harvard Six Cities study and the American Cancer Society study of particulate air pollution and 

mortality. Health Effects Institute. Cambridge, MA.http://pubs.healtheffects.org/view.php?id=6. 012281

December 2009 7-106 
Krewski D; Jerrett M; Burnett RT; Ma R; Hughes E; Shi Y; Turner MC; Pope AC III; Thurston G; Calle EE; Thun MJ 

(2009). Extended follow-up and spatial analysis of the American Cancer Society study linking particulate air 

pollution and mortality. Health Effects Institute. Cambridge, MA. Report Nr. 140. 191193

Kulkarni N; Pierse N; Rushton L; Grigg J (2006). Carbon in airway macrophages and lung function in children. N Engl J 

Med, 355: 21-30. 089257

Kundu JK; Surh YJ (2008). Inflammation: Gearing the journey to cancer. Mutat Res, 659: 15-30. 198840

Kunzli N; Bridevaux P-O; Liu S; Garcia-Esteban R; Schindler C; Gerbase M; Sunyer J; keidel D; Rochat T (2009). Trafficrelated air pollution correlates with adult-onset asthma among never-smokers. Thorax, 64: 664-670. 191949

Kunzli N; Jerrett M; Mack WJ; Beckerman B; LaBree L; Gilliland F; Thomas D; Peters J; Hodis HN (2005). Ambient air 

pollution and atherosclerosis in Los Angeles. Environ Health Perspect, 113: 201-206. 087387

Kunzli N; Perez L; Lurmann F; Hricko A; Penfold B; McConnell R (2008). An Attributable Risk Model for Exposures 

Assumed to Cause Both Chronic Disease and its Exacerbations. Epidemiology, 19: 179-185. 129258

Lacasaña M; Esplugues A; Ballester F (2005). Exposure to ambient air pollution and prenatal and early childhood health 

effects. Eur J Epidemiol, 20: 183-199. 155914

Laden F; Schwartz J; Speizer FE; Dockery DW (2006). Reduction in fine particulate air pollution and mortality: extended 

follow-up of the Harvard Six Cities study. Am J Respir Crit Care Med, 173: 667-672. 087605

Lee BE; Ha EH; Park HS; Kim YJ; Hong YC; Kim H; Lee JT (2003). Exposure to air pollution during different gestational 

phases contributes to risks of low birth weight. Hum Reprod, 18: 638-643. 043202

Lee JS; Kim KI; Baek SH (2008). Nuclear receptors and coregulators in inflammation and cancer . Cancer Lett, 267: 189-

196. 198811

Leem J-H; Kaplan BM; Shim YK; Pohl HR; Gotway CA; Bullard SM; Rogers JF; Smith MM; Tylenda CA (2006). 

Exposures to air pollutants during pregnancy and preterm delivery. Environ Health Perspect, 114: 905-910. 089828

Lemos M; Mohallen S; Macchione M; Dolhnikoff M; Assun■ao J; Godleski J; Saldiva P (2006). Chronic exposure to urban 

air Pollution induces structural alterations in murine pulmonary and coronary arteries. Inhal Toxicol, 18: 247-253. 

088594

Li Y-J; Kawada T; Matsumoto A; Azuma A; Kudoh S; Takizawa H; Sugawara I (2007). Airway inflammatory responses to 

oxidative stress induced by low-dose diesel exhaust particle exposure differ between mouse strains. Exp Lung Res, 

33: 227-244. 155929

Lichtenfels AJFC; Gomes JB; Pieri PC; Miraglia SGEK; Hallak J; Saldiva PHN (2007). Increased levels of air pollution 

and a decrease in the human and mouse male-to-female ration in Sao Paulo, Brazil. Fertil Steril, 87: 230-232. 

097041

Lin C-M; Li C-Y; Mao I-F (2004). Increased risks of term low-birth-weight infants in a petrochemical industrial city with 

high air pollution levels. Arch Environ Occup Health, 59: 663-668. 089827

Lin CA; Pereira LAA; Nishioka DC; Conceicao GMS; Graga ALF; Saldiva PHN (2004). Air pollution and neonatal deaths 

in Sao Paulo, Brazil. Braz J Med Biol Res, 37: 765-770. 095787

Lipfert FW; Baty JD; Miller JP; Wyzga RE (2006). PM2.5 constituents and related air quality variables as predictors of 

survival in a cohort of U.S. military veterans. Inhal Toxicol, 18: 645-657. 088756

Lipfert FW; Wyzga RE; Baty JD; Miller JP (2006). Traffic density as a surrogate measure of environmental exposures in 

studies of air pollution health effects: long-term mortality in a cohort of US veterans. Atmos Environ, 40: 154-169. 

088218

Lipfert FW; Zhang J; Wyzga RE (2000). Infant mortality and air pollution: a comprehensive analysis of US data for 1990. J 

Air Waste Manag Assoc, 50: 1350-1366. 004103

Lippmann M; Gordon T; Chen LC (2005). Effects of subchronic exposures to concentrated ambient particles in mice. IX. 

Integral assessment and human health implications of subchronic exposures of mice to CAPs. Inhal Toxicol, 17: 

255-261. 087452

Liu L-JS; Curjuric I; Keidel D; Heldstab J; Kunzli N; Bayer-Oglesby L; Ackermann-Liebrich U; Schindler C; SAPALDIA 

team (2007). Characterization of source-specific air pollution exposure for a large population-based Swiss cohort 

(SAPALDIA). Environ Health Perspect, 115: 1638-1645. 093093

December 2009 7-107 
Liu S; Krewski D; Shi Y; Chen Y; Burnett R (2007). Association between maternal exposure to ambient air pollutants 

during pregnancy and fetal growth restriction. J Expo Sci Environ Epidemiol, 17: 426-432. 090429

Liu Y-Q; Keane M; Ensell M; Miller W; Kashon M; Ong T-m; Mauderly J; Lawson D; Gautam M; Zielinska B; Whitney 

K; Eberhardt J; Wallace W (2005). In vitro genotoxicity of exhaust emissions of diesel and gasoline engine vehicles 

operated on a unified driving cycle. J Environ Monit, 7: 60-66. 097019

Loomis D; Castillejos M; Gold DR; McDonnell W; Borja-Aburto VH (1999). Air pollution and infant mortality in Mexico 

City. Epidemiology, 10: 118-123. 087288

Lopes FD; Pinto TS; Arantes-Costa FM; Moriya HT; Biselli PJ; Ferraz LF; Lichtenfels AJ; Saldiva PH; Mauad T; Martins 

MA (2009). Exposure to ambient levels of particles emitted by traffic worsens emphysema in mice. Environ Res, 

109: 544-551. 190430

Lund AK; Knuckles TL; Obot Akata C; Shohet R; McDonald JD; Gigliotti A; Seagrave JC; Campen MJ (2007). Gasoline 

exhaust emissions induce vascular remodeling pathways involved in atherosclerosis. Toxicol Sci, 95: 485-94. 

125741

Maciejczyk P; Chen LC (2005). Effects of subchronic exposures to concentrated ambient particles (CAPs) in mice: VIII 

source-related daily variations in in vitro responses to CAPs. Inhal Toxicol, 17: 243-253. 087456

Mackman N; Tilley RE; Key NS (2007). Role of the extrinsic pathway of blood coagulation in hemostasis and thrombosis. 

Arterioscler Thromb Vasc Biol, 27: 1687-1693. 156723

Maheswaran R; Haining RP; Brindley P; Law J; Pearson T; Fryers PR; Wise S; Campbell MJ (2005). Outdoor air pollution 

and stroke in Sheffield, United Kingdom: A small-area level geographical study. Stroke, 36: 239-243. 088683

Maheswaran R; Haining RP; Brindley P; Law J; Pearson T; Fryers PR; Wise S; Campbell MJ (2005). Outdoor air pollution, 

mortality, and hospital admissions from coronary heart disease in Sheffield, UK: A small-area level ecological 

study. Eur Heart J, 26: 2543-2549. 090769

Maisonet M; Bush TJ; Correa A; Jaakkola JJK (2001). Relation between ambient air pollution and low birth weight in the 

northeastern United States. Environ Health Perspect, 1093: 351-356. 016624

Maisonet M; Correa A; Misra D; Jaakkola JJ (2004). A review of the literature on the effects of ambient air pollution on 

fetal growth. Environ Res, 95: 106-115. 156725

Mannes T; Jalaludin B; Morgan G; Lincoln D; Sheppeard V; Corbett S (2005). Impact of ambient air pollution on birth 

weight in Sydney, Australia. Occup Environ Med, 62: 524-530. 087895

Matsumoto A; Hiramatsu K; Li Y; Azuma A; Kudoh S; Takizawa H; Sugawara I (2006). Repeated exposure to low-dose 

diesel exhaust after allergen challenge exaggerates asthmatic responses in mice. Clin Immunol, 121: 227-235. 

098017

Mauad T; Rivero DH; de Oliveira RC; Lichtenfels AJ; Guimaraes ET; de Andre PA; Kasahara DI; Bueno HM; Saldiva PH 

(2008). Chronic exposure to ambient levels of urban particles affects mouse lung development. Am J Respir Crit 

Care Med, 178: 721-728. 156743

McConnell R; Berhane K; Gilliland F; London SJ; Vora H; Avol E; Gauderman WJ; Margolis HG; Lurmann F; Thomas 

DC; Peters JM (1999). Air pollution and bronchitic symptoms in southern California children with asthma. Environ 

Health Perspect, 107: 757-760. 007028

McConnell R; Berhane K; Gilliland F; Molitor J; Thomas D; Lurmann F; Avol E; Gauderman WJ; Peters JM (2003). 

Prospective study of air pollution and bronchitic symptoms in children with asthma. Am J Respir Crit Care Med, 

168: 790-797. 049490

McDonald JD; Reed MD; Campen MJ; Barrett EG; Seagrave J; Mauderly JL (2007). Health effects of inhaled gasoline 

engine emissions. Inhal Toxicol, 19 Suppl 1: 107-116. 156746

McDonnell WF; Nishino-Ishikawa N; Petersen FF; Chen LH; Abbey DE (2000). Relationships of mortality with the fine 

and coarse fractions of long-term ambient PM10 concentrations in nonsmokers. J Expo Sci Environ Epidemiol, 10: 

427-436. 010319

Medeiros A; Gouveia N (2005). Relacao entre baixo peso ao nascer e a poluicao do ar no Municipio de Sao Paulo 

[Relationship between low birthweight and air pollution in the city of Sao Paulo, Brazil]. Rev Saude Publica, 39: 

965-972. 089824

December 2009 7-108 
Ménache MG; Miller FJ; Raabe OG (1995). Particle inhalability curves for humans and small laboratory animals. Ann 

Occup Hyg, 39: 317-328. 006533

Meng YY; Wilhelm M; Rull RP; English P; Ritz B (2007). Traffic and outdoor air pollution levels near residences and 

poorly controlled asthma in adults. Ann Allergy Asthma Immunol, 98: 455-463. 093275

Miller KA; Siscovick DS; Sheppard L; Shepherd K; Sullivan JH; Anderson GL; Kaufman JD (2007). Long-term exposure 

to air pollution and incidence of cardiovascular events in women. N Engl J Med, 356: 447-458. 090130

Mogensen CE (1984). Microalbuminuria predicts clinical proteinuria and early mortality in maturity-onset diabetes. N Engl 

J Med, 310: 356-360. 156769

Mohallem SV; de Araujo Lobo DJ; Pesquero CR; Assuncao JV; de Andre PA; Saldiva PH; Dolhnikoff M (2005). Decreased 

fertility in mice exposed to environmental air pollution in the city of Sao Paulo. Environ Res, 98: 196-202. 088657

Mollet NR; Cademartiri F; de Feyter PJ (2005). Non-invasive multislice CT coronary imaging. Heart, 91: 401-407. 155988

Montauban van Swijndregt AD; De Lange EE; De Groot E; Ackerstaff RG (1999). An in vivo evaluation of the 

reproducibility of intima-media thickness measurements of the carotid artery segments using B-mode ultrasound. 

Ultrasound Med Biol, 25: 323-330. 156777

Morgenstern V; Zutavern A; Cyrys J; Brockow I; Koletzko S; Kramer U; Behrendt H; Herbarth O; von Berg A; Bauer CP; 

Wichmann HE; Heinrich J (2008). Atopic diseases, allergic sensitization, and exposure to traffic-related air 

pollution in children. Am J Respir Crit Care Med, 177: 1331-1337. 156782

Mortimer K; Neugebauer R; Lurmann F; Alcorn S; Balmes J; Tager I (2008). Early-Lifetime exposure to air pollution and 

allergic sensitization in children with asthma. J Asthma, 45: 874-881. 187280

Naess O; Nafstad P; Aamodt G; Claussen B; Rosland P (2007). Relation between concentration of air pollution and causespecific mortality: four-year exposures to nitrogen dioxide and particulate matter pollutants in 470 neighborhoods 

in Oslo, Norway. Am J Epidemiol, 165: 435-443. 090736

National Kidney Foundation (2008). Chronic Kidney Disease: Evaluation, Classification, and Stratification. The National 

Kidney Foundation . Kidney Disease Outcomes Quality Initiative. 156796

Newman AB; Sutton-Tyrrell K; Vogt MT; Kuller LH (1993). Morbidity and mortality in hypertensive adults with a low 

ankle/arm blood pressure index. JAMA, 270: 487-489. 156805

Nordling E; Berglind N; Melen E; Emenius G; Hallberg J; Nyberg F; Pershagen G; Svartengren M; Wickman M; Bellander 

T (2008). Traffic-related air pollution and childhood respiratory symptoms, function and allergies. Epidemiology, 

19: 401-8. 097998

O'Leary DH; Polak JF; Kronmal RA; Kittner SJ; Bond MG; Wolfson SK Jr; Bommer W; Price TR; Gardin JM; Savage PJ 

(1992). Distribution and correlates of sonographically detected carotid artery disease in the Cardiovascular Health 

Study. The CHS Collaborative Research Group. Stroke, 23: 1752-1760. 156825

O'Leary DH; Polak JF; Kronmal RA; Manolio TA; Burke GL; Wolfson SK Jr (1999). Carotid-artery intima and media 

thickness as a risk factor for myocardial infarction and stroke in older adults. Cardiovascular Health Study 

Collaborative Research Group. N Engl J Med, 340: 14-22. 156826

O'Neill MS; Diez-Roux AV; Auchincloss AH; Franklin TG; Jacobs Jnr DR; Astor BC; Dvonch JT; Kaufman J (2007). 

Airborne particulate matter exposure and urinary albumin excretion: The Multi-Ethnic Study of Atherosclerosis. 

Occup Environ Med, 65: 534-540. 156006

O'Rourke RA; Brundage BH; Froelicher VF; Greenland P; Grundy SM; Hachamovitch R; Pohost GM; Shaw LJ; Weintraub 

WS; Winters WL; Forrester JS; Douglas PS; Faxon DP; Fisher JD; Gregoratos G; Hochman JS; Hutter AM; Kaul S; 

Wolk MJ (2000). American College of Cardiology/American Heart Association Expert Consensus document on 

electron-beam computed tomography for the diagnosis and prognosis of coronary artery disease. Circulation, 102: 

126-140. 192159

Oei HH; Vliegenthart R; Hak AE; Iglesias del Sol A; Hofman A; Oudkerk M; Witteman JC (2002). The association 

between coronary calcification assessed by electron beam computed tomography and measures of extracoronary 

atherosclerosis: the Rotterdam Coronary Calcification Study. J Am Coll Cardiol, 39: 1745-1751. 156820

Oftedal B; Brunekreef B; Nystad W; Madsen C; Walker S-E; Nafstad P (2008). Residential outdoor air pollution and lung 

function in schoolchildren. Epidemiology, 19: 129-137. 093202

December 2009 7-109 
Oftedal B; Brunekreef B; Nystad W; Nafstad P (2007). Residential outdoor air pollution and allergen sensitization in 

schoolchildren in Oslo, Norway. Clin Exp Allergy, 37: 1632. 191948

Oh S-M; Chung K-H (2006). Identification of mammalian cell genotoxins in respirable diesel exhaust particles by 

bioassay-directed chemical analysis. Toxicol Lett, 161: 226-235. 088296

Ono N; Oshio S; Niwata Y; Yoshida S; Tsukue N; Sugawara I; Takano H; Takeda K (2007). Prenatal exposure to diesel 

exhaust impairs mouse spermatogenesis. Inhal Toxicol, 19: 275-281. 156007

Ozkaynak H; Thurston GD (1987). Associations between 1980 US mortality rates and alternative measures of airborne 

particle concentration. Risk Anal, 7: 449-461. 072960

Palli D; Saieva C; Munnia A; Peluso M; Grechi D; Zanna I; Caini S; Decarli A; Sera F; Masala G (2008). DNA adducts 

and PM10 exposure in traffic-exposed workers and urban residents from the EPIC-Florence City study. Sci Total 

Environ, 403: 105-112. 156837

Parker JD; Mendola P; Woodruff TJ (2008). Preterm birth after the Utah Valley Steel Mill closure: a natural experiment. 

Epidemiology, 19: 820-823. 156013

Parker JD; Woodruff TJ (2008). Influences of study design and location on the relationship between particulate matter air 

pollution and birthweight. Paediatr Perinat Epidemiol, 22: 214-227. 156846

Parker JD; Woodruff TJ; Basu R; Schoendorf KC (2005). Air pollution and birth weight among term infants in California. 

Pediatrics, 115: 121-128. 087462

Pedersen M; Vinzents P; Petersen JH; Kleinjans JC; Plas G; Kirsch-Volders M; Dostal M; Rossner P; Beskid O; Sram RJ; 

Merlo DF; Knudsen LE (2006). Cytogenetic effects in children and mothers exposed to air pollution assessed by the 

frequency of micronuclei and fluorescence in situ hybridization (FISH): a family pilot study in the Czech Republic. 

Mutat Res Fund Mol Mech Mutagen, 608: 112-120. 156848

Penard-Morand C; Charpin D; Raherison C; Kopferschmitt C; Caillaud D; Lavaud F; Annesi-Maesano I (2005). Long-term 

exposure to background air pollution related to respiratory and allergic health in schoolchildren. Clin Exp Allergy, 

35: 1279-1287. 087951

Penna MLF; Duchiade MP (1991). Air pollution and infant mortality from pneumonia in the Rio de Janeiro metropolitan 

area. Bull Pan Am Health Organ, 25: 47-54. 073325

Pereira LAA; Loomis D; Conceicao GMS; Braga ALF; Arcas RM; Kishi HS; Singer JM; Bohm GM; Saldiva PHN (1998). 

Association between air pollution and intrauterine mortality in Sao Paulo, Brazil. Environ Health Perspect, 106: 

325-329. 007264

Peters A; Skorkovsky J; Kotesovec F; Brynda J; Spix C; Wichmann HE; Heinrich J (2000). Associations between mortality 

and air pollution in central Europe. Environ Health Perspect, 108: 283-287. 001756

Pierse N; Rushton L; Harris RS; Kuehni CE; Silverman M; Grigg J (2006). Locally-generated particulate pollution and 

respiratory symptoms in young children. Thorax, 61: 216-220. 088757

Pignoli P; Tremoli E; Poli A; Oreste P; Paoletti R (1986). Intimal plus medial thickness of the arterial wall: a direct 

measurement with ultrasound imaging. Circulation, 74: 1399-1406. 156026

Pinkerton KE; Zhou Y; Teague SV; Peake JL; Walther RC; Kennedy IM; Leppert VJ; Aust AE (2004). Reduced lung cell 

proliferation following short-term exposure to ultrafine soot and iron particles in neonatal rats: key to impaired lung 

growth? Inhal Toxicol, 1: 73-81. 087465

Pires-Neto RC; Lichtenfels AJ; Soares SR; Macchione M; Saldiva PHN; Dolhnikoff M (2006). Effects of Sao Paulo air 

pollution on the upper airways of mice. Environ Res, 101: 356-361. 096734

Poma A; Limongi T; Pisani C; Granato V; Picozzi P (2006). Genotoxicity induced by fine urban air particulate matter in the 

macrophages cell line RAW 264.7. Toxicol In Vitro, 20: 1023-1029. 096903

Pope CA III (1989). Respiratory disease associated with community air pollution and a steel mill, Utah Valley. Am J Public 

Health, 79: 623-628. 044461

Pope CA III; Burnett RT (2007). Confounding in air pollution epidemiology: the broader context. Epidemiology, 18: 424-

426. 090928

Pope CA III; Burnett RT; Thun MJ; Calle EE; Krewski D; Ito K; Thurston GD (2002). Lung cancer, cardiopulmonary 

mortality, and long-term exposure to fine particulate air pollution. JAMA, 287: 1132-1141. 024689

December 2009 7-110 
Pope CA III; Thun MJ; Namboodiri MM; Dockery DW; Evans JS; Speizer FE; Heath CW Jr (1995). Particulate air 

pollution as a predictor of mortality in a prospective study of US adults. Am J Respir Crit Care Med, 151: 669-674. 

045159

Pope III CA; Burnett RT; Thurston GD; Thun MJ; Calle EE; Krewski D; Godleski JJ (2004). Cardiovascular mortality and 

long-term exposure to particulate air pollution: epidemiological evidence of general pathophysiological pathways of 

disease. Circulation, 109: 71-77. 055880

Pope III CA; Ezzati M; Dockery DW (2009). Fine-particulate air pollution and life expectancy in the United States. N Engl 

J Med, 360: 376-386. 190107

Puett RC; Schwartz J; Hart JE; Yanosky JD; Speizer FE; Suh H; Paciorek CJ; Neas LM; Laden F (2008). Chronic 

particulate exposure, mortality, and coronary heart disease in the nurses' health study. Am J Epidemiol, 168: 1161-

1168. 156891

Raizenne M; Neas LM; Damokosh AI; Dockery DW; Spengler JD; Koutrakis P; Ware JH; Speizer FE (1996). Health 

effects of acid aerosols on North American children: pulmonary function. Environ Health Perspect, 104: 506-514. 

077268

Ramos C; Cisneros J; Gonzalez-Avila G; Becerril C; Ruiz V; Montaño M (2009). Increase of matrix metalloproteinases in 

woodsmoke-induced lung emphysema in guinea pigs. Inhal Toxicol, 21: 119-132. 190116

Reed MD; Barrett EG; Campen MJ; Divine KK; Gigliotti AP; McDonald JD; Seagrave JC; Mauderly JL; Seilkop SK; 

Swenberg JA (2008). Health effects of subchronic inhalation exposure to gasoline engine exhaust. Inhal Toxicol, 

20: 1125-1143. 156903

Reed MD; Campen MJ; Gigliotti AP; Harrod KS; McDonald JD; Seagrave JC; Mauderly JL; Seilkop SK (2006). Health 

effects of subchronic exposure to environmental levels of hardwood smoke. Inhal Toxicol, 18: 523-539. 156043

Reed MD; Gigliotti AP; McDonald JD; Seagrave JC; Seilkop SK; Mauderly JL (2004). Health effects of subchronic 

exposure to environmental levels of diesel exhaust. Inhal Toxicol, 16: 177-193. 055625

Resnick HE; Lindsay RS; McDermott MM; Devereux RB; Jones KL; Fabsitz RR; Howard BV (2004). Relationship of 

high and low ankle brachial index to all-cause and cardiovascular disease mortality: the Strong Heart Study. 

Circulation, 109: 733-739. 156048

Reymao MSF; Cury PM; Lichtenfels AJFC; Lemos M; Battlehner CN; Conceicao GMS; Capelozzi VL; Montes GS; Junior 

MF; Martins MA; Bohm GM; Saldiva PHN (1997). Urban air pollution enhances the formation of urethaneinduced lung tumors in mice. Environ Res, 74: 150-158. 084653

Rich DQ; Demissie K; Lu SE; Kamat L; Wartenberg D; Rhoads GG (2009). Ambient air pollutant concentrations during 

pregnancy and the risk of fetal growth restriction. J Epidemiol Community Health, 63: 488-496. 180122

Ritz B; Wilhelm M (2008). Ambient air pollution and adverse birth outcomes: methodologic issues in an emerging field. 

Basic Appl Ecol, 102: 182-190. 156914

Ritz B; Wilhelm M; Hoggatt KJ; Ghosh JK (2007). Ambient air pollution and preterm birth in the environment and 

pregnancy outcomes study at the University of California, Los Angeles. Am J Epidemiol, 166: 1045-1052. 096146

Ritz B; Wilhelm M; Zhao Y (2006). Air pollution and infant death in southern California, 1989-2000. Pediatrics, 118: 493-

502. 089819

Ritz B; Yu F; Chapa G; Fruin S (2000). Effect of air pollution on preterm birth among children born in Southern California 

between 1989 and 1993. Epidemiology, 11: 502-511. 012068

Ritz B; Yu F; Fruin S; Chapa G; Shaw GM; Harris JA (2002). Ambient air pollution and risk of birth defects in Southern 

California. Am J Epidemiol, 155: 17-25. 023227

Rocha ESIR; Lichtenfels AJ; Amador Pereira LA; Saldiva PH (2008). Effects of ambient levels of air pollution generated 

by traffic on birth and placental weights in mice. Fertil Steril, 90: 1921-1924. 096685

Rogers JF; Dunlop AL (2006). Air pollution and very low birth weight infants: a target population? Pediatrics, 118: 156-

164. 091232

Rojas-Martinez R; Perez-Padilla R; Olaiz-Fernandez G; Mendoza-Alvarado L; Moreno-Macias H; Fortoul T; McDonnell 

W; Loomis D; Romieu I (2007). Lung function growth in children with long-term exposure to air pollutants in 

Mexico City. Am J Respir Crit Care Med, 176: 377-384. 091064

December 2009 7-111 
Roman HA; Walker KD; Walsh TL; Conner L; Richmond HM; Hubbell BJ; Kinney PL (2008). Expert judgment 

assessment of the mortality impact of changes in ambient fine particulate matter in the U.S. Environ Sci Technol, 

42: 2268-2274. 156921

Romieu I; Garcia-Esteban R; Sunyer J; Rios C; Alcaraz-Zubeldia M; Velasco SR; Holguin F (2008). The effect of 

supplementation with omega-3 polyunsaturated fatty acids on markers of oxidative stress in elderly exposed to 

PM(2.5). Environ Health Perspect, 116: 1237-1242. 156922

Romieu I; Ramirez-Aguilar M; Moreno-Macias H; Barraza-Villarreal A; Miller P; Hernandez-Cadena L; Carbajal-Arroyo 

LA; Hernandez-Avila M (2004). Infant mortality and air pollution: modifying effect by social class. J Occup 

Environ Hyg, 46: 1210-1216. 093074

Röösli M; Braun-Fahrlander C; Kunzli N; Oglesby L; Theis G; Camenzind M; Mathys P; Staehelin J (2000). Spatial 

variability of different fractions of particulate matter within an urban environment and between urban and rural 

sites. J Air Waste Manag Assoc, 50: 1115-1124. 010296

Röösli M; Kunzli N; Braun-Fahrlander C; Egger M (2005). Years of life lost attributable to air pollution in Switzerland: 

dynamic exposure-response model. Int J Epidemiol, 34: 1029-1035. 156923

Röösli M; Theis G; Kunzli N; Staehelin J; Mathys P; Oglesby L; Camenzind M; Braun-Fahrlander C (2001). Temporal and 

spatial variation of the chemical composition of PM10 at urban and rural sites in the Basel area, Switzerland. Atmos 

Environ, 35: 3701-3713. 108738

Rosenlund M; Bellander T; Nordquist T; Alfredsson L (2009). Traffic-generated air pollution and myocardial infarction. 

Epidemiology, 20: 265-71. 190309

Rosenlund M; Berglind N; Pershagen G; Hallqvist J; Jonson T; Bellander T (2006). Long-term exposure to urban air 

pollution and myocardial infarction. Epidemiology, 17: 383-390. 089796

Ross R (1999). Atherosclerosis--an inflammatory disease. N Engl J Med, 340: 115-126. 156926

Rubes J; Selevan SG; Evenson DP; Zudova D; Vozdova M; Zudova Z; Robbins WA; Perreault SD (2005). Episodic air 

pollution is associated with increased DNA fragmentation in human sperm without other changes in semen quality. 

Hum Reprod, 20: 2776-2783. 078091

Ruggenenti P; Remuzzi G (2006). Time to abandon microalbuminuria? Kidney Int, 70: 1214-1222. 156933

Sagiv SK; Mendola P; Loomis D; Herring AH; Neas LM; Savitz DA; Poole C (2005). A time-series analysis of air 

pollution and preterm birth in Pennsylvania, 1997-2001. Environ Health Perspect, 113: 602-606. 087468

Salam MT; Millstein J; Li Y-F; Lurmann FW; Margolis HG; Gilliland FD (2005). Birth outcomes and prenatal exposure to 

ozone, carbon monoxide, and particulate matter: results from the Children's Health Study. Environ Health Perspect, 

113: 1638-1644. 087885

Salonen JT; Salonen R (1991). Ultrasonographically assessed carotid morphology and the risk of coronary heart disease. 

Arterioscler Thromb Vasc Biol, 11: 1245-1249. 156938

Sato H; Suzuki Kazuo T; Sone H; Yamano Y; Kagawa J; Aoki Y (2003). DNA-adduct formation in lungs, nasal mucosa, 

and livers of rats exposed to urban roadside air in Kawasaki City, Japan. Environ Res, 93: 36-44. 096615

Schatz M; Zeiger RS; Hoffman CP (1990). Intrauterine growth is related to gestational pulmonary function in pregnant 

asthmatic women. Kaiser-Permanente Asthma and Pregnancy Study Group. Chest, 98: 389-392. 156073

Schikowski T; Sugiri D; Ranft U; Gehring U; Heinrich J; Wichmann HE; Kramer U (2005). Long-term air pollution 

exposure and living close to busy roads are associated with COPD in women. Respir Res, 22: 152-161. 088637

Schindler C; Keidel D; Gerbase MW; Zemp E; Bettschart R; Brandli O; Brutsche MH; Burdet L; Karrer W; Knopfli B; 

Pons M; Rapp R; Bayer-Oglesby L; Kunzli N; Schwartz J; Liu L-JS; Ackermann-Liebrich U; Rochat T; the 

SAPALDIA Team (2009). Improvements in PM10 exposure and reduced rates of respiratory symptoms in a cohort 

of swiss adults (SAPALDIA). Am J Respir Crit Care Med, 179: 579-587. 191950

Schwartz J; Coull B; Laden F; Ryan L (2008). The effect of dose and timing of dose on the association between airborne 

particles and survival. Environ Health Perspect, 116: 64-69. 156963

Seagrave J; McDonald JD; Reed MD; Seilkop SK; Mauderly JL (2005). Responses to subchronic inhalation of low 

concentrations of diesel exhaust and hardwood smoke measured in rat bronchoalveolar lavage fluid. Inhal Toxicol, 

17: 657-670. 088000

December 2009 7-112 
Selevan SG; Borkovec L; Slott VL; Zudova Z; Rubes J; Evenson DP; Perreault SD (2000). Semen quality and reproductive 

health of young Czech men exposed to seasonal air pollution. Environ Health Perspect, 108: 887-894. 012578

Sevastyanova O; Binkova B; Topinka J; Sram RJ; Kalina I; Popov T; Novakova Z; Farmer PB (2007). In vitro genotoxicity 

of PAH mixtures and organic extract from urban air particles part II: human cell lines. Mutat Res Fund Mol Mech 

Mutagen, 620: 123-134. 156969

Sharma AK; Jensen KA; Rank J; White PA; Lundstedt S; Gagne R; Jacobsen NR; Kristiansen J; Vogel U; Wallin H (2007). 

Genotoxicity, inflammation and physico-chemical properties of fine particle samples from an incineration energy 

plant and urban air. Mutat Res Fund Mol Mech Mutagen, 633: 95-111. 156975

Shaw LJ; Raggi P; Schisterman E; Berman DS; Callister TQ (2003). Prognostic value of cardiac risk factors and coronary 

artery calcium screening for all-cause mortality. Radiology, 228: 826-833. 156083

Shemesh J; Tenenbaum A; Fisman EZ; Apter S; Rath S; Rozenman J; Itzchak Y; Motro M (1996). Absence of coronary 

calcification on double-helical CT scans: predictor of angiographically normal coronary arteries in elderly women? 

Radiology, 199: 665-668. 156085

Shinkura R; Fujiyama C; Akiba S (1999). Relationship between ambient sulfur dioxide levels and neonatal mortality near 

the Mt Sakurajima volcano in Japan. J Epidemiol, 9: 344-349. 090050

Slama R; Darrow L; Parker J; Woodruff TJ; Strickland M; Nieuwenhuijsen M; Glinianaia S; Hoggatt KJ; Kannan S; Hurley 

F; Kalinka J; Sram R; Brauer M; Wilhelm M; Heinrich J; Ritz B (2008). Meeting report: atmospheric pollution and 

human reproduction. Environ Health Perspect, 116: 791-798. 156985

Slama R; Morgenstern V; Cyrys J; Zutavern A; Herbarth O; Wichmann HE; Heinrich J; LISA Study Group (2007). Trafficrelated atmospheric pollutants levels during pregnancy and offspring's term birth weight: a study relying on a landuse regression exposure model. Environ Health Perspect, 115: 1283-1292. 093216

Smilde TJ; Wollersheim H; Van Langen H; Stalenhoef AF (1997). Reproducibility of ultrasonographic measurements of 

different carotid and femoral artery segments in healthy subjects and in patients with increased intima-media 

thickness. Clin Sci (Lond), 93: 317-324. 156988

Sokol RZ; Kraft P; Fowler IM; Mamet R; Kim E; Berhane KT (2006). Exposure to environmental ozone alters semen 

quality. Environ Health Perspect, 114: 360-5. 098539

Solomon P; Baumann K; Edgerton E; Tanner R; Eatough D; Modey W; Marin H; Savoie D; Natarajan S; Meyer MB 

(2003). Comparison of integrated samplers for mass and composition during the 1999 Atlanta supersites project. J 

Geophys Res, 108: 8423. 156994

Somers CM; McCarry BE; Malek F; Quinn JS (2004). Reduction of particulate air pollution lowers the risk of heritable 

mutations in mice. Science, 304: 1008-1010. 078098

Somers CM; Yauk CL; White PA; Parfett CLJ; Quinn JS (2002). Air pollution induces heritable DNA mutations. PNAS, 

99: 15904-15907. 078100

Song CL; Zhou YC; Huang RJ; Wang YQ; Huang QF; Lu G; Liu KM (2007). Influence of ethanol-diesel blended fuels on 

diesel exhaust emissions and mutagenic and genotoxic activities of particulate extracts. J Hazard Mater, 149: 355-

363. 155306

Sorensen M; Schins RPF; Hertel O; Loft S (2005). Transition metals in personal samples of PM25 and oxidative stress in 

human volunteers. Cancer Epidemiol Biomarkers Prev, 14: 1340-1343. 083053

Sram R; Beskid O; Binkova B; Chvatalova I; Lnenickova Z; Milcova A; Solansky I; Tulupova E; Bavorova H; Ocadlikova 

D (2007). Chromosomal aberrations in environmentally exposed population in relation to metabolic and DNA 

repair genes polymorphisms. Mutat Res Fund Mol Mech Mutagen, 620: 22-33. 188457

Sram RJ; Beskid O; Rössnerova A; Rössner P; Lnenickova Z; Milcova A; Solansky I; Binkova B (2007). Environmental 

exposure to carcinogenic polycyclic aromatic hydrocarbons--the interpretation of cytogenetic analysis by FISH. 

Toxicol Lett, 172: 12-20. 192084

Sram RJ; Binkova B; Dejmek J; Bobak M (2005). Ambient air pollution and pregnancy outcomes: a review of the 

literature. Environ Health Perspect, 113: 375-382. 087442

Stanojevic S; Wade A; Stocks J; Hankinson J; Coates AL; Pan H; Rosenthal M; Corey M; Lebecque P; Cole TJ (2008). 

Reference ranges for spirometry across all ages: a new approach. Am J Respir Crit Care Med, 177: 253-260. 

157007

December 2009 7-113 
Stocker R; Keaney Jr JF (2004). Role of oxidative modifications in atherosclerosis. Physiol Rev, 84: 1381-1478. 157013

Strom KA; Garg BD; Johnson JT; D'Arcy JB; Smiler KL (1990). Inhaled particle retention in rats receiving low exposures 

of diesel exhaust. J Toxicol Environ Health, 29: 377-398. 157020

Sugamata M; Ihara T; Takano H; Oshio S; Takeda K (2006). Maternal diesel exhaust exposure damages newborn murine 

brains. Eisei Kagaku, 52: 82-84. 097166

Sugiri D; Ranft U; Schikowski T; Kramer U (2006). The influence of large-scale airborne particle decline and trafficrelated exposure on children's lung function. Environ Health Perspect, 114: 282-288. 088760

Suh YJ; Ha EH; Park H; Kim YJ; Kim H; Hong YC (2008). GSTM1 polymorphism along with PM10 exposure contributes 

to the risk of preterm delivery. Mutat Res, 656: 62-67. 192077

Sun Q; Wang A; Jin X; Natanzon A; Duquaine D; Brook RD; Aguinaldo JG; Fayad ZA; Fuster V; Lippmann M; Chen LC; 

Rajagopalan S (2005). Long-term air pollution exposure and acceleration of atherosclerosis and vascular 

inflammation in an animal model. JAMA, 294: 3003-3010. 087952

Sun Q; Yue P; Deiuliis JA; Lumeng CN; Kampfrath T; Mikolaj MB; Cai Y; Ostrowski MC; Lu B; Parthasarathy S; Brook 

RD; Moffatt-Bruce SD; Chen LC; Rajagopalan S (2009). Ambient air pollution exaggerates adipose inflammation 

and insulin resistance in a mouse model of diet-induced obesity. Circulation, 119: 538-546. 190487

Sun Q; Yue P; Kirk RI; Wang A; Moatti D; Jin X; Lu B; Schecter AD; Lippmann M; Gordon T; Chen LC; Rajagopalan S 

(2008). Ambient air particulate matter exposure and tissue factor expression in atherosclerosis. Inhal Toxicol, 20: 

127-137. 157033

Sun Q; Yue P; Ying Z; Cardounel AJ; Brook RD; Devlin R; Hwang JS; Zweier JL; Chen LC; Rajagopalan S (2008). Air 

Pollution Exposure Potentiates Hypertension Through Reactive Oxygen Species-Mediated Activation of 

Rho/ROCK. Arterioscler Thromb Vasc Biol, 28: 1760-1766. 157032

Suwa T; Hogg JC; Quinlan KB; Ohgami A; Vincent R; Van Eeden SF (2002). Particulate air pollution induces progression 

of atherosclerosis. J Am Coll Cardiol, 39: 935-942. 028588

Tarantini L; Bonzini M; Apostoli P; Pegoraro V; Bollati V; Marinelli B; Cantone L; Rizzo G; Hou L; Schwartz J; Bertazzi 

PA; Baccarelli A (2009). Effects of particulate matter on genomic DNA methylation content and iNOS promoter 

methylation. Environ Health Perspect, 117: 217-222. 192010

Tesfaigzi Y; McDonald JD; Reed MD; Singh SP; De Sanctis GT; Eynott PR; Hahn FF; Campen MJ; Mauderly JL (2005). 

Low-level subchronic exposure to wood smoke exacerbates inflammatory responses in allergic rats. Toxicol Sci, 

88: 505-513. 156116

Tesfaigzi Y; Singh SP; Foster JE; Kubatko J; Barr EB; Fine PM; McDonald JD; Hahn FF; Mauderly JL (2002). Health 

effects of subchronic exposure to low levels of wood smoke in rats. Toxicol Sci, 65: 115-125. 025575

Thurlbeck WM (1982). Postnatal human lung growth. Thorax, 37: 564-571. 093260

Tokiwa H; Sera N; Nakanishi Y (2005). Involvement of alveolar macrophages in the formation of 8-oxodeoxyguanosine 

associated with exogenous particles in human lungs. Inhal Toxicol, 17: 577-585. 191952

Tong S; Colditz P (2004). Air pollution and sudden infant death syndrome: a literature review. Paediatr Perinat Epidemiol, 

18: 327-335. 087883

Tovalin H; Valverde M; Morandi MT; Blanco S; Whitehead L; Rojas E (2006). DNA damage in outdoor workers 

occupationally exposed to environmental air pollutants. Occup Environ Med, 63: 230-236. 091322

Tozuka Y; Watanabe N; Ohsawa M; Toriba A; Kizu R; Hayakawa K (2004). Transfer of polycyclic aromatic hydrocarbons 

to fetuses and breast milk of rats exposed to diesel exhaust. Eisei Kagaku, 50: 497-502. 090864

Tsai S-S; Chen C-C; Hsieh H-J; Chang C-C; Yang C-Y (2006). Air pollution and postneonatal mortality in a tropical city: 

Kaohsiung, Taiwan. Inhal Toxicol, 18: 185-189. 090709

Tsukue N; Tsubone H; Suzuki AK (2002). Diesel exhaust affects the abnormal delivery in pregnant mice and the growth of 

their young. Inhal Toxicol, 14: 635-651. 030593

Tsukue N; Yoshida S; Sugawara I; Taked K (2004). Effect of diesel exhaust on development of fetal reproductive function 

in ICR female mice. Eisei Kagaku, 50: 174-80. 096643

December 2009 7-114 
U.S. EPA (1996). Air quality criteria for particulate matter. U.S. Environmental Protection Agency. Research Triangle Park, 

NC. EPA/600/P-95/001aF-cF. 079380

U.S. EPA (2002). Health assessment document for diesel engine exhaust. U.S. Environmental Protection Agency. 

Washington, DC. 042866

U.S. EPA (2004). Air quality criteria for particulate matter. U.S. Environmental Protection Agency. Research Triangle Park, 

NC. EPA/600/P-99/002aF-bF. 056905

U.S. EPA (2005). Guidelines for carcinogen risk assessment, Final Report. Risk Assessment Forum, U.S. Environmental 

Protection Agency. Washington, DC. EPA/630/P-03/001F. 

http://cfpub.epa.gov/ncea/cfm/recordisplay.cfm?deid=116283. 086237

U.S. EPA (2006). Provisional Assessment of Recent Studies on Health Effects of Particulate Matter Exposure. U.S. 

Environmental Protection Agency. Research Triangle Park, NC. 157071

van der Meer IM; Bots ML; Hofman A; del Sol AI; van der Kuip DA; Witteman JC (2004). Predictive value of noninvasive 

measures of atherosclerosis for incident myocardial infarction: the Rotterdam Study. Circulation, 109: 1089-1094. 

156129

Van Hee VC; Adar SD; Szpiro AA; Barr RG; Bluemke DA; Diez Roux AV; Gill EA; Sheppard L; Kaufman JD (2009). 

Exposure to traffic and left ventricular mass and function: the Multi-Ethnic Study of Atherosclerosis. Am J Respir 

Crit Care Med, 179: 827-834. 192110

Veras MM; Damaceno-Rodrigues NR; Caldini EG; Maciel Ribeiro AA; Mayhew TM; Saldiva PH; Dolhnikoff M (2008). 

Particulate urban air pollution affects the functional morphology of mouse placenta. Biol Reprod, 79: 578-584. 

190493

Veras MM; Damaceno-Rodrigues NR; Guimarães Silva RM; Scoriza JN; Saldiva PH; Caldini EG; Dolhnikoff M (2009). 

Chronic exposure to fine particulate matter emitted by traffic affects reproductive and fetal outcomes in mice. 

Environ Res, 109: 536-543. 190496

Villeneuve PJ; Goldberg MS; Krewski D; Burnett RT; Chen Y (2002). Fine particulate air pollution and all-cause mortality 

within the Harvard six-cities study: variations in risk by period of exposure. Ann Epidemiol, 12: 568-576. 042576

Vineis P; Hoek G; Krzyzanowski M; Vigna-Taglianti F; Veglia F; Airoldi L; Autrup H; Dunning A; Garte S; Hainaut P; 

Malaveille C; Matullo G; Overvad K; Raaschou-Nielsen O; Clavel-Chapelon F; Linseisen J; Boeing H; 

Trichopoulou A; Palli D; Peluso M; Krogh V; Tumino R; Panico S; Bueno-De-Mesquita HB; Peeters PH; Lund EE; 

Gonzalez CA; Martinez C; Dorronsoro M; Barricarte A; Cirera L; Quiros JR; Berglund G; Forsberg B; Day NE; 

Key TJ; Saracci R; Kaaks R; Riboli E (2006). Air pollution and risk of lung cancer in a prospective study in 

Europe. Int J Cancer, 119: 169-174. 192089

Vinzents PS; Moller P; Sorensen M; Knudsen LE; Herte LQ; Jensen FP; Schibye B; Loft S (2005). Personal exposure to 

ultrafine particles and oxidative DNA damage. Environ Health Perspect, 113: 1485-1490. 087482

Vogt MT; Cauley JA; Newman AB; Kuller LH; Hulley SB (1993). Decreased ankle/arm blood pressure index and mortality 

in elderly women. JAMA, 270: 465-469. 157100

Wallenborn JG; Evansky P; Shannahan JH; Vallanat B; Ledbetter AD; Schladweiler MC; Richards JH; Gottipolu RR; 

Nyska A; Kodavanti UP (2008). Subchronic inhalation of zinc sulfate induces cardiac changes in healthy rats. 

Toxicol Appl Pharmacol, 232: 69-77. 191171

Walsh CR; Cupples LA; Levy D; Kiel DP; Hannan M; Wilson PW; O'Donnell CJ (2002). Abdominal aortic calcific 

deposits are associated with increased risk for congestive heart failure: the Framingham Heart Study. Am Heart J, 

144: 733-739. 157103

Watanabe N (2005). Decreased number of sperms and Sertoli cells in mature rats exposed to diesel exhaust as fetuses. 

Toxicol Lett, 155: 51-58. 087985

Wayne LG; Chambers LA (1968). Biological effects of urban air pollution: V a study of effects of Los Angeles atmosphere 

on laboratory rodents. Arch Environ Occup Health, 16: 871-885. 038537

Weisenberger DJ; Campan M; Long TI; Kim M; Woods C; Fiala E; Ehrlich M; Laird PW (2005). Analysis of repetitive 

element DNA methylation by MethyLight. Nucleic Acids Res, 33: 6823-36. 192101

December 2009 7-115 
Weitz JI; Byrne J; Clagett GP; Farkouh ME; Porter JM; Sackett DL; Strandness DE Jr; Taylor LM (1996). Diagnosis and 

treatment of chronic arterial insufficiency of the lower extremities: a critical review. Circulation, 94: 3026-3049. 

156150

Wendelhag I; Gustavsson T; Suurkula M; Berglund G; Wikstrand J (1991). Ultrasound measurement of wall thickness in 

the carotid artery: fundamental principles and description of a computerized analysing system. Clinical Physiol, 11: 

565-577. 157135

Wendelhag I; Wiklund O; Wikstrand J (1993). Atherosclerotic changes in the femoral and carotid arteries in familial 

hypercholesterolemia. Ultrasonographic assessment of intima-media thickness and plaque occurrence. Arterioscler 

Thromb Vasc Biol, 13: 1404-1411. 157136

Wheeler AJ; Villeneuve P; Smith-Doiron M; Mahmud M; Dales R; Brook JR (2006). Children's exposure to ambient air 

pollution: the application of different exposure assessment methodologies. Epidemiology, 17: S33-S34. 103905

Wilhelm M; Ritz B (2005). Local variations in CO and particulate air pollution and adverse birth outcomes in Los Angeles 

County, California, USA. Environ Health Perspect, 113: 1212-1221. 088668

Willekes C; Brands PJ; Willigers JM; Hoeks AP; Reneman RS (1999). Assessment of local differences in intima-media 

thickness in the human common carotid artery. J Vasc Res, 36: 222-228. 157147

Williams L; Ulrich C; Larson T; Wener M; Wood B; Campbell P; Potter J; McTiernan A; De Roos A (2009). Proximity to 

Traffic, Inflammation, and Immune Function among Women in the Seattle, Washington, Area. Environ Health 

Perspect, 117: 373. 191945

Wilson PW; Kauppila LI; O'Donnell CJ; Kiel DP; Hannan M; Polak JM; Cupples LA (2001). Abdominal aortic calcific 

deposits are an important predictor of vascular morbidity and mortality. Circulation, 103: 1529-1534. 156159

Wilson WE; Mage DT; Grant LD (2000). Estimating separately personal exposure to ambient and nonambient particulate 

matter for epidemiology and risk assessment: why and how. J Air Waste Manag Assoc, 50: 1167-1183. 010288

Witteman JC; Kok FJ; van Saase JL; Valkenburg HA (1986). Aortic calcification as a predictor of cardiovascular mortality. 

Lancet, 2: 1120-1122. 156161

Woodruff TJ; Darrow LA; Parker JD (2008). Air pollution and postneonatal infant mortality in the United States, 1999-

2002. Environ Health Perspect, 116: 110-115. 098386

Woodruff TJ; Grillo J; Schoendorf KC (1997). The relationship between selected causes of postneonatal infant mortality 

and particulate air pollution in the United States. Environ Health Perspect, 105: 608-612. 084271

Woodruff TJ; Parker JD; Schoendorf KC (2006). Fine particulate matter (PM25) air pollution and selected causes of 

postneonatal infant mortality in California. Environ Health Perspect, 114: 785-790. 088758

Xu J; Lee ET; Devereux RB; Umans JG; Bella JN; Shara NM; Yeh J; Fabsitz RR; Howard BV (2008). A longitudinal study 

of risk factors for incident albuminuria in diabetic American Indians: the Strong Heart Study. Am J Kidney Dis, 51: 

415-424. 157157

Yang AS; Estécio MR; Doshi K; Kondo Y; Tajara EH; Issa JP (2004). A simple method for estimating global DNA 

methylation using bisulfite PCR of repetitive DNA elements. Nucleic Acids Res, 32: e38. 192102

Yang C-Y; Hsieh H-J; Tsai S-S; Wu T-N; Chiu H-F (2006). Correlation between air pollution and postneonatal mortality in 

a subtropical city: Taipei, Taiwan. J Toxicol Environ Health A Curr Iss, 69: 2033-2040. 090760

Yang C-Y; Tseng Y-T; Chang C-C (2003). Effects of air pollution on birthweight among children born between 1995 and 

1997 in Kaohsiung, Taiwan. J Toxicol Environ Health A Curr Iss, 66: 807-816. 087886

Yatera K; Hsieh J; Hogg James C; Tranfield E; Suzuki H; Shih C-H; Behzad Ali R; Vincent R; van Eeden Stephan F 

(2008). Particulate matter air pollution exposure promotes recruitment of monocytes into atherosclerotic plaques. 

Am J Physiol Heart Circ Physiol, 294: H944-H953. 157162

Yauk C; Polyzos A; Rowan-Carroll A; Somers CM; Godschalk RW; Van Schooten FJ; Berndt ML; Pogribny IP; Koturbash 

I; Williams A; Douglas GR; Kovalchuk O (2008). Germ-line mutations, DNA damage, and global hypermethylation 

in mice exposed to particulate air pollution in an urban/industrial location. PNAS, 105: 605-610. 157164

Yauk CL; Quinn JS (1996). Multilocus DNA fingerprinting reveals high rate of heritable genetic mutation in herring gulls 

nesting in an industrialized urban site. PNAS, 93: 12137-12141. 089093

December 2009 7-116 
Ying Z; Kampfrath T; Thurston G; Farrar B; Lippmann M; Wang A; Sun Q; Chen LC; Rajagopalan S (2009). Ambient 

particulates alter vascular function through induction of reactive oxygen and nitrogen species. Toxicol Sci, 111: 80-

88. 190111

Yokota S; Mizuo K; Moriya N; Oshio S; Sugawara I; Takeda K (2009). Effect of prenatal exposure to diesel exhaust on 

dopaminergic system in mice. Neurosci Lett, 449: 38-41. 190518

Yoshida S; Ono N; Tsukue N; Oshio S; Umeda T; Takano H; Takeda K (2006). In utero exposure to diesel exhaust 

increased accessory reproductive gland weight and serum testosterone concentration in male mice. Environ Sci J 

Integr Environ Res, 13: 139-147. 097015

Yoshida S; Takedab K (2004). The effects of diesel exhaust on murine male reproductive function. Eisei Kagaku, 50: 210-

214. 097760

Yoshida S; Yoshida M; Sugawara I; Takeda K (2006). Mice strain differences in effects of fetal exposure to diesel exhaust 

gas on male gonadal differentiation. Environ Sci J Integr Environ Res, 13: 117-123. 156170

Zanobetti A; Bind MAC; Schwartz J (2008). Particulate air pollution and survival in a COPD cohort. Environ Health 

Perspect, 7: 48. 156177

Zanobetti A; Schwartz J (2007). Particulate air pollution, progression, and survival after myocardial infarction. Environ 

Health Perspect, 115: 769-775. 091247

Zeger S; Dominici F; McDermott A; Samet J (2008). Mortality in the Medicare population and chronic exposure to fine 

particulate air pollution in urban centers (2000-2005). Environ Health Perspect, 116: 1614-1619. 191951

Zeger S; McDermott A; Dominici F; Samet J (2007). Mortality in the medicare population and chronic exposure to fine 

particulate air pollution. Johns Hopkins University. Baltimore.http://www.bepress.com/jhubiostat/paper133. 157176

Zeman KL; Bennett WD (2006). Growth of the small airways and alveoli from childhood to the adult lung measured by 

aerosol-derived airway morphometry. J Appl Physiol, 100: 965-971. 157178

Zhang Z; Che W; Liang Y; Wu M; Li N; Shu Y; Liu F; Wu D (2007). Comparison of cytotoxicity and genotoxicity induced 

by the extracts of methanol and gasoline engine exhausts. Toxicol In Vitro, 21: 1058-1065. 157186

December 2009 7-117