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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

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 um 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 

and referring to total particle mass under 2.5 and between 2.5 and 10 micrometers, respectively. 

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

A

Particles span many sizes and shapes and 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 pollution also varies by time of year and location and 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 

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.1 

 

A 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.
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6.1.1.2 Health Effects of PM

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

ambient concentrations of PM.B The information in this section is based on the information and 

conclusions in the Integrated Science Assessment (ISA) for Particulate Matter (December 2009) 

prepared by EPA’s Office of Research and Development (ORD).

C

The ISA concludes that ambient concentrations of PM are associated with a number of 

adverse health effects.DThe ISA characterizes the weight of evidence for different health effects 

associated with three PM size ranges: PM2.5, PM10-2.5, and UFPs. The discussion below 

highlights the ISA’s conclusions pertaining to these three size fractions of PM, considering 

variations in health effects associated with both short-term and long-term exposure periods.

6.1.1.2.1 Effects Associated with Short-term Exposure to PM2.5

The ISA concludes that cardiovascular effects and premature mortality are causally 

associated with short-term exposure to PM2.5.

2

 It also concludes that respiratory effects are 

likely to be causally associated with short-term exposure to PM2.5, including respiratory

emergency department (ED) visits and hospital admissions for chronic obstructive pulmonary

disease (COPD), respiratory infections, and asthma; and exacerbation of respiratory symptoms in 

asthmatic children.

6.1.1.2.2 Effects Associated with Long-term Exposure to PM2.5

The ISA concludes that there are causal associations between long-term exposure to 

PM2.5 and cardiovascular effects, such as the development/progression of cardiovascular disease 

(CVD), and premature mortality, particularly from cardiovascular causes.3 It also concludes that 

long-term exposure to PM2.5 is likely to be causally associated with respiratory effects, such as 

reduced lung function growth, increased respiratory symptoms, and asthma development. The 

ISA characterizes the evidence as suggestive of a causal relationship for associations between 

long-term PM2.5 exposure and reproductive and developmental outcomes, such as low birth 

weight and infant mortality. It also characterizes the evidence as suggestive of a causal 

relationship between PM2.5 and cancer incidence, mutagenicity, and genotoxicity.

 

B

 Personal exposure includes contributions from many different types of particles, from many sources, and in many 

different environments. Total personal exposure to PM includes both ambient and nonambient components and 

collectively these components may contribute to adverse health effects.

C

 The ISA is available at http://cfpub.epa.gov/ncea/cfm/recordisplay.cfm?deid=216546

D 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 1.5 of the ISA. 
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6.1.1.2.3 Effects Associated with PM10-2.5

The ISA summarizes evidence related to short-term exposure to PM10-2.5. PM10-2.5 is the 

fraction of PM10 particles that is larger than PM2.5.

4

 The ISA concludes that available evidence 

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

cardiovascular effects. It also concludes that the available evidence is suggestive of a causal 

relationship between short-term exposures to PM10-2.5 and respiratory effects, including 

respiratory-related ED visits and hospitalizations. The ISA also concludes that the available 

literature suggests a causal relationship between short-term exposures to PM10-2.5 and mortality. 

Data are inadequate to draw conclusions regarding health effects associated with long-term 

exposure to PM10-2.5.

5

6.1.1.2.4 Effects Associated with Ultrafine Particles

The ISA concludes that the evidence is suggestive of a causal relationship between shortterm exposures to UFPs and cardiovascular effects, including changes in heart rhythm and 

vasomotor function (the ability of blood vessels to expand and contract).6 

The ISA also concludes that there is suggestive evidence of a causal relationship between 

short-term UFP exposure and respiratory effects. The types of respiratory effects examined in 

epidemiologic studies include respiratory symptoms and asthma hospital admissions, the results 

of which are not entirely consistent. There is evidence from toxicological and controlled human 

exposure studies that exposure to UFPs may increase lung inflammation and produce small 

asymptomatic changes in lung function. Data are inadequate to draw conclusions regarding 

health effects associated with long-term exposure to UFPs.7

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 
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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

Exposure to ambient ozone contributes to a wide range of adverse health effects.E These 

health effects are well documented and are critically assessed in the EPA ozone air quality

criteria document (ozone AQCD) and EPA staff paper.8,9 We are relying on the data and 

conclusions in the ozone AQCD and staff paper, regarding the health effects associated with 

ozone exposure.

Ozone-related health effects include lung function decrements, respiratory symptoms, 

aggravation of asthma, increased hospital and emergency room visits, increased asthma 

medication usage, and a variety of other respiratory effects. Cellular-level effects, such as 

inflammation of lungs, have been documented as well. In addition, there is suggestive evidence 

of a contribution of ozone to cardiovascular-related morbidity and highly suggestive evidence 

that short-term ozone exposure directly or indirectly contributes to non-accidental and 

cardiopulmonary-related mortality, but additional research is needed to clarify the underlying 

mechanisms causing these effects. In a recent report on the estimation of ozone-related 

premature mortality published by the National Research Council (NRC), a panel of experts and 

reviewers concluded that short-term exposure to ambient ozone is likely to contribute to 

premature deaths and that ozone-related mortality should be included in estimates of the health 

benefits of reducing ozone exposure.10 People who appear to be more susceptible to effects 

associated with exposure to ozone include children, asthmatics and the elderly. Those with 

greater exposures to ozone, for instance due to time spent outdoors (e.g., children and outdoor 

workers), are also of concern.

 

E

 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.
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Based on a large number of scientific studies, EPA has identified several key health 

effects associated with exposure to levels of ozone found today in many areas of the country. 

Short-term (1 to 3 hours) and prolonged exposures (6 to 8 hours) to ambient ozone 

concentrations have been linked to lung function decrements, respiratory symptoms, increased 

hospital admissions and emergency room visits for respiratory problems.11, 12, 13, 14, 15, 16 

Repeated exposure to ozone can increase susceptibility to respiratory infection and lung 

inflammation and can aggravate preexisting respiratory diseases, such as asthma.17, 18, 19, 20, 21

Repeated exposure to sufficient concentrations of ozone can also cause inflammation of the lung, 

impairment of lung defense mechanisms, and possibly irreversible changes in lung structure, 

which over time could affect premature aging of the lungs and/or the development of chronic 

respiratory illnesses, such as emphysema and chronic bronchitis.22, 23, 24, 25

Children and outdoor workers tend to have higher ozone exposure because they typically 

are active outside, working, playing and exercising, during times of day and seasons (e.g., the 

summer) when ozone levels are highest.26 For example, summer camp studies have reported 

statistically significant reductions in lung function in children who are active outdoors.27, 28, 29, 30, 

31, 32, 33, 34

 Further, children are more at risk of experiencing health effects from ozone exposure 

than adults because their respiratory systems are still developing. These individuals (as well as 

people with respiratory illnesses, such as asthma, especially asthmatic children) can experience 

reduced lung function and increased respiratory symptoms, such as chest pain and cough, when 

exposed to relatively low ozone levels during prolonged periods of moderate exertion.35, 36, 37, 38

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 

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 EPA Integrated Science Assessment 

for Sulfur Oxides.39 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 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 
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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 concentrationresponse 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 

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

Information on the health effects of NO2 can be found in the EPA Integrated Science 

Assessment (ISA) for Nitrogen Oxides.40 The EPA has 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 ISA concludes that the strongest evidence for such a relationship comes from 

epidemiologic studies of respiratory effects including symptoms, emergency department visits, 

and hospital admissions. Based on both short- and long-term studies, the ISA concludes that

associations of NO2 with respiratory health effects are stronger among a number of groups; these 

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

older adults. The ISA also draws two broad conclusions regarding airway responsiveness 

following NO2 exposure. First, the ISA concludes that NO2 exposure may enhance the 
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sensitivity to allergen-induced decrements in lung function and increase the allergen-induced 

airway inflammatory response following 30-minute exposures of asthmatics to NO2

concentrations as low as 0.26 ppm. 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 asthmatic subjects. Small but significant increases in non-specific airway

hyperresponsiveness were reported following 1-hour exposures of asthmatics to 0.1 ppm 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. 

Although the weight of evidence supporting a causal relationship is somewhat less certain 

than that associated with respiratory morbidity, NO2 has also been linked to other health 

endpoints. These include all-cause (nonaccidental) mortality, hospital admissions or emergency

department visits for cardiovascular disease, and decrements in lung function growth associated 

with chronic exposure.

6.1.4 Health Effects of Carbon Monoxide

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

Integrated Science Assessment (ISA) for Carbon Monoxide.

41

 The ISA concludes that ambient 

concentrations of CO are associated with a number of adverse health effects.FThis section 

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

CO.

G

 

Human clinical 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 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. 

 

F

 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. 

G

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

human exposure studies report inconsistent neural and behavioral effects following low-level CO 

exposures. The 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 epidemiologic and animal toxicological studies cited in the ISA have 

evaluated associations between CO exposure and 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 associations between perinatal CO exposure and decrements in 

birth weight, as well as other developmental outcomes. The ISA concludes these studies are 

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

effects and birth outcomes.

Epidemiologic studies provide evidence of effects on respiratory morbidity such as 

changes in pulmonary function, respiratory symptoms, and hospital admissions associated with 

ambient CO concentrations. 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 pulmonary vascular 

remodeling and oxidative injury. The 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 ISA concludes that the epidemiologic evidence is suggestive of a causal 

relationship between short-term exposures to 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 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.42,43,44 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 
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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.H,45 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.46,47 

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.48,49 The most sensitive noncancer effect observed in humans, based on current data, is 

the depression of the absolute lymphocyte count in blood.50,51 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.52,53,54,55 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.56,I

6.1.5.2 1,3-Butadiene

EPA has characterized 1,3-butadiene as carcinogenic to humans by inhalation.57,58 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.59,60,61 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.

62 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.63 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 

 

H 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.

I 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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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.75 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.76,77,78

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.79,80,81 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.82 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.83 Finally, a study of embalmers 

reported formaldehyde exposures to be associated with an increased risk of myeloid leukemia 

but not brain cancer.84

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

Toxics Substances and Disease Registry in 199985 and supplemented in 2010,

86 and by the World 

Health Organization.87 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 2010.88 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 

201189 (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.90 The URE in IRIS for acetaldehyde is 2.2 × 10-6per μg/m3.

91

 Acetaldehyde is 

reasonably anticipated to be a human carcinogen by the U.S. DHHS in the 12th Report on 
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Carcinogens and is classified as possibly carcinogenic to humans (Group 2B) by the IARC.92,93 

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.94 In short-term (4 week) rat studies, degeneration of 

olfactory epithelium was observed at various concentration levels of acetaldehyde exposure.95,96 

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.97 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.

98

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

carcinogenicity in humans.99 

Lesions to the lungs and upper respiratory tract of rats, rabbits, and hamsters have been 

observed after subchronic exposure to acrolein.100 The Agency has developed an RfC for 

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

101

 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.102 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.103 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 mice104) 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.105 
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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.106

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

photochemical oxidants, including PAN.107 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.108

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.109110 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.111 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.112 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).113,114 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.115 Chronic (long term) exposure of workers and rodents to naphthalene has been 

reported to cause cataracts and retinal damage.116 EPA released an external review draft of a 

reassessment of the inhalation carcinogenicity of naphthalene based on a number of recent 
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animal carcinogenicity studies.117 The draft reassessment completed external peer review.118 

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.119 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.120 

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

abnormal cell changes and growth in respiratory and nasal tissues.121 The current EPA IRIS 

assessment includes noncancer data on hyperplasia and metaplasia in nasal tissue that form the 

basis of the inhalation RfC of 3 μg/m3.

122

 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.J

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.K,123 Other 

systematic reviews of relevant literature are cited were appropriate.

 

J U.S. EPA Integrated Risk Information System (IRIS) database is available at: www.epa.gov/iris

K

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.
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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. 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 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.
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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.124 Although the ATUS does not indicate their mode of travel, the 

majority of trips undertaken nationally is by motor vehicle.125 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 

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 
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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.L 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.126,127

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.128

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.

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 

 

L 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.
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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.129,130,131 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 

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 
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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 

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.133 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 
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effects tended to be small, while those with null effects tended to be larger, suggesting the 

potential for publication bias in the available literature.134

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.135 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 

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.136

EPA is pursuing a two-part strategy to address visibility impairment. First, EPA 

developed the regional haze program (64 FR 35714) which was put in place in July 1999 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. Second, EPA has 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. Figure 6-2 shows the location of the 156 Mandatory

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

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 

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, 

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 
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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.137

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 Integrated Science Assessment for Oxides of Nitrogen and Sulfur- Ecological 

Criteria.

138,139,140

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. 

Excesses of nitrogen or sulfur can lead to acidification, nutrient enrichment, and eutrophication 

of aquatic ecosystems.141 

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 and sediments, 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 
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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.142 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. 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.143 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 

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.144

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 
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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 

edible fish and shellfish, and when toxins become airborne, causing respiratory problems due to 

inhalation. According to a NOAA 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.145

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.

146

 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. Little research has been conducted on the effects associated 

with mixtures of contaminants found in ambient PM. 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 reenter 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.147 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.148,149 Upon 

entering the soil environment, PM pollutants can alter ecological processes of energy flow and 
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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.150

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 

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.151,152 Over fifty percent of the 

mercury in the Chesapeake Bay has been attributed to atmospheric deposition.153 Overall, the 

National Science and Technology Council identifies atmospheric deposition as the primary

source of mercury to aquatic systems.154 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.155,156 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.157 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.158 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.159 

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.160,161 Analyses of PAH deposition in Chesapeake and 

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

water predominate.162,163 Sediment concentrations of PAHs are high enough in some segments 
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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.

164

 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.165 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 

influence of local sources on deposition levels.166 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.167 

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

(SVOC) emissions in the United Kingdom deposit on soil.168 An analysis of PAH concentrations 

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

background.169

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

There are a number of environmental or public welfare effects associated with the 

presence of ozone in the ambient air.170 In this section we discuss the impact of ozone on plants, 

including trees, agronomic crops and urban ornamentals.

The Air Quality Criteria Document for Ozone and related Photochemical Oxidants notes 

that, “ozone affects vegetation throughout the United States, impairing crops, native vegetation, 

and ecosystems more than any other air pollutant.”

171

 Like carbon dioxide (CO2) and other 

gaseous substances, ozone enters plant tissues primarily through apertures (stomata) in leaves in 

a process called “uptake.”

172

 Once sufficient levels of ozone (a highly reactive substance), or its 

reaction products, reaches the interior of plant cells, it can inhibit or damage essential cellular 

components and functions, including enzyme activities, lipids, and cellular membranes, 

disrupting the plant's osmotic (i.e., water) balance and energy utilization patterns.173,174 If 

enough tissue becomes damaged from these effects, a plant’s capacity to fix carbon to form 
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carbohydrates, which are the primary form of energy used by plants, is reduced,175 while plant 

respiration increases. With fewer resources available, the plant reallocates existing resources 

away from root growth and storage, above ground growth or yield, and reproductive processes, 

toward leaf repair and maintenance, leading to reduced growth and/or reproduction. Studies 

have shown that plants stressed in these ways may exhibit a general loss of vigor, which can lead 

to secondary impacts that modify plants' responses to other environmental factors. Specifically, 

plants may become more sensitive to other air pollutants, more susceptible to disease, insect 

attack, harsh weather (e.g., drought, frost) and other environmental stresses. Furthermore, there 

is evidence that ozone can interfere with the formation of mycorrhiza, essential symbiotic fungi 

associated with the roots of most terrestrial plants, by reducing the amount of carbon available 

for transfer from the host to the symbiont.176,177

This ozone damage may or may not be accompanied by visible injury on leaves, and 

likewise, visible foliar injury may or may not be a symptom of the other types of plant damage 

described above. When visible injury is present, it is commonly manifested as chlorotic or 

necrotic spots, and/or increased leaf senescence (accelerated leaf aging). Because ozone damage 

can consist of visible injury to leaves, it can also reduce the aesthetic value of ornamental 

vegetation and trees in urban landscapes, and negatively affects scenic vistas in protected natural 

areas. 

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

concentration level and the duration of the exposure. Ozone effects also tend to accumulate over 

the growing season of the plant, so that even lower concentrations experienced for a longer 

duration have the potential to create chronic stress on sensitive vegetation. Not all plants, 

however, are equally sensitive to ozone. Much of the variation in sensitivity between individual 

plants or whole species is related to the plant’s ability to regulate the extent of gas exchange via 

leaf stomata (e.g., avoidance of ozone uptake through closure of stomata)178,179,180 Other 

resistance mechanisms may involve the intercellular production of detoxifying substances. 

Several biochemical substances capable of detoxifying ozone have been reported to occur in 

plants, including the antioxidants ascorbate and glutathione. After injuries have occurred, plants 

may be capable of repairing the damage to a limited extent.181

Because of the differing sensitivities among plants to ozone, ozone pollution can also 

exert a selective pressure that leads to changes in plant community composition. Given the range 

of plant sensitivities and the fact that numerous other environmental factors modify plant uptake 

and response to ozone, it is not possible to identify threshold values above which ozone is 

consistently toxic for all plants. The next few paragraphs present additional information on 

ozone damage to trees, ecosystems, agronomic crops and urban ornamentals.

Assessing the impact of ground-level ozone on forests in the United States involves 

understanding the risks to sensitive tree species from ambient ozone concentrations and 

accounting for the prevalence of those species within the forest. As a way to quantify the risks to 

particular plants from ground-level ozone, scientists have developed ozone-exposure/treeresponse functions by exposing tree seedlings to different ozone levels and measuring reductions 

in growth as “biomass loss.” Typically, seedlings are used because they are easy to manipulate 

and measure their growth loss from ozone pollution. The mechanisms of susceptibility to ozone 
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within the leaves of seedlings and mature trees are identical, though the magnitude of the effect 

may be higher or lower depending on the tree species.182

Some of the common tree species in the United States that are sensitive to ozone are 

black cherry (Prunus serotina), tulip-poplar (Liriodendron tulipifera), and eastern white pine 

(Pinus strobus). Ozone-exposure/tree-response functions have been developed for each of these 

tree species, as well as for aspen (Populus tremuliodes), and ponderosa pine (Pinus ponderosa). 

Other common tree species, such as oak (Quercus spp.) and hickory (Carya spp.), are not nearly 

as sensitive to ozone. Consequently, with knowledge of the distribution of sensitive species and 

the level of ozone at particular locations, it is possible to estimate a “biomass loss” for each 

species across their range. 

Ozone also has been conclusively shown to cause discernible injury to forest trees.183,184 

In terms of forest productivity and ecosystem diversity, ozone may be the pollutant with the 

greatest potential for regional-scale forest impacts. Studies have demonstrated repeatedly that 

ozone concentrations commonly observed in polluted areas can have substantial impacts on plant 

function.185,186

Because plants are at the base of the food web in many ecosystems, changes to the plant 

community can affect associated organisms and ecosystems (including the suitability of habitats 

that support threatened or endangered species and below ground organisms living in the root 

zone). Ozone impacts at the community and ecosystem level vary widely depending upon 

numerous factors, including concentration and temporal variation of tropospheric ozone, species 

composition, soil properties and climatic factors.187 In most instances, responses to chronic or 

recurrent exposure in forested ecosystems are subtle and not observable for many years. These 

injuries can cause stand-level forest decline in sensitive ecosystems.188,189,190 It is not yet 

possible to predict ecosystem responses to ozone with much certainty; however, considerable 

knowledge of potential ecosystem responses has been acquired through long-term observations 

in highly damaged forests in the United States.

Air pollution can have noteworthy cumulative impacts on forested ecosystems by

affecting regeneration, productivity, and species composition.191 In the U.S., ozone in the lower 

atmosphere is one of the pollutants of primary concern. Ozone injury to forest plants can be 

diagnosed by examination of plant leaves. Foliar injury is usually the first visible sign of injury

to plants from ozone exposure and indicates impaired physiological processes in the leaves.192

However, not all impaired plants will exhibit visible symptoms.

Laboratory and field experiments have also shown reductions in yields for agronomic 

crops exposed to ozone, including vegetables (e.g., lettuce) and field crops (e.g., cotton and 

wheat). The most extensive field experiments, conducted under the National Crop Loss 

Assessment Network (NCLAN) examined 15 species and numerous cultivars. The NCLAN 

results show that “several economically important crop species are sensitive to ozone levels 

typical of those found in the United States.”193 In addition, economic studies have shown 

reduced economic benefits as a result of predicted reductions in crop yields associated with 

observed ozone levels.194,195,196
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Urban ornamentals represent an additional vegetation category likely to experience some 

degree of negative effects associated with exposure to ambient ozone levels. It is estimated that 

more than $20 billion (1990 dollars) are spent annually on landscaping using ornamentals, both 

by private property owners/tenants and by governmental units responsible for public areas.197 

This is therefore a potentially costly environmental effect. However, in the absence of adequate 

exposure-response functions and economic damage functions for the potential range of effects 

relevant to these types of vegetation, no direct quantitative analysis has been conducted.

6.2.3.1 Recent Ozone Visible Foliar Injury Data for the U.S

In the U.S. the national-level visible foliar injury indicator is based on data from the U.S. 

Department of Agriculture (USDA) Forest Service Forest Inventory and Analysis (FIA) 

program. As part of its Phase 3 program, formerly known as Forest Health Monitoring, FIA 

examines ozone injury to ozone-sensitive plant species at ground monitoring sites in forest land 

across the country. For this indicator, forest land does not include woodlots and urban trees. 

Sites are selected using a systematic sampling grid, based on a global sampling design.198,199 At 

each site that has at least 30 individual plants of at least three ozone-sensitive species and enough 

open space to ensure that sensitive plants are not protected from ozone exposure by the forest 

canopy, FIA looks for damage on the foliage of ozone-sensitive forest plant species. Because 

ozone injury is cumulative over the course of the growing season, examinations are conducted in 

July and August, when ozone injury is typically highest. Monitoring of ozone injury to plants by

the USDA Forest Service has expanded over time from monitoring sites in 10 states in 1994 to 

nearly 1,000 monitoring sites in 41 states in 2002. 

There is considerable regional variation in ozone-related visible foliar injury to sensitive 

plants in the U.S. The U.S. EPA has developed an environmental indicator based on data from 

the USDA FIA program which examines ozone injury to ozone-sensitive plant species at ground 

monitoring sites in forest land across the country. The data underlying the indicator in Figure 

6-3 are based on averages of all observations collected in 2002, the latest year for which data are 

publicly available at the time the study was conducted, and is broken down by U.S. EPA Region. 

Ozone damage to forest plants is classified using a subjective five-category biosite index based 

on expert opinion, but designed to be equivalent from site to site. Ranges of biosite values 

translate to no injury, low or moderate foliar injury (visible foliar injury to highly sensitive or 

moderately sensitive plants, respectively), and high or severe foliar injury, which would be 

expected to result in tree-level or ecosystem-level responses, respectively.

200,201

The highest percentages of observed high and severe foliar injury, those which are most 

likely to be associated with tree or ecosystem-level responses, are primarily found in the MidAtlantic and Southeast regions. In EPA Region 3 (which comprises the States of Pennsylvania, 

West Virginia, Virginia, Delaware, Maryland and Washington D.C.), 12 percent of ozonesensitive plants showed signs of high or severe foliar damage, and in Region 2 (States of New 

York, New Jersey), and Region 4 (States of North Carolina, South Carolina, Kentucky, 

Tennessee, Georgia, Florida, Alabama, and Mississippi) the values were 10 and 7 percent, 

respectively. The sum of high and severe ozone injury ranged from 2 to 4 percent in EPA 

Region 1 (the six New England States), Region 7 (States of Missouri, Iowa, Nebraska and 

Kansas), and Region 9 (States of California, Nevada, Hawaii and Arizona). The percentage of 

sites showing some ozone damage was about 45 percent in each of these EPA Regions. 
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Figure 6-3 Ozone Injury to Forest Plants in U.S. by EPA Regions, 2002ab

6.2.3.2 Indicator Limitations

The categories for the biosite index are subjective and may not necessarily be directly 

related to biomass loss or physiological damage to plants in a particular area. Ozone may have 

other adverse impacts on plants (e.g., reduced productivity) that do not show signs of visible 

foliar injury.

202

 The presence of diagnostic visible ozone injury on indicator plants does provide 

evidence that ozone is having an impact in an area. However, absence of ozone injury in an area 

does not necessarily mean that there is no impact from ozone exposure.

Field and laboratory studies were reviewed to identify the forest plant species in each 

region that are sensitive to ozone air pollution and exhibit diagnostic injury. Other forest plant 

species, or even genetic variants of the same species, may not show symptoms at ozone levels 

that cause effects on the selected ozone-sensitive species. 

Because species distributions vary regionally, different ozone-sensitive plant species 

were examined in different parts of the country. These target species could vary with respect to 

ozone sensitivity, which might account for some of the apparent differences in ozone injury

among regions of the U.S. Ozone damage to foliage may be reduced under conditions of low 

soil moisture, but most of the variability in the index (70 percent) was explained by ozone 

concentration.203 
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Though FIA has extensive spatial coverage based on a robust sample design, not all 

forested areas in the U.S. are monitored for ozone injury. Even though the biosite data have been 

collected over multiple years, most biosites were not monitored over the entire period, so these 

data cannot provide more than a baseline for future trends.

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.204 In laboratory experiments, a wide range of tolerance to VOCs has been 

observed.205 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.206

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.207,208,209 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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