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Chapter 6 Health and Environmental Effects Associated with 

Exposure to Criteria and Toxic Pollutants

6.1 Health Effects of Criteria and Toxic Pollutants

Motor vehicles emit pollutants that contribute to ambient concentrations of ozone, PM, 

NO2, SO2, CO, and air toxics. A discussion of the health effects of these pollutants is included in 

this section of the RIA. Children may be more vulnerable to air pollution and other 

environmental exposures than adults because their bodily systems are still developing, they

breathe more in proportion to their body size than adults and their behavior can expose them 

more to chemicals than adults. Early lifestages (e.g., children) are thought to be more 

susceptible to tumor development than adults when exposed to carcinogenic chemicals that act 

through a mutagenic mode of action.1

6.1.1 Particulate Matter

6.1.1.1 Background

Particulate matter (PM) is a highly complex mixture of solid particles and liquid droplets 

distributed among numerous atmospheric gases which interact with solid and liquid phases. 

Particles range in size from those smaller than 1 nanometer (10-9 meter) to over 100 micrometer 

(μm, or 10-6 meter) in diameter (for reference, a typical strand of human hair is 70 μm in 

diameter and a grain of salt is about 100 μm). Atmospheric particles can be grouped into several 

classes according to their aerodynamic and physical sizes, including ultrafine particles (<0.1 

μm), accumulation mode or ‘fine’ particles (< 1 to 3 μm), and coarse particles (>1 to 3 μm). For 

regulatory purposes, fine particles are measured as PM2.5 and inhalable or thoracic coarse 

particles are measured as PM10-2.5, corresponding to their size (diameter) range in micrometers. 

The EPA currently has standards that measure PM2.5 and PM10.

2

Particles span many sizes and shapes and may consist of hundreds of different chemicals. 

Particles are emitted directly from sources and are also formed through atmospheric chemical 

reactions; the former are often referred to as “primary” particles, and the latter as “secondary” 

particles. Particle concentration and composition varies by time of year and location, and in 

addition to differences in source emissions, is affected by several weather-related factors, such as 

temperature, clouds, humidity, and wind. A further layer of complexity comes from particles’

ability to shift between solid/liquid and gaseous phases, which is influenced by concentration and 

meteorology, especially temperature.

Fine particles are produced primarily by combustion processes and by transformations of 

gaseous emissions (e.g., sulfur oxides (SOX), nitrogen oxides (NOX) and volatile organic 

compounds (VOCs)) in the atmosphere. The chemical and physical properties of PM2.5 may vary

greatly with time, region, meteorology and source category. Thus, PM2.5 may include a complex 

mixture of different components including sulfates, nitrates, organic compounds, elemental 
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carbon and metal compounds. These particles can remain in the atmosphere for days to weeks 

and travel through the atmosphere hundreds to thousands of kilometers.3 

6.1.1.2 Health Effects of PM

Scientific studies show ambient PM is associated with a broad range of health effects. 

These health effects are discussed in detail in the December 2009 Integrated Science Assessment 

for Particulate Matter (PM ISA).

4

 The PM ISA summarizes health effects evidence associated 

with both short- and long-term exposures to PM2.5, PM10-2.5, and ultrafine particles. The PM ISA 

concludes that human exposures to ambient PM2.5 concentrations are associated with a number 

of adverse health effects and characterizes the weight of evidence for these health outcomes.5 

The discussion below highlights the PM ISA’s conclusions pertaining to health effects associated 

with both short- and long-term PM exposures. Further discussion of health effects associated 

with PM2.5 can also be found in the rulemaking documents for the most recent review of the PM 

NAAQS completed in 2012.

6,7

The EPA concludes that a causal relationship exists between both long- and short-term 

exposures to PM2.5 and premature mortality and cardiovascular effects and a likely causal 

relationship exists between long- and short-term PM2.5 exposures and respiratory effects. 

Further, there is evidence suggestive of a causal relationship between long-term PM2.5 exposures 

and other health effects, including developmental and reproductive effects (e.g., low birth 

weight, infant mortality) and carcinogenic, mutagenic, and genotoxic effects (e.g., lung cancer 

mortality).8

As summarized in the Final PM NAAQS rule, and discussed extensively in the 2009 PM 

ISA, the scientific evidence available since the completion of the 2006 PM NAAQS review

significantly strengthens the link between long- and short-term exposure to PM2.5 and premature 

mortality, while providing indications that the magnitude of the PM2.5- mortality association with 

long-term exposures may be larger than previously estimated.9,10 The strongest evidence comes 

from recent studies investigating long-term exposure to PM2.5 and cardiovascular-related

mortality. The evidence supporting a causal relationship between long-term PM2.5 exposure and 

mortality also includes consideration of new studies that demonstrated an improvement in 

community health following reductions in ambient fine particles.11

Several studies evaluated in the 2009 PM ISA have examined the association between 

cardiovascular effects and long-term PM2.5 exposures in multi-city studies conducted in the U.S. 

and Europe. While studies were not available in the 2006 PM NAAQS review with regard to 

long-term exposure and cardiovascular-related morbidity, studies published since then have 

provided new evidence linking long-term exposure to PM2.5 with an array of cardiovascular 

effects such as heart attacks, congestive heart failure, stroke, and mortality. This evidence is 

coherent with studies of short-term exposure to PM2.5 that have observed associations with a 

continuum of effects ranging from subtle changes in indicators of cardiovascular health to 

serious clinical events, such as increased hospitalizations and emergency department visits due to 

cardiovascular disease and cardiovascular mortality.12
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As detailed in the 2009 PM ISA, extended analyses of studies available in the 2006 PM 

NAAQS review as well as epidemiological studies conducted in the U.S. and abroad published 

since then provide stronger evidence of respiratory-related morbidity effects associated with 

long-term PM2.5 exposure. The strongest evidence for respiratory-related effects is from studies 

that evaluated decrements in lung function growth (in children), increased respiratory symptoms, 

and asthma development. The strongest evidence from short-term PM2.5 exposure studies has 

been observed for increased respiratory-related emergency department visits and hospital 

admissions for chronic obstructive pulmonary disease (COPD) and respiratory infections.13

The body of scientific evidence detailed in the 2009 PM ISA is still limited with respect 

to associations between long-term PM2.5 exposures and developmental and reproductive effects 

as well as cancer, mutagenic, and genotoxic effects, but is somewhat expanded from the 2006 

review The strongest evidence for an association between PM2.5 and developmental and 

reproductive effects comes from epidemiological studies of low birth weight and infant 

mortality, especially due to respiratory causes during the post-neonatal period (i.e., 1 month to 12 

months of age). With regard to cancer effects, ‘‘[m]ultiple 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.’’14,15

Specific groups within the general population are at increased risk for experiencing 

adverse health effects related to PM exposures.16,17,18,19 The evidence detailed in the 2009 PM 

ISA expands our understanding of previously identified at-risk populations and lifestages (i.e., 

children, older adults, and individuals with pre-existing heart and lung disease) and supports the 

identification of additional at-risk populations (e.g., persons with lower socioeconomic status, 

genetic differences). Additionally, there is emerging, though still limited, evidence for additional 

potentially at-risk populations and lifestages, such as those with diabetes, people who are obese, 

pregnant women, and the developing fetus.22

For PM10-2.5, the 2009 PM ISA concluded that available evidence was suggestive of a 

causal relationship between short-term exposures to PM10-2.5 and cardiovascular effects (e.g., 

hospital admissions and ED visits, changes in cardiovascular function), respiratory effects (e.g, 

ED visits and hospital admissions, increase in markers of pulmonary inflammation), and 

premature mortality. Data were inadequate to draw conclusions regarding the relationships 

between long-term exposure to PM10-2.5 and various health effects. 23,24,25

For ultrafine particles, the 2009 PM ISA concluded that the evidence was suggestive of a 

causal relationship between short-term exposures and cardiovascular effects, including changes 

in heart rhythm and vasomotor function (the ability of blood vessels to expand and contract). It 

also concluded that there was evidence suggestive of a causal relationship between short-term 

exposure to ultrafine particles and respiratory effects, including lung function and pulmonary

inflammation, with limited and inconsistent evidence for increases in ED visits and hospital 

admissions. Data were inadequate to draw conclusions regarding the relationship between shortterm exposure to ultrafine particle and additional health effects including premature mortality as 

well as long-term exposure to ultrafine particles and all health outcomes evaluated.

26,27
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6.1.2 Ozone

6.1.2.1 Background

Ground-level ozone pollution is typically formed through reactions involving VOCs and 

NOX in the lower atmosphere in the presence of sunlight. These pollutants, often referred to as 

ozone precursors, are emitted by many types of pollution sources such as highway and nonroad 

motor vehicles and engines, power plants, chemical plants, refineries, makers of consumer and 

commercial products, industrial facilities, and smaller area sources. 

The science of ozone formation, transport, and accumulation is complex. Ground-level 

ozone is produced and destroyed in a cyclical set of chemical reactions, many of which are 

sensitive to temperature and sunlight. When ambient temperatures and sunlight levels remain 

high for several days and the air is relatively stagnant, ozone and its precursors can build up and 

result in more ozone than typically occurs on a single high-temperature day. Ozone and its 

precursors can be transported hundreds of miles downwind of precursor emissions, resulting in 

elevated ozone levels even in areas with low VOC or NOX emissions. 

The highest levels of ozone are produced when both VOC and NOX emissions are present 

in significant quantities on clear summer days. Relatively small amounts of NOX enable ozone 

to form rapidly when VOC levels are relatively high, but ozone production is quickly limited by 

removal of the NOX. Under these conditions NOX reductions are highly effective in reducing 

ozone while VOC reductions have little effect. Such conditions are called “NOX-limited.” 

Because the contribution of VOC emissions from biogenic (natural) sources to local ambient 

ozone concentrations can be significant, even some areas where man-made VOC emissions are 

relatively low can be NOX-limited.

Ozone concentrations in an area also can be lowered by the reaction of nitric oxide (NO) 

with ozone, forming nitrogen dioxide (NO2). As the air moves downwind and the cycle 

continues, the NO2 forms additional ozone. The importance of this reaction depends, in part, on 

the relative concentrations of NOX, VOC, and ozone, all of which change with time and location. 

When NOX levels are relatively high and VOC levels relatively low, NOX forms inorganic 

nitrates (i.e., particles) but relatively little ozone. Such conditions are called “VOC-limited.” 

Under these conditions, VOC reductions are effective in reducing ozone, but NOX reductions can 

actually increase local ozone under certain circumstances. Even in VOC-limited urban areas, 

NOX reductions are not expected to increase ozone levels if the NOX reductions are sufficiently 

large. Rural areas are usually NOX-limited, due to the relatively large amounts of biogenic VOC 

emissions in such areas. Urban areas can be either VOC- or NOX-limited, or a mixture of both, 

in which ozone levels exhibit moderate sensitivity to changes in either pollutant.

6.1.2.2 Health Effects of Ozone

This section provides a summary of the health effects associated with exposure to 

ambient concentrations of ozone.

28

 The information in this section is based on the information

and conclusions in the February 2013 Integrated Science Assessment for Ozone (Ozone ISA) 
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prepared by EPA’s Office of Research and Development (ORD).

29 The Ozone ISA concludes 

that human exposures to ambient concentrations of ozone are associated with a number of 

adverse health effects and characterizes the weight of evidence for these health effects. 30 The 

discussion below highlights the Ozone ISA’s conclusions pertaining to health effects associated 

with both short-term and long-term periods of exposure to ozone.

For short-term exposure to ozone, the Ozone ISA concludes that respiratory effects, 

including lung function decrements, pulmonary inflammation, exacerbation of asthma, 

respiratory-related hospital admissions, and mortality, are causally associated with ozone 

exposure. It also concludes that cardiovascular effects, including decreased cardiac function and 

increased vascular disease, and total mortality are likely to be causally associated with short-term 

exposure to ozone and that evidence is suggestive of a causal relationship between central 

nervous system effects and short-term exposure to ozone. 

For long-term exposure to ozone, the Ozone ISA concludes that respiratory effects, 

including new onset asthma, pulmonary inflammation and injury, are likely to be a causally 

related with ozone exposure. The Ozone ISA characterizes the evidence as suggestive of a 

causal relationship for associations between long-term ozone exposure and cardiovascular 

effects, reproductive and developmental effects, central nervous system effects and total 

mortality. The evidence is inadequate to infer a causal relationship between chronic ozone 

exposure and increased risk of lung cancer.

Finally, interindividual variation in human responses to ozone exposure can result in 

some groups being at increased risk for detrimental effects in response to exposure. The Ozone 

ISA identified several groups that are at increased risk for ozone-related health effects. These 

groups are people with asthma, children and older adults, individuals with reduced intake of 

certain nutrients (i.e., Vitamins C and E), outdoor workers, and individuals having certain 

genetic variants related to oxidative metabolism or inflammation. Ozone exposure during 

childhood can have lasting effects through adulthood. Such effects include altered function of 

the respiratory and immune systems. Children absorb higher doses (normalized to lung surface 

area) of ambient ozone, compared to adults, due to their increased time spent outdoors, higher 

ventilation rates relative to body size, and a tendency to breathe a greater fraction of air through 

the mouth. Children also have a higher asthma prevalence compared to adults. Additional 

children’s vulnerability and susceptibility factors are listed in Section XII.G of the preamble.

6.1.3 Nitrogen Oxides and Sulfur Oxides

6.1.3.1 Background

Sulfur dioxide (SO2), a member of the sulfur oxide (SOX) family of gases, is formed from 

burning fuels containing sulfur (e.g., coal or oil), extracting gasoline from oil, or extracting 

metals from ore. Nitrogen dioxide (NO2) is a member of the nitrogen oxide (NOX) family of 

gases. Most NO2 is formed in the air through the oxidation of nitric oxide (NO) emitted when 

fuel is burned at a high temperature. SO2 andNO2 and their gas phase oxidation products can 

dissolve in water droplets and further oxidize to form sulfuric and nitric acid which react with 
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ammonia to form sulfates and nitrates, both of which are important components of ambient PM. 

The health effects of ambient PM are discussed in Section 6.1.1.2. NOX along with VOCs are 

the two major precursors of ozone. The health effects of ozone are covered in Section 6.1.2.2.

6.1.3.2 Health Effects of Sulfur Oxides

This section provides an overview of the health effects associated with SO2. Additional 

information on the health effects of SO2 can be found in the 2008 Integrated Science Assessment 

for Sulfur Oxides (SO2 ISA).

31

 Following an extensive evaluation of health evidence from 

epidemiologic and laboratory studies, the U.S. EPA has concluded that there is a causal 

relationship between respiratory health effects and short-term exposure to SO2. The immediate 

effect of SO2 on the respiratory system in humans is bronchoconstriction. Asthmatics are more 

sensitive to the effects of SO2 likely resulting from preexisting inflammation associated with this 

disease. In addition to those with asthma (both children and adults), potentially sensitive groups 

include all children and the elderly. In laboratory studies involving controlled human exposures 

to SO2, respiratory effects have consistently been observed following 5-10 min exposures at SO2

concentrations ≥ 0.4 ppm in asthmatics engaged in moderate to heavy levels of exercise, with 

more limited evidence of respiratory effects among exercising asthmatics exposed to 

concentrations as low as 0.2-0.3 ppm. A clear concentration-response relationship has been 

demonstrated in these studies following exposures to SO2 at concentrations between 0.2 and 1.0 

ppm, both in terms of increasing severity of respiratory symptoms and decrements in lung 

function, as well as the percentage of asthmatics adversely affected. 

In epidemiologic studies, respiratory effects have been observed in areas where the mean 

24-hour SO2 levels range from 1 to 30 ppb, with maximum 1 to 24-hour average SO2 values 

ranging from 12 to 75 ppb. Important new multicity studies and several other studies have found 

an association between 24-hour average ambient SO2 concentrations and respiratory symptoms 

in children, particularly those with asthma. Generally consistent associations also have been 

observed between ambient SO2 concentrations and emergency department visits and 

hospitalizations for all respiratory causes, particularly among children and older adults (≥ 65 

years), and for asthma. A limited subset of epidemiologic studies has examined potential 

confounding by copollutants using multipollutant regression models. These analyses indicate 

that although copollutant adjustment has varying degrees of influence on the SO2 effect 

estimates, the effect of SO2 on respiratory health outcomes appears to be generally robust and 

independent of the effects of gaseous and particulate copollutants, suggesting that the observed 

effects of SO2 on respiratory endpoints occur independent of the effects of other ambient air 

pollutants. 

Consistent associations between short-term exposure to SO2 and mortality have been 

observed in epidemiologic studies, with larger effect estimates reported for respiratory mortality 

than for cardiovascular mortality. While this finding is consistent with the demonstrated effects 

of SO2 on respiratory morbidity, uncertainty remains with respect to the interpretation of these 

associations due to potential confounding by various copollutants. The U.S. EPA has therefore 

concluded that the overall evidence is suggestive of a causal relationship between short-term 
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exposure to SO2 and mortality. Significant associations between short-term exposure to SO2 and 

emergency department visits and hospital admissions for cardiovascular diseases have also been 

reported. However, these findings have been inconsistent across studies and do not provide 

adequate evidence to infer a causal relationship between SO2 exposure and cardiovascular 

morbidity.

6.1.3.3 Health Effects of Nitrogen Oxides

The most recent review of the health effects of oxides of nitrogen completed by the EPA 

can be found in the 2008 Integrated Science Assessment for Nitrogen Oxides (NOX ISA).

32

 The 

EPA concluded that the findings of epidemiologic, controlled human exposure, and animal 

toxicological studies provide evidence that is sufficient to infer a likely causal relationship 

between respiratory effects and short-term NO2 exposure. The 2008 NOX ISA concluded that the 

strongest evidence for such a relationship comes from epidemiologic studies of respiratory

effects including increased respiratory symptoms, emergency department visits, and hospital 

admissions. Based on both short- and long-term exposure studies, the 2008 NOX ISA concluded 

that individuals with preexisting pulmonary conditions (e.g., asthma or COPD), children, and 

older adults are potentially at greater risk of NO2-related respiratory effects. Based on findings 

from controlled human exposure studies, the 2008 NOX ISA also drew two broad conclusions 

regarding airway responsiveness following NO2 exposure. First, the NOX ISA concluded that 

NO2 exposure may enhance the sensitivity to allergen-induced decrements in lung function and 

increase the allergen-induced airway inflammatory response following 30-minute exposures of 

asthmatic adults to NO2 concentrations as low as 260 ppb. Second, exposure to NO2 has been 

found to enhance the inherent responsiveness of the airway to subsequent nonspecific challenges 

in controlled human exposure studies of healthy and asthmatic adults. Small but statistically

significant increases in nonspecific airway hyperresponsiveness were reported for asthmatic 

adults following 30-minute exposures to 200-300 ppb NO2 and following 1-hour exposures of 

asthmatics to 100 ppb NO2. Enhanced airway responsiveness could have important clinical 

implications for asthmatics since transient increases in airway responsiveness following NO2

exposure have the potential to increase symptoms and worsen asthma control. Together, the 

epidemiologic and experimental data sets form a plausible, consistent, and coherent description 

of a relationship between NO2 exposures and an array of adverse health effects that range from 

the onset of respiratory symptoms to hospital admission. 

In evaluating a broader range of health effects, the 2008 NOX ISA concluded evidence 

was “suggestive but not sufficient to infer a causal relationship” between short-term NO2

exposure and premature mortality and between long-term NO2 exposure and respiratory effects. 

The latter was based largely on associations observed between long-term NO2 exposure and 

decreases in lung function growth in children. Furthermore, the 2008 NOX ISA concluded that 

evidence was “inadequate to infer the presence or absence of a causal relationship” between 

short-term NO2 exposure and cardiovascular effects as well as between long-term NO2 exposure 

and cardiovascular effects, reproductive and developmental effects, premature mortality, and 

cancer.33 The conclusions for these health effect categories were informed by uncertainties in the 

evidence base such as the independent effects of NO2 exposure within the broader mixture of 
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traffic-related pollutants, limited evidence from experimental studies, and/or an overall limited 

literature base.

6.1.4 Health Effects of Carbon Monoxide

Information on the health effects of carbon monoxide (CO) can be found in the January

2010 Integrated Science Assessment for Carbon Monoxide (CO ISA).

34

 The CO ISA concludes 

that ambient concentrations of CO are associated with a number of adverse health effects.35 This 

section provides a summary of the health effects associated with exposure to ambient 

concentrations of CO.

36

 

Controlled human exposure studies of subjects with coronary artery disease show a 

decrease in the time to onset of exercise-induced angina (chest pain) and electrocardiogram 

changes following CO exposure. In addition, epidemiologic studies show associations between 

short-term CO exposure and cardiovascular morbidity, particularly increased emergency room 

visits and hospital admissions for coronary heart disease (including ischemic heart disease, 

myocardial infarction, and angina). Some epidemiologic evidence is also available for increased 

hospital admissions and emergency room visits for congestive heart failure and cardiovascular 

disease as a whole. The CO ISA concludes that a causal relationship is likely to exist between 

short-term exposures to CO and cardiovascular morbidity. It also concludes that available data 

are inadequate to conclude that a causal relationship exists between long-term exposures to CO 

and cardiovascular morbidity. 

Animal studies show various neurological effects with in-utero CO exposure. Controlled 

human exposure studies report central nervous system and behavioral effects following low-level 

CO exposures, although the findings have not been consistent across all studies. The CO ISA 

concludes the evidence is suggestive of a causal relationship with both short- and long-term 

exposure to CO and central nervous system effects.

A number of studies cited in the CO ISA have evaluated the role of CO exposure in birth 

outcomes such as preterm birth or cardiac birth defects. The epidemiologic studies provide 

limited evidence of a CO-induced effect on preterm births and birth defects, with weak evidence 

for a decrease in birth weight. Animal toxicological studies have found perinatal CO exposure to 

affect birth weight, as well as other developmental outcomes. The CO ISA concludes the 

evidence is suggestive of a causal relationship between long-term exposures to CO and 

developmental effects and birth outcomes.

Epidemiologic studies provide evidence of associations between ambient CO 

concentrations and respiratory morbidity such as changes in pulmonary function, respiratory

symptoms, and hospital admissions. A limited number of epidemiologic studies considered 

copollutants such as ozone, SO2, and PM in two-pollutant models and found that CO risk 

estimates were generally robust, although this limited evidence makes it difficult to disentangle 

effects attributed to CO itself from those of the larger complex air pollution mixture. Controlled 

human exposure studies have not extensively evaluated the effect of CO on respiratory

morbidity. Animal studies at levels of 50-100 ppm CO show preliminary evidence of altered 
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pulmonary vascular remodeling and oxidative injury. The CO ISA concludes that the evidence 

is suggestive of a causal relationship between short-term CO exposure and respiratory morbidity, 

and inadequate to conclude that a causal relationship exists between long-term exposure and 

respiratory morbidity. 

Finally, the CO ISA concludes that the epidemiologic evidence is suggestive of a causal 

relationship between short-term concentrations of CO and mortality. Epidemiologic studies 

provide evidence of an association between short-term exposure to CO and mortality, but limited 

evidence is available to evaluate cause-specific mortality outcomes associated with CO exposure. 

In addition, the attenuation of CO risk estimates which was often observed in copollutant models 

contributes to the uncertainty as to whether CO is acting alone or as an indicator for other 

combustion-related pollutants. The CO ISA also concludes that there is not likely to be a causal 

relationship between relevant long-term exposures to CO and mortality.

6.1.5 Health Effects of Air Toxics

6.1.5.1 Benzene

The EPA’s IRIS database lists benzene as a known human carcinogen (causing leukemia) 

by all routes of exposure, and concludes that exposure is associated with additional health 

effects, including genetic changes in both humans and animals and increased proliferation of 

bone marrow cells in mice.37,38,39 EPA states in its IRIS database that data indicate a causal 

relationship between benzene exposure and acute lymphocytic leukemia and suggest a 

relationship between benzene exposure and chronic non-lymphocytic leukemia and chronic 

lymphocytic leukemia. EPA’s IRIS documentation for benzene also lists a range of 2.2 x 10-6to 

7.8 x 10-6as the unit risk estimate (URE) for benzene.40,41 The International Agency for 

Research on Carcinogens (IARC) has determined that benzene is a human carcinogen and the 

U.S. Department of Health and Human Services (DHHS) has characterized benzene as a known 

human carcinogen.42,43

A number of adverse noncancer health effects including blood disorders, such as 

preleukemia and aplastic anemia, have also been associated with long-term exposure to 

benzene.44,45 The most sensitive noncancer effect observed in humans, based on current data, is 

the depression of the absolute lymphocyte count in blood.46,47 EPA’s inhalation reference 

concentration (RfC) for benzene is 30 μg/m3. The RfC is based on suppressed absolute 

lymphocyte counts seen in humans under occupational exposure conditions. In addition, recent 

work, including studies sponsored by the Health Effects Institute (HEI), provides evidence that 

biochemical responses are occurring at lower levels of benzene exposure than previously 

known.48,49,50,51 EPA’s IRIS program has not yet evaluated these new data. EPA does not 

currently have an acute reference concentration for benzene. The Agency for Toxic Substances 

and Disease Registry (ATSDR) Minimal Risk Level (MRL) for acute exposure to benzene is 29 

μg/m3for 1-14 days exposure.52,53
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6.1.5.2 1,3-Butadiene

EPA has characterized 1,3-butadiene as carcinogenic to humans by inhalation.54,55 The 

IARC has determined that 1,3-butadiene is a human carcinogen and the U.S. DHHS has 

characterized 1,3-butadiene as a known human carcinogen.56,57,58 There are numerous studies 

consistently demonstrating that 1,3-butadiene is metabolized into genotoxic metabolites by

experimental animals and humans. The specific mechanisms of 1,3-butadiene-induced 

carcinogenesis are unknown; however, the scientific evidence strongly suggests that the 

carcinogenic effects are mediated by genotoxic metabolites. Animal data suggest that females 

may be more sensitive than males for cancer effects associated with 1,3-butadiene exposure; 

there are insufficient data in humans from which to draw conclusions about sensitive 

subpopulations. The URE for 1,3-butadiene is 3 × 10-5per μg/m3.

59 1,3-butadiene also causes a 

variety of reproductive and developmental effects in mice; no human data on these effects are 

available. The most sensitive effect was ovarian atrophy observed in a lifetime bioassay of 

female mice.60 Based on this critical effect and the benchmark concentration methodology, an 

RfC for chronic health effects was calculated at 0.9 ppb (approximately 2 μg/m3).

6.1.5.3 Ethanol

EPA is planning to develop an assessment of the health effects of exposure to ethanol, a 

compound which is not currently listed on EPA’s IRIS database. Extensive health effects data 

are available for ingestion of ethanol, while data on inhalation exposure effects are sparse. In 

developing the assessment, EPA is evaluating pharmacokinetic models as a means of 

extrapolating across species (animal to human) and across exposure routes (oral to inhalation) to 

better characterize the health hazards and dose-response relationships for low levels of ethanol 

exposure in the environment.

6.1.5.4 Formaldehyde

In 1991, EPA concluded that formaldehyde is a carcinogen based on nasal tumors in 

animal bioassays.61 An Inhalation Unit Risk for cancer and a Reference Dose for oral noncancer 

effects were developed by the Agency and posted on the Integrated Risk Information System 

(IRIS) database. Since that time, the National Toxicology Program (NTP) and International 

Agency for Research on Cancer (IARC) have concluded that formaldehyde is a known human 

carcinogen.62,63,64

The conclusions by IARC and NTP reflect the results of epidemiologic research 

published since 1991 in combination with previous animal, human and mechanistic evidence. 

Research conducted by the National Cancer Institute reported an increased risk of 

nasopharyngeal cancer and specific lymphohematopoietic malignancies among workers exposed 

to formaldehyde.65,66,67 A National Institute of Occupational Safety and Health study of garment 

workers also reported increased risk of death due to leukemia among workers exposed to 

formaldehyde.68 Extended follow-up of a cohort of British chemical workers did not report 

evidence of an increase in nasopharyngeal or lymphohematopoietic cancers, but a continuing 

statistically significant excess in lung cancers was reported.69 Finally, a study of embalmers 
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reported formaldehyde exposures to be associated with an increased risk of myeloid leukemia 

but not brain cancer.70

Health effects of formaldehyde in addition to cancer were reviewed by the Agency for 

Toxics Substances and Disease Registry in 199971 and supplemented in 2010,

72 and by the World 

Health Organization.73 These organizations reviewed the literature concerning effects on the 

eyes and respiratory system, the primary point of contact for inhaled formaldehyde, including 

sensory irritation of eyes and respiratory tract, pulmonary function, nasal histopathology, and 

immune system effects. In addition, research on reproductive and developmental effects and 

neurological effects were discussed.

EPA released a draft Toxicological Review of Formaldehyde – Inhalation Assessment 

through the IRIS program for peer review by the National Research Council (NRC) and public 

comment in June 2010.74 The draft assessment reviewed more recent research from animal and 

human studies on cancer and other health effects. The NRC released their review report in April 

201175 (http://www.nap.edu/catalog.php?record_id=13142). The EPA is currently revising the 

draft assessment in response to this review.

6.1.5.5 Acetaldehyde

Acetaldehyde is classified in EPA’s IRIS database as a probable human carcinogen, 

based on nasal tumors in rats, and is considered toxic by the inhalation, oral, and intravenous 

routes.76 The URE in IRIS for acetaldehyde is 2.2 × 10-6per μg/m3.

77

 Acetaldehyde is 

reasonably anticipated to be a human carcinogen by the U.S. DHHS in the 12th Report on 

Carcinogens and is classified as possibly carcinogenic to humans (Group 2B) by the IARC.78,79 

EPA is currently conducting a reassessment of cancer risk from inhalation exposure to 

acetaldehyde.

The primary noncancer effects of exposure to acetaldehyde vapors include irritation of 

the eyes, skin, and respiratory tract.80 In short-term (4 week) rat studies, degeneration of 

olfactory epithelium was observed at various concentration levels of acetaldehyde exposure.81,82 

Data from these studies were used by EPA to develop an inhalation reference concentration of 9 

μg/m3. Some asthmatics have been shown to be a sensitive subpopulation to decrements in 

functional expiratory volume (FEV1 test) and bronchoconstriction upon acetaldehyde 

inhalation.83 The agency is currently conducting a reassessment of the health hazards from 

inhalation exposure to acetaldehyde. 

6.1.5.6 Acrolein

EPA most recently evaluated the toxicological and health effects literature related to 

acrolein in 2003 and concluded that the human carcinogenic potential of acrolein could not be 

determined because the available data were inadequate. No information was available on the 

carcinogenic effects of acrolein in humans and the animal data provided inadequate evidence of 

carcinogenicity.

84

 The IARC determined in 1995 that acrolein was not classifiable as to its 

carcinogenicity in humans.85 
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Lesions to the lungs and upper respiratory tract of rats, rabbits, and hamsters have been 

observed after subchronic exposure to acrolein.86 The Agency has developed an RfC for acrolein 

of 0.02 μg/m3 and an RfD of 0.5 μg/kg-day.

87

 EPA is considering updating the acrolein 

assessment with data that have become available since the 2003 assessment was completed.

Acrolein is extremely acrid and irritating to humans when inhaled, with acute exposure 

resulting in upper respiratory tract irritation, mucus hypersecretion and congestion. The intense 

irritancy of this carbonyl has been demonstrated during controlled tests in human subjects, who 

suffer intolerable eye and nasal mucosal sensory reactions within minutes of exposure.88 These 

data and additional studies regarding acute effects of human exposure to acrolein are 

summarized in EPA’s 2003 IRIS Human Health Assessment for acrolein.89 Studies in humans 

indicate that levels as low as 0.09 ppm (0.21 mg/m3) for five minutes may elicit subjective 

complaints of eye irritation with increasing concentrations leading to more extensive eye, nose 

and respiratory symptoms. Acute exposures in animal studies report bronchial hyperresponsiveness. Based on animal data (more pronounced respiratory irritancy in mice with 

allergic airway disease in comparison to non-diseased mice90) and demonstration of similar 

effects in humans (e.g., reduction in respiratory rate), individuals with compromised respiratory

function (e.g., emphysema, asthma) are expected to be at increased risk of developing adverse 

responses to strong respiratory irritants such as acrolein. EPA does not currently have an acute 

reference concentration for acrolein. The available health effect reference values for acrolein 

have been summarized by EPA and include an ATSDR MRL for acute exposure to acrolein of 

7 μg/m3for 1-14 days exposure; and Reference Exposure Level (REL) values from the 

California Office of Environmental Health Hazard Assessment (OEHHA) for one-hour and 8-

hour exposures of 2.5 μg/m3and 0.7 μg/m3, respectively.91 

6.1.5.7 PAN

PAN (peroxy acetyl nitrate) has not been evaluated by EPA’s IRIS program. Information 

regarding the potential carcinogenicity of PAN is limited. As noted in the EPA air quality 

criteria document for ozone and related photochemical oxidants, cytogenetic studies indicate that 

PAN is not a potent mutagen, clastogen (a compound that can cause breaks in chromosomes), or 

DNA-damaging agent in mammalian cells either in vivo or in vitro. Some studies suggest that 

PAN may be a weak bacterial mutagen at concentrations much higher than exist in present urban 

atmospheres.92

Effects of ground-level smog causing intense eye irritation have been attributed to 

photochemical oxidants, including PAN.93 Animal toxicological information on the inhalation 

effects of the non-ozone oxidants has been limited to a few studies on PAN. Acute exposure to 

levels of PAN can cause changes in lung morphology, behavioral modifications, weight loss, and 

susceptibility to pulmonary infections. Human exposure studies indicate minor pulmonary

function effects at high PAN concentrations, but large inter-individual variability precludes 

definitive conclusions.94
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6.1.5.8 Polycyclic Organic Matter

The term polycyclic organic matter (POM) defines a broad class of compounds that 

includes the polycyclic aromatic hydrocarbon compounds (PAHs). One of these compounds, 

naphthalene, is discussed separately below. POM compounds are formed primarily from 

combustion and are present in the atmosphere in gas and particulate form. Cancer is the major 

concern from exposure to POM. Epidemiologic studies have reported an increase in lung cancer 

in humans exposed to diesel exhaust, coke oven emissions, roofing tar emissions, and cigarette 

smoke; all of these mixtures contain POM compounds.9596 Animal studies have reported 

respiratory tract tumors from inhalation exposure to benzo[a]pyrene and alimentary tract and 

liver tumors from oral exposure to benzo[a]pyrene.97 In 1997 EPA classified seven PAHs 

(benzo[a]pyrene, benz[a]anthracene, chrysene, benzo[b]fluoranthene, benzo[k]fluoranthene, 

dibenz[a,h]anthracene, and indeno[1,2,3-cd]pyrene) as Group B2, probable human carcinogens.98 

Since that time, studies have found that maternal exposures to PAHs in a population of pregnant 

women were associated with several adverse birth outcomes, including low birth weight and 

reduced length at birth, as well as impaired cognitive development in preschool children (3 years 

of age).99,100 These and similar studies are being evaluated as a part of the ongoing IRIS 

assessment of health effects associated with exposure to benzo[a]pyrene.

6.1.5.9 Naphthalene

Naphthalene is found in small quantities in gasoline and diesel fuels. Naphthalene 

emissions have been measured in larger quantities in both gasoline and diesel exhaust compared 

with evaporative emissions from mobile sources, indicating it is primarily a product of 

combustion. Acute (short-term) exposure of humans to naphthalene by inhalation, ingestion, or 

dermal contact is associated with hemolytic anemia and damage to the liver and the nervous 

system.101 Chronic (long term) exposure of workers and rodents to naphthalene has been 

reported to cause cataracts and retinal damage.102 EPA released an external review draft of a 

reassessment of the inhalation carcinogenicity of naphthalene based on a number of recent 

animal carcinogenicity studies.103 The draft reassessment completed external peer review.104 

Based on external peer review comments received, a revised draft assessment that considers all 

routes of exposure, as well as cancer and noncancer effects, is under development. The external 

review draft does not represent official agency opinion and was released solely for the purposes 

of external peer review and public comment. The National Toxicology Program listed 

naphthalene as "reasonably anticipated to be a human carcinogen" in 2004 on the basis of 

bioassays reporting clear evidence of carcinogenicity in rats and some evidence of 

carcinogenicity in mice.105 California EPA has released a new risk assessment for naphthalene, 

and the IARC has reevaluated naphthalene and re-classified it as Group 2B: possibly

carcinogenic to humans.106 

Naphthalene also causes a number of chronic non-cancer effects in animals, including 

abnormal cell changes and growth in respiratory and nasal tissues.107 The current EPA IRIS 

assessment includes noncancer data on hyperplasia and metaplasia in nasal tissue that form the 
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basis of the inhalation RfC of 3 μg/m3.

108

 The ATSDR MRL for acute exposure to naphthalene 

is 0.6 mg/kg/day.

6.1.5.10 Other Air Toxics

In addition to the compounds described above, other compounds in gaseous hydrocarbon 

and PM emissions from vehicles will be affected by this proposal. Mobile source air toxic 

compounds that would potentially be impacted include ethylbenzene, propionaldehyde, toluene, 

and xylene. Information regarding the health effects of these compounds can be found in EPA’s 

IRIS database.109

6.1.6 Traffic-associated health effects

In addition to health concerns resulting from specific air pollutants, a large number of 

studies have examined the health status of populations near major roadways. These studies 

frequently have employed exposure metrics that are not specific to individual pollutants, but 

rather reflect the large number of different pollutants found in elevation near major roads.

In this section of the RIA, information on health effects associated with air quality near 

major roads or traffic in general is summarized. Generally, the section makes use of publications 

that systematically review literature on a given health topic. In particular, this section makes 

frequent reference of a report of by the Health Effects Institute (HEI) Panel on the Health Effects 

of Traffic-Related Air Pollution, published in 2010 as a review of relevant studies.110,111 Other 

systematic reviews of relevant literature are cited were appropriate.

6.1.6.1 Populations near major roads

Numerous studies have estimated the size and demographics of populations that live near 

major roads. Other studies have estimated the number of schools near major roads, and the 

populations of students in such schools.

Every two years, the U.S. Census Bureau’s American Housing Survey (AHS) has 

reported whether housing units are within 300 feet of an “airport, railroad, or highway with four 

or more lanes.” The 2009 survey reports that over 22 million homes, or 17 percent of all housing 

units in the U.S., were located in such areas. Assuming that populations and housing units are in 

the same locations, this corresponds to a population of more than 50 million U.S. residents in 

close proximity to high-traffic roadways or other transportation sources. According to the 

Central Intelligence Agency’s World Factbook, in 2010, the United States had 6,506,204 km or 

roadways, 224,792 km of railways, and 15,079 airports. As such, highways represent the 

overwhelming majority of transportation facilities described by this factor in the AHS.

The AHS reports are published every two years, and until 2011 recorded whether homes 

were located near highways with four or more lanes, railroads, or airports. As such, trends in the 

AHS can be reported to describe whether a greater or lesser proportion of homes are located near 

major roads over time. Figure 6-1 depicts trends in the number and proportion of homes located 
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near major transportation sources, which generally indicate large roadways. As the figure 

indicates, since 2005, there has been a substantial increase in the number and percentage of 

homes located near major transportation sources. As such, the population in close proximity to 

these sources, which may be affected by near-road air quality and health concerns, appears to 

have increased over time.

0.0%

2.0%

4.0%

6.0%

8.0%

10.0%

12.0%

14.0%

16.0%

18.0%

10

12

14

16

18

20

22

24

1997 1999 2001 2003 2005 2007 2009

P

ercent N

e

arInfrastructure

Million H

o

m

es

Year of American Housing Survey

Million Housing Units

Near 4+ Lane Highway,

Railroad, or Airport

Figure 6-1 Trends in Populations Near Large Highways, Railroads, and Airports

Furthermore, according to data from the 2008 American Time Use Survey (ATUS), 

conducted by the Bureau of Labor Statistics (BTS), Americans spend more than an hour 

traveling each day, on average.112 Although the ATUS does not indicate their mode of travel, the 

majority of trips undertaken nationally is by motor vehicle.113 As such, daily travel activity 

brings nearly all residents into a high-exposure microenvironment for part of the day. 

6.1.6.2 Premature mortality

The HEI panel report concluded that evidence linking traffic-associated air pollution with 

premature mortality from all causes was “suggestive but not sufficient” to infer a causal 

relationship. This conclusion was based largely on several long-term studies that “qualitatively” 

examined whether or not someone was exposed to traffic-associated air pollution. In addition, 

based on several short-term studies of exposure, the panel concluded that there was evidence that 
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there was “suggestive but not sufficient” evidence to infer a causal relation between trafficrelated exposure and cardiovascular mortality. 

6.1.6.3 Cardiovascular effects

6.1.6.3.1 Cardiac physiology

Exposure to traffic-associated pollutants has been associated with changes in cardiac 

physiology, including cardiac function. One common measure of cardiac function is heart rate 

variability (HRV), an indicator of the heart’s ability to respond to variations in stress, reflecting 

the nervous system’s ability to regulate the heart.114 Reduced HRV is associated with adverse 

cardiovascular events, such as myocardial infarction, in heart disease patients. The HEI panel 

concluded that available evidence provides evidence for a causal association between exposure 

to traffic-related pollutants and reduced control of HRV by the nervous system. Overall, the 

panel concluded that the evidence was “suggestive but not sufficient” to infer a causal relation 

between traffic-related pollutants and cardiac function. Studies suggest that the HRV changes 

from traffic-related air pollution result in changes to heart rhythms, which can lead to 

arrhythmia.115,116

6.1.6.3.2 Heart attack and atherosclerosis

The HEI panel concluded that epidemiologic evidence of the association between trafficrelated pollutants and heart attacks and atherosclerosis was “suggestive but not sufficient” to 

infer a causal association. In addition, the panel concluded that the toxicology studies they

reviewed provided “suggestive evidence that exposure to traffic emissions, including ambient 

and laboratory-generated [PM] and diesel- and gasoline-engine exhaust, alters cardiovascular 

function.” The panel noted there are few studies of human volunteers exposed to real-world 

traffic mixture, which were not entirely consistent. The panel notes that the studies provide 

consistent evidence for exposure to PM and impaired cardiovascular responses. In addition to 

the HEI study, several other reviews of available evidence conclude that there is evidence 

supporting a causal association between traffic-related air pollution and cardiovascular 

disease.117

A number of mechanisms for cardiovascular disease are highlighted in the HEI and AHA 

report, including modified blood vessel endothelial function (e.g, the ability to dilate), 

atherosclerosis, and oxidative stress. The HEI review cites “two well executed studies” in which 

hospitalization for acute myocardial infarction (i.e., heart attack) were associated with traffic 

exposures and a prospective study finding higher rates of arterial hardening and coronary heart 

disease near traffic.
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6.1.6.4 Respiratory effects

6.1.6.4.1 Asthma

Pediatric asthma and asthma symptoms are the effects that have been evaluated by the 

largest number of studies in the epidemiologic literature on the topic. In general, studies 

consistently show effects of residential or school exposure to traffic and asthma symptoms, and 

the effects are frequently statistically significant. Studies have employed both short-term and 

long-term exposure metrics, and a range of different respiratory measures. HEI Special Report 

17 (HEI Panel on the Health Effects of Traffic-Related Air Pollution, 2010) concluded that there 

is sufficient evidence for a causal association between exposure to traffic-related air pollution 

and exacerbation of asthma symptoms in children. 

While there is general consistency in studies examining asthma incidence in children, the 

available studies employ different definitions of asthma (e.g., self-reported vs. hospital records), 

methods of exposure assessment, and population age ranges. As such, the overall evidence, 

while supportive of an association between traffic exposure and new onset asthma, are less 

consistent than for asthma symptoms. The HEI report determined that there is “sufficient” or 

“suggestive” evidence of a causal relationship between exposure to traffic-related air pollution 

and incident (new onset) asthma in children (HEI Panel on the Health Effects of Traffic-Related 

Air Pollution, 2010). A recent meta-analysis of studies on incident asthma and air pollution in 

general, based on studies dominated by traffic-linked exposure metrics, also concluded that 

available evidence that exposures is consistent with a effect of exposure on asthma incidence 

(Anderson et al., 2011). The study reported excess main risk estimates for different pollutants 

ranging from 7-16 percent per 10 g/m3of long-term exposure (random effects models). Other 

qualitative reviews (Salam et al., 2008; Braback and Forsberg, 2009) conclude that available 

evidence is consistent with the hypothesis that traffic-associated air pollutants are associated with 

incident asthma.

6.1.6.4.2 Chronic obstructive pulmonary disease (COPD)

The HEI panel reviewed available studies examining COPD in the context of trafficassociated air pollution. Because of how the panel selected studies for inclusion in review, there 

were only two studies that they used to review the available evidence. Both studies reported 

some positive associations, but not for all traffic metrics. The small number of studies and lack 

of consistency across traffic metrics led the panel to conclude that there is insufficient evidence 

for traffic-associated air pollution causing COPD.

6.1.6.4.3 Allergy

There are numerous human and animal experimental studies that provides strongly 

suggestive evidence that traffic-related air pollutants can enhance allergic responses to common 

allergens.118,119,120 However, in its review of 16 epidemiologic studies that address traffic-related 

air pollution’s effect on allergies, the HEI expert panel (HEI, 2010) reported that only two such 

studies showed consistently positive associations. As a result, despite the strong experimental 
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evidence, the panel concluded that there is “inadequate/insufficient” evidence of an association 

between allergy and traffic-associated air pollution. As noted above, the HEI panel considered 

toxicological evidence only based on whether or not they provide mechanistic support for 

observations and inferences derived from epidemiology.

6.1.6.4.4 Lung function

There are numerous measurements of breathing (spirometry) that indicate the presence or 

degree of airway disease, such as asthma and chronic obstructive pulmonary disease (COPD). 

Forced vital capacity (FVC) is measured when a patient maximally fills their lungs and then 

blows their hardest in completely exhaling. The peak expiratory flow (PEF) is the maximum air 

flow achievable during exhalation. The forced expiratory volume in the first second of 

exhalation is referred to as FEV1. FEV1 and PEF reflect the function of the large airways. FVC 

and FEV1, along with their ratio (FVC/FEV1) are used to classify airway obstruction in asthma 

and COPD. Measurements of air flow at various times during forced exhalation, such as 25 

percent, 50 percent, and 75 percent, are also used. The flow at 75 percent of forced exhalation 

(FEF75) reflects the status of small airways, which asthma and COPD affect. 

The HEI panel concluded that the available literature suggests that long-term exposure to 

traffic-related air pollution is associated with reduced lung function in adolescents and young 

adults and that lung function is lower in populations in areas with high traffic-related air 

pollutant levels. However, the panel noted the difficulty of disentangling traffic-specific 

exposures from urban air pollution in general. The studies reviewed that were more specifically

oriented toward traffic were not consistent in their findings. As a result, the panel found that the 

evidence linking lung function and traffic exposure is “inadequate and insufficient” to infer a 

causal relationship. 

6.1.6.5 Reproductive and developmental effects

Several studies have reported associations between traffic-related air pollution and 

adverse birth outcomes, such as preterm birth and low birth weight. At the time of the HEI 

review, the panel concluded that evidence for adverse birth outcomes being causally associated 

with traffic-related exposures was “inadequate and insufficient.” Only four studies met the 

panel’s inclusion criteria, and had limited geographic coverage. One study provided evidence of 

small but consistently increased risks using multiple exposure metrics. No studies were at the 

time available that examined traffic-specific exposures and congenital abnormalities. Since then, 

several studies investigating birth outcomes have been published, but no new systematic reviews. 

One new meta-analysis of air pollution and congenital abnormalities has been published, though 

none of the reviewed studies includes traffic-specific exposure information.

The HEI panel also reviewed toxicological studies of traffic-related air pollutants and 

fertility. While numerous studies examining animal or human exposure and sperm count have 

been published, the panel concluded that the generally high exposure concentrations employed in 

the studies limited the applicability to typical ambient concentrations. Because there was no 
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overlap in the effects studied by epidemiology and toxicology studies, no synthesis review of the 

combined literature was undertaken.

Since the HEI panel’s publication, a systematic review and meta-analysis of air pollution 

and congenital abnormalities was published.121 In that review, only one study directly included 

nearby traffic in its exposure analysis. As such, there are so systematic reviews that specifically

address traffic’s impact on congenital abnormalities.

6.1.6.6 Cancer

6.1.6.6.1 Childhood cancer

A number of studies examining various types of childhood cancer have been published 

with mixed results. The HEI panel concluded that the available epidemiologic evidence was 

“inadequate and insufficient” to infer a causal relationship between traffic-related air pollution 

and childhood cancer. An earlier review article on the topic noted that studies reporting positive 

effects tended to be small, while those with null effects tended to be larger, suggesting the 

potential for publication bias in the available literature.122

6.1.6.6.2 Adult cancer

Several studies have examined the risk of adult lung cancers in relation to exposure to 

traffic-related air pollutants. The HEI panel evaluated four such studies, and rated the available 

evidence as “inadequate and insufficient” to infer a causal relation for non-occupational lung 

cancer.

6.1.6.7 Neurological effects

The HEI panel found that current toxicologic and epidemiologic literature on the 

neurotoxicity of traffic-related air pollution was inadequate for their evaluation. The panel noted 

that there were a number of toxicologic studies of traffic-associated pollutants, but found them to 

have diverse exposure protocols, animal models, and endpoints, making them unsuitable for 

systematic evaluation.

6.2 Environmental Effects of Criteria and Toxic Pollutants

6.2.1 Visibility Degradation

Visibility can be defined as the degree to which the atmosphere is transparent to visible 

light.123 Visibility impairment is caused by light scattering and absorption by suspended 

particles and gases. Visibility is important because it has direct significance to people’s 

enjoyment of daily activities in all parts of the country. Individuals value good visibility for the 

well-being it provides them directly, where they live and work, and in places where they enjoy 

recreational opportunities. Visibility is also highly valued in significant natural areas, such as 
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national parks and wilderness areas, and special emphasis is given to protecting visibility in these 

areas. For more information on visibility see the final 2009 PM ISA.124

EPA is working to address visibility impairment. In 1999, EPA finalized the regional 

haze program (64 FR 35714) to protect the visibility in Mandatory Class I Federal areas. There 

are 156 national parks, forests and wilderness areas categorized as Mandatory Class I Federal 

areas (62 FR 38680-38681, July 18, 1997). These areas are defined in CAA section 162 as those 

national parks exceeding 6,000 acres, wilderness areas and memorial parks exceeding 5,000 

acres, and all international parks which were in existence on August 7, 1977. Figure 6-2 shows 

the location of the 156 Mandatory Class I Federal areas.

Figure 6-2 Mandatory Class I Federal Areas in the U.S.

EPA has also concluded that PM2.5 causes adverse effects on visibility in other areas that 

are not protected by the Regional Haze Rule, depending on PM2.5 concentrations and other 

factors that control their visibility impact effectiveness such as dry chemical composition and 

relative humidity (i.e., an indicator of the water composition of the particles). EPA revised the 

PM2.5 standards in December 2012 and established a target level of protection that is expected to 

be met through attainment of the existing secondary standards for PM2.5. 

6.2.1.1 Visibility Monitoring

In conjunction with the U.S. National Park Service, the U.S. Forest Service, other Federal 

land managers, and State organizations in the U.S., the U.S. EPA has supported visibility

monitoring in national parks and wilderness areas since 1988. The monitoring network was 

originally established at 20 sites, but it has now been expanded to 110 sites that represent all but 
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one of the 156 Mandatory Federal Class I areas across the country (see Figure 6-2). This longterm visibility monitoring network is known as IMPROVE (Interagency Monitoring of Protected 

Visual Environments).

IMPROVE provides direct measurement of fine particles that contribute to visibility

impairment. The IMPROVE network employs aerosol measurements at all sites, and optical and 

scene measurements at some of the sites. Aerosol measurements are taken for PM10 and PM2.5

mass, and for key constituents of PM2.5, such as sulfate, nitrate, organic and elemental carbon

OC and EC), soil dust, and several other elements. Measurements for specific aerosol 

constituents are used to calculate "reconstructed" aerosol light extinction by multiplying the mass 

for each constituent by its empirically-derived scattering and/or absorption efficiency, with 

adjustment for the relative humidity. Knowledge of the main constituents of a site's light 

extinction "budget" is critical for source apportionment and control strategy development. In 

addition to this indirect method of assessing light extinction, there are optical measurements 

which directly measure light extinction or its components. Such measurements are made 

principally with either a nephelometer to measure light scattering, some sites also include an 

aethalometer for light absorption, or at a few sites using a transmissometer, which measures total 

light extinction. Scene characteristics are typically recorded using digital or video photography

and are used to determine the quality of visibility conditions (such as effects on color and 

contrast) associated with specific levels of light extinction as measured under both direct and 

aerosol-related methods. Directly measured light extinction is used under the IMPROVE 

protocol to cross check that the aerosol-derived light extinction levels are reasonable in 

establishing current visibility conditions. Aerosol-derived light extinction is used to document 

spatial and temporal trends and to determine how changes in atmospheric constituents would 

affect future visibility conditions.

Annual average visibility conditions (reflecting light extinction due to both anthropogenic 

and non-anthropogenic sources) vary regionally across the U.S. Visibility is typically worse in 

the summer months and the rural East generally has higher levels of impairment than remote 

sites in the West. Figures 9-9 through 9-11 in the PM ISA detail the percent contributions to 

particulate light extinction for ammonium nitrate and sulfate, EC and OC, and coarse mass and 

fine soil, by season.125

6.2.2 Particulate Matter Deposition

Particulate matter contributes to adverse effects on vegetation and ecosystems, and to 

soiling and materials damage. These welfare effects result predominately from exposure to 

excess amounts of specific chemical species, regardless of their source or predominant form 

(particle, gas or liquid). The following characterizations of the nature of these environmental

effects are based on information contained in the 2009 PM ISA and the 2005 PM Staff Paper as 

well as the 2008 Integrated Science Assessment for Oxides of Nitrogen and Sulfur- Ecological 

Criteria (secondary NOX/SOX ISA).

126,127,128
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6.2.2.1 Deposition of Nitrogen and Sulfur

Nitrogen and sulfur interactions in the environment are highly complex. Both nitrogen 

and sulfur are essential, and sometimes limiting, nutrients needed for growth and productivity of 

ecosystem components (e.g. algae, plants). In terrestrial and aquatic ecosystems excesses of 

nitrogen or sulfur can lead to acidification and nutrient enrichment.129 

The process of acidification affects both freshwater aquatic and terrestrial ecosystems. 

Acid deposition causes acidification of sensitive surface waters. The effects of acid deposition 

on aquatic systems depend largely upon the ability of the ecosystem to neutralize the additional 

acid. As acidity increases, aluminum leached from soils, flows into lakes and streams and can be 

toxic to both terrestrial and aquatic biota. The lower pH concentrations and higher aluminum 

levels resulting from acidification make it difficult for some fish and other aquatic organisms to 

survive, grow, and reproduce. Research on effects of acid deposition on forest ecosystems has 

come to focus increasingly on the biogeochemical processes that affect uptake, retention, and 

cycling of nutrients within these ecosystems. Decreases in available base cations from soils are 

at least partly attributable to acid deposition. Base cation depletion is a cause for concern 

because of the role these ions play in acid neutralization, and because calcium, magnesium and 

potassium are essential nutrients for plant growth and physiology. Changes in the relative 

proportions of these nutrients, especially in comparison with aluminum concentrations, have 

been associated with declining forest health.

At current ambient levels, risks to vegetation from short-term exposures to dry deposited 

particulate nitrate or sulfate are low. However, when found in acid or acidifying deposition, such 

particles do have the potential to cause direct leaf injury. Specifically, the responses of forest 

trees to acid precipitation (rain, snow) include accelerated weathering of leaf cuticular surfaces, 

increased permeability of leaf surfaces to toxic materials, water, and disease agents; increased 

leaching of nutrients from foliage; and altered reproductive processes—all which serve to 

weaken trees so that they are more susceptible to other stresses (e.g., extreme weather, pests, 

pathogens). Acid deposition with levels of acidity associated with the leaf effects described 

above are currently found in some locations in the eastern U.S.130 Even higher concentrations of 

acidity can be present in occult depositions (e.g., fog, mist or clouds) which more frequently

impacts higher elevations. Thus, the risk of leaf injury occurring from acid deposition in some 

areas of the eastern U.S. is high. Nitrogen deposition has also been shown to impact ecosystems 

in the western U.S. For example, a study conducted in the Columbia River Gorge National 

Scenic Area (CRGNSA), located along a portion of the Oregon/Washington border, indicates 

that lichen communities in the CRGNSA have shifted to a higher proportion of nitrophilous 

species and the nitrogen content of lichen tissue is elevated.131 Lichens are sensitive indicators 

of nitrogen deposition effects to terrestrial ecosystems and the lichen studies in the Columbia 

River Gorge clearly show that ecological effects from air pollution are occurring.

Some of the most significant detrimental effects associated with excess nitrogen 

deposition are those associated with a condition known as nitrogen saturation. Nitrogen 

saturation is the condition in which nitrogen inputs from atmospheric deposition and other 
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sources exceed the biological requirements of the ecosystem. The effects associated with 

nitrogen saturation include: (1) decreased productivity, increased mortality, and/or shifts in plant 

community composition, often leading to decreased biodiversity in many natural habitats 

wherever atmospheric reactive nitrogen deposition increases significantly above background and 

critical thresholds are exceeded; (2) leaching of excess nitrate and associated base cations from 

soils into streams, lakes, and rivers, and mobilization of soil aluminum; and (3) fluctuation of 

ecosystem processes such as nutrient and energy cycles through changes in the functioning and 

species composition of beneficial soil organisms.132

In the U.S. numerous forests now show severe symptoms of nitrogen saturation. These 

forests include: the northern hardwoods and mixed conifer forests in the Adirondack and 

Catskill Mountains of New York; the red spruce forests at Whitetop Mountain, Virginia, and 

Great Smoky Mountains National Park, North Carolina; mixed hardwood watersheds at Fernow 

Experimental Forest in West Virginia; American beech forests in Great Smoky Mountains 

National Park, Tennessee; mixed conifer forests and chaparral watersheds in southern California 

and the southwestern Sierra Nevada in Central California; the alpine tundra/subalpine conifer 

forests of the Colorado Front Range; and red alder forests in the Cascade Mountains in 

Washington.

Excess nutrient inputs into aquatic ecosystems (i.e. streams, rivers, lakes, estuaries or 

oceans) either from direct atmospheric deposition, surface runoff, or leaching from nitrogen 

saturated soils into ground or surface waters can contribute to conditions of severe water oxygen 

depletion; eutrophication and algae blooms; altered fish distributions, catches, and physiological 

states; loss of biodiversity; habitat degradation; and increases in the incidence of disease.

Atmospheric deposition of nitrogen is a significant source of total nitrogen to many

estuaries in the United States. The amount of nitrogen entering estuaries that is ultimately

attributable to atmospheric deposition is not well-defined. On an annual basis, atmospheric 

nitrogen deposition may contribute significantly to the total nitrogen load, depending on the size 

and location of the watershed. In addition, episodic nitrogen inputs, which may be ecologically 

important, may play a more important role than indicated by the annual average concentrations. 

Estuaries in the U.S. that suffer from nitrogen enrichment often experience a condition known as 

eutrophication. Symptoms of eutrophication include changes in the dominant species of 

phytoplankton, low levels of oxygen in the water column, fish and shellfish kills, outbreaks of 

toxic alga, and other population changes which can cascade throughout the food web. In 

addition, increased phytoplankton growth in the water column and on surfaces can attenuate light

causing declines in submerged aquatic vegetation, which serves as an important habitat for many 

estuarine fish and shellfish species.

Severe and persistent eutrophication often directly impacts human activities. For 

example, losses in the nation’s fishery resources may be directly caused by fish kills associated 

with low dissolved oxygen and toxic blooms. Declines in tourism occur when low dissolved 

oxygen causes noxious smells and floating mats of algal blooms create unfavorable aesthetic 

conditions. Risks to human health increase when the toxins from algal blooms accumulate in 
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edible fish and shellfish, and when toxins become airborne, causing respiratory problems due to 

inhalation. According to a National Oceanic and Atmospheric Administration report, more than 

half of the nation’s estuaries have moderate to high expressions of at least one of these symptoms 

– an indication that eutrophication is well developed in more than half of U.S. estuaries.133

6.2.2.2 Deposition of Heavy Metals

Heavy metals, including cadmium, copper, lead, chromium, mercury, nickel and zinc, 

have the greatest potential for impacting forest growth.

134

 Investigation of trace metals near 

roadways and industrial facilities indicate that a substantial load of heavy metals can accumulate 

on vegetative surfaces. Copper, zinc, and nickel have been documented to cause direct toxicity

to vegetation under field conditions. Trace metals are associated with ambient PM, however 

little research has been conducted on the effects associated with mixtures of contaminants. 

While metals typically exhibit low solubility, limiting their bioavailability and direct toxicity, 

chemical transformations of metal compounds occur in the environment, particularly in the 

presence of acidic or other oxidizing species. These chemical changes influence the mobility

and toxicity of metals in the environment. Once taken up into plant tissue, a metal compound 

can undergo chemical changes, exert toxic effects on the plant itself, accumulate and be passed 

along to herbivores or can re-enter the soil and further cycle in the environment. Although there 

has been no direct evidence of a physiological association between tree injury and heavy metal 

exposures, heavy metals have been implicated because of similarities between metal deposition 

patterns and forest decline. This hypothesized relationship/correlation was further explored in 

high elevation forests in the northeastern U.S. These studies measured levels of a group of 

intracellular compounds found in plants that bind with metals and are produced by plants as a 

response to sublethal concentrations of heavy metals. These studies indicated a systematic and 

significant increase in concentrations of these compounds associated with the extent of tree 

injury. These data strongly imply that metal stress causes tree injury and contributes to forest 

decline in the northeastern United States.135 Contamination of plant leaves by heavy metals can 

lead to elevated soil levels. Trace metals absorbed into the plant frequently bind to the leaf 

tissue, and then are lost when the leaf drops. As the fallen leaves decompose, the heavy metals 

are transferred into the soil.136,137 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. Many of the most important effects 

occur in the soil. 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. These 

organisms are essential participants in the nutrient cycles that make elements available for plant 

uptake. Changes in the soil environment that influence the role of the bacteria and fungi in 

nutrient cycling determine plant and ultimately ecosystem response.138

The environmental sources and cycling of mercury are currently of particular concern due 

to the bioaccumulation and biomagnification of this metal in aquatic ecosystems and the potent

toxic nature of mercury in the forms in which is it ingested by people and other animals. 

Mercury is unusual compared with other metals in that it largely partitions into the gas phase (in 

elemental form), and therefore has a longer residence time in the atmosphere than a metal found 
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predominantly in the particle phase. This property enables mercury to travel far from the 

primary source before being deposited and accumulating in the aquatic ecosystem. The major 

source of mercury in the Great Lakes is from atmospheric deposition, accounting for 

approximately eighty percent of the mercury in Lake Michigan.139,140 Over fifty percent of the 

mercury in the Chesapeake Bay has been attributed to atmospheric deposition.141 Overall, the 

National Science and Technology Council identifies atmospheric deposition as the primary

source of mercury to aquatic systems.142 Forty-four states have issued health advisories for the 

consumption of fish contaminated by mercury; however, most of these advisories are issued in 

areas without a mercury point source.

Elevated levels of zinc and lead have been identified in streambed sediments, and these 

elevated levels have been correlated with population density and motor vehicle use.143,144,145 

Zinc and nickel have also been identified in urban water and soils. In addition, platinum, 

palladium, and rhodium, metals found in the catalysts of modern motor vehicles, have been 

measured at elevated levels along roadsides.146 Plant uptake of platinum has been observed at 

these locations.

6.2.2.3 Deposition of Polycyclic Organic Matter

Polycyclic organic matter (POM) is a byproduct of incomplete combustion and consists 

of organic compounds with more than one benzene ring and a boiling point greater than or equal 

to 100 degrees centigrade.147 Polycyclic aromatic hydrocarbons (PAHs) are a class of POM that 

contains compounds which are known or suspected carcinogens.

Major sources of PAHs include mobile sources. PAHs in the environment may be 

present as a gas or adsorbed onto airborne particulate matter. Since the majority of PAHs are 

adsorbed onto particles less than 1.0 μm in diameter, long range transport is possible. However, 

studies have shown that PAH compounds adsorbed onto diesel exhaust particulate and exposed 

to ozone have half lives of 0.5 to 1.0 hours.

148

 

Since PAHs are insoluble, the compounds generally are particle reactive and accumulate 

in sediments. Atmospheric deposition of particles is believed to be the major source of PAHs to 

the sediments of Lake Michigan.149,150 Analyses of PAH deposition in Chesapeake and 

Galveston Bay indicate that dry deposition and gas exchange from the atmosphere to the surface 

water predominate.151,152 Sediment concentrations of PAHs are high enough in some segments 

of Tampa Bay to pose an environmental health threat. EPA funded a study to better characterize 

the sources and loading rates for PAHs into Tampa Bay.

153

 PAHs that enter a water body 

through gas exchange likely partition into organic rich particles and can be biologically recycled, 

while dry deposition of aerosols containing PAHs tend to be more resistant to biological 

recycling.154 Thus, dry deposition is likely the main pathway for PAH concentrations in 

sediments while gas/water exchange at the surface may lead to PAH distribution into the food 

web, leading to increased health risk concerns.

Trends in PAH deposition levels are difficult to discern because of highly variable 

ambient air concentrations, lack of consistency in monitoring methods, and the significant 
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influence of local sources on deposition levels.155 Van Metre et al. noted PAH concentrations in 

urban reservoir sediments have increased by 200-300 percent over the last forty years and 

correlate with increases in automobile use.156 

Cousins et al. estimate that more than ninety percent of semi-volatile organic compound 

(SVOC) emissions in the United Kingdom deposit on soil.157 An analysis of PAH concentrations 

near a Czechoslovakian roadway indicated that concentrations were thirty times greater than 

background.158

6.2.2.4 Materials Damage and Soiling

The effects of the deposition of atmospheric pollution, including ambient PM, on 

materials are related to both physical damage and impaired aesthetic qualities. The deposition of 

PM (especially sulfates and nitrates) can physically affect materials, adding to the effects of 

natural weathering processes, by potentially promoting or accelerating the corrosion of metals, 

by degrading paints, and by deteriorating building materials such as concrete and limestone. 

Only chemically active fine particles or hygroscopic coarse particles contribute to these physical 

effects. In addition, the deposition of ambient PM can reduce the aesthetic appeal of buildings 

and culturally important articles through soiling. Particles consisting primarily of carbonaceous 

compounds cause soiling of commonly used building materials and culturally important items 

such as statues and works of art.

6.2.3 Plant and Ecosystem Effects of Ozone

The welfare effects of ozone can be observed across a variety of scales, i.e. subcellular, 

cellular, leaf, whole plant, population and ecosystem. Ozone effects that begin at small spatial

scales, such as the leaf of an individual plant, when they occur at sufficient magnitudes (or to a 

sufficient degree) can result in effects being propagated along a continuum to larger and larger 

spatial scales. For example, effects at the individual plant level, such as altered rates of leaf gas 

exchange, growth and reproduction can, when widespread, result in broad changes in 

ecosystems, such as productivity, carbon storage, water cycling, nutrient cycling, and community

composition.

Ozone can produce both acute and chronic injury in sensitive species depending on the 

concentration level and the duration of the exposure.159 In those sensitive species, ozone effects 

also tend to accumulate over the growing season of the plant, so that even low concentrations 

experienced for a longer duration have the potential to create chronic stress on vegetation. 

Ozone damage to sensitive species includes visible injury to leaves and impaired photosynthesis, 

both of which can lead to reduced plant growth and reproduction, resulting in reduced crop 

yields, forestry production, and use of sensitive ornamentals in landscaping. In addition, the 

impairment of photosynthesis, the process by which the plant makes carbohydrates (its source of 

energy and food), can lead to a subsequent reduction in root growth and carbohydrate storage 

below ground, resulting in other, more subtle plant and ecosystems impacts. These latter impacts 

include increased susceptibility of plants to insect attack, disease, harsh weather, interspecies 

competition and overall decreased plant vigor. The adverse effects of ozone on areas with 
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sensitive species could potentially lead to species shifts and loss from the affected ecosystems, 

resulting in a loss or reduction in associated ecosystem goods and services. Lastly, visible ozone 

injury to leaves can result in a loss of aesthetic value in areas of special scenic significance like 

national parks and wilderness areas. 

The Integrated Science Assessment (ISA) for Ozone presents more detailed information 

on how ozone effects vegetation and ecosystems.160 The ISA concludes that ambient 

concentrations of ozone are associated with a number of adverse welfare effects and 

characterizes the weight of evidence for different effects associated with ozone.161 The ISA 

concludes that visible foliar injury effects on vegetation, reduced vegetation growth, reduced 

productivity in terrestrial ecosystems, reduced yield and quality of agricultural crops, and 

alteration of below-ground biogeochemical cycles are causally associated with exposure to 

ozone. It also concludes that reduced carbon sequestration in terrestrial ecosystems, alteration of 

terrestrial ecosystem water cycling, and alteration of terrestrial community composition are 

likely to be causally associated with exposure to ozone. 

6.2.4 Environmental Effects of Air Toxics

Emissions from producing, transporting and combusting fuel contribute to ambient levels 

of pollutants that contribute to adverse effects on vegetation. Volatile organic compounds 

(VOCs), some of which are considered air toxics, have long been suspected to play a role in 

vegetation damage.162 In laboratory experiments, a wide range of tolerance to VOCs has been 

observed.163 Decreases in harvested seed pod weight have been reported for the more sensitive 

plants, and some studies have reported effects on seed germination, flowering and fruit ripening. 

Effects of individual VOCs or their role in conjunction with other stressors (e.g., acidification, 

drought, temperature extremes) have not been well studied. In a recent study of a mixture of 

VOCs including ethanol and toluene on herbaceous plants, significant effects on seed production, 

leaf water content and photosynthetic efficiency were reported for some plant species.164

Research suggests an adverse impact of vehicle exhaust on plants, which has in some 

cases been attributed to aromatic compounds and in other cases to nitrogen oxides.165,166,167 The 

impacts of VOCs on plant reproduction may have long-term implications for biodiversity and 

survival of native species near major roadways. Most of the studies of the impacts of VOCs on 

vegetation have focused on short-term exposure and few studies have focused on long-term 

effects of VOCs on vegetation and the potential for metabolites of these compounds to affect 

herbivores or insects. 
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References

1 U.S. Environmental Protection Agency. (2005). Supplemental guidance for assessing susceptibility

from early-life exposure to carcinogens. Washington, DC: Risk Assessment Forum. EPA/630/

R-03/003F. http://www.epa.gov/raf/publications/pdfs/childrens_supplement_final.pdf.

2 Regulatory definitions of PM size fractions, and information on reference and equivalent methods for measuring 

PM in ambient air, are provided in 40 CFR Parts 50, 53, and 58. With regard to national ambient air quality 

standards (NAAQS) which provide protection against health and welfare effects, the 24-hour PM10 standard 

provides protection against effects associated with short-term exposure to thoracic coarse particles (i.e., PM10-2.5).

3 U.S. EPA (2009). Integrated Science Assessment for Particulate Matter (Final Report). U.S. Environmental 

Protection Agency, Washington, DC, EPA/600/R-08/139F, 2009.

4 U.S. EPA. (2009). Integrated Science Assessment for Particulate Matter (Final Report). U.S. Environmental 

Protection Agency, Washington, DC, EPA/600/R-08/139F. 

5 The causal framework draws upon the assessment and integration of evidence from across epidemiological, 

controlled human exposure, and toxicological studies, and the related uncertainties that ultimately influence our 

understanding of the evidence. This framework employs a five-level hierarchy that classifies the overall weight of 

evidence and causality using the following categorizations: causal relationship, likely to be causal relationship, 

suggestive of a causal relationship, inadequate to infer a causal relationship, and not likely to be a causal relationship 

(U.S. EPA. (2009). Integrated Science Assessment for Particulate Matter (Final Report). U.S. Environmental 

Protection Agency, Washington, DC, EPA/600/R-08/139F, Table 1–3). 

6

78 FR 3086 (January 15, 2013) pages 3103-3104.

7

 77 FR 38890 (June 29, 2012) pages 38906-38911.

8 These causal inferences are based not only on the more expansive epidemiological evidence available in this 

review but also reflect consideration of important progress that has been made to advance our understanding of a 

number of potential biologic modes of action or pathways for PM-related cardiovascular and respiratory effects 

(U.S. EPA. (2009). Integrated Science Assessment for Particulate Matter (Final Report). U.S. Environmental 

Protection Agency, Washington, DC, EPA/600/R-08/139F, chapter 5).

9

78 FR 3086 (January 15, 2013) pages 3103-3104.

10 U.S. EPA. (2009). Integrated Science Assessment for Particulate Matter (Final Report). U.S. Environmental 

Protection Agency, Washington, DC, EPA/600/R-08/139F, chapter 6 (section 6.5) and chapter 7 (Section 7.6).

11 78 FR 3103 (January 15, 2013).

12 78 FR 3103 (January 15, 2013).

13 78 FR 3103-3104 (January 15, 2013).

14 78 FR 3104 (January 15, 2013).

15 U.S. EPA. (2009). Integrated Science Assessment for Particulate Matter (Final Report). U.S. Environmental 

Protection Agency, Washington, DC, EPA/600/R-08/139F. pg 2-13.

16 U.S. EPA. (2009). Integrated Science Assessment for Particulate Matter (Final Report). U.S. Environmental 

Protection Agency, Washington, DC, EPA/600/R-08/139F. Chapter 8 and Chapter 2.

17 77 FR 38890,(June 29, 2012) page 38910.

18 78 FR 3086 (January 15, 2013) page 3104.

19 U.S. EPA. (2011). Policy Assessment for the Review of the PM NAAQS. U.S. Environmental Protection 

Agency, Washington, DC, EPA/452/R-11-003. section 2.2.1.
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22 U.S. EPA. (2009). Integrated Science Assessment for Particulate Matter (Final Report). U.S. Environmental 

Protection Agency, Washington, DC, EPA/600/R-08/139F. Chapter 8 and Chapter 2 (Section 2.4.1).

23 U.S. EPA. (2009). Integrated Science Assessment for Particulate Matter (Final Report). U.S. Environmental 

Protection Agency, Washington, DC, EPA/600/R-08/139F. Section 2.3.4 and Table 2-6.

24 78 FR 3167-8 (January 15, 2013).

25 77 FR 38947-51 (June 29, 2012).

26 U.S. EPA. (2009). Integrated Science Assessment for Particulate Matter (Final Report). U.S. Environmental 

Protection Agency, Washington, DC, EPA/600/R-08/139F. Section 2.3.5 and Table 2-6.

27 78 FR 3121 (January 15, 2013).

28

 Human exposure to ozone varies over time due to changes in ambient ozone concentration and because people 

move between locations which have notable different ozone concentrations. Also, the amount of ozone delivered to 

the lung is not only influenced by the ambient concentrations but also by the individuals breathing route and rate.

29

 U.S. EPA. Integrated Science Assessment of Ozone and Related Photochemical Oxidants (Final Report). U.S. 

Environmental Protection Agency, Washington, DC, EPA/600/R-10/076F, 2013. The ISA is available at 

http://cfpub.epa.gov/ncea/isa/recordisplay.cfm?deid=247492#Download.

30 The ISA evaluates evidence and draws conclusions on the causal relationship between relevant pollutant 

exposures and health effects, assigning one of five “weight of evidence” determinations: causal relationship, likely 

to be a causal relationship, suggestive of a causal relationship, inadequate to infer a causal relationship, and not 

likely to be a causal relationship. For more information on these levels of evidence, please refer to Table II in the 

Preamble of the ISA. 

31 U.S. EPA (2008). Integrated Science Assessment (ISA) for Sulfur Oxides – Health Criteria (Final Report). 

EPA/600/R-08/047F. Washington, DC,: U.S.EPA. Retrieved on March 19, 2009 from 

http://cfpub.epa.gov/ncea/cfm/recordisplay.cfm?deid=198843.

32 U.S. EPA (2008). Integrated Science Assessment for Oxides of Nitrogen – Health Criteria (Final Report).

EPA/600/R-08/071. Washington, DC,: U.S.EPA. Retrieved on March 19, 2009 from 

http://cfpub.epa.gov/ncea/cfm/recordisplay.cfm?deid=194645.

33 U.S. EPA (2008). Integrated Science Assessment for Oxides of Nitrogen – Health Criteria (Final Report).

EPA/600/R-08/071. Washington, DC: U.S.EPA.

34 U.S. EPA, (2010). Integrated Science Assessment for Carbon Monoxide (Final Report). U.S. Environmental 

Protection Agency, Washington, DC, EPA/600/R-09/019F, 2010. 

http://cfpub.epa.gov/ncea/cfm/recordisplay.cfm?deid=218686.

35

 The ISA evaluates the health evidence associated with different health effects, assigning one of five “weight of 

evidence” determinations: causal relationship, likely to be a causal relationship, suggestive of a causal relationship, 

inadequate to infer a causal relationship, and not likely to be a causal relationship. For definitions of these levels of 

evidence, please refer to Section 1.6 of the ISA. 

36

 Personal exposure includes contributions from many sources, and in many different environments. Total 

personal exposure to CO includes both ambient and nonambient components; and both components may contribute 

to adverse health effects.

37 U.S. EPA. (2000). Integrated Risk Information System File for Benzene. This material is available electronically 

at: http://www.epa.gov/iris/subst/0276.htm.

38 International Agency for Research on Cancer. (1982). IARC monographs on the evaluation of carcinogenic risk of 

chemicals to humans, Volume 29, Some industrial chemicals and dyestuffs, International Agency for Research on 

Cancer, World Health Organization, Lyon, France 1982.
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39 Irons, R.D.; Stillman, W.S.; Colagiovanni, D.B.; Henry, V.A. (1992). Synergistic action of the benzene metabolite 

hydroquinone on myelopoietic stimulating activity of granulocyte/macrophage colony-stimulating factor in vitro, 

Proc. Natl. Acad. Sci. 89:3691-3695. 

40 A unit risk estimate is defined as the increase in the lifetime risk of an individual who is exposed for a lifetime to 

1 μg/m3 benzene in air.

41 U.S. EPA. (2000). Integrated Risk Information System File for Benzene. This material is available electronically 

at: http://www.epa.gov/iris/subst/0276.htm.

42 International Agency for Research on Cancer (IARC). 1987. Monographs on the evaluation of carcinogenic risk 

of chemicals to humans, Volume 29, Supplement 7, Some industrial chemicals and dyestuffs, World Health 

Organization, Lyon, France.

43 U.S. Department of Health and Human Services National Toxicology Program. (2011). 12th Report on 

Carcinogens available at: http://ntp.niehs.nih.gov/?objectid=03C9AF75-E1BF-FF40-DBA9EC0928DF8B15. 

44 Aksoy, M. (1989). Hematotoxicity and carcinogenicity of benzene. Environ. Health Perspect. 82: 193-197. 

EPA-HQ-OAR-2011-0135.

45 Goldstein, B.D. (1988). Benzene toxicity. Occupational medicine. State of the Art Reviews. 3: 541-554.

46 Rothman, N., G.L. Li, M. Dosemeci, W.E. Bechtold, G.E. Marti, Y.Z. Wang, M. Linet, L.Q. Xi, W. Lu, M.T. 

Smith, N. Titenko-Holland, L.P. Zhang, W. Blot, S.N. Yin, and R.B. Hayes. (1996). Hematotoxicity among Chinese 

workers heavily exposed to benzene. Am. J. Ind. Med. 29: 236-246.

47 U.S. EPA (2002). Toxicological Review of Benzene (Noncancer Effects). Environmental Protection Agency, 

Integrated Risk Information System (IRIS), Research and Development, National Center for Environmental 

Assessment, Washington DC. This material is available electronically at http://www.epa.gov/iris/subst/0276.htm.

48 Qu, O.; Shore, R.; Li, G.; Jin, X.; Chen, C.L.; Cohen, B.; Melikian, A.; Eastmond, D.; Rappaport, S.; Li, H.; Rupa, 

D.; Suramaya, R.; Songnian, W.; Huifant, Y.; Meng, M.; Winnik, M.; Kwok, E.; Li, Y.; Mu, R.; Xu, B.; Zhang, 

X.; Li, K. (2003). HEI Report 115, Validation & Evaluation of Biomarkers in Workers Exposed to Benzene in 

China.

49 Qu, Q., R. Shore, G. Li, X. Jin, L.C. Chen, B. Cohen, et al. (2002). Hematological changes among Chinese 

workers with a broad range of benzene exposures. Am. J. Industr. Med. 42: 275-285.

50 Lan, Qing, Zhang, L., Li, G., Vermeulen, R., et al. (2004). Hematotoxically in Workers Exposed to Low Levels of 

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51 Turtletaub, K.W. and Mani, C. (2003). Benzene metabolism in rodents at doses relevant to human exposure from 

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Atlanta, GA: U.S. Department of Health and Human Services, Public Health Service. 

http://www.atsdr.cdc.gov/ToxProfiles/tp3.pdf.

53 A minimal risk level (MRL) is defined as an estimate of the daily human exposure to a hazardous substance that is 

likely to be without appreciable risk of adverse noncancer health effects over a specified duration of exposure.

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for Environmental Assessment, Washington Office, Washington, DC. Report No. EPA600-P-98-001F. This 

document is available electronically at http://www.epa.gov/iris/supdocs/buta-sup.pdf.

55 U.S. EPA. (2002) “Full IRIS Summary for 1,3-butadiene (CASRN 106-99-0)” Environmental Protection Agency, 

Integrated Risk Information System (IRIS), Research and Development, National Center for Environmental 

Assessment, Washington, DC http://www.epa.gov/iris/subst/0139.htm.
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56 International Agency for Research on Cancer (IARC). (1999). Monographs on the evaluation of carcinogenic risk 

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58 NTP. (2011). Report on Carcinogens, Twelfth Edition. Research Triangle Park, NC: U.S. Department of Health 

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– Inhalation Assessment: In Support of Summary Information on the Integrated Risk Information System (IRIS). 
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External Review Draft. EPA/635/R-10/002A. U.S. Environmental Protection Agency, Washington DC [online]. 

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75 NRC (National Research Council). 2011. Review of the Environmental Protection Agency’s Draft IRIS 

Assessment of Formaldehyde. Washington DC: National Academies Press. 

http://books.nap.edu/openbook.php?record_id=13142.

76 U.S. EPA (1991). Integrated Risk Information System File of Acetaldehyde. Research and Development, 

National Center for Environmental Assessment, Washington, DC. This material is available electronically at 

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77 U.S. EPA (1991). Integrated Risk Information System File of Acetaldehyde. This material is available 

electronically at http://www.epa.gov/iris/subst/0290.htm.

78 NTP. (2011). Report on Carcinogens, Twelfth Edition. Research Triangle Park, NC: U.S. Department of Health 

and Human Services, Public Health Service, National Toxicology Program. 499 pp.

79 International Agency for Research on Cancer (IARC). (1999). Re-evaluation of some organic chemicals, 

hydrazine, and hydrogen peroxide. IARC Monographs on the Evaluation of Carcinogenic Risk of Chemical to 

Humans, Vol 71. Lyon, France.

80 U.S. EPA (1991). Integrated Risk Information System File of Acetaldehyde. This material is available 

electronically at http://www.epa.gov/iris/subst/0290.htm.

81 U.S. EPA. (2003). Integrated Risk Information System File of Acrolein. Research and Development, National 

Center for Environmental Assessment, Washington, DC. This material is available electronically at 

http://www.epa.gov/iris/subst/0364.htm.

82 Appleman, L.M., R.A. Woutersen, and V.J. Feron. (1982). Inhalation toxicity of acetaldehyde in rats. I. Acute and 

subacute studies. Toxicology. 23: 293-297.

83 Myou, S.; Fujimura, M.; Nishi K.; Ohka, T.; and Matsuda, T. (1993). Aerosolized acetaldehyde induces 

histamine-mediated bronchoconstriction in asthmatics. Am. Rev. Respir.Dis.148(4 Pt 1): 940-943.

84 U.S. EPA. (2003). Integrated Risk Information System File of Acrolein. Research and Development, National 

Center for Environmental Assessment, Washington, DC. This material is available at 

http://www.epa.gov/iris/subst/0364.htm.

85 International Agency for Research on Cancer (IARC). (1995). Monographs on the evaluation of carcinogenic risk 

of chemicals to humans, Volume 63. Dry cleaning, some chlorinated solvents and other industrial chemicals, World 

Health Organization, Lyon, France.

86 U.S. EPA. (2003). Integrated Risk Information System File of Acrolein. Office of Research and Development, 

National Center for Environmental Assessment, Washington, DC. This material is available at 

http://www.epa.gov/iris/subst/0364.htm.

87 U.S. EPA. (2003). Integrated Risk Information System File of Acrolein. Office of Research and Development, 

National Center for Environmental Assessment, Washington, DC. This material is available at 

http://www.epa.gov/iris/subst/0364.htm.

88 U.S. EPA. (2003). Toxicological review of acrolein in support of summary information on Integrated Risk 

Information System (IRIS) National Center for Environmental Assessment, Washington, DC. EPA/635/R-03/003. p. 

10. Available online at: http://www.epa.gov/ncea/iris/toxreviews/0364tr.pdf.

89 U.S. EPA. (2003). Toxicological review of acrolein in support of summary information on Integrated Risk 

Information System (IRIS) National Center for Environmental Assessment, Washington, DC. EPA/635/R-03/003. 

Available online at: http://www.epa.gov/ncea/iris/toxreviews/0364tr.pdf.

90 Morris JB, Symanowicz PT, Olsen JE, et al. (2003). Immediate sensory nerve-mediated respiratory responses to 

irritants in healthy and allergic airway-diseased mice. J Appl Physiol 94(4):1563-1571.
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91 U.S. EPA. (2009). Graphical Arrays of Chemical-Specific Health Effect Reference Values for Inhalation 

Exposures (Final Report). U.S. Environmental Protection Agency, Washington, DC, EPA/600/R-09/061, 2009. 

http://cfpub.epa.gov/ncea/cfm/recordisplay.cfm?deid=211003.

92 U.S. EPA. (2006). Air quality criteria for ozone and related photochemical oxidants (Ozone CD). Research 

Triangle Park, NC: National Center for Environmental Assessment; report no. EPA/600/R-05/004aF-cF.3v. page 5-

78 Available at http://cfpub.epa.gov/ncea/. 

93 U.S. EPA Air Quality Criteria for Ozone and Related Photochemical Oxidants (Final). U.S. Environmental 

Protection Agency, Washington, D.C., EPA 600/R-05/004aF-cF, 2006. page 5-63. This document is available in . 

This document may be accessed electronically at: http://www.epa.gov/ttn/naaqs/standards/ozone/s_o3_cr_cd.html.

94 U.S. EPA Air Quality Criteria for Ozone and Related Photochemical Oxidants (Final). U.S. Environmental 

Protection Agency, Washington, D.C., EPA 600/R-05/004aF-cF, 2006. page 5-78. This document is available in . 

This document may be accessed electronically at: http://www.epa.gov/ttn/naaqs/standards/ozone/s_o3_cr_cd.html.

95 Agency for Toxic Substances and Disease Registry (ATSDR). (1995). Toxicological profile for Polycyclic 

Aromatic Hydrocarbons (PAHs). Atlanta, GA: U.S. Department of Health and Human Services, Public Health 

Service. Available electronically at http://www.atsdr.cdc.gov/ToxProfiles/TP.asp?id=122&tid=25.

96 U.S. EPA (2002). Health Assessment Document for Diesel Engine Exhaust. EPA/600/8-90/057F Office of 

Research and Development, Washington DC. http://cfpub.epa.gov/ncea/cfm/recordisplay.cfm?deid=29060. 

97 International Agency for Research on Cancer (IARC). (2012). Monographs on the Evaluation of the 

Carcinogenic Risk of Chemicals for Humans, Chemical Agents and Related Occupations. Vol. 100F. Lyon, France.

98 U.S. EPA (1997). Integrated Risk Information System File of indeno(1,2,3-cd)pyrene. Research and Development, 

National Center for Environmental Assessment, Washington, DC. This material is available electronically at 

http://www.epa.gov/ncea/iris/subst/0457.htm.

99 Perera, F.P.; Rauh, V.; Tsai, W-Y.; et al. (2002). Effect of transplacental exposure to environmental pollutants on 

birth outcomes in a multiethnic population. Environ Health Perspect. 111: 201-205.

100 Perera, F.P.; Rauh, V.; Whyatt, R.M.; Tsai, W.Y.; Tang, D.; Diaz, D.; Hoepner, L.; Barr, D.; Tu, Y.H.; Camann, 

D.; Kinney, P. (2006). Effect of prenatal exposure to airborne polycyclic aromatic hydrocarbons on 

neurodevelopment in the first 3 years of life among inner-city children. Environ Health Perspect 114: 1287-1292.

101 U. S. EPA. 1998. Toxicological Review of Naphthalene (Reassessment of the Inhalation Cancer Risk), 

Environmental Protection Agency, Integrated Risk Information System, Research and Development, National 

Center for Environmental Assessment, Washington, DC. This material is available electronically at 

http://www.epa.gov/iris/subst/0436.htm.

102 U. S. EPA. 1998. Toxicological Review of Naphthalene (Reassessment of the Inhalation Cancer Risk), 

Environmental Protection Agency, Integrated Risk Information System, Research and Development, National 

Center for Environmental Assessment, Washington, DC. This material is available electronically at 

http://www.epa.gov/iris/subst/0436.htm.

103 U. S. EPA. (1998). Toxicological Review of Naphthalene (Reassessment of the Inhalation Cancer Risk), 

Environmental Protection Agency, Integrated Risk Information System, Research and Development, National 

Center for Environmental Assessment, Washington, DC. This material is available electronically at 

http://www.epa.gov/iris/subst/0436.htm. 

104 Oak Ridge Institute for Science and Education. (2004). External Peer Review for the IRIS Reassessment of the 

Inhalation Carcinogenicity of Naphthalene. August 2004. 

http://cfpub.epa.gov/ncea/cfm/recordisplay.cfm?deid=84403.

105 NTP. (2011). Report on Carcinogens, Twelfth Edition. Research Triangle Park, NC: U.S. Department of Health 

and Human Services, Public Health Service, National Toxicology Program. 499 pp.
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106 International Agency for Research on Cancer (IARC). (2002). Monographs on the Evaluation of the 

Carcinogenic Risk of Chemicals for Humans. Vol. 82. Lyon, France.

107 U. S. EPA. (1998). Toxicological Review of Naphthalene, Environmental Protection Agency, Integrated Risk 

Information System, Research and Development, National Center for Environmental Assessment, Washington, DC. 

This material is available electronically at http://www.epa.gov/iris/subst/0436.htm.

108 U.S. EPA. (1998). Toxicological Review of Naphthalene. Environmental Protection Agency, Integrated Risk 

Information System (IRIS), Research and Development, National Center for Environmental Assessment, 

Washington, DC http://www.epa.gov/iris/subst/0436.htm.

109 U.S. EPA Integrated Risk Information System (IRIS) database is available at: www.epa.gov/iris.

110 It should be noted that there are no peer reviewed EPA-authored reviews of traffic-related health studies. The 

HEI panel primarily used epidemiology studies for inferring whether there was sufficient evidence of a causal 

association exists between a particular health effect and traffic-related air pollution, In its weight-of-evidence 

determinations, the panel also placed “considerable weight” on controlled human exposure studies. However, it 

restricted consideration of other toxicological studies to whether or not the studies provided “general mechanistic 

support” for the inferences of causality made on the basis of epidemiology.

111 Health Effects Institute Panel on the Health Effects of Traffic-Related Air Pollution. (2010). Traffic-related air 

pollution: a critical review of the literature on emissions, exposure, and health effects. HEI Special Report 17. 

[Online at http://www.healtheffects.org].

112 Bureau of Labor Statistics. (2009). American Time Use Survey. [Online at http://www.bls.gov/tus].

113 Bureau of Transportation Statistics. (2003). Highlights of the 2001 National Household Travel Survey. Report 

BTS03-05. [Online at http://www.bts.gov].

114 The autonomic nervous system (ANS) consists of sympathetic and parasympathetic components. The 

sympathetic ANS signals body systems to “fight or flight.” The parasympathetic ANS signals the body to “rest and 

digest.” In general, HRV is indicative of parasympathetic control of the heart.

115 Baja, E.S.; Schwartz, J.D.; Wellenius, G.A.; Coull, B.A.; Zanobetti, A.; Vokonas, P.S.; Suh, H.H. (2010). 

Traffic-related air pollution and QT interval: modification by diabetes, obesity, and oxidate stress gene 

polymorphisms in the Normative Aging Study. Environ Health Perspect 118: 840-846. doi:10.1289/ehp.0901396.

116 Zanobettia, A.; Stone, P.H.; Speizer, F.E.; Schwarz, J.D.; Coull, B.A.; Suh, H.H.; Nearing, B.D.; Mittleman, 

M.A.; Verrier, R.L.; Gold, D.R. (2009). T-wave alterans, air pollution and traffic in high-risk subjects. Am J 

Cardiol 104: 665-670. doi:10.1016/j.amjcard.2009.04.046.

117 Brook, R.D.; Rajagopalan, S.; Pope, C.A.; Brook, J.R.; Bhatnagar, A.; Diez-Rouz, A.V.; Holguin, F.; Hong, Y.; 

Luepker, R.V.; et la. (2010). Particulate matter air pollution and cardiovascular disease: an update to the scientific 

statement from the American Heart Association. Circulation 121: 2331-2378. 

doi:10.1161/CIR.0b013e3181dbece1.

118 Bastain, T.M.; Gilliland, F.D.; Li, Y.; Saxon, A.; Diaz-Sanchez, D. (2003) Intraindividual reproducibility of nazal 

allergic responses to diesel exhaust particles indicates a susceptible phenotype. Clinical Immunol 109: 130-136.

119 Gilliland, F.D.; Li, Y.; Diaz-Sanchez, D. (2004) Effect of glutathione-S-transferase M1 and P1 genotypes on 

xenobiotic enhancement of allergic responses: randomized, placebo-controlled crossover study. Lancet 363:119-

125.

120 Svartengren, M., Strand, V.; Bylin, G. Järup, L.; Pershagen, G. (2000) Short-term exposure to air pollution in a 

road tunnel enhances the asthmatic response to allergen. Eur Respir J 15: 716-724.

121 Vrijheid, M.; Martinez, D.; Manzanares, S.; Dadvand, P.; Schembari, A.; Rankin, F.; Nieuwenhuijsen, M. (2011). 

Ambient air pollution and risk of congenital anomalies: a systematic review and meta-analysis. Environ Health 

Perspect 119: 598-606. doi:10.1289/ehp.1002946.
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122 Raaschou-Nielson, O.; Reynolds, P. (2006). Air pollution and childhood cancer: a review of the epidemiological 

literature. Int J Cancer 118: 2920-2929. doi: 10.1002/ijc.21787 [Online at http://dx.doi.org].

123 National Research Council, (1993). Protecting Visibility in National Parks and Wilderness Areas. National 

Academy of Sciences Committee on Haze in National Parks and Wilderness Areas. National Academy Press, 

Washington, DC. This book can be viewed on the National Academy Press Website at 

http://www.nap.edu/books/0309048443/html/.

124 U.S. EPA (2009). Integrated Science Assessment for Particulate Matter (Final Report). U.S. Environmental 

Protection Agency, Washington, DC, EPA/600/R-08/139F, 2009.

125 U.S. EPA (2009). Integrated Science Assessment for Particulate Matter (Final Report). U.S. Environmental 

Protection Agency, Washington, DC, EPA/600/R-08/139F, 2009. pg 9-19 through 9-23. 

126 U.S. EPA (2009). Integrated Science Assessment for Particulate Matter (Final Report). U.S. Environmental 

Protection Agency, Washington, DC, EPA/600/R-08/139F, 2009. 

127 U.S. EPA. (2005). Review of the National Ambient Air Quality Standard for Particulate Matter: Policy 

Assessment of Scientific and Technical Information, OAQPS Staff Paper. EPA-452/R-05-005. 

128 U.S. EPA. (2008). Integrated Science Assessment for Oxides of Nitrogen and Sulfur- Ecological Criteria (Final). 

U.S. EPA, Washington D.C., EPA/600/R-08/082F. 

129 U.S. EPA. (2008). Integrated Science Assessment for Oxides of Nitrogen and Sulfur- Ecological Criteria (Final). 

U.S. EPA, Washington D.C., EPA/600/R-08/082F. 

130 Environmental Protection Agency (2003). Response Of Surface Water Chemistry to the Clean Air Act 

Amendments of 1990. National Health and Environmental Effects Research Laboratory, Office of Research and 

Development, U.S. Environmental Protection Agency. Research Triangle Park, NC. EPA 620/R–03/001. 

131 Fenn, M.E. and Blubaugh, T.J. (2005). Winter Deposition of Nitrogen and Sulfur in the Eastern Columbia River 

Gorge National Scenic Area, USDA Forest Service. 

132 Galloway, J. N.; Cowling, E. B. (2002). Reactive nitrogen and the world: 200 years of change. Ambio 31: 64–71. 

133 Bricker, Suzanne B., et al., (1999). National Estuarine Eutrophication Assessment, Effects of Nutrient 

Enrichment in the Nation’s Estuaries, National Ocean Service, National Oceanic and Atmospheric Administration, 

September, 1999. 

134 Smith, W.H. (1991). “Air pollution and Forest Damage.” Chemical Engineering News, 69(45): 30-43. 

135 Gawel, J.E.; Ahner, B.A.; Friedland, A.J.; and Morel, F.M.M. (1996). “Role for heavy metals in forest decline 

indicated by phytochelatin measurements.” Nature, 381: 64-65. 

136 Cotrufo, M.F.; DeSanto, A.V.; Alfani, A.; et al. (1995). “Effects of urban heavy metal pollution on organic matter 

decomposition in Quercus ilix L. woods.” Environmental Pollution, 89: 81-87. 

137 Niklinska, M.; Laskowski, R.; Maryanski, M. (1998). “Effect of heavy metals and storage time on two types of 

forest litter: basal respiration rate and exchangeable metals.” Ecotoxicological Environmental Safety, 41: 8-18. 

138 U.S. EPA (2009). Integrated Science Assessment for Particulate Matter (Final Report). U.S. Environmental 

Protection Agency, Washington, DC, EPA/600/R-08/139F, 2009. Section 9.4.5.2. 

139 Mason, R.P. and Sullivan, K.A. (1997). “Mercury in Lake Michigan.” Environmental Science & Technology, 31: 

942-947. (from Delta Report “Atmospheric deposition of toxics to the Great Lakes”). 

140 Landis, M.S. and Keeler, G.J. (2002). “Atmospheric mercury deposition to Lake Michigan during the Lake 

Michigan Mass Balance Study.” Environmental Science & Technology, 21: 4518-24. 

141 U.S. EPA. (2000). EPA453/R-00-005, “Deposition of Air Pollutants to the Great Waters: Third Report to 

Congress,” Office of Air Quality Planning and Standards, Research Triangle Park, North Carolina.
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142 National Science and Technology Council (NSTC) (1999). “The Role of Monitoring Networks in the 

Management of the Nation’s Air Quality.” 

143 Callender, E. and Rice, K.C. (2000). “The Urban Environmental Gradient: Anthropogenic Influences on the 

Spatial and Temporal Distributions of Lead and Zinc in Sediments.” Environmental Science & Technology, 34: 232-

238. 

144 Rice, K.C. (1999). “Trace Element Concentrations in Streambed Sediment Across the Conterminous United 

States.” Environmental Science & Technology, 33: 2499-2504. 

145 U.S. EPA. (2013) Integrated Science Assessment for Lead (Final Report). U.S. Environmental Protection 

Agency, Washington, DC, EPA/600/R-10/075F, 2013.

146 Ely, JC; Neal, CR; Kulpa, CF; et al. (2001). “Implications of Platinum-Group Element Accumulation along U.S. 

Roads from Catalytic-Converter Attrition.” Environ. Sci. Technol. 35: 3816-3822. 

147 U.S. EPA. (1998). EPA454/R-98-014, “Locating and Estimating Air Emissions from Sources of Polycyclic 

Organic Matter,” Office of Air Quality Planning and Standards, Research Triangle Park, North Carolina. 

148 U.S. EPA. (1998). EPA454/R-98-014, “Locating and Estimating Air Emissions from Sources of Polycyclic 

Organic Matter,” Office of Air Quality Planning and Standards, Research Triangle Park, North Carolina. 

149 Simcik, M.F.; Eisenreich, S.J.; Golden, K.A.; et al. (1996). “Atmospheric Loading of Polycyclic Aromatic 

Hydrocarbons to Lake Michigan as Recorded in the Sediments.” Environmental Science and Technology, 30: 3039-

3046. 

150 Simcik, M.F.; Eisenreich, S.J.; and Lioy, P.J. (1999). “Source apportionment and source/sink relationship of 

PAHs in the coastal atmosphere of Chicago and Lake Michigan.” Atmospheric Environment, 33: 5071-5079. 

151 Arzayus, K.M.; Dickhut, R.M.; and Canuel, E.A. (2001). “Fate of Atmospherically Deposited Polycyclic 

Aromatic Hydrocarbons (PAHs) in Chesapeake Bay.” Environmental Science & Technology, 35, 2178-2183. 

152 Park, J.S.; Wade, T.L.; and Sweet, S. (2001). “Atmospheric distribution of polycyclic aromatic hydrocarbons and 

deposition to Galveston Bay, Texas, USA.” Atmospheric Environment, 35: 3241-3249. 

153 Poor, N.; Tremblay, R.; Kay, H.; et al. (2002). “Atmospheric concentrations and dry deposition rates of 

polycyclic aromatic hydrocarbons (PAHs) for Tampa Bay, Florida, USA.” Atmospheric Environment 38: 6005-

6015. 

154 Arzayus, K.M.; Dickhut, R.M.; and Canuel, E.A. (2001). “Fate of Atmospherically Deposited Polycyclic 

Aromatic Hydrocarbons (PAHs) in Chesapeake Bay.” Environmental Science & Technology, 35, 2178-2183. 

155 U.S. EPA. (2000). EPA453/R-00-005, “Deposition of Air Pollutants to the Great Waters: Third Report to 

Congress,” Office of Air Quality Planning and Standards, Research Triangle Park, North Carolina. 

156 Van Metre, P.C.; Mahler, B.J.; and Furlong, E.T. (2000). “Urban Sprawl Leaves its PAH Signature.” 

Environmental Science & Technology, 34: 4064-4070. 

157 Cousins, I.T.; Beck, A.J.; and Jones, K.C. (1999). “A review of the processes involved in the exchange of semivolatile organic compounds across the air-soil interface.” The Science of the Total Environment, 228: 5-24. 

158 Tuhackova, J. et al. (2001). Hydrocarbon deposition and soil microflora as affected by highway traffic. 

Environmental Pollution, 113: 255-262. 

159 Only a small percentage of all the plant species growing within the U.S. (over 43,000 species have been 

catalogued in the USDA PLANTS database) have been studied with respect to ozone sensitivity (U.S. EPA. Review 

of the National Ambient Air Quality Standards for Ozone: Policy Assessment of Scientific and Technical 

Information. OAQPS Staff Paper. U.S. Environmental Protection Agency, Washington, DC, EPA-452/R-07-007, 

2007. p.7-7). Thus, ozone impacts could be occurring in areas where plant species sensitive to ozone have not yet 

been studied or identified. In addition, the concentration at which ozone levels overwhelm a plant’s ability to 
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detoxify or compensate for oxidant exposure varies (U.S. EPA. Welfare Risk and Exposure Assessment for Ozone. 

Second External Review Draft. EPA-452/P-14-003a, 2014. Figure 6-2, p. 6-5). Thus, whether a plant is classified as 

sensitive or tolerant depends in part on the exposure levels being considered.

160 U.S. EPA. Integrated Science Assessment of Ozone and Related Photochemical Oxidants (Final Report). U.S. 

Environmental Protection Agency, Washington, DC, EPA/600/R-10/076F, 2013. The ISA is available at 

http://cfpub.epa.gov/ncea/isa/recordisplay.cfm?deid=247492#Download.

161 The Ozone ISA evaluates the evidence associated with different ozone related health and welfare effects, 

assigning one of five “weight of evidence” determinations: causal relationship, likely to be a causal relationship, 

suggestive of a causal relationship, inadequate to infer a causal relationship, and not likely to be a causal 

relationship. For more information on these levels of evidence, please refer to Table II of the ISA. 

162 U.S. EPA. (1991). Effects of organic chemicals in the atmosphere on terrestrial plants. EPA/600/3-91/001.

163 Cape JN, ID Leith, J Binnie, J Content, M Donkin, M Skewes, DN Price AR Brown, AD Sharpe. (2003). Effects 

of VOCs on herbaceous plants in an open-top chamber experiment. Environ. Pollut. 124:341-343.

164 Cape JN, ID Leith, J Binnie, J Content, M Donkin, M Skewes, DN Price AR Brown, AD Sharpe. (2003). Effects 

of VOCs on herbaceous plants in an open-top chamber experiment. Environ. Pollut. 124:341-343.

165 Viskari E-L. (2000). Epicuticular wax of Norway spruce needles as indicator of traffic pollutant deposition. 

Water, Air, and Soil Pollut. 121:327-337.

166 Ugrekhelidze D, F Korte, G Kvesitadze. (1997). Uptake and transformation of benzene and toluene by plant 

leaves. Ecotox. Environ. Safety 37:24-29.

167 Kammerbauer H, H Selinger, R Rommelt, A Ziegler-Jons, D Knoppik, B Hock. (1987). Toxic components of 

motor vehicle emissions for the spruce Picea abies. Environ. Pollut. 48:235-243.