Chapter 2. 

 

Integrative Health and Welfare Effects Overview 

The subsequent chapters of this ISA will present the most policy-relevant information related 

to this review of the NAAQS for PM. This chapter integrates the key findings from the disciplines 

evaluated in this current assessment of the PM scientific literature, which includes the atmospheric 

sciences, ambient air data analyses, exposure assessment, dosimetry, health studies 

(e.g., toxicological, controlled human exposure, and epidemiologic), and welfare effects. The EPA 

framework for causal determinations described in Chapter 1 has been applied to the body of 

scientific evidence in order to collectively examine the health or welfare effects attributed to PM 

exposure in a two-step process. 

As described in Chapter 1, EPA assesses the results of recent relevant publications, building 

upon evidence available during the previous NAAQS reviews, to draw conclusions on the causal 

relationships between relevant pollutant exposures and health or environmental effects. This ISA 

uses a five-level hierarchy that classifies the weight of evidence for causation: 

 Causal relationship 

 Likely to be a causal relationship 

 Suggestive of a causal relationship 

 Inadequate to infer a causal relationship 

 Not likely to be a causal relationship 

Beyond judgments regarding causality are questions relevant to quantifying health or 

environmental risks based on our understanding of the quantitative relationships between pollutant 

exposures and health or welfare effects. Once a determination is made regarding the causal 

relationship between the pollutant and outcome category, important questions regarding quantitative 

relationships include: 

 What is the concentration-response or dose-response relationship? 

 Under what exposure conditions (amount deposited, dose or concentration, duration and 

pattern) are effects observed? 

 What populations appear to be differentially affected (i.e., more susceptible) to effects? 

 What elements of the ecosystem (e.g., types, regions, taxonomic groups, populations, 

functions, etc.) appear to be affected, or are more sensitive to effects? 

To address these questions, in the second step of the EPA framework, the entirety of 

quantitative evidence is evaluated to identify and characterize potential concentration-response 

relationships. This requires evaluation of levels of pollutant and exposure durations at which effects 

were observed for exposed populations including potentially susceptible populations. 

This chapter summarizes and integrates the newly available scientific evidence that best 

informs consideration of the policy-relevant questions that frame this assessment, presented in 

Chapter 1. Section 2.1 discusses the trends in ambient concentrations and sources of PM and 

provides a brief summary of ambient air quality. Section 2.2 presents the evidence regarding 

personal exposure to ambient PM in outdoor and indoor microenvironments, and it discusses the 

 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 2-1 
relationship between ambient PM concentrations and exposure to PM from ambient sources. 

Section 2.3 integrates the evidence for studies that examine the health effects associated with shortand long-term exposure to PM and discusses important uncertainties identified in the interpretation 

of the scientific evidence. Section 2.4 provides a discussion of policy-relevant considerations, such 

as potentially susceptible populations, lag structure, and the PM concentration-response relationship, 

and PM sources and constituents linked to health effects. Section 2.5 summarizes the evidence for 

welfare effects related to PM exposure. Finally, Section 2.6 provides all of the causal determinations 

reached for each of the health outcomes and PM exposure durations evaluated in this ISA. 

2.1. Concentrations and Sources of Atmospheric PM 

2.1.1. Ambient PM Variability and Correlations 

Recently, advances in understanding the spatiotemporal distribution of PM mass and its 

constituents have been made, particularly with regard to PM2.5 and its components as well as 

ultrafine particles (UFPs). Emphasis in this ISA is placed on the period from 2005-2007, 

incorporating the most recent validated EPA Air Quality System (AQS) data. The AQS is EPA’s 

repository for ambient monitoring data reported by the national, and state and local air monitoring 

networks. Measurements of PM2.5 and PM10 are reported into AQS, while PM10-2.5 concentrations 

are obtained as the difference between PM10 and PM2.5 (after converting PM10 concentrations from 

STP to local conditions; Section 3.5). Note, however, that a majority of U.S. counties were

represented in AQS because their population fell below the regulatory monitoring threshold. 

Moreover, monitors reporting to AQS were not uniformly distributed across the U.S. or within 

counties, and conclusions drawn from AQS data may not apply equally to all parts of a geographic 

region. Furthermore, biases can exist for some PM constituents (and hence total mass) owing to 

volatilization losses of nitrates and other semi-volatile compounds, and, conversely, to retention of 

particle-bound water by hygroscopic species. The degree of spatial variability in PM was likely to be 

region-specific and strongly influenced by local sources and meteorological and topographic 

conditions. 

 not 

2.1.1.1. Spatial Variability across the U.S. 

AQS data for daily average concentrations of PM2.5 for 2005-2007 showed considerable 

variability across the U.S. (Section 3.5.1.1). Counties with the highest average concentrations of 

PM2.5 (>18 μg/m3

) were reported for several counties in the San Joaquin Valley and inland southern 

California as well as Jefferson County, AL (containing Birmingham) and Allegheny County, PA 

(containing Pittsburgh). Relatively few regulatory monitoring sites have the appropriate co-located 

monitors for computing PM10-2.5, resulting in poor geographic coverage on a national scale 

(Figure 3-10). Although the general understanding of PM differential settling leads to an expectation 

of greater spatial heterogeneity in the PM10-2.5 fraction, deposition of particles as a function of size 

depends strongly on local meteorological conditions. Better geographic coverage is available for 

PM10, where the highest reported annual average concentrations (>50 μg/m3

) occurred in southern 

California, southern Arizona and central New Mexico. The size distribution of PM varied 

substantially by location, with a generally larger fraction of PM10 mass in the PM10-2.5 size range in 

western cities (e.g., Phoenix and Denver) and a larger fraction of PM10 in the PM2.5 size range in 

eastern U.S. cities (e.g., Pittsburgh and Philadelphia). UFPs are not measured as part of AQS or any 

other routine regulatory network in the U.S. Therefore, limited information is available regarding 

regional variability in the spatiotemporal distribution of UFPs. 

Spatial variability in PM2.5 components obtained from the Chemical Speciation Network 

(CSN) varied considerably by species from 2005-2007 (Figures 3-12 through 3-18). The highest 

annual average organic carbon (OC) concentrations were observed in the western and southeastern 

U.S. OC concentrations in the western U.S. peaked in the fall and winter, while OC concentrations in 

the Southeast peaked anytime between spring and fall. Elemental carbon (EC) exhibited less 

seasonality than OC and showed lowest seasonal variability in the eastern half of the U.S. The 

December 2009 2-2 
highest annual average EC concentrations were present in Los Angeles, Pittsburgh, New York, and 

El Paso. Concentrations of sulfate (SO4

2–) were higher in the eastern U.S. as a result of higher SO2

emissions in the East compared with the West. There is also considerable seasonal variability with 

higher SO4

2– concentrations in the summer months when the oxidation of SO2 proceeds at a faster 

rate than during the winter. Nitrate (NO3

–

) concentrations were highest in California and during the 

winter in the Upper Midwest. In general, NO3

–

 was higher in the winter across the country, in part as 

a result of temperature-driven partitioning and volatilization. Exceptions existed in Los Angeles and 

Riverside, CA, where high NO3

– concentrations appeared year-round. There is variation in both 

PM2.5 mass and composition among cities, some of which might be due to regional differences in 

meteorology, sources, and topography. 

2.1.1.2. Spatial Variability on the Urban and Neighborhood Scales 

In general, PM2.5 has a longer atmospheric lifetime than PM10-2.5. As a result, PM2.5 is more 

homogeneously distributed than PM10-2.5, whose concentrations more closely reflect proximity to 

local sources (Section 3.5.1.2). Because PM10 encompasses PM10-2.5 in addition to PM2.5, it also 

exhibits more spatial heterogeneity than PM2.5. Urban- and neighborhood-scale variability in PM 

mass and composition was examined by focusing on 15 metropolitan areas, which were chosen 

based on their geographic distribution and coverage in recent health effects studies. The urban areas 

selected were Atlanta, Birmingham, Boston, Chicago, Denver, Detroit, Houston, Los Angeles, New 

York, Philadelphia, Phoenix, Pittsburgh, Riverside, Seattle and St. Louis. Inter-monitor correlation 

remained higher over long distances for PM2.5 as compared with PM10 in these 15 urban areas. To a 

large extent, greater variation in PM2.5 and PM10 concentrations within cities was observed in areas 

with lower ratios of PM2.5 to PM10. When the data was limited to only sampler pairs with less than 

4 km separation (i.e., on a neighborhood scale), inter-sampler correlations remained higher for PM2.5

than for PM10. The average inter-sampler correlation was 0.93 for PM2.5, while it dropped to 0.70 for 

PM10 (Section 3.5.1.3). Insufficient data were available in the 15 metropolitan areas to perform 

similar analyses for PM10-2.5 using co-located, low volume FRM monitors. 

As previously mentioned, UFPs are not measured as part of AQS or any other routine 

regulatory network in the U.S. Therefore, information about the spatial variability of UFPs is sparse; 

however, their number concentrations are expected to be highly spatially and temporally variable. 

This has been shown on the urban scale in studies in which UFP number concentrations drop off 

quickly with distance from roads compared to accumulation mode particle numbers. 

2.1.2. Trends and Temporal Variability 

Overall, PM2.5 concentrations decreased from 1999 (the beginning of nationwide monitoring 

for PM2.5) to 2007 in all ten EPA Regions, with the 3-yr avg of the 98th percentile of 24-h PM2.5

concentrations dropping 10% over this time period. However from 2002-2007, concentrations of 

PM2.5 were nearly constant with decreases observed in only some EPA Regions (Section 3.5.2.1). 

Concentrations of PM2.5 components were only available for 2002-2007 using CSN data and showed 

little decline over this time period. This trend in PM2.5 components is consistent with trends in PM2.5

mass concentration observed after 2002 (shown in Figures 3-44 through 3-47). Concentrations of 

PM10 also declined from 1988 to 2007 in all ten EPA Regions. 

Using hourly PM observations in the 15 metropolitan areas, diel variation showed average 

hourly peaks that differ by size fraction and region (Section 3.5.2.3). For both PM2.5 and PM10, a 

morning peak was typically observed starting at approximately 6:00 a.m., corresponding with the 

start of morning rush hour. There was also an evening concentration peak that was broader than the 

morning peak and extended into the overnight period, reflecting the concentration increase caused by 

the usual collapse of the mixing layer after sundown. The magnitude and duration of these peaks 

varied considerably by metropolitan area investigated. 

UFPs were found to exhibit similar two-peaked diel patterns in Los Angeles and the San 

Joaquin Valley of CA and Rochester, NY as well as in Kawasaki City, Japan, and Copenhagen, 

Denmark. The morning peak in UFPs likely represents primary source emissions, such as rush-hour 

traffic, while the afternoon peak likely represents the combination of primary source emissions and 

nucleation of new particles. 

December 2009 2-3 
2.1.3. Correlations between Copollutants 

Correlations between PM and gaseous copollutants, including SO2, NO2, carbon monoxide 

(CO) and O3, varied both seasonally and spatially between and within metropolitan areas 

(Section 3.5.3). On average, PM2.5 and PM10 were correlated with each other better than with the 

gaseous copollutants. Although data are limited for PM10-2.5, the available data suggest a stronger 

correlation between PM10 and PM10-2.5 than between PM2.5 and PM10-2.5 on a national basis.There 

was relatively little seasonal variability in the mean correlation between PM in both size fractions 

and SO2 and NO2. CO, however, showed higher correlations with PM2.5 and PM10 on average in

winter compared with the other seasons. This seasonality results in part because a larger fraction of 

PM is primary in origin during the winter. To the extent that this primary component of PM is 

associated with common combustion sources of NO

 the 

sis 

s 

 designated as an FEM. 

tion 

n 

2 and CO, then higher correlations with these 

gaseous copollutants are to be expected. Increased atmospheric stability in colder months also results 

in higher correlations between primary pollutants (Section 3.5). 

The correlation between daily maximum 8-h avg O3 and 24-h avg PM2.5 showed the highest 

degree of seasonal variability with positive correlations on average in summer (avg = 0.56) and 

negative correlations on average in the winter (avg = -0.30). During the transition seasons, spring 

and fall, correlations were mixed but on average were still positive. PM2.5 is both primary and 

secondary in origin, whereas O3 is only secondary. Photochemical production of O3 and secondary 

PM in the planetary boundary layer (PBL) is much slower during the winter than during other 

seasons. Primary pollutant concentrations (e.g., primary PM2.5 components, NO and NO2) in many 

urban areas are elevated in winter as the result of heating emissions, cold starts and low mixing 

heights. O3 in the PBL during winter is mainly associated with air subsiding from above the 

boundary layer following the passage of cold fronts, and this subsiding air has much lower PM 

concentrations than are present in the PBL. Therefore, a negative association between O3 and PM2.5 

is frequently observed in the winter. During summer, both O3 and secondary PM2.5 are produced in 

the PBL and in the lower free troposphere at faster rates compared to winter, and so they tend to be 

positively correlated. 

2.1.4. Measurement Techniques 

The federal reference methods (FRMs) for PM2.5 and PM10 are based on criteria outlined in 

the Code of Federal Regulations. They are, however, subject to several limitations that should be 

kept in mind when using compliance monitoring data for health studies. For example, FRM 

techniques are subject to the loss of semi-volatile species such as organic compounds and 

ammonium nitrate (especially in the West). Since FRMs based on gravimetry use 24-h integrated

filter samples to collect PM mass, no information is available for variations over shorter averaging 

times from these instruments. However, methods have been developed to measure real-time PM 

mass concentrations. Real-time (or continuous and semi-continuous) measurement techniques are 

also available for PM species, such as particle into liquid sampler (PILS) for multiple ions analy

and aerosol mass spectrometer (AMS) for multiple components analysis (Section 3.4.1). Advance

have also been achieved in PM organic speciation. New 24-h FRMs and Federal Equivalent Methods 

(FEMs) based on gravimetry and continuous FEMs for PM10-2.5 are available. FRMs for PM10-2.5 rely 

on calculating the difference between co-located PM10 and PM2.5 measurements while a 

dichotomous sampler is

2.1.5. PM Formation in the Atmosphere and Removal 

PM in the atmosphere contains both primary (i.e., emitted directly by sources) and secondary 

components, which can be anthropogenic or natural in origin. Secondary PM components can be 

produced by the oxidation of precursor gases such as SO2 and NOX to acids followed by 

neutralization with ammonia (NH3) and the partial oxidation of organic compounds. In addition to 

being emitted as primary particles, UFPs are produced by the nucleation of H2SO4 vapor, H2O 

vapor, and perhaps NH3 and certain organic compounds. Over most of the earth’s surface, nuclea

is probably the major mechanism forming new UFPs. New UFP formation has been observed i

environments ranging from relatively unpolluted marine and continental environments to polluted 

December 2009 2-4 
urban areas as an ongoing background process and during nucleation events. However, as noted 

above, a large percentage of UFPs come from combustion-related sources such as motor vehicles. 

Developments in the chemistry of formation of secondary organic aerosol (SOA) indicate that 

oligomers are likely a major component of OC in aerosol samples. Recent observations also suggest 

that small but significant quantities of SOA are formed from the oxidation of isoprene in addition to 

the oxidation of terpenes and organic hydrocarbons with six or more carbon atoms. Gasoline engines 

have been found to emit a mix of nucleation-mode heavy and large polycyclic aromatic 

hydrocarbons on which unspent fuel and trace metals can condense, while diesel particles are 

composed of a soot nucleus on which sulfates and hydrocarbons can condense. To the extent that the 

primary component of organic aerosol is overestimated in emissions from combustion sources, the 

semi-volatile components are underestimated. This situation results from the lack of capture of 

evaporated semi-volatile components upon dilution in common emissions tests. As a result, neartraffic sources of precursors to SOA would be underestimated. The oxidation of these precursors 

results in more oxidized forms of SOA than previously considered, in both near source urban 

environments and further downwind. Primary organic aerosol can also be further oxidized to forms 

that have many characteristics in common with oxidized SOA formed from gaseous precursors. 

Organic peroxides constitute a significant fraction of SOA and represent an important class of 

reactive oxygen species (ROS) that have high oxidizing potential. More information on sources, 

emissions and deposition of PM are included in Section 3.3. 

Wet and dry deposition are important processes for removing PM and other pollutants from the 

atmosphere on urban, regional, and global scales. Wet deposition includes incorporation of particles 

into cloud droplets that fall as rain (rainout) and collisions with falling rain (washout). Other 

hydrometeors (snow, ice) can also serve the same purpose. Dry deposition involves transfer of 

particles through gravitational settling and/or by impaction on surfaces by turbulent motions. The 

effects of deposition of PM on ecosystems and materials are discussed in Section 2.5 and in 

Chapter 9. 

2.1.6. Source Contributions to PM 

Results of receptor modeling calculations indicate that PM2.5 is produced mainly by 

combustion of fossil fuel, either by stationary sources or by transportation. A relatively small number 

of broadly defined source categories, compared to the total number of chemical species that typically 

are measured in ambient monitoring source receptor studies, account for the majority of the observed 

PM mass. Some ambiguity is inherent in identifying source categories. For example, quite different 

mobile sources such as trucks, farm equipment, and locomotives rely on diesel engines and ancillary 

data is often required to resolve these sources. A compilation of study results shows that secondary 

SO4

2– (derived mainly from SO2 emitted by Electricity Generating Units [EGUs]), NO3–

 (from the 

oxidation of NOx emitted mainly from transportation sources and EGUs), and primary mobile source 

categories, constitute most of PM2.5 (and PM10) in the East. PM10-2.5 is mainly primary in origin, 

having been emitted as fully formed particles derived from abrasion and crushing processes, soil 

disturbances, plant and insect fragments, pollens and other microorganisms, desiccation of marine 

aerosol emitted from bursting bubbles, and hygroscopic fine PM expanding with humidity to coarse 

mode. Gases such as HNO3 can also condense directly onto preexisting coarse particles. Suspended 

primary coarse PM can contain Fe, Si, Al, and base cations from soil, plant and insect fragments, 

pollen, fungal spores, bacteria, and viruses, as well as fly ash, brake lining particles, debris, and 

automobile tire fragments. Quoted uncertainties in the source apportionment of constituents in 

ambient aerosol samples typically range from 10 to 50%. An intercomparison of source 

apportionment techniques indicated that the same major source categories of PM2.5 were consistently 

identified by several independent groups working with the same data sets. Soil-, sulfate-, residual 

oil-, and salt-associated mass were most clearly identified by the groups. Other sources with more 

ambiguous signatures, such as vegetative burning and traffic-related emissions were less consistently 

identified. 

Spatial variability in source contributions across urban areas is an important consideration in 

assessing the likelihood of exposure error in epidemiologic studies relating health outcomes to 

sources. Concepts similar to those for using ambient concentrations as surrogates for personal 

exposures apply here. Some source attribution studies for PM2.5 indicate that intra-urban variability 

increases in the following order: regional sources (e.g., secondary SO4

2– originating from EGUs) 

< area sources (e.g., on-road mobile sources) < point sources (e.g., metals from stacks of smelters). 

December 2009 2-5 
Although limited information was available for PM10-2.5, it does indicate a similar ordering, but 

without a regional component (resulting from the short lifetime of PM10-2.5 compared to transport 

times on the regional scale). More discussion on source contributions to PM is available in 

Section 3.6. 

2.1.7. Policy-Relevant Background 

The background concentrations of PM that are useful for risk and policy assessments, which 

inform decisions about the NAAQS are referred to as policy-relevant background (PRB) 

concentrations. PRB concentrations have historically been defined by EPA as those concentrations 

that would occur in the U.S. in the absence of anthropogenic emissions in continental North America 

defined here as the U.S., Canada, and Mexico. For this document, PRB concentrations include 

contributions from natural sources everywhere in the world and from anthropogenic sources outside 

continental North America. Background concentrations so defined facilitated separation of pollution 

that can be controlled by U.S. regulations or through international agreements with neighboring 

countries from those that were judged to be generally uncontrollable by the U.S. Over time, 

consideration of potential broader ranging international agreements may lead to alternative 

determinations of which PM source contributions should be considered by EPA as part of PRB. 

Contributions to PRB concentrations of PM include both primary and secondary natural and 

anthropogenic components. For this document, PRB concentrations of PM2.5 for the continental U.S. 

were estimated using EPA’s Community Multi-scale Air Quality (CMAQ) modeling system, a 

deterministic, chemical-transport model (CTM), using output from GEOS-Chem a global-scale 

model for CMAQ boundary conditions. PRB concentrations of PM2.5 were estimated to be less than 

1 μg/m3

 on an annual basis, with maximum daily average values in a range from 3.1 to 20 μg/m3 and 

having a peak of 63 μg/m3

 at the nine national park sites across the U.S. used to evaluate model 

performance for this analysis. A description of the models and evaluation of their performance is 

given in Section 3.6 and further details about the calculations of PRB concentrations are given in 

Section 3.7. 

2.2. Human Exposure 

This section summarizes the findings from the recent exposure assessment literature. This 

summary is intended to support the interpretation of the findings from epidemiologic studies and 

reflects the material presented in Section 3.8. Attention is given to how concentration metrics can be 

used in exposure assessment and what errors and uncertainties are incurred for different approaches. 

Understanding of exposure errors is important because exposure error can potentially bias an 

estimate of a health effect or increase the size of confidence intervals around a health effect estimate. 

2.2.1. Spatial Scales of PM Exposure Assessment 

Assessing population-level exposure at the urban scale is particularly relevant for time-series 

epidemiologic studies, which provide information on the relationship between health effects and 

community-average exposure, rather than an individual’s exposure. PM concentrations measured at a 

central-site ambient monitor are used as surrogates for personal PM exposure. However, the 

correlation between the PM concentration measured at central-site ambient monitor(s) and the 

unknown true community average concentration depends on the spatial distribution of PM, the 

location of the monitoring site(s) chosen to represent the community average, and division of the 

community by terrain features or local sources into several sub-communities that differ in the 

temporal pattern of pollution. Concentrations of SO4

2– and some components of SOA measured at 

central-site monitors are expected to be uniform in urban areas because of the regional nature of their 

sources. However, this is not true for primary components like EC whose sources are strongly 

spatially variable in urban areas. 

At micro-to-neighborhood scales, heterogeneity of sources and topography contribute to 

variability in exposure. This is particularly true for PM10-2.5 and for UFPs, which have spatially 

December 2009 2-6 
variable urban sources and loss processes (mainly gravitational settling for PM10-2.5 and coagulation 

for UFPs) that also limit their transport from sources more readily than for PM2.5. Personal activity 

patterns also vary across urban areas and across regions. Some studies, conducted mainly in Europe, 

have found personal PM2.5 and PM10 exposures for pedestrians in street canyons to be higher than 

ambient concentrations measured by urban central site ambient monitors. Likewise, 

microenvironmental UFP concentrations were observed to be substantially higher in near-road 

environments, street canyons, and tunnels when compared with urban background concentrations. 

In-vehicle UFP and PM2.5 exposures can also be important. As a result, concentrations measured by 

ambient monitors likely do not reflect the contributions of UFP or PM2.5 exposures to individuals 

while commuting. 

There is significant variability within and across regions of the country with respect to indoor 

exposures to ambient PM. Infiltrated ambient PM concentrations depend in part on the ventilation 

properties of the building or vehicle in which the person is exposed. PM infiltration factors depend 

on particle size, chemical composition, season, and region of the country. Infiltration can best be 

modeled dynamically rather than being represented by a single value. Season is important to PM 

infiltration because it affects the ventilation practices (e.g., open windows) used. In addition, ambient 

temperature and humidity conditions affect the transport, dispersion, and size distribution of PM. 

Residential air exchange rates have been observed to be higher in the summer for regions with low 

air conditioning usage. Regional differences in air exchange rates (Southwest < Southeast 

< Northeast < Northwest) also reflect ventilation practices. Differential infiltration occurs as a 

function of PM size and composition (the latter of which is described below). PM infiltration is 

larger for accumulation mode particles than for UFPs and PM10-2.5. Differential infiltration by size 

fraction can affect exposure estimates if not accurately characterized. 

2.2.2. Exposure to PM Components and Copollutants 

Emission inventories and source apportionment studies suggest that sources of PM exposure 

vary by region. Comparison of studies performed in the eastern U.S. with studies performed in the 

western U.S. suggest that the contribution of SO4

2– to exposure is higher for the East (16-46%) 

compared with the West (~4%) and that motor vehicle emissions and secondary NO3

–

 are larger 

sources of exposure for the West (~9%) as compared with the East (~4%). Results of source 

apportionment studies of exposure to SO4

2– indicate that SO42– exposures are mainly attributable to 

ambient sources. Source apportionment for OC and EC is difficult because they originate from both 

indoor and outdoor sources. Exposure to OC of indoor and outdoor origin can be distinguished by 

the presence of aliphatic C-H groups generated indoors, since outdoor concentrations of aliphatic 

C-H are low. Studies of personal exposure to ambient trace metal have shown significant variation 

among cities and over seasons. This is in response to geographic and seasonal variability in sources 

including incinerator operation, fossil fuel combustion, biomass combustion (wildfires), and the 

resuspension of crustal materials in the built environment. Differential infiltration is also affected by 

variations in particle composition and volatility. For example, EC infiltrates more readily than OC. 

This can lead to outdoor-indoor differentials in PM composition. 

Some studies have explored the relationship between PM and copollutant gases and suggested 

that certain gases can serve as surrogates for describing exposure to other air pollutants. The findings 

indicate that ambient concentrations of gaseous copollutants can act as surrogates for personal 

exposure to ambient PM. Several studies have concluded that ambient concentrations of O3, NO2, 

and SO2 are associated with the ambient component of personal exposure to total PM2.5. If 

associations between ambient gases and personal exposure to PM2.5 of ambient origin exist, such 

associations are complex and vary by season and location. 

2.2.3. Implications for Epidemiologic Studies 

In epidemiologic studies, exposure may be estimated using various approaches, most of which 

rely on measurements obtained using central site monitors. The magnitude and direction of the 

biases introduced through error in exposure measurement depend on the extent to which the error is 

associated with the measured PM concentration. In general, when exposure error is not strongly 

correlated with the measured PM concentration, bias is toward the null and effect estimates are 

December 2009 2-7 
underestimated. Moreover, lack of information regarding exposure measurement error can also add 

uncertainty to the health effects estimate. 

One important factor to be considered is the spatial variation in PM concentrations. The degree 

of urban-scale spatial variability in PM concentrations varies across the country and by size fraction. 

PM2.5 concentrations are relatively well-correlated across monitors in the urban areas examined for 

this assessment. The limited available evidence indicates that there is greater spatial variability in 

PM10-2.5 concentrations than PM2.5 concentrations, resulting in increased exposure error for the larger 

size fraction. Likewise, studies have shown UFPs to be more spatially variable across urban areas 

compared to PM2.5. Even if PM2.5, PM10-2.5, or UFP concentrations measured at sites within an urban 

area are generally highly correlated, significant spatial variation in their concentrations can occur on 

any given day. In addition, there can be differential exposure errors for PM components (e.g., SO4

2–, 

OC, EC). Current information suggests that UFPs, PM10-2.5, and some PM components are more 

spatially variable than PM2.5. Spatial variability of these PM indicators adds uncertainty to exposure 

estimates. 

Overall, recent studies generally confirm and build upon the key conclusions of the 2004 PM 

AQCD: separation of total PM exposures into ambient and nonambient components reduces 

potential uncertainties in the analysis and interpretation of PM health effects data; and ambient PM 

concentration can be used as a surrogate for ambient PM exposure in community time-series 

epidemiologic studies because the change in ambient PM concentration should be reflected in the 

change in the health risk coefficient. The use of the community average ambient PM2.5 concentration 

as a surrogate for the community average personal exposure to ambient PM2.5 is not expected to 

change the principal conclusions from time-series and most panel epidemiologic studies that use 

community average health and pollution data. Several recent studies support this by showing how 

the ambient component of personal exposure to PM2.5 could be estimated using various tracer and 

source apportionment techniques and by showing that the ambient component is highly correlated 

with ambient concentrations of PM2.5. These studies show that the non-ambient component of 

personal exposure to PM2.5 is largely uncorrelated with ambient PM2.5 concentrations. A few panel 

epidemiologic studies have included personal as well as ambient monitoring data, and generally 

reported associations with all types of PM measurements. Epidemiologic studies of long-term 

exposure typically exploit the differences in PM concentration across space, as well as time, to 

estimate the effect of PM on the health outcome of interest. Long-term exposure estimates are most 

accurate for pollutants that do not vary substantially within the geographic area studied. 

2.3. Health Effects 

This section evaluates the evidence from toxicological, controlled human exposure, and 

epidemiologic studies that examined the health effects associated with short- and long-term exposure 

to PM (i.e., PM2.5, PM10-2.5 and UFPs). The results from the health studies evaluated in combination 

with the evidence from atmospheric chemistry and exposure assessment studies contribute to the 

causal determinations made for the health outcomes discussed in this assessment (a description of 

the causal framework can be found in Section 1.5.4). In the following sections a discussion of the 

causal determinations will be presented by PM size fraction and exposure duration (i.e., short- or 

long-term exposure) for the health effects for which sufficient evidence was available to conclude a 

causal, likely to be causal or suggestive relationship. Although not presented in depth in this chapter, 

a detailed discussion of the underlying evidence used to formulate each causal determination can be 

found in Chapters 6 and 7. 

December 2009 2-8 
2.3.1. Exposure to PM2.5 

2.3.1.1. Effects of Short-Term Exposure to PM2.5 

Table 2-1. Summary of causal determinations for short-term exposure to PM2.5 . 

Size Fraction Outcome Causality Determination 

Cardiovascular Effects Causal 

PM2.5 Respiratory Effects Likely to be causal 

Mortality Causal 

Cardiovascular Effects 

Epidemiologic studies that examined the effect of PM2.5 on cardiovascular emergency 

department (ED) visits and hospital admissions reported consistent positive associations 

(predominantly for ischemic heart disease [IHD] and congestive heart failure [CHF]), with the 

majority of studies reporting increases ranging from 0.5 to 3.4% per 10 μg/m3

 increase in PM2.5.

These effects were observed in study locations with mean1 24-h avg PM2.5 concentrations ranging 

from 7-18 μg/m3 (Section 6.2.10). The largest U.S.-based multicity study evaluated, Medicare Air 

Pollution Study (MCAPS), provided evidence of regional heterogeneity (e.g., the largest excess risks 

occurred in the Northeast [1.08%]) and seasonal variation (e.g., the largest excess risks occurred 

during the winter season [1.49%]) in PM2.5 cardiovascular disease (CVD) risk estimates, which is 

consistent with the null findings of several single-city studies conducted in the western U.S. These 

associations are supported by multicity epidemiologic studies that observed consistent positive 

associations between short-term exposure to PM2.5 and cardiovascular mortality and also reported 

regional and seasonal variability in risk estimates. The multicity studies evaluated reported 

consistent increases in cardiovascular mortality ranging from 0.47 to 0.85% in study locations with 

mean 24-h avg PM2.5 concentrations above 12.8 μg/m3 (Table 6-15). 

Controlled human exposure studies have demonstrated PM2.5-induced changes in various 

measures of cardiovascular function among healthy and health-compromised adults. The most 

consistent evidence is for altered vasomotor function following exposure to diesel exhaust (DE) or 

CAPs with O3 (Section 6.2.4.2). Although these findings provide biological plausibility for the 

observations from epidemiologic studies, the fresh DE used in the controlled human exposure 

studies evaluated contains gaseous components (e.g., CO, NOx), and therefore, the possibility that 

some of the changes in vasomotor function might be due to gaseous components cannot be ruled out. 

Furthermore, the prevalence of UFPs in fresh DE limits the ability to conclusively attribute the 

observed effects to either the UF fraction or PM2.5 as a whole. An evaluation of toxicological studies 

found evidence for altered vessel tone and microvascular reactivity, which provide coherence and 

biological plausibility for the vasomotor effects that have been observed in both the controlled 

human exposure and epidemiologic studies (Section 6.2.4.3). However, most of these toxicological 

studies exposed animals via intratracheal (IT) instillation or using relatively high inhalation 

concentrations. 

In addition to the effects observed on vasomotor function, myocardial ischemia has been 

observed across disciplines through PM2.5 effects on ST-segment depression, with toxicological 

studies providing biological plausibility by demonstrating reduced blood flow during ischemia 

(Section 6.2.3). There is also a growing body of evidence from controlled human exposure and 

toxicological studies demonstrating PM2.5-induced changes on heart rate variability (HRV) and 

 

1 In this context mean represents the arithmetic mean of 24-h avg PM concentrations. 

December 2009 2-9 
markers of systemic oxidative stress (Sections 6.2.1 and 6.2.9, respectively). Additional but 

inconsistent effects of PM2.5 on blood pressure (BP), blood coagulation markers, and markers of 

systemic inflammation have also been reported across disciplines. Toxicological studies have 

provided biologically plausible mechanisms (e.g., increased right ventricular pressure and 

diminished cardiac contractility) for the associations observed between PM2.5 and CHF in 

epidemiologic studies. 

Together, the collective evidence from epidemiologic, controlled human exposure, and 

toxicological studies is sufficient to conclude that a causal relationship exists between shortterm exposures to PM2.5 and cardiovascular effects. 

Respiratory Effects 

The recent epidemiologic studies evaluated report consistent positive associations between 

short-term exposure to PM2.5 and respiratory ED visits and hospital admissions for chronic 

obstructive pulmonary disease (COPD) and respiratory infections (Section 6.3). Positive associations 

were also observed for asthma ED visits and hospital admissions for adults and children combined, 

but effect estimates are imprecise and not consistently positive for children alone. Most studies 

reported effects in the range of ~1% to 4% increase in respiratory hospital admissions and ED visits 

and were observed in study locations with mean 24-h avg PM2.5 concentrations ranging from 

6.1-22 μg/m3

. Additionally, multicity epidemiologic studies reported consistent positive associations 

between short-term exposure to PM2.5 and respiratory mortality as well as regional and seasonal 

variability in risk estimates. The multicity studies evaluated reported consistent, precise increases in 

respiratory mortality ranging from 1.67 to 2.20% in study locations with mean 24-h avg PM2.5

concentrations above 12.8 μg/m3

 (Table 6-15). Evidence for PM2.5-related respiratory effects was 

also observed in panel studies, which indicate associations with respiratory symptoms, pulmonary 

function, and pulmonary inflammation among asthmatic children. Although not consistently 

observed, some controlled human exposure studies have reported small decrements in various 

measures of pulmonary function following controlled exposures to PM2.5 (Section 6.3.2.2). 

Controlled human exposure studies using adult volunteers have demonstrated increased 

markers of pulmonary inflammation following exposure to a variety of different particle types; 

oxidative responses to DE and wood smoke; and exacerbations of allergic responses and allergic 

sensitization following exposure to DE particles (Section 6.3). Toxicological studies have provided 

additional support for PM2.5-related respiratory effects through inhalation exposures of animals to 

CAPs, DE, other traffic-related PM and wood smoke. These studies reported an array of respiratory 

effects including altered pulmonary function, mild pulmonary inflammation and injury, oxidative 

responses, airway hyperresponsiveness (AHR) in allergic and non-allergic animals, exacerbations of 

allergic responses, and increased susceptibility to infections (Section 6.3). 

Overall, the evidence for an effect of PM2.5 on respiratory outcomes is somewhat restricted by 

limited coherence between some of the findings from epidemiologic and controlled human exposure 

studies for the specific health outcomes reported and the sub-populations in which those health 

outcomes occur. Epidemiologic studies have reported variable results among specific respiratory 

outcomes, specifically in asthmatics (e.g., increased respiratory symptoms in asthmatic children, but 

not increased asthma hospital admissions and ED visits) (Section 6.3.8). Additionally, respiratory 

effects have not been consistently demonstrated following controlled exposures to PM2.5 among 

asthmatics or individuals with COPD. Collectively, the epidemiologic, controlled human exposure, 

and toxicological studies evaluated demonstrate a wide range of respiratory responses, and although 

results are not fully consistent and coherent across studies the evidence is sufficient to conclude that 

a causal relationship is likely to exist between short-term exposures to PM2.5 and 

respiratory effects.

Mortality 

An evaluation of the epidemiologic literature indicates consistent positive associations 

between short-term exposure to PM2.5 and all-cause, cardiovascular-, and respiratory-related 

mortality (Section 6.5.2.2.). The evaluation of multicity studies found that consistent and precise risk 

estimates for all-cause (nonaccidental) mortality that ranged from 0.29 to 1.21% per 10 μg/m3

December 2009 2-10 
increase in PM2.5 at lags of 1 and 0-1 days. In these study locations, mean 24-h avg PM2.5

concentrations were 12.8 μg/m3

 and above (Table 6-15). Cardiovascular-related mortality risk 

estimates were found to be similar to those for all-cause mortality; whereas, the risk estimates for 

respiratory-related mortality were consistently larger (i.e., 1.01-2.2%) using the same lag periods and 

averaging indices. The studies evaluated that examined the relationship between short-term exposure 

to PM2.5 and cardiovascular effects (Section 6.2) provide coherence and biological plausibility for 

PM2.5-induced cardiovascular mortality, which represents the largest component of total 

(nonaccidental) mortality (~ 35%) (American Heart Association, 2009, 198920). However, as noted 

in Section 6.3, there is limited coherence between some of the respiratory morbidity findings from 

epidemiologic and controlled human exposure studies for the specific health outcomes reported and 

the subpopultions in which those health outcomes occur, complicating the interpretation of the PM2.5

respiratory mortality effects observed. Regional and seasonal patterns in PM2.5 risk estimates were 

observed with the greatest effect estimates occurring in the eastern U.S. and during the spring. Of the 

studies evaluated only Burnett et al. (2004, 086247), a Canadian multicity study, analyzed gaseous 

pollutants and found mixed results, with possible confounding of PM2.5 risk estimates by NO2. Although the recently evaluated U.S.-based multicity studies did not analyze potential confounding 

of PM2.5 risk estimates by gaseous pollutants, evidence from the limited number of single-city 

studies evaluated in the 2004 PM AQCD (U.S. EPA, 2004, 056905) suggest that gaseous 

copollutants do not confound the PM2.5-mortality association. This is further supported by studies 

that examined the PM10-mortality relationship. An examination of effect modifiers (e.g., 

demographic and socioeconomic factors), specifically air conditioning use as an indicator for 

decreased pollutant penetration indoors, has suggested that PM2.5 risk estimates increase as the 

percent of the population with access to air conditioning decreases. Collectively, the epidemiologic 

literature provides evidence that a causal relationship exists between short-term exposures 

to PM2.5 and mortality.

2.3.1.2. Effects of Long-Term Exposure to PM2.5 

Table 2-2. Summary of causal determinations for long-term exposure to PM2.5 . 

Size Fraction Outcome Causality Determination 

Cardiovascular Effects Causal 

Respiratory Effects Likely to be causal 

Mortality Causal 

Reproductive and Developmental Suggestive 

PM2.5

Cancer, Mutagenicity, and Genotoxicity Suggestive 

Cardiovascular Effects 

The strongest evidence for cardiovascular health effects related to long-term exposure to PM2.5

comes from large, multicity U.S.-based studies, which provide consistent evidence of an association 

between long-term exposure to PM2.5 and cardiovascular mortality (Section 7.2.10). These 

associations are supported by a large U.S.-based epidemiologic study (i.e., Women’s Health Initiative 

[WHI] study) that reports associations between PM2.5 and CVDs among post-menopausal women 

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

) (Section 7.2). However, epidemiologic 

studies that examined subclinical markers of CVD report inconsistent findings. Epidemiologic 

studies have also provided some evidence for potential modification of the PM2.5-CVD association 

when examining individual-level data, specifically smoking status and the use of antiDecember 2009 2-11 
hyperlipidemics. Although epidemiologic studies have not consistently detected effects on markers 

of atherosclerosis due to long-term exposure to PM2.5, toxicological studies have provided strong 

evidence for accelerated development of atherosclerosis in ApoE-/- mice exposed to CAPs and have 

shown effects on coagulation, experimentally-induced hypertension, and vascular reactivity (Section 

7.2.1.2). Evidence from toxicological studies provides biological plausibility and coherence with 

studies of short-term exposure and cardiovascular morbidity and mortality, as well as with studies 

that examined long-term exposure to PM2.5 and cardiovascular mortality. Taken together, the 

evidence from epidemiologic and toxicological studies is sufficient to conclude that a causal 

relationship exists between long-term exposures to PM2.5 and cardiovascular effects.

Respiratory Effects 

Recent epidemiologic studies conducted in the U.S. and abroad provide evidence of 

associations between long-term exposure to PM2.5 and decrements in lung function growth, 

increased respiratory symptoms, and asthma development in study locations with mean PM2.5

concentrations ranging from 13.8 to 30 μg/m3 during the study periods (Section 7.3.1.1 and Section 

7.3.2.1). These results are supported by studies that observed associations between long-term 

exposure to PM10 and an increase in respiratory symptoms and reductions in lung function growth in

areas where PM

xidative 

asia 

2.5

10 is dominated by PM2.5. However, the evidence to support an association with 

long-term exposure to PM2.5 and respiratory mortality is limited (Figure 7-7). Subchronic and 

chronic toxicological studies of CAPs, DE, roadway air and woodsmoke provide coherence and 

biological plausibility for the effects observed in the epidemiologic studies. These toxicological 

studies have presented some evidence for altered pulmonary function, mild inflammation, o

responses, immune suppression, and histopathological changes including mucus cell hyperpl

(Section 7.3). Exacerbated allergic responses have been demonstrated in animals exposed to DE and 

wood smoke. In addition, pre- and postnatal exposure to ambient levels of urban particles was found 

to affect lung development in an animal model. This finding is important because impaired lung 

development is one mechanism by which PM exposure may decrease lung function growth in 

children. Collectively, the evidence from epidemiologic and toxicological studies is sufficient to 

conclude that a causal relationship is likely to exist between long-term exposures to PM

and respiratory effects. 

Mortality 

The recent epidemiologic literature reports associations between long-term PM2.5 exposure 

and increased risk of mortality. Mean PM2.5 concentrations ranged from 13.2 to 29 μg/m3

 during the 

study period in these areas (Section 7.6). When evaluating cause-specific mortality, the strongest 

evidence can be found when examining associations between PM2.5 and cardiovascular mortality, 

and positive associations were also reported between PM2.5 and lung cancer mortality (Figure 7-7). 

The cardiovascular mortality association has been confirmed further by the extended Harvard Six 

Cities and American Cancer Society studies, which both report strong associations between longterm exposure to PM2.5 and cardiopulmonary and IHD mortality (Figure 7-7). Additional new 

evidence from a study that used the WHI cohort found a particularly strong association between 

long-term exposure to PM2.5 and CVD mortality in post-menopausal women. Fewer studies have 

evaluated the respiratory component of cardiopulmonary mortality, and, as a result, the evidence to 

support an association with long-term exposure to PM2.5 and respiratory mortality is limited (Figure 

7-7). The evidence for cardiovascular and respiratory morbidity due to short- and long-term exposure 

to PM2.5 provides biological plausibility for cardiovascular- and respiratory-related mortality. 

Collectively, the evidence is sufficient to conclude that a causal relationship exists between 

long-term exposures to PM2.5 and mortality.

December 2009 2-12 
Reproductive and Developmental Effects 

Evidence is accumulating for PM2.5 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

 (Section 7.4), with effects becoming more precise and 

consistently positive in locations with mean PM2.5 concentrations of 15 μg/m3 and above 

(Section 7.4). Exposure to PM2.5 was usually associated with greater reductions in birth weight than 

exposure to PM10. The evidence from a few U.S. studies that investigated PM10 effects on fetal 

growth, which reported similar decrements in birth weight, provide consistency for the PM2.5

associations observed and strengthen the interpretation that particle exposure may be causally related 

to reductions in birth weight. The epidemiologic literature does not consistently report associations 

between long-term exposure to PM and preterm birth, growth restriction, birth defects or decreased 

sperm quality. Toxicological evidence supports an association between PM2.5 and PM10 exposure and 

adverse reproductive and developmental outcomes, but provide little mechanistic information or 

biological plausibility for an association between long-term PM exposure and adverse birth 

outcomes (e.g., low birth weight or infant mortality). 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 causal relationship between longterm exposures to PM2.5 and reproductive and developmental outcomes. 

Cancer, Mutagenicity, and Genotoxicity 

Multiple epidemiologic studies have shown a consistent positive association between PM2.5

and lung cancer mortality, but studies have generally not reported associations between PM2.5 and 

lung cancer incidence (Section 7.5). Animal toxicological studies have examined the potential 

relationship between PM and cancer, but have not focused on specific size fractions of PM. Instead 

they have examined ambient PM, wood smoke, and DEP. A number of 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. 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 examined epigenetic effects, and 

demonstrate that PM induces some changes in methylation. However, it has yet to be determined 

how these alterations in the genome could influence the initiation and promotion of cancer. 

Additionally, inflammation and immune suppression induced by exposure to PM may confer 

susceptibility to cancer. Collectively, the evidence from epidemiologic studies, primarily those of 

lung cancer mortality, along with the toxicological studies that show some evidence of the mutagenic 

and genotoxic effects of PM is suggestive of a causal relationship between long-term 

exposures to PM2.5 and cancer. 

2.3.2. Integration of PM2.5 Health Effects 

In epidemiologic studies, short-term exposure to PM2.5 is associated with a broad range of 

respiratory and cardiovascular effects, as well as mortality. For cardiovascular effects and mortality, 

the evidence supports the existence of a causal relationship with short-term PM2.5 exposure; while 

the evidence indicates that a causal relationship is likely to exist between short-term PM2.5 exposure 

and respiratory effects. The effect estimates from recent and older U.S. and Canadian-ba

epidemiologic studies that examined the relationship between short-term exposure to PM

sed 

2.5 and 

health outcomes with mean 24-h avg PM2.5 concentrations <17 μg/m3

 are shown in Figure 2-1. A 

number of different health effects are included in Figure 2-1 to provide an integration of the range of 

effects by mean concentration, with a focus on cardiovascular and respiratory effects and all-cause 

(nonaccidental) mortality (i.e., health effects categories with at least a suggestive causal 

determination). A pattern of consistent positive associations with mortality and morbidity effects can 

be seen in this figure. Mean PM2.5 concentrations ranged from 6.1 to 16.8 μg/m3

.in these study 

locations. 

December 2009 2-13 
Study Outcome Meana 98tha Effect Estimate (95% CI)

Chimonas & Gessner (2007, 093261) Asthma HA 6.1 - - - 

LRI HA 6.1 - - - 

Lisabeth et al. (2008, 155939) Ischemic Stroke/TIA HA 7.0e 23.6f

Slaughter et al. (2005, 073854) Asthma Exacerbation 7.3e - - - 

Rabinovitch et al. (2006, 088031) Asthma Medication Use 7.4 17.2f

Chen et al. (2004, 087262) COPD HA 7.7 - - - 

Chen et al. (2005, 087555) Respiratory HA 7.7 - - - 

Fung et al. (2006, 089789) Respiratory HA 7.7 - - - 

Villeneuve et al. (2003, 055051) Nonaccidental Mortality 7.9 - - - 

CVD ED Visits 8.5 27.3f Stieb et al. (2000, 011675) 

Respiratory ED Visits 8.5 27.3f

Hemhrgc Stroke HA 8.5 24.0f

Ischemic Stroke HA 8.5 24.0f Villeneuve et al. (2006, 090191) 

TIA HA 8.5 24.0f

Lin et al. (2005, 087828) RTI HA 9.6 - - - 

Respiratory Symptoms (any) 9.8c 25.8f Mar et al. (2004, 057309) 

Respiratory Symptoms (any) 9.8c 25.8f

Rich et al. (2005, 079620) Ventricular Arrhythmia 9.8e - - - 

Dockery et al. (2005, 078995) Ventricular Arrhythmia 10.3e - - - 

Rabinovitch et al. (2004, 096753) Asthma Exacerbation 10.6d 29.3f

Pope et al. (2006, 091246) IHD HA 10.7c - - - 

CVD HA 10.8 29.6f Slaughter et al. (2005, 073854) 

Respiratory ED Visits 10.8 29.6f

Pope et al. (2008, 191969) CHF HA 10.8 44.5d

MI HA 11.1e Zanobetti and Schwartz (2006, 090195) - - - 

Pneumonia HA 11.1e - - - 

Peters et al. (2001, 016546) MI 12.1 28.2f

Delfino et al. (1997, 082687) Respiratory HA (summer) 12.1 31.2f

Sullivan et al. (2005, 050854) MI 12.8 - - - 

Burnett et al. (2004, 086247) Nonaccidental Mortality 12.8 38.0f

Respiratory HA 12.9d 34.2f Bell et al. (2008, 156266) 

CVD HA 12.9d 34.2f

Wilson et al. (2007, 157149)

m CVD Mortality 13.0 31.6f

Zanobetti & Schwartz (2009, 188462) Nonaccidental Mortality 13.2d 34.3f

Burnett and Goldberg (2003, 042798) Nonaccidental Mortality 13.3 38.9f

CBVD HA 13.3 34.8f

PVD HA 13.3 34.8f

IHD HA 13.3 34.8f

Dysrhythmia HA 13.3 34.8f

CHF HA 13.3 34.8f

COPD HA 13.3 34.8f

Dominici et al. (2006, 088398) 

RTI HA 13.3 34.8f

Fairley (2003, 042850) Nonaccidental Mortality 13.6 59.0f

Zhang et al. (2009, 191970) ST Segment Depression 13.9j 37.6f

O’Connor et al. (2008, 156818) Wheeze/Cough 14.0c 39.0g

Klemm and Mason (2003, 042801) Nonaccidental Mortality 14.7e,i - - - 

Franklin et al. (2008, 097426) Nonaccidental mortality 14.8 43.0f

NYDOH (2006, 090132) Asthma ED Visits 15.0k - - - 

Ito et al. (2007, 156594) Asthma HA 15.1 39.0f

Franklin et al. (2007, 091257) Non-accidental Mortality 15.6 45.8f

Rich et al. (2006, 089814) Ventricular Arrhythmia 16.2e - - - 

Symons et al. (2006, 091258) CHF HA 16.5d 50.1f

Sheppard (2003, 042826) Asthma HA 16.7 46.6f

NYDOH (2006, 090132) Asthma ED Visits 16.7l - - - 

Respiratory HA (summer) 16.8 47.4f Burnett et al. (1997, 084194) 

CVD HA (summer) 16.8 47.4f

a

 μg/m3

b

 Study did not present mean; median presented. 

c

 Mean estimated from data in study. 

d

 Mean value slightly different from those reported in the published 

study or not reported in the published study; mean was either provided 

by study authors or calculated from data provided by study authors. e

 Mean value not reported in study; median presented. 

f

 98th percentile of PM2.5 distribution was either provided by study 

authors or calculated from data provided by study authors. g

 98th estimated from data in study. 

h

 Averaged annual values for years in study from data 

provided by study author. 

i

 Air quality data obtained from original study by 

Schwartz et al. (1996, 077325) j

 Mean PM2.5 concentration reported is for lag 0-2. k

 Bronx; TEOM data. 

l Manhattan; TEOM data. 

m Study does not present an overall effect estimate; the 

vertical lines represent the effect estimate for each of the 

areas of Phoenix examined. Relative Risk / Odds Ratio 

Figure 2-1. Summary of effect estimates (per 10 μg/m3) by increasing concentration from U.S. 

studies examining the association between short-term exposure to PM2.5 and 

cardiovascular and respiratory effects, and mortality, conducted in locations 

where the reported mean 24-h avg PM2.5 concentrations were <17 μg/m3. 

December 2009 2-14 
Long-term exposure to PM2.5 has been associated with health outcomes similar to those found 

in the short-term exposure studies, specifically for respiratory and cardiovascular effects and 

mortality. As found for short-term PM2.5 exposure, the evidence indicates that a causal relationship 

exists between long-term PM2.5 exposure and cardiovascular effects and mortality, and that a causal 

relationship is likely to exist between long-term PM2.5 exposure and effects on the respiratory 

system. 

Figure 2-2 highlights the findings of epidemiologic studies where the long-term mean PM2.5

concentrations were ≤ 29 μg/m3

. A range of health outcomes are displayed (including cardiovascular 

mortality, all-cause mortality, infant mortaltiy, and bronchitis) ordered by mean concentration. The 

range of mean PM2.5 concentrations in these studies was 10.7-29 μg/m3

 during the study periods. 

Additional studies not included in this figure that focus on subclinical outcomes, such as changes in 

lung function or atherosclerotic markers also report effects in areas with similar concentrations 

(Sections 7.2 and 7.3). Although not highlighted in the summary figure, long-term PM2.5 exposure 

studies also provide evidence for reproductive and developmental effects (i.e., low birth weight) and 

cancer (i.e., lung cancer mortality) in response to to exposure to PM2.5. 

Study Outcome Mean+ Effect Estimate (95% CI)

Zeger et al. (2008, 191951) All-Cause Mortality, Central U.S. 10.7 

Kim et al. (2004, 087383) Bronchitis (Children) 12.0 

Zeger et al. (2008, 191951) All-Cause Mortality, Western U.S. 13.1 

Miller et al. (2007, 090130) CVD Morbidity or Mortality 13.5 

Eftim et al. (2008, 099104) All-Cause Mortality, ACS Sites 13.6 

Goss et al. (2004, 055624) All-Cause Mortality 13.7 

McConnell et al. (2003, 049490) Bronchitis (Children) 13.8 

Zeger et al. (2008, 191951) All-Cause Mortality, Eastern U.S. 14.0 

Krewski et al. (2009, 191193) All-Cause Mortality 14.0 

Eftim et al. (2008, 099104) All-Cause Mortality, Harv 6-Cities 14.1 

Lipfert et al. (2006, 088756) All-Cause Mortality 14.3 

Dockery et al. (1996, 046219) Bronchitis (Children) 14.5 

Woodruff et al. (2008, 098386) Infant Mortality (Respiratory) 14.8 

Laden et al. (2006, 087605) All-Cause Mortality 16.4* 

Woodruff et al. (2008, 098386) Infant Mortality (Respiratory) 19.2 

Enstrom (2005, 087356) All-Cause Mortality 23.4 

Chen et al. (2005, 087942) CHD Mortality, Females 29.0 

CHD Mortality, Males 29.0 

 * Mean estimated from data in study 

 + μg/m3

Relative Risk 

Figure 2-2. Summary of effect estimates (per 10 μg/m3) by increasing concentration from U.S. 

studies examining the association between long-term exposure to PM2.5 and 

cardiovascular and respiratory effects, and mortality.

 

The observations from both the short- and long-term exposure studies are supported by 

experimental findings of PM2.5-induced subclinical and clinical cardiovascular effects. 

Epidemiologic studies have shown an increase in ED visits and hospital admissions for IHD upon 

exposure to PM2.5. These effects are coherent with the changes in vasomotor function and STsegment depression observed in both toxicological and controlled human exposure studies. It has 

been postulated that exposure to PM2.5 can lead to myocardial ischemia through an effect on the 

autonomic nervous system or by altering vasomotor function. PM-induced systemic inflammation, 

oxidative stress and/or endothelial dysfunction may contribute to altered vasomotor function. These 

effects have been demonstrated in recent animal toxicological studies, along with altered 

microvascular reactivity, altered vessel tone, and reduced blood flow during ischemia. Toxicological 

studies demonstrating increased right ventricular pressure and diminished cardiac contractility also 

provide biological plausibility for the associations observed between PM2.5 and CHF in 

epidemiologic studies. 

Thus, the overall evidence from the short-term epidemiologic, controlled human exposure, and 

toxicological studies evaluated provide coherence and biological plausibility for cardiovascular 

effects related to myocardial ischemia and CHF. Coherence in the cardiovascular effects observed 

December 2009 2-15 
can be found in long-term exposure studies, especially for CVDs among post-menopausal women. 

Additional studies provide limited evidence for subclinical measures of atherosclerosis in 

epidemiologic studies with stronger evidence from toxicological studies that have demonstrated 

accelerated development of atherosclerosis in ApoE-/- mice exposed to PM2.5 CAPs along with 

effects on coagulation, experimentally-induced hypertension, and vascular reactivity. Repeated acute 

responses to PM may lead to cumulative effects that manifest as chronic disease, such as 

atherosclerosis. Contributing factors to atherosclerosis development include systemic inflammation, 

endothelial dysfunction, and oxidative stress all of which are associated with PM2.5 exposure. 

However, it has not yet been determined whether PM initiates or promotes atherosclerosis. The 

evidence from both short- and long-term exposure studies on cardiovascular morbidity provide 

coherence and biological plausibility for the cardiovascular mortality effects observed when 

examining both exposure durations. In addition, cardiovascular hospital admission and mortality 

studies that examined the PM10 concentration-response relationship found evidence of a log-linear 

no-threshold relationship between PM exposure and cardiovascular-related morbidity (Section 6.2) 

and mortality (Section 6.5). 

Epidemiologic studies have also reported respiratory effects related to short-term exposure to 

PM2.5, which include increased ED visits and hospital admissions, as well as alterations in lung 

function and respiratory symptoms in asthmatic children. These respiratory effects were found to be 

generally robust to the inclusion of gaseous pollutants in copollutant models with the strongest 

evidence from the higher powered studies (Figure 6-9 and Figure 6-15). Consistent positive 

associations were also reported between short-term exposure to PM2.5 and respiratory mortality in 

epidemiologic studies. However, uncertainties exist in the PM2.5-respiratory mortality associations 

reported due to the limited number of studies that examined potential confounders of the PM2.5-

respiratory mortality relationship, and the limited information regarding the biological plausibility of 

the clinical and subclinical respiratory outcomes observed in the epidemiologic and controlled 

human exposure studies (Section 6.3) resulting in the progression to PM2.5-induced respiratory 

mortality. Important new findings, which support the PM2.5-induced respiratory effects mentioned 

above, include associations with post-neonatal (between 1 mo and 1 yr of age) respiratory mortality. 

Controlled human exposure studies provide some support for the respiratory findings from 

epidemiologic studies, with demonstrated increases in pulmonary inflammation following short-term 

exposure. However, there is limited and inconsistent evidence of effects in response to controlled 

exposures to PM2.5 on respiratory symptoms or pulmonary function among healthy adults or adults 

with respiratory disease. Long-term exposure epidemiologic studies provide additional evidence for 

PM2.5-induced respiratory morbidity, but little evidence for an association with respiratory mortality. 

These epidemiologic morbidity studies have found decrements in lung function growth, as well as 

increased respiratory symptoms, and asthma. Toxicological studies provide coherence and biological 

plausibility for the respiratory effects observed in response to short and long-term exposures to PM 

by demonstrating a wide array of biological responses including: altered pulmonary function, mild 

pulmonary inflammation and injury, oxidative responses, and histopathological changes in animals 

exposed by inhalation to PM2.5 derived from a wide variety of sources. In some cases, prolonged 

exposures led to adaptive responses. Important evidence was also found in an animal model for 

altered lung development following pre- and post-natal exposure to urban air, which may provide a 

mechanism to explain the reduction in lung function growth observed in children in response to 

long-term exposure to PM. 

Additional respiratory-related effects have been tied to allergic responses. Epidemiologic 

studies have provided evidence for increased hospital admissions for allergic symptoms (e.g., 

allergic rhinitis) in response to short- and long-term exposure to PM2.5. Panel studies also positively 

associate long-term exposure to PM2.5 and PM10 with indicators of allergic sensitization. Controlled 

human exposure and toxicological studies provide coherence for the exacerbation of allergic 

symptoms, by showing that PM2.5 can promote allergic responses and intensify existing allergies. 

Allergic responses require repeated exposures to antigen over time and co-exposure to an adjuvant 

(possibly DE particles or UF CAPs) can enhance this response. Allergic sensitization often underlies 

allergic asthma, characterized by inflammation and AHR. In this way, repeated or chronic exposures 

involving multifactorial responses (immune system activation, oxidative stress, inflammation) can 

lead to irreversible outcomes. Epidemiologic studies have also reported evidence for increased 

hospital admissions for respiratory infections in response to both short- and long-term exposures to 

PM2.5. Toxicological studies suggest that PM impairs innate immunity, which is the first line of 

December 2009 2-16 
defense against infection, providing coherence for the respiratory infection effects observed in 

epidemiologic studies. 

The difference in effects observed across studies and between cities may be attributed, at least 

in part, to the differences in PM composition across the U.S. Differences in PM toxicity may result 

from regionally varying PM composition and size distribution, which in turn reflects differences in 

sources and PM volatility. A person’s exposure to ambient PM will also vary due to regional 

differences in personal activity patterns, microenvironmental characteristics and the spatial 

variability of PM concentrations in urban areas. Regional differences in PM2.5 composition are 

outlined briefly in Section 2.1 above and in more detail in Section 3.5. An examination of data from 

the CSN indicates that East-West gradients exist for a number of PM components. Specifically, SO4

2-

concentrations are higher in the East, OC constitutes a larger fraction of PM in the West, and NO3

-

concentrations are highest in the valleys of central California and during the winter in the Midwest. 

However, the available evidence and the limited amount of city-specific speciated PM2.5 data does 

not allow conclusions to be drawn that specifically differentiate effects of PM in different locations. 

It remains a challenge to determine relationships between specific constituents, combinations 

of constituents, or sources of PM2.5 and the various health effects observed. Source apportionment 

studies of PM2.5 have attempted to decipher some of these relationships and in the process have 

identified associations between multiple sources and various respiratory and cardiovascular health 

effects, as well as mortality. Although different source apportionment methods have been used across 

these studies, the methods used have been evaluated and found generally to identify the same 

sources and associations between sources and health effects (Section 6.6). While uncertainty 

remains, it has been recognized that many sources and components of PM2.5 contribute to health 

effects. Overall, the results displayed in Table 6-18 indicate that many constituents of PM2.5 can be 

linked with multiple health effects, and the evidence is not yet sufficient to allow differentiation of 

those constituents or sources that are more closely related to specific health outcomes. 

Variability in the associations observed across PM2.5 epidemiologic studies may be due in part 

to exposure error related to the use of county-level air quality data. Because western U.S. counties 

tend to be much larger and more topographically diverse than eastern U.S. counties, the day-to-day 

variations in concentration at one site, or even for the average of several sites, may not correlate well 

with the day-to-day variations in all parts of the county. For example, site-to-site correlations as a 

function of distance between sites (Section 3.5.1.2) fall off rapidly with distance in Los Angeles, but 

high correlations extend to larger distances in eastern cities such as Boston and Pittsburgh. These 

differences may be attributed to a number of factors including topography, the built environment, 

climate, source characteristics, ventilation usage, and personal activity patterns. For instance, 

regional differences in climate and infrastructure can affect time spent outdoors or indoors, air 

conditioning usage, and personal activity patterns. Characteristics of housing stock may also cause 

regional differences in effect estimates because new homes tend to have lower infiltration factors 

than older homes. Biases and uncertainties in exposure estimates resulting from these aspects can, in 

turn, cause bias and uncertainty in associated health effects estimates. 

The new evidence reviewed in this ISA greatly expands upon the evidence available in the 

2004 PM AQCD particularly in providing greater understanding of the underlying mechanisms for 

PM2.5 induced cardiovascular and respiratory effects for both short- and long-term exposures. Recent 

studies have provided new evidence linking long-term exposure to PM2.5 with cardiovascular 

outcomes that has expanded upon the continuum of effects ranging from the more subtle subclinical 

measures to cardiopulmonary mortality. 

December 2009 2-17 
2.3.3. Exposure to PM10-2.5

2.3.3.1. Effects of Short-Term Exposure to PM10-2.5

Table 2-3. Summary of causal determinations for short-term exposure to PM10-2.5 . 

Size Fraction Outcome Causality Determination 

Cardiovascular Effects Suggestive 

PM10-2.5 Respiratory Effects Suggestive 

Mortality Suggestive 

Cardiovascular Effects 

Generally positive associations were reported between short-term exposure to PM10-2.5 and 

hospital admissions or ED visits for cardiovascular causes. These results are supported by a large 

U.S. multicity study of older adults that reported PM10-2.5 associations with CVD hospital 

admissions, and only a slight reduction in the PM10-2.5 risk estimate when included in a copollutant 

model with PM2.5 (Section 6.2.10). The PM10-2.5 associations with cardiovascular hospital admissions 

and ED visits were observed in study locations with mean 24-h avg PM10-2.5 concentrations ranging 

from 7.4 to 13 μg/m3

. These results are supported by the associations observed between PM10-2.5 and 

cardiovascular mortality in areas with 24-h avg PM10-2.5 concentrations ranging from 6.1-16.4 μg/m3

(Section 6.2.11). The results of the epidemiologic studies were further confirmed by studies that 

examined dust storm events, which contain high concentrations of crustal material, and found an 

increase in cardiovascular-related ED visits and hospital admissions. Additional epidemiologic 

studies have reported PM10-2.5 associations with other cardiovascular health effects including 

supraventricular ectopy and changes in HRV (Section 6.2.1.1). Although limited in number, studies 

of controlled human exposures provide some evidence to support the alterations in HRV observed in 

the epidemiologic studies (Section 6.2.1.2). The few toxicological studies that examined the effect of 

PM10-2.5 on cardiovascular health effects used IT instillation due to the technical challenges in 

exposing rodents via inhalation to PM10-2.5, and, as a result, provide only limited evidence on the 

biological plausibility of PM10-2.5 induced cardiovascular effects. The potential for PM10-2.5 to elicit 

an effect is supported by dosimetry studies, which show that a large proportion of inhaled particles 

the 3-6 micron (d

in 

ts. 

ae) range can reach and deposit in the lower respiratory tract, particularly the 

tracheobronchial (TB) airways (Figures 4-3 and 4-4). Collectively, the evidence from epidemiologic 

studies, along with the more limited evidence from controlled human exposure and toxicological 

studies is suggestive of a causal relationship between short-term exposures to PM10-2.5

and cardiovascular effec

Respiratory Effects 

A number of recent epidemiologic studies conducted in Canada and France found consistent, 

positive associations between respiratory ED visits and hospital admissions and short-term exposure 

to PM10-2.5 in studies with mean 24-h avg concentrations ranging from 5.6-16.2 μg/m3 (Section 6.3.8). 

In these studies, the strongest relationships were observed among children, with less consistent 

evidence for adults and older adults (i.e., ≥ 65). In a large multicity study of older adults, PM10-2.5

was positively associated with respiratory hospital admissions in both single and copollutant models 

with PM2.5. In addition, a U.S.-based multicity study found evidence for an increase in respiratory 

mortality upon short-term exposure to PM10-2.5, but these associations have not been consistently 

December 2009 2-18 
observed in single-city studies (Section 6.3.9). A limited number of epidemiologic studies have 

focused on specific respiratory morbidity outcomes, and found no evidence of an association with 

lower respiratory symptoms, wheeze, and medication use (Section 6.3.1.1). While controlled human 

exposure studies have not observed an effect on lung function or respiratory symptoms in healthy or 

asthmatic adults in response to short-term exposure to PM10-2.5, healthy volunteers have exhibited an 

increase in markers of pulmonary inflammation. Toxicological studies using inhalation exposures are 

still lacking, but pulmonary injury has been observed in animals after IT instillation exposure 

(Section 6.3.5.3). In some cases, PM10-2.5 was found to be more potent than PM2.5 and effects were 

not attributable to endotoxin. Both rural and urban PM10-2.5 have induced inflammation and injury 

responses in rats or mice exposed via IT instillation, making it difficult to distinguish the health 

effects of PM10-2.5 from different environments. Overall, epidemiologic studies, along with the 

limited number of controlled human exposure and toxicological studies that examined PM10-2.5 

respiratory effects provide evidence that is suggestive of a causal relationship between shortterm exposures to PM 10-2.5 and respiratory effects. 

Mortality 

The majority of studies evaluated in this review provide some evidence for mortality 

associations with PM10-2.5 in areas with mean 24-h avg concentrations ranging from 6.1-16.4 μg/m3. However, uncertainty surrounds the PM10-2.5 associations reported in the studies evaluated due to the 

different methods used to estimate PM10-2.5 concentrations across studies (e.g., direct measurement 

of PM10-2.5 using dichotomous samplers, calculating the difference between PM10 and PM2.5

concentrations). In addition, only a limited number of PM10-2.5 studies have investigated potential 

confounding by gaseous copollutants or the influence of model specification on PM10-2.5 risk 

estimates. 

A new U.S.-based multicity study, which estimated PM10-2.5 concentrations by calculating the 

difference between the county-average PM10 and PM2.5, found associations between PM10-2.5 and 

mortality across the U.S., including evidence for regional variability in PM10-2.5 risk estimates 

(Section 6.5.2.3). Additionally, the U.S.-based multicity study provides preliminary evidence for 

greater effects occurring during the warmer months (i.e., spring and summer). A multicity Canadian 

study provides additional evidence for an association between short-term exposure to PM10-2.5 and 

mortality (Section 6.5.2.3). Although consistent positive associations have been observed across both 

multi- and single-city studies, more data are needed to adequately characterize the chemical and 

biological components that may modify the potential toxicity of PM10-2.5 and compare the different 

methods used to estimate exposure. Overall, the evidence evaluated is suggestive of a causal 

relationship between short-term exposures to PM10-2.5 and mortality.

2.3.4. Integration of PM10-2.5 Effects 

Epidemiologic, controlled human exposure, and toxicological studies have provided evidence that is 

suggestive for relationships between short-term exposure to PM10-2.5 and cardiovascular effects, 

respiratory effects, and mortality. Conclusions regarding causation for the various health effects and 

outcomes were made for PM10-2.5 as a whole regardless of origin, since PM10-2.5-related effects have 

been demonstrated for a number of different environments (e.g., cities reflecting a wide range of 

environmental conditions). Associations between short-term exposure to PM10-2.5 and cardiovascular 

and respiratory effects, and mortality have been observed in locations with mean PM10-2.5

concentrations ranging from 5.6 to 33.2 μg/m3

, and maximum PM10-2.5 concentrations ranging from 

24.6 to 418.0 μg/m3) (Figure 2-3). A number of different health effects are included in Figure 2-3 to 

provide an integration of the range of effects by mean concentration, with a focus on cardiovascular 

and respiratory effects, and mortality (i.e., health effects categories with at least a suggestive causal 

determination). To date, a sufficient amount of evidence does not exist in order to draw conclusions 

regarding the health effects and outcomes associated with long-term exposure to PM10-2.5. 

In epidemiologic studies, associations between short-term exposure to PM10-2.5 and 

cardiovascular outcomes (i.e., IHD hospital admissions, supraventricular ectopy, and changes in 

HRV) have been found that are similar in magnitude to those observed in PM2.5 studies. Controlled 

human exposure studies have also observed alterations in HRV, providing consistency and coherence 

December 2009 2-19 
for the effects observed in the epidemiologic studies. To date, only a limited number of toxicological 

studies have been conducted to examine the effects of PM10-2.5 on cardiovascular effects. All of these 

studies involved IT instillation due to the technical challenges of using PM10-2.5 for rodent inhalation 

studies. As a result, the toxicological studies evaluated provide limited biological plausibility for the 

PM10-2.5 effects observed in the epidemiologic and controlled human exposure studies. 

Study Outcome Meana Maxa Effect Estimate (95% CI)

Chen et al. (2004, 087262) COPD HA 5.6 24.6 

Fung et al. (2006, 089789) RD HA 5.6 27.1 

Chen et al. (2005, 087942) RD HA 5.6 24.6 

Villeneuve et al. (2003, 055051) Nonaccidental Mortality 6.1 72.0 

Lipfert et al. (2000, 004088) CVD Mortality 6.9d 28.3 

Peters et al. (2001, 016546) MI 7.4 - - - 

Tolbert et al. (2007, 090316) CVD ED Visits 9.0 50.3 

RD ED Visits 9.0 50.3 

Klemm et al. (2003, 042801) Nonaccidental Mortality 9.0b 30.0 

Metzger et al. (2007, 092856) Ventricular Arrhythmia 9.6 50.3 

Asthma ED Visits 9.7 34.2e

COPD ED Visits 9.7 34.2e

RD ED Visits 9.7 34.2e

Pneumonia ED Visits 9.7 34.2e

Peel et al. (2005, 056305) 

URI ED Visits 9.7 34.2e

CHF ED Visits 9.7d 34.2e Metzger et al. (2004, 044222) 

IHD ED Visits 9.7d 34.2e

Klemm et al. (2004, 056585) Nonaccidental Mortality 9.9 25.2 

Symptoms (any) 10.8c 50.9e Mar et al. (2004, 057309) 

Asthma Symptoms 10.8c 50.9e

Lin et al. (2005, 087828) RTI HA 10.9 45.0 

Burnett et al. (2004, 086247) Non-accidental Mortality 11.4 151.0 

CVD HA 11.5d Burnett et al. (1997, 084194) 56.1 

Respiratory HA 11.5d 56.1 

Fairley (2003, 042850) Nonaccidental Mortality 11.7d 55.2 

Zanobetti & Schwartz (2009, 188462) Nonaccidental Mortality 11.8 88.3e

Lin et al. (2002, 026067) Asthma HA (boys) 12.2 68.0 

Lin et al. (2002, 026067; 2004, 056067) Asthma HA (girls) 12.2 68.0 

RD HA 12.3d 81.3e Peng et al. (2008, 156850) 

CVD HA 12.3d 81.3e

Burnett and Goldberg (2003, 042798) Nonaccidental Mortality 12.6 99.0 

Nonaccidental Mortality 13.3d 50.0 

CHF HA 13.3d 50.0 

IHD HA 13.3d 50.0 

COPD HA 13.3d 50.0 

Ito (2003, 042856) 

Pneumonia HA 13.3d 50.0 

Thurston et al. (1994, 043921) Respiratory HA 14.4c 33.0 

Sheppard (2003, 042826) Asthma HA 16.2 88.0 

Ostro et al. (2003, 042824) CVD Mortality 30.5 418.0 

Mar et al. (2003, 042841) CVD Mortality 33.2 158.6 a

 μg/m3

b

 Study did not present mean; median presented. 

c

 Mean estimated from data in study. 

d

 Mean value slightly different from those reported in the published study; mean was either 

provided by study authors or calculated from data provided by study authors. e

 Maximum PM10-2.5 concentration provided by study authors or calculated from data provided by 

study authors. 

Relative Risk / Odds Ratio 

Figure 2-3. Summary of U.S. studies examining the association between short-term 

exposure to PM10-2.5 and cardiovascular morbidity/mortality and respiratory 

morbidity/mortality. All effect estimates have been standardized to reflect a 

10 μg/m3 increase in mean 24-h avg PM10-2.5 concentration and ordered by 

increasing concentration. 

Limited evidence is available from epidemiologic studies for respiratory health effects and 

outcomes in response to short-term exposure to PM10-2.5. An increase in respiratory hospital 

admissions and ED visits has been observed, but primarily in studies conducted in Canada and 

Europe. In addition, associations are not reported for lower respiratory symptoms, wheeze, or 

medication use. Controlled human exposure studies have not observed an effect on lung function or 

respiratory symptoms in healthy or asthmatic adults, but healthy volunteers have exhibited 

pulmonary inflammation. The toxicological studies (all IT instillation) provide evidence of 

December 2009 2-20 
pulmonary injury and inflammation. In some cases, PM10-2.5 was found to be more potent than PM2.5

and effects were not solely attributable to endotoxin. 

Currently, a national network is not in place to monitor PM10-2.5 concentrations. As a result, 

uncertainties surround the concentration at which the observed associations occur. Ambient 

concentrations of PM10-2.5 are generally determined by the subtraction of PM10 and PM2.5

measurements, using various methods. For example, some epidemiologic studies estimate PM10-2.5

by taking the difference between collocated PM10 and PM2.5 monitors while other studies have taken 

the difference between county average PM10 and PM2.5 concentrations. Moreover, there are potential 

differences among operational flow rates and temperatures for PM10 and PM2.5 monitors used to

calculate PM10-2.5. Therefore, there is greater error in ambient exposure to PM10-2.5 compared to 

PM2.5. This would tend to increase uncertainty and make it more difficult to detect effects of PM10-2.5

2.3.2). 

2.3.

in epidemiologic studies. In addition, the various differences between eastern and western U.S. 

counties can lead to exposure misclassification, and the potential underestimation of effects in 

western counties (as discussed for PM2.5 in Section

It is also important to note that the chemical composition of PM10-2.5 can vary considerably by 

location, but city-specific speciated PM10-2.5 data are limited. PM10-2.5 may contain Fe, Si, Al, and 

base cations from soil, plant and insect fragments, pollen, fungal spores, bacteria, and viruses, as 

well as fly ash, brake lining particles, debris, and automobile tire fragments. 

The 2004 PM AQCD presented the limited amount of evidence available that examined the 

potential association between exposure to PM10-2.5 and health effects and outcomes. The current 

evidence, primarily from epidemiologic studies, builds upon the results from the 2004 PM AQCD 

and indicates that short-term exposure to PM10-2.5 is associated with effects on both the 

cardiovascular and respiratory systems. However, variability in the chemical and biological 

composition of PM10-2.5, limited evidence regarding effects of the various components of PM10-2.5, 

and lack of clearly defined biological mechanisms for PM10-2.5-related effects are important sources 

of uncertainty. 

2.3.5. Exposure to UFPs 

5.1. Effects of Short-Term Exposure to UFPs 

Table 2-4. Summary of causal determinations for short-term exposure to UFPs. 

Size Fraction Outcome Causality Determination 

Cardiovascular Effects Suggestive 

UFPs 

Respiratory Effects Suggestive 

Cardiovascular Effects 

Controlled human exposure studies provide the majority of the evidence for cardiovascular 

health effects in response to short-term exposure to UFPs. While there are a limited number of 

studies that have examined the association between UFPs and cardiovascular morbidity, there is a 

larger body of evidence from studies that exposed subjects to fresh DE, which is typically dominated 

by UFPs. These studies have consistently demonstrated changes in vasomotor function following 

exposure to atmospheres containing relatively high concentrations of particles (Section 6.2.4.2). 

Markers of systemic oxidative stress have also been observed to increase after exposure to various 

particle types that are predominantly in the UFP size range. In addition, alterations in HRV 

parameters have been observed in response to controlled human exposure to UF CAPs, with 

inconsistent evidence for changes in markers of blood coagulation following exposure to UF CAPs 

December 2009 2-21 
and DE (Sections 6.2.1.2 and 6.2.8.2). A few toxicological studies have also found consistent 

changes in vasomotor function, which provides coherence with the effects demonstrated in the 

controlled human exposure studies (Section 6.2.4.3). Additional UFP-induced effects observed in 

toxicological studies include alterations in HRV, with less consistent effects observed for systemic 

inflammation and blood coagulation. Only a few epidemiologic studies have examined the effect of 

UFPs on cardiovascular morbidity and collectively they found inconsistent evidence for an 

association between UFPs and CVD hospital admissions, but some positive associations for 

subclinical cardiovascular measures (i.e., arrhythmias and supraventricular beats) (Section 6.2.2.1). 

These studies were conducted in the U.S. and Europe in areas with mean particle number 

concentration ranging from ~8,500 to 36,000 particles/cm3

. However, UFP number concentrations 

are highly variable (i.e., concentrations drop off quickly from the road compared to accumulation 

mode particles), and therefore, more subject to exposure error than accumulation mode particles. In 

conclusion, the evidence from the studies evaluated is suggestive of a causal relationship 

between short-term exposures to UFPs and cardiovascular effects.

Respiratory Effects 

A limited number of epidemiologic studies have examined the potential association between 

short-term exposure to UFPs and respiratory morbidity. Of the studies evaluated, there is limited, and 

inconsistent evidence for an association between short-term exposure to UFPs and respiratory 

symptoms, as well as asthma hospital admissions in locations a median particle number 

concentration of ~6,200 to a mean of 38,000 particles/cm3 (Section 6.3.10). The spatial and temporal 

variability of UFPs also affects these associations. Toxicological studies have reported respiratory 

effects including oxidative, inflammatory, and allergic responses using a number of different UFP 

types (Section 6.3). Although controlled human exposure studies have not extensively examined the 

effect of UFPs on respiratory outcomes, a few studies have observed small UFP-induced 

asymptomatic decreases in pulmonary function. Markers of pulmonary inflammation have been 

observed to increase in healthy adults following controlled exposures to UFPs, particularly in studies 

using fresh DE. However, it is important to note that for both controlled human exposure and animal 

toxicological studies of exposures to fresh DE, the relative contributions of gaseous copollutants to 

the respiratory effects observed remain unresolved. Thus, the current collective evidence is 

suggestive of a causal relationship between short-term exposures to UFPs and 

respiratory effects. 

2.3.6. Integration of UFP Effects 

The controlled human exposure studies evaluated have consistently demonstrated effects on 

vasomotor function and systemic oxidative stress with additional evidence for alterations in HRV 

parameters in response to exposure to UF CAPs. The toxicological studies provide coherence for the 

changes in vasomotor function observed in the controlled human exposure studies. Epidemiologic 

studies are limited because a national network is not in place to measure UFP in the U.S. UFP 

concentrations are spatially and temporally variable, which would increase uncertainty and make it 

difficult to detect associations between health effects and UFPs in epidemiologic studies. In addition, 

data on the composition of UFPs, the spatial and temporal evolution of UFP size distribution and 

chemical composition, and potential effects of UFP constituents are sparse. 

More limited evidence is available regarding the effect of UFPs on respiratory effects. 

Controlled human exposure studies have not extensively examined the effect of UFPs on respiratory 

measurements, but a few studies have observed small decrements in pulmonary function and 

increases in pulmonary inflammation. Additional effects including oxidative, inflammatory, and proallergic outcomes have been demonstrated in toxicological studies. Epidemiologic studies have 

found limited and inconsistent evidence for associations between UFPs and respiratory effects. 

Overall, a limited number of studies have examined the association between exposure to UFPs 

and morbidity and mortality. Of the studies evaluated, controlled human exposure and toxicological 

studies provide the most evidence for UFP-induced cardiovascular and respiratory effects; however, 

many studies focus on exposure to DE. As a result, it is unclear if the effects observed are due to 

UFP, larger particles (i.e., PM2.5), or the gaseous components of DE. Additionally, UF CAPs systems 

December 2009 2-22 
are limited as the atmospheric UFP composition is modified when concentrated, which adds 

uncertainty to the health effects observed in controlled human exposure studies (Section 1.5.3). 

2.4. Policy Relevant Considerations 

2.4.1. Potentially Susceptible Populations 

Upon evaluating the association between short- and long-term exposure to PM and various 

health outcomes, studies also attempted to identify populations that are more susceptible to PM (i.e., 

populations that have a greater likelihood of experiencing health effects related to exposure to an air 

pollutant (e.g., PM) due to a variety of factors including, but not limited to: genetic or developmental 

factors, race, gender, life stage, lifestyle (e.g., smoking status and nutrition) or preexisting disease; as 

well as, population-level factors that can increase an individual's exposure to an air pollutant (e.g., 

PM) such as socioeconomic status [SES], which encompasses reduced access to health care, low 

educational attainment, residential location, and other factors). These studies did so by conducting 

stratified analyses; by examining effects in individuals with an underlying health condition; or by 

developing animal models that mimic the pathophysiologic conditions associated with an adverse 

health effect. In addition, numerous studies that focus on only one potentially susceptible population 

provide supporting evidence on whether a population is susceptible to PM exposure. These studies 

identified a multitude of factors that could potentially contribute to whether an individual is 

susceptible to PM (Table 8-2). Although studies have primarily used exposures to PM2.5 or PM10, the 

available evidence suggests that the identified factors may also enhance susceptibility to PM10-2.5. 

The examination of susceptible populations to PM exposure allows for the NAAQS to provide an 

adequate margin of safety for both the general population and for susceptible populations. 

During specific periods of life (i.e., childhood and advanced age), individuals may be more 

susceptible to environmental exposures, which in turn can render them more susceptible to PMrelated health effects. An evaluation of age-related health effects suggests that older adults have 

heightened responses for cardiovascular morbidity with PM exposure. In addition, epidemiologic 

and toxicological studies provide evidence that indicates children are at an increased risk of PMrelated respiratory effects. It should be noted that the health effects observed in children could be 

initiated by exposures to PM that occurred during key windows of development, such as in utero. 

Epidemiologic studies that focus on exposures during development have reported inconsistent 

findings (Section 7.4), but a recent toxicological study suggests that inflammatory responses in 

pregnant women due to exposure to PM could result in health effects in the developing fetus. 

Epidemiologic studies have also examined whether additional factors, such as gender, race, or 

ethnicity modify the association between PM and morbidity and mortality outcomes. Although 

gender and race do not seem to modify PM risk estimates, limited evidence from two studies 

conducted in California suggest that Hispanic ethnicity may modify the association between PM and 

mortality. 

Recent epidemiologic and toxicological studies provided evidence that individuals with null 

alleles or polymorphisms in genes that mediate the antioxidant response to oxidative stress (i.e., 

GSTM1), regulate enzyme activity (i.e., MTHFR and cSHMT), or regulate levels of procoagulants 

(i.e., fibrinogen) are more susceptible to PM exposure. However, some studies have shown that 

polymorphisms in genes (e.g., HFE) can have a protective effect against effects of PM exposure. 

Additionally, preliminary evidence suggests that PM exposure can impart epigenetic effects (i.e., 

DNA methylation); however, this requires further investigation. 

Collectively, the evidence from epidemiologic and toxicological, and to a lesser extent, 

controlled human exposure studies, indicate increased susceptibility of individuals with underlying 

CVDs and respiratory illnesses (i.e., asthma) to PM exposure. Controlled human exposure and 

toxicological studies provide additional evidence for increased PM-related cardiovascular effects in 

individuals with underlying respiratory health conditions. 

Recently studies have begun to examine the influence of preexisting chronic inflammatory 

conditions, such as diabetes and obesity, on PM-related health effects. These studies have found 

some evidence for increased associations for cardiovascular outcomes along with pathophysiologic 

alterations in markers of inflammation, oxidative stress, and acute phase response. However, more 

December 2009 2-23 
research is needed to thoroughly examine the affect of PM exposure on obese individuals and to 

identify the biological pathway(s) that could increase the susceptibility of diabetic and obese 

individuals to PM. 

There is also evidence that SES, measured using surrogates such as educational attainment or 

residential location, modifies the association between PM and morbidity and mortality outcomes. In 

addition, nutritional status, another surrogate measure of SES, has been shown to have protective 

effects against PM exposure in individuals that have a higher intake of some vitamins and nutrients. 

Overall, the epidemiologic, controlled human exposure, and toxicological studies evaluated in 

this review provide evidence for increased susceptibility for various populations, including children 

and older adults, people with pre-existing cardiopulmonary diseases, and people with lower SES. 

2.4.2. Lag Structure of PM-Morbidity and PM-Mortality Associations 

Epidemiologic studies have evaluated the time-frame in which exposure to PM can impart a 

health effect. PM exposure-response relationships can potentially be influenced by a multitude of 

factors, such as the underlying susceptibility of an individual (e.g., age, pre-existing diseases), which 

could increase or decrease the lag times observed. 

An attempt has been made to identify whether certain lag periods are more strongly associated 

with specific health outcomes. The epidemiologic evidence evaluated in the 2004 PM AQCD 

supported the use of lags of 0-1 days for cardiovascular effects and longer moving averages or 

distributed lags for respiratory diseases (U.S. EPA, 2004, 056905). However, currently, little 

consensus exists as to the most appropriate a priori lag times to use when examining morbidity and 

mortality outcomes. As a result, many investigators have chosen to examine the lag structure of 

associations between PM concentration and health outcome instead of focusing on a priori lag times. 

This approach is informative because if effects are cumulative, higher overall risks may exist than 

would be observed for any given single-day lag. 

2.4.

2.4.

2.1. PM-Cardiovascular Morbidity Associations 

Most of the studies evaluated that examined the association between cardiovascular hospital 

admissions and ED visits report associations with short-term PM exposure at lags 0- to 2-days, with 

more limited evidence for shorter durations (i.e., hours) between exposure and response for some 

health effects (e.g., onset of MI) (Section 6.2.10). However, these studies have rarely examined 

alternative lag structures. Controlled human exposure and toxicological studies provide biological 

plausibility for the health effects observed in the epidemiologic studies at immediate or concurrent 

day lags. Although the majority of the evidence supports shorter lag times for cardiovascular health 

effects, a recent study has provided preliminary evidence suggesting that longer lag times (i.e., 14-

day distributed lag model) may be plausible for non-ischemic cardiovascular conditions 

(Section 6.2.10). Panel studies of short-term exposure to PM and cardiovascular endpoints have also 

examined the time frame from exposure to health effect using a wide range of lag times. Studies of 

ECG changes indicating ischemia show effects at lags from several hours to 2 days, while lag times 

ranging from hours to several week moving averages have been observed in studies of arrhythmias, 

vasomotor function and blood markers of inflammation, coagulation and oxidative stress 

(Section 6.2). The longer lags observed in these panel studies may be explained if the effects of PM 

are cumulative. Although few studies of cumulative effects have been conducted, toxicological 

studies have demonstrated PM-dependent progression of atherosclerosis. It should be noted that PM 

exposure could also lead to an acute event (e.g., infarction or stroke) in individuals with 

atherosclerosis that may have progressed in response to cumulative PM exposure. Therefore, effects 

have been observed at a range of lag periods from a few hours to several days with no clear evidence 

for any lag period having stronger associations then another. 

2.2. PM-Respiratory Morbidity Associations 

Generally, recent studies of respiratory hospital admissions that evaluate multiple lags, have 

found effect sizes to be larger when using longer moving averages or distributed lag models. For 

example, when examining hospital admissions for all respiratory diseases among older adults, the 

strongest associations were observed when using PM concentrations 2 days prior to the hospital 

December 2009 2-24 
admission (Section 6.3.8). Longer lag periods were also found to be most strongly associated with 

asthma hospital admissions and ED visits in children (3-5 days) with some evidence for more 

immediate effects in older adults (lags of 0 and 1 day), but these observations were not consistent 

across studies (Section 6.3.8). These variable results could be due to the biological complexity of 

asthma, which inhibits the identification of a specific lag period. The longer lag times identified in 

the epidemiologic studies evaluated are biologically plausible considering that PM effects on allergic 

sensitization and lung immune defenses have been observed in controlled human exposure and 

toxicological studies. These effects could lead to respiratory illnesses over a longer time course (e.g., 

within several days respiratory infection may become evident, resulting in respiratory symptoms or a 

hospital admission). However, inflammatory responses, which contribute to some forms of asthma, 

may result in symptoms requiring medical care within a shorter time frame (e.g., 0-1 days). 

2.4.2.3. PM-Mortality Associations 

Epidemiologic studies that focused on the association between short-term PM exposure and 

mortality (i.e., all-cause, cardiovascular, and respiratory) mostly examined a priori lag structures of 

either 1 or 0-1 days. Although mortality studies do not often examine alternative lag structures, the 

selection of the aforementioned a priori lag days has been confirmed in additional studies, with the 

strongest PM-mortality associations consistently being observed at lag 1 and 0-1-days (Section 6.5). 

However, of note is recent evidence for larger effect estimates when using a distributed lag model. 

Epidemiologic studies that examined the association between long-term exposure to PM and 

mortality have also attempted to identify the latency period from PM exposure to death 

(Section 7.6.4). Results of the lag comparisons from several cohort studies indicate that the effects of 

changes in exposure on mortality are seen within five years, with the strongest evidence for effects 

observed within the first two years. Additionally, there is evidence, albeit from one study, that the 

mortality effect had larger cumulative effects spread over the follow-up year and three preceding 

years. 

2.4.3. PM Concentration-Response Relationship 

An important consideration in characterizing the PM-morbidity and mortality association is 

whether the concentration-response relationship is linear across the full concentration range that is 

encountered or if there are concentration ranges where there are departures from linearity 

(i.e., nonlinearity). In this ISA studies have been identified that attempt to characterize the shape of 

the concentration-response curve along with possible PM “thresholds” (i.e., levels which PM 

concentrations must exceed in order to elicit a health response). The epidemiologic studies evaluated 

that examined the shape of the concentration-response curve and the potential presence of a 

threshold have focused on cardiovascular hospital admissions and ED visits and mortality associated 

with short-term exposure to PM10 and mortality associated with long-term exposure to PM2.5. 

A limited number of studies have been identified that examined the shape of the PMcardiovascular hospital admission and ED visit concentration-response relationship. Of these studies, 

some conducted an exploratory analysis during model selection to determine if a linear curve most 

adequately represented the concentration-response relationship; whereas, only one study conducted 

an extensive analysis to examine the shape of the concentration-response curve at different 

concentrations (Section 6.2.10.10). Overall, the limited evidence from the studies evaluated supports 

the use of a no-threshold, log-linear model, which is consistent with the observations made in studies 

that examined the PM-mortality relationship. 

Although multiple studies have previously examined the PM-mortality concentration-response 

relationship and whether a threshold exists, more complex statistical analyses continue to be 

developed to analyze this association. Using a variety of methods and models, most of the studies 

evaluated support the use of a no-threshold, log-linear model; however, one study did observe 

heterogeneity in the shape of the concentration-response curve across cities (Section 6.5). Overall, 

the studies evaluated further support the use of a no-threshold log-linear model, but additional issues 

such as the influence of heterogeneity in estimates between cities, and the effect of seasonal and 

regional differences in PM on the concentration-response relationship still require further 

investigation. 

December 2009 2-25 
In addition to examining the concentration-response relationship between short-term exposure 

to PM and mortality, Schwartz et al. (2008, 156963) conducted an analysis of the shape of the 

concentration-response relationship associated with long-term exposure to PM. Using a variety of 

statistical methods, the concentration-response curve was found to be indistinguishable from linear, 

and, therefore, little evidence was observed to suggest that a threshold exists in the association 

between long-term exposure to PM2.5 and the risk of death (Section 7.6). 

2.4.4. PM Sources and Constituents Linked to Health Effects 

Recent epidemiologic, toxicological, and controlled human exposure studies have evaluated 

the health effects associated with ambient PM constituents and sources, using a variety of 

quantitative methods applied to a broad set of PM constituents, rather than selecting a few 

constituents a priori (Section 6.6). There is some evidence for trends and patterns that link particular 

ambient PM constituents or sources with specific health outcomes, but there is insufficient evidence 

to determine whether these patterns are consistent or robust. 

For cardiovascular effects, multiple outcomes have been linked to a PM2.5 crustal/soil/road 

dust source, including cardiovascular mortality and ST-segment changes. Additional studies have 

reported associations between other sources (i.e., traffic and wood smoke/vegetative burning) and 

cardiovascular outcomes (i.e., mortality and ED visits). Studies that only examined the effects of 

individual PM2.5 constituents found evidence for an association between EC and cardiovascular 

hospital admissions and cardiovascular mortality. Many studies have also observed associations 

between other sources (i.e., salt, secondary SO4

2–/long-range transport, other metals) and 

cardiovascular effects, but at this time, there does not appear to be a consistent trend or pattern of 

effects for those factors. 

There is less consistent evidence for associations between PM sources and respiratory health 

effects, which may be partially due to the fact that fewer source apportionment studies have been 

conducted that examined respiratory-related outcomes (e.g., hospital admissions) and measures (e.g., 

lung function). However, there is some evidence for associations between respiratory ED visits and 

decrements in lung function with secondary SO4

2– PM2.5. In addition, crustal/soil/road dust and 

traffic sources of PM have been found to be associated with increased respiratory symptoms in 

asthmatic children and decreased PEF in asthmatic adults. Inconsistent results were observed in 

those PM2.5 studies that used individual constituents to examine associations with respiratory 

morbidity and mortality, although Cu, Pb, OC, and Zn were related to respiratory health effects in 

two or more studies. 

A few studies have identified PM2.5 sources associated with total mortality. These studies 

found an association between mortality and the PM2.5 sources: secondary SO4

2–/long-range 

transport, traffic, and salt. In addition, studies have evaluated whether the variation in associations

between PM

e association. 

2.5 and mortality or PM10 and mortality reflects differences in PM2.5 constituents. PM10-

mortality effect estimates were greater in areas with a higher proportion of Ni in PM2.5, but the 

overall PM10-mortality association was diminished when New York City was excluded in sensitivity

analyses in two of the studies. V was also found to modify PM10-mortality effect estimates. When 

examining the effect of species-to-PM2.5 mass proportion on PM2.5-mortality effect estimates, Ni, 

but not V, was also found to modify th

Overall, the results indicate that many constituents of PM can be linked with differing health 

effects and the evidence is not yet sufficient to allow differentiation of those constituents or sources 

that are more closely related to specific health outcomes. These findings are consistent with the 

conclusions of the 2004 PM AQCD (U.S. EPA, 2004, 056905) (i.e., that a number of source types, 

including motor vehicle emissions, coal combustion, oil burning, and vegetative burning, are 

associated with health effects). Although the crustal factor of fine particles was not associated with 

mortality in the 2004 PM AQCD (U.S. EPA, 2004, 056905), recent studies have suggested that PM 

(both PM2.5 and PM10-2.5) from crustal, soil or road dust sources or PM tracers linked to these sources 

are associated with cardiovascular effects. In addition, PM2.5 secondary SO4

2– has been associated 

with both cardiovascular and respiratory effects. 

December 2009 2-26 
2.5. Welfare Effects 

This section presents key conclusions and scientific judgments regarding causality for welfare 

effects of PM as discussed in Chapter 9. The effects of particulate NOX and SOX have recently been 

evaluated in the ISA for Oxides of Nitrogen and Sulfur – Ecological Criteria (U.S. EPA, 2008, 

157074). That ISA focused on the effects from deposition of gas- and particle-phase pollutants 

related to ambient NOX and SOX concentrations that can lead to acidification and nutrient 

enrichment. Thus, emphasis in Chapter 9 is placed on the effects of airborne PM, including NOX and 

SOX, on visibility and climate, and on the effects of deposition of PM constituents other than NOX

and SOX, primarily metals and carbonaceous compounds. EPA’s framework for causality, described 

in Chapter 1, was applied and the causal determinations are highlighted. 

Table 2-5. Summary of causality determination for welfare effects. 

Welfare Effects Causality Determination 

Effects on Visibility Causal 

Effects on Climate Causal 

Ecological Effects Likely to be causal 

Effects on Materials Causal 

2.5.1. Summary of Effects on Visibility 

Visibility impairment is caused by light scattering and absorption by suspended particles and 

gases. There is strong and consistent evidence that PM is the overwhelming source of visibility 

impairment in both urban and remote areas. EC and some crustal minerals are the only commonly 

occurring airborne particle components that absorb light. All particles scatter light, and generally 

light scattering by particles is the largest of the four light extinction components (i.e., absorption and 

scattering by gases and particles). Although a larger particle scatters more light than a similarly 

shaped smaller particle of the same composition, the light scattered per unit of mass is greatest for 

particles with diameters from ~0.3-1.0 μm. 

For studies where detailed data on particle composition by size are available, accurate 

calculations of light extinction can be made. However, routinely available PM speciation data can be 

used to make reasonable estimates of light extinction using relatively simple algorithms that multiply 

the concentrations of each of the major PM species by its dry extinction efficiency and by a water 

growth term that accounts for particle size change as a function of relative humidity for hygroscopic 

species (e.g., sulfate, nitrate, and sea salt). This permits the visibility impairment associated with 

each of the major PM components to be separately approximated from PM speciation monitoring 

data. 

Direct optical measurement of light extinction measured by transmissometer, or by combining 

the PM light scattering measured by integrating nephelometers with the PM light absorption 

measured by an aethalometer, offer a number of advantages compared to algorithm estimates of light 

extinction based on PM composition and relative humidity data. The direct measurements are not 

subject to the uncertainties associated with assumed scattering and absorption efficiencies used in the 

PM algorithm approach. The direct measurements have higher time resolution (i.e., minutes to 

hours), which is more commensurate with visibility effects compared with calculated light extinction 

using routinely available PM speciation data (i.e., 24-h duration). 

Particulate sulfate and nitrate have comparable light extinction efficiencies (haze impacts per 

unit mass concentration) at any relative humidity value. Their light scattering per unit mass 

concentration increases with increasing relative humidity, and at sufficiently high humidity values 

(RH>85%) they are the most efficient particulate species contributing to haze. Particulate sulfate is 

December 2009 2-27 
the dominant source of regional haze in the eastern U.S. (>50% of the particulate light extinction) 

and an important contributor to haze elsewhere in the country (>20% of particulate light extinction). 

Particulate nitrate is a minor component of remote-area regional haze in the non-California western 

and eastern U.S., but an important contributor in much of California and in the upper Midwestern 

U.S., especially during winter when it is the dominant contributor to particulate light extinction. 

EC and OC have the highest dry extinction efficiencies of the major PM species and are 

responsible for a large fraction of the haze, especially in the northwestern U.S., though absolute 

concentrations are as high in the eastern U.S. Smoke plume impacts from large wildfires dominate 

many of the worst haze periods in the western U.S. Carbonaceous PM is generally the largest 

component of urban excess PM2.5 (i.e., the difference between urban and regional background 

concentration). Western urban areas have more than twice the average concentrations of 

carbonaceous PM than remote areas sites in the same region. In eastern urban areas PM2.5 is 

dominated by about equal concentrations of carbonaceous and sulfate components, though the 

usually high relative humidity in the East causes the hydrated sulfate particles to be responsible for 

about twice as much of the urban haze as that caused by the carbonaceous PM. 

PM2.5 crustal material (referred to as fine soil) and PM10-2.5 are significant contributors to haze 

for remote areas sites in the arid southwestern U.S. where they contribute a quarter to a third of the 

haze, with PM10-2.5 usually contributing twice that of fine soil. Coarse mass concentrations are as 

high in the Central Great Plains as in the deserts though there are no corresponding high 

concentrations of fine soil as in the Southwest. Also the relative contribution to haze by the high 

coarse mass in the Great Plains is much smaller because of the generally higher haze values caused 

by the high concentrations of sulfate and nitrate PM in that region. 

Visibility has direct significance to people’s enjoyment of daily activities and their overall 

sense of wellbeing. For example, psychological research has demonstrated that people are 

emotionally affected by poor VAQ such that their overall sense of wellbeing is diminished. Urban 

visibility has been examined in two types of studies directly relevant to the NAAQS review process: 

urban visibility preference studies and urban visibility valuation studies. Both types of studies are 

designed to evaluate individuals’ desire for good VAQ where they live, using different metrics. 

Urban visibility preference studies examine individuals’ preferences by investigating the amount of 

visibility degradation considered unacceptable, while economic studies examine the value an 

individual places on improving VAQ by eliciting how much the individual would be willing to pay 

for different amounts of VAQ improvement. 

There are three urban visibility preference studies and two additional pilot studies that have 

been conducted to date that provide useful information on individuals’ preferences for good VAQ in 

the urban setting. The completed studies were conducted in Denver, Colorado, two cities in British 

Columbia, Canada, and Phoenix, AZ. The additional studies were conducted in Washington, DC. The 

range of median preference values for an acceptable amount of visibility degradation from the 4 

urban areas was approximately 19-33 dv. Measured in terms of visual range (VR), these median 

acceptable values were between approximately 59 and 20 km. 

The economic importance of urban visibility has been examined by a number of studies 

designed to quantify the benefits (or willingness to pay) associated with potential improvements in 

urban visibility. Urban visibility valuation research was described in the 2004 PM AQCD (U.S. EPA, 

2004, 056905) and the 2005 PM Staff Paper (U.S. EPA, 2005, 090209). Since the mid-1990s, little 

new information has become available regarding urban visibility valuation (Section 9.2.4). 

Collectively, the evidence is sufficient to conclude that a causal relationship exists 

between PM and visibility impairment. 

2.5.2. Summary of Effects on Climate 

Aerosols affect climate through direct and indirect effects. The direct effect is primarily 

realized as planet brightening when seen from space because most aerosols scatter most of the 

visible spectrum light that reaches them. The Intergovernmental Panel on Climate Change (IPCC) 

Fourth Assessment Report (AR4) (IPCC, 2007, 092765), hereafter IPCC AR4, reported that the 

radiative forcing from this direct effect was -0.5 (±0.4) W/m2

 and identified the level of scientific 

understanding of this effect as 'Medium-low'. The global mean direct radiative forcing effect from 

individual components of aerosols was estimated for the first time in the IPCC AR4 where they were 

reported to be (all in W/m2

 units): -0.4 (±0.2) for sulfate, -0.05 (±0.05) for fossil fuel-derived organic 

December 2009 2-28 
carbon, +0.2 (±0.15) for fossil fuel-derived black carbon (BC), +0.03 (±0.12) for biomass burning, 

-0.1 (±0.1) for nitrates, and -0.1 (±0.2) for mineral dust. Global loadings of anthropogenic dust and 

nitrates remain very troublesome to estimate, making the radiative forcing estimates for these 

constituents particularly uncertain. 

Numerical modeling of aerosol effects on climate has sustained remarkable progress since the 

time of the 2004 PM AQCD (U.S. EPA, 2004, 056905), PM AQCD, though model solutions still 

display large heterogeneity in their estimates of the direct radiative forcing effect from 

anthropogenic aerosols. The clear-sky direct radiative forcing over ocean due to anthropogenic 

aerosols is estimated from satellite instruments to be on the order of -1.1 (±0.37) W/m2

 while model 

estimates are -0.6 W/m2. The models' low bias over ocean is carried through for the global average: 

global average direct radiative forcing from anthropogenic aerosols is estimated from measurements 

to range from -0.9 to -1.9 W/m2

, larger than the estimate of -0.8 W/m2 from the models. 

Aerosol indirect effects on climate are primarily realized as an increase in cloud brightness 

(termed the 'first indirect' or Twomey effect), changes in precipitation, and possible changes in cloud 

lifetime. The IPCC AR4 reported that the radiative forcing from the Twomey effect was -0.7 (range: 

-1.1 to +4) and identified the level of scientific understanding of this effect as “Low” in part owing 

to the very large unknowns concerning aerosol size distributions and important interactions with 

clouds. Other indirect effects from aerosols are not considered to be radiative forcing. 

Taken together, direct and indirect effects from aerosols increase Earth's shortwave albedo or 

reflectance thereby reducing the radiative flux reaching the surface from the Sun. This produces net 

climate cooling from aerosols. The current scientific consensus reported by IPCC AR4 is that the 

direct and indirect radiative forcing from anthropogenic aerosols computed at the top of the 

atmosphere, on a global average, is about -1.3 (range: -2.2 to -0.5) W/m2

. While the overall global 

average effect of aerosols at the top of the atmosphere and at the surface is negative, absorption and 

scattering by aerosols within the atmospheric column warms the atmosphere between the Earth's 

surface and top of the atmosphere. In part, this is owing to differences in the distribution of aerosol 

type and size within the vertical atmospheric column since aerosol type and size distributions 

strongly affect the aerosol scattering and reradiation efficiencies at different altitudes and 

atmospheric temperatures. And, although the magnitude of the overall negative radiative forcing at 

the top of the atmosphere appears large in comparison to the analogous IPCC AR4 estimate of 

positive radiative forcing from anthropogenic GHG of about +2.9 (± 0.3) W/m2

, the horizontal, 

vertical, and temporal distributions and the physical lifetimes of these two very different radiative 

forcing agents are not similar; therefore, the effects do not simply off-set one another. 

Overall, the evidence is sufficient to conclude that a causal relationship exists between 

PM and effects on climate, including both direct effects on radiative forcing and indirect 

effects that involve cloud feedbacks that influence precipitation formation and cloud 

lifetimes. 

2.5.3. Summary of Ecological Effects of PM 

Ecological effects of PM include direct effects to metabolic processes of plant foliage; 

contribution to total metal loading resulting in alteration of soil biogeochemistry and microbiology, 

plant growth and animal growth and reproduction; and contribution to total organics loading 

resulting in bioaccumulation and biomagnification across trophic levels. These effects were wellcharacterized in the 2004 PM AQCD (U.S. EPA, 2004, 056905). Thus, the summary below builds 

upon the conclusions provided in that review. 

PM deposition comprises a heterogeneous mixture of particles differing in origin, size, and 

chemical composition. Exposure to a given concentration of PM may, depending on the mix of 

deposited particles, lead to a variety of phytotoxic responses and ecosystem effects. Moreover, many 

of the ecological effects of PM are due to the chemical constituents (e.g., metals, organics, and ions) 

and their contribution to total loading within an ecosystem. 

Investigations of the direct effects of PM deposition on foliage have suggested little or no 

effects on foliar processes, unless deposition levels were higher than is typically found in the 

ambient environment. However, consistent and coherent evidence of direct effects of PM has been 

found in heavily polluted areas adjacent to industrial point sources such as limestone quarries, 

cement kilns, and metal smelters (Sections 9.4.3 and 9.4.5.7). Where toxic responses have been 

December 2009 2-29 
documented, they generally have been associated with the acidity, trace metal content, surfactant 

properties, or salinity of the deposited materials. 

An important characteristic of fine particles is their ability to affect the flux of solar radiation 

passing through the atmosphere, which can be considered in both its direct and diffuse components. 

Foliar interception by canopy elements occurs for both up- and down-welling radiation. Therefore, 

the effect of atmospheric PM on atmospheric turbidity influences canopy processes both by radiation 

attenuation and by changing the efficiency of radiation interception in the canopy through 

conversion of direct to diffuse radiation. Crop yields can be sensitive to the amount of radiation 

received, and crop losses have been attributed to increased regional haze in some areas of the world 

such as China (Section 9.4.4). On the other hand, diffuse radiation is more uniformly distributed 

throughout the canopy and may increase canopy photosynthetic productivity by distributing radiation 

to lower leaves. The enrichment in photosynthetically active radiation (PAR) present in diffuse 

radiation may offset a portion of the effect of an increased atmospheric albedo due to atmospheric 

particles. Further research is needed to determine the effects of PM alteration of radiative flux on the 

growth of vegetation in the U.S. 

The deposition of PM onto vegetation and soil, depending on its chemical composition, can 

produce responses within an ecosystem. The ecosystem response to pollutant deposition is a direct 

function of the level of sensitivity of the ecosystem and its ability to ameliorate resulting change. 

Many of the most important ecosystem effects of PM deposition occur in the soil. Upon entering the 

soil environment, PM pollutants can alter ecological processes of energy flow and nutrient cycling, 

inhibit nutrient uptake, change ecosystem structure, and affect ecosystem biodiversity. The soil 

environment is one of the most dynamic sites of biological interaction in nature. It is inhabited by 

microbial communities of bacteria, fungi, and actinomycetes, in addition to plant roots and soil 

macro-fauna. These organisms are essential participants in the nutrient cycles that make elements 

available for plant uptake. Changes in the soil environment can be important in determining plant 

and ultimately ecosystem response to PM inputs. 

There is strong and consistent evidence from field and laboratory experiments that metal 

components of PM alter numerous aspects of ecosystem structure and function. Changes in the soil 

chemistry, microbial communities and nutrient cycling, can result from the deposition of trace 

metals. Exposures to trace metals are highly variable, depending on whether deposition is by wet or 

dry processes. Although metals can cause phytotoxicity at high concentrations, few heavy metals 

(e.g., Cu, Ni, Zn) have been documented to cause direct phytotoxicity under field conditions. 

Exposure to coarse particles and elements such as Fe and Mg are more likely to occur via dry 

deposition, while fine particles, which are more often deposited by wet deposition, are more likely to 

contain elements such as Ca, Cr, Pb, Ni, and V. Ecosystems immediately downwind of major 

emissions sources can receive locally heavy deposition inputs. Phytochelatins produced by plants as 

a response to sublethal concentrations of heavy metals are indicators of metal stress to plants. 

Increased concentrations of phytochelatins across regions and at greater elevation have been 

associated with increased amounts of forest injury in the northeastern U.S. 

Overall, the ecological evidence is sufficient to conclude that a causal relationship is likely 

to exist between deposition of PM and a variety of effects on individual organisms and 

ecosystems, based on information from the previous review and limited new findings in 

this review. However, in many cases, it is difficult to characterize the nature and magnitude of 

effects and to quantify relationships between ambient concentrations of PM and ecosystem response 

due to significant data gaps and uncertainties as well as considerable variability that exists in the 

components of PM and their various ecological effects. 

2.5.4. Summary of Effects on Materials 

Building materials (metals, stones, cements, and paints) undergo natural weathering processes 

from exposure to environmental elements (wind, moisture, temperature fluctuations, sunlight, etc.). 

Metals form a protective film of oxidized metal (e.g., rust) that slows environmentally induced 

corrosion. However, the natural process of metal corrosion is enhanced by exposure to 

anthropogenic pollutants. For example, formation of hygroscopic salts increases the duration of 

surface wetness and enhances corrosion. 

A significant detrimental effect of particle pollution is the soiling of painted surfaces and other 

building materials. Soiling changes the reflectance of opaque materials and reduces the transmission 

December 2009 2-30 
of light through transparent materials. Soiling is a degradation process that requires remediation by 

cleaning or washing, and, depending on the soiled surface, repainting. Particulate deposition can 

result in increased cleaning frequency of the exposed surface and may reduce the usefulness of the 

soiled material. 

Attempts have been made to quantify the pollutant exposure levels at which materials damage 

and soiling have been perceived. However, to date, insufficient data are available to advance the 

knowledge regarding perception thresholds with respect to pollutant concentration, particle size, and 

chemical composition. Nevertheless, the evidence is sufficient to conclude that a causal 

relationship exists between PM and effects on materials. 

2.6. Summary of Health Effects and Welfare Effects 

Causal Determinations 

This chapter has provided an overview of the underlying evidence used in making the causal 

determinations for the health and welfare effects and PM size fractions evaluated. This review builds 

upon the main conclusions of the last PM AQCD (U.S. EPA, 2004, 056905): 

 “A growing body of evidence both from epidemiological and toxicological studies... 

supports the general conclusion that PM2.5 (or one or more PM2.5 components), acting 

alone and/or in combination with gaseous copollutants, are likely causally related to 

cardiovascular and respiratory mortality and morbidity.” (pg 9-79) 

 “A much more limited body of evidence is suggestive of associations between short-term 

(but not long-term) exposures to ambient coarse-fraction thoracic particles... and various 

mortality and morbidity effects observed at times in some locations. This suggests that 

PM10-2.5, or some constituent component(s) of PM10-2.5, may contribute under some 

circumstances to increased human health risks... with somewhat stronger evidence for... 

associations with morbidity (especially respiratory) endpoints than for mortality.” (pg 

9-79 and 9-80) 

 “Impairment of visibility in rural and urban areas is directly related to ambient 

concentrations of fine particles, as modulated by particle composition, size, and 

hygroscopic characteristics, and by relative humidity.” (pg 9-99) 

 “Available evidence, ranging from satellite to in situ measurements of aerosol effects on 

incoming solar radiation and cloud properties, is strongly indicative of an important role 

in climate for aerosols, but this role is still poorly quantified.” (pg 9-111) 

The evaluation of the epidemiologic, toxicological, and controlled human exposure studies 

published since the completion of the 2004 PM AQCD have provided additional evidence for 

PM-related health effects. Table 2-6 provides an overview of the causal determinations for all PM 

size fractions and health effects. Causal determinations for PM and welfare effects, including 

visibility, climate, ecological effects, and materials are included in Table 2-7. Detailed discussions of 

the scientific evidence and rationale for these causal determinations are provided in the subsequent 

chapters of this ISA.

December 2009 2-31 
Table 2-6. Summary of PM causal determinations by exposure duration and health outcome. 

Size Fraction Exposure Outcome Causality Determination 

Cardiovascular Effects Causal 

Respiratory Effects Likely to be causal 

Central Nervous System Inadequate 

Short-term 

Mortality Causal 

Cardiovascular Effects Causal 

Respiratory Effects Likely to be Causal 

Mortality Causal 

Reproductive and Developmental Suggestive 

PM2.5

Long-term 

Cancer, Mutagenicity, Genotoxicity Suggestive 

Cardiovascular Effects Suggestive 

Respiratory Effects Suggestive 

Central Nervous System Inadequate 

Short-term 

Mortality Suggestive 

Cardiovascular Effects Inadequate 

Respiratory Effects Inadequate 

Mortality Inadequate 

Reproductive and Developmental Inadequate 

PM10-2.5

Long-term 

Cancer, Mutagenicity, Genotoxicity Inadequate 

Cardiovascular Effects Suggestive 

Respiratory Effects Suggestive 

Central Nervous System Inadequate 

Short-term 

Mortality Inadequate 

Cardiovascular Effects Inadequate 

Respiratory Effects Inadequate 

Mortality Inadequate 

Reproductive and Developmental Inadequate 

UFPs 

Long-term 

Cancer, Mutagenicity, Genotoxicity Inadequate 

December 2009 2-32 
Table 2-7. Summary of PM causal determinations for welfare effects 

Welfare Effects Causality Determination 

Effects on Visibility Causal 

Effects on Climate Causal 

Ecological Effects Likely to be causal 

Effects on Materials Causal 

December 2009 2-33 
Chapter 2 References 

American Heart Association (2009). Cardiovascular disease statistics. Retrieved 17-NOV-09, from 

http://www.americanheart.org/presenter.jhtml?identifier=4478. 198920

ATSDR (2006). A study of ambient air contaminants and asthma in New York City: Part A and B. Agency for Toxic 

Substances and Disease Registry; Public Health Service; U.S. Department of Health and Human Services. Atlanta, 

GA.http://permanent.access.gpo.gov/lps88357/ASTHMA_BRONX_FINAL_REPORT.pdf. 090132

Bell ML; Ebisu K; Peng RD; Walker J; Samet JM; Zeger SL; Dominic F (2008). Seasonal and regional short-term effects 

of fine particles on hospital admissions in 202 U.S. counties, 1999-2005. Am J Epidemiol, 168: 1301-1310. 156266

Burnett RT; Cakmak S; Brook JR; Krewski D (1997). The role of particulate size and chemistry in the association between 

summertime ambient air pollution and hospitalization for cardiorespiratory diseases. Environ Health Perspect, 105: 

614-620. 084194

Burnett RT; Goldberg MS (2003). Size-fractionated particulate mass and daily mortality in eight Canadian cities. Health 

Effects Institute. Boston, MA. 042798

Burnett RT; Stieb D; Brook JR; Cakmak S; Dales R; Raizenne M; Vincent R; Dann T (2004). Associations between shortterm changes in nitrogen dioxide and mortality in Canadian cities. Arch Environ Occup Health, 59: 228-236. 

086247

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

Chen Y; Yang Q; Krewski D; Burnett RT; Shi Y; McGrail KM (2005). The effect of coarse ambient particulate matter on 

first, second, and overall hospital admissions for respiratory disease among the elderly. Inhal Toxicol, 17: 649-655. 

087555

Chen Y; Yang Q; Krewski D; Shi Y; Burnett RT; McGrail K (2004). Influence of relatively low level of particulate air 

pollution on hospitalization for COPD in elderly people. Inhal Toxicol, 16: 21-25. 087262

Chimonas MA; Gessner BD (2007). Airborne particulate matter from primarily geologic, non-industrial sources at levels 

below National Ambient Air Quality Standards is associated with outpatient visits for asthma and quick-relief 

medication prescriptions among children less than 20 years old enrolled in Medicaid in Anchorage, Alaska. Environ 

Res, 103: 397-404. 093261

Delfino RJ; Murphy-Moulton AM; Burnett RT; Brook JR; Becklake MR (1997). Effects of air pollution on emergency 

room visits for respiratory illnesses in Montreal, Quebec. Am J Respir Crit Care Med, 155: 568-576. 082687

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; Luttmann-Gibson H; Rich DQ; Link MS; Mittleman MA; Gold DR; Koutrakis P; Schwartz JD; Verrier RL 

(2005). Association of air pollution with increased incidence of ventricular tachyarrhythmias recorded by implanted 

cardioverter defibrillators. Environ Health Perspect, 113: 670-674. 078995

Dominici F; Peng RD; Bell ML; Pham L; McDermott A; Zeger SL; Samet JL (2006). Fine particulate air pollution and 

hospital admission for cardiovascular and respiratory diseases. JAMA, 295: 1127-1134. 088398

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

 

 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 2-34 
Fairley D (2003). Mortality and air pollution for Santa Clara County, California, 1989-1996, In: Revised analyses of timeseries studies of air pollution and health. Special report. Health Effects Institute. Boston, 

MA.http://www.healtheffects.org/Pubs/TimeSeries.pdf . 042850

Franklin M; Koutrakis P; Schwartz J (2008). The role of particle composition on the association between PM2.5 and 

mortality. Epidemiology, 19: 680-689. 097426

Franklin M; Zeka A; Schwartz J (2007). Association between PM2.5 and all-cause and specific-cause mortality in 27 US 

communities. J Expo Sci Environ Epidemiol, 17: 279-287. 091257

Fung KY; Khan S; Krewski D; Chen Y (2006). Association between air pollution and multiple respiratory hospitalizations 

among the elderly in Vancouver, Canada. Inhal Toxicol, 18: 1005-1011. 089789

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

IPCC (2007). Climate Change 2007: The Physical Science Basis. Contribution of Working Group I to the Fourth 

Assessment Report (AR4) of the Intergovernmental Panel on Climate Change . Cambridge, UK and New York, NY: 

Intergovernmental Panel on Climate Change, Cambridge University Press. 092765

Ito K (2003). Associations of particulate matter components with daily mortality and morbidity in Detroit, Michigan, In: 

Revised analyses of time-series studies of air pollution and health. Special report. Health Effects Institute. Boston, 

MA. R828112. http: //www.healtheffects.org/Pubs/TimeSeries.pdf . 042856

Ito K; Thurston GD; Silverman RA (2007). Characterization of PM2.5, gaseous pollutants, and meteorological interactions 

in the context of time-series health effects models. J Expo Sci Environ Epidemiol, 17 Suppl 2: S45-S60. 156594

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

Klemm RJ; Lipfert FW; Wyzga RE; Gust C (2004). Daily mortality and air pollution in Atlanta: two years of data from 

ARIES. Inhal Toxicol, 16 Suppl 1: 131-141. 056585

Klemm RJ; Mason R (2003). Replication of reanalysis of Harvard Six-City mortality study. In HEI Special Report: Revised 

Analyses of Time-Series Studies of Air Pollution and Health, Part II (pp. 165-172). Boston, MA: Health Effects 

Institute. 042801

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

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

Lin M; Chen Y; Burnett RT; Villeneuve PJ; Krewski D (2002). The influence of ambient coarse particulate matter on 

asthma hospitalization in children: case-crossover and time-series analyses. Environ Health Perspect, 110: 575-581. 

026067

Lin M; Stieb DM; Chen Y (2005). Coarse particulate matter and hospitalization for respiratory infections in children 

younger than 15 years in Toronto: a case-crossover analysis. Pediatrics, 116: 235-240. 087828

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; Morris SC; Wyzga RE (2000). Daily mortality in the Philadelphia metropolitan area and size-classified 

particulate matter. J Air Waste Manag Assoc, 50: 1501-1513. 004088

Lisabeth LD; Escobar JD; Dvonch JT; Sanchez BN; Majersik JJ; Brown DL; Smith MA; Morgenstern LB (2008). Ambient 

air pollution and risk for ischemic stroke and transient ischemic attack. Ann Neurol, 64: 53-59. 155939

Mar TF; Larson TV; Stier RA; Claiborn C; Koenig JQ (2004). An analysis of the association between respiratory symptoms 

in subjects with asthma and daily air pollution in Spokane, Washington. Inhal Toxicol, 16: 809-815. 057309

Mar TF; Norris GA; Larson TV; Wilson WE; Koenig JQ (2003). Air pollution and cardiovascular mortality in Phoenix, 

1995-1997. Health Effects Institute. Cambridge, MA. 042841

December 2009 2-35 
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

Metzger KB; Klein M; Flanders WD; Peel JL; Mulholland JA; Langberg JJ; Tolbert PE (2007). Ambient air pollution and 

cardiac arrhythmias in patients with implantable defibrillators. Epidemiology, 18: 585-592. 092856

Metzger KB; Tolbert PE; Klein M; Peel JL; Flanders WD; Todd KH; Mulholland JA; Ryan PB; Frumkin H (2004). 

Ambient air pollution and cardiovascular emergency department visits. Epidemiology, 15: 46-56. 044222

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

O'Connor GT; Neas L; Vaughn B; Kattan M; Mitchell H; Crain EF; Evans R 3rd; Gruchalla R; Morgan W; Stout J; Adams 

GK; Lippmann M (2008). Acute respiratory health effects of air pollution on children with asthma in US inner 

cities. J Allergy Clin Immunol, 121: 1133-1139. 156818

Ostro BD; Broadwin R; Lipsett MJ (2003). Coarse particles and daily mortality in Coachella Valley, California. Health 

Effects Institute. Boston, MA. 042824

Peel JL; Tolbert PE; Klein M; Metzger KB; Flanders WD; Knox T; Mulholland JA; Ryan PB; Frumkin H (2005). Ambient 

air pollution and respiratory emergency department visits. Epidemiology, 16: 164-174. 056305

Peng RD; Chang HH; Bell ML; McDermott A; Zeger SL; Samet JM; Dominici F (2008). Coarse particulate matter air 

pollution and hospital admissions for cardiovascular and respiratory diseases among Medicare patients. JAMA, 

299: 2172-2179. 156850

Peters A; Dockery DW; Muller JE; Mittleman MA (2001). Increased particulate air pollution and the triggering of 

myocardial infarction. Circulation, 103: 2810-2815. 016546

Pope C; Renlund D; Kfoury A; May H; Horne B (2008). Relation of heart failure hospitalization to exposure to fine 

particulate air pollution. Am J Cardiol, 102: 1230-1234. 191969

Pope CA III; Muhlestein JB; May HT; Renlund DG; Anderson JL; Horne BD (2006). Ischemic heart disease events 

triggered by short-term exposure to fine particulate air pollution. Circulation, 114: 2443-2448. 091246

Rabinovitch N; Strand M; Gelfand EW (2006). Particulate levels are associated with early asthma worsening in children 

with persistent disease. Am J Respir Crit Care Med, 173: 1098-1105. 088031

Rabinovitch N; Zhang LN; Murphy JR; Vedal S; Dutton SJ; Gelfand EW (2004). Effects of wintertime ambient air 

pollutants on asthma exacerbations in urban minority children with moderate to severe disease. J Allergy Clin 

Immunol, 114: 1131-1137. 096753

Renwick LC; Brown D; Clouter A; Donaldson K (2004). Increased inflammation and altered macrophage chemotactic 

responses caused by two ultrafine particle types. Occup Environ Med, 61: 442-447. 056067

Rich DQ; Kim MH; Turner JR; Mittleman MA; Schwartz J; Catalano PJ; Dockery DW (2006). Association of ventricular 

arrhythmias detected by implantable cardioverter defibrillator and ambient air pollutants in the St Louis, Missouri 

metropolitan area. Occup Environ Med, 63: 591-596. 089814

Rich DQ; Schwartz J; Mittleman MA; Link M; Luttmann-Gibson H; Catalano PJ; Speizer FE; Dockery DW (2005). 

Association of short-term ambient air pollution concentrations and ventricular arrhythmias. Am J Epidemiol, 161: 

1123-1132. 079620

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

Schwartz J; Dockery DW; Neas LM (1996). Is daily mortality associated specifically with fine particles? J Air Waste 

Manag Assoc, 46: 927-939. 077325

Sheppard L; Levy D; Norris G; Larson TV; Koenig JQ (2003). Effects of ambient air pollution and nonelderly asthma 

hospital admissions in Seattle, Washington, 1987-1994. Epidemiology, 10: 23-30. 042826

Slaughter JC; Kim E; Sheppard L; Sullivan JH; Larson TV; Claiborn C (2005). Association between particulate matter and 

emergency room visits, hospital admissions and mortality in Spokane, Washington. J Expo Sci Environ Epidemiol, 

15: 153-159. 073854

December 2009 2-36 
Stieb DM; Beveridge RC; Brook JR; Smith-Doiron M; Burnett RT; Dales RE; Beaulieu S; Judek S; Mamedov A (2000). 

Air pollution, aeroallergens and cardiorespiratory emergency department visits in Saint John, Canada. J Expo Sci 

Environ Epidemiol, 10: 461-477. 011675

Sullivan J; Sheppard L; Schreuder A; Ishikawa N; Siscovick D; Kaufman J (2005). Relation between short-term fineparticulate matter exposure and onset of myocardial infarction. Epidemiology, 16: 41-48. 050854

Symons JM; Wang L; Guallar E; Howell E; Dominici F; Schwab M; Ange BA; Samet J; Ondov J; Harrison D; Geyh A 

(2006). A case-crossover study of fine particulate matter air pollution and onset of congestive heart failure symptom 

exacerbation leading to hospitalization. Am J Epidemiol, 164: 421-433. 091258

Thurston GD; Ito K; Hayes CG; Bates DV; Lippmann M (1994). Respiratory hospital admissions and summertime haze air 

pollution in Toronto, Ontario: consideration of the role of acid aerosols. Environ Res, 65: 271-290. 043921

Tolbert PE; Klein M; Peel JL; Sarnat SE; Sarnat JA (2007). Multipollutant modeling issues in a study of ambient air quality 

and emergency department visits in Atlanta. J Expo Sci Environ Epidemiol, 17: S29-S35. 090316

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). Review of the national ambient air quality standards for particulate matter: Policy assessment of 

scientific and technical information OAQPS staff paper. U.S. Environmental Protection Agency. Washington, DC. 

EPA/452/R-05-005a. http://www.epa.gov/ttn/naaqs/standards/pm/data/pmstaffpaper_20051221.pdf . 090209

U.S. EPA (2008). Integrated science assessment for oxides of nitrogen and sulfur: Ecological criteria. U.S. Environmental 

Protection Agency. Research Triangle Park, NC. EPA/600/R-08/082F. 157074

Villeneuve PJ; Burnett RT; Shi Y; Krewski D; Goldberg MS; Hertzman C; Chen Y; Brook J (2003). A time-series study of 

air pollution, socioeconomic status, and mortality in Vancouver, Canada. J Expo Sci Environ Epidemiol, 13: 427-

435. 055051

Villeneuve PJ; Chen L; Stieb D; Rowe BH (2006). Associations between outdoor air pollution and emergency department 

visits for stroke in Edmonton, Canada. Eur J Epidemiol, 21: 689-700. 090191

Wilson WE; Mar TF; Koenig JQ (2007). Influence of exposure error and effect modification by socioeconomic status on 

the association of acute cardiovascular mortality with particulate matter in Phoenix. J Expo Sci Environ Epidemiol, 

17: S11-S19. 157149

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

Zanobetti A; Schwartz J (2006). Air pollution and emergency admissions in Boston, MA. J Epidemiol Community Health, 

60: 890-895. 090195

Zanobetti A; Schwartz J (2009). The effect of fine and coarse particulate air pollution on mortality: A national analysis. 

Environ Health Perspect, 117: 1-40. 188462

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

Zhang Z; Whitsel E; Quibrera P; Smith R; Liao D; Anderson G; Prineas R (2009). Ambient fine particulate matter exposure 

and myocardial ischemia in the Environmental Epidemiology of Arrhythmogenesis in the Women's Health Initiative 

(EEAWHI) study. Environ Health Perspect, 117: 751-756. 191970

December 2009 2-37