Document ID: EPA-HQ-OAR-2011-0135-2792
Agency: epa
Document Type: Supporting & Related Material
Title: 
Posted Date: 2013-05-17T04:00Z

Health and Environmental Effects Associated with Exposure to Criteria
and Toxic Pollutants 

Health Effects of Criteria and Toxic Pollutants

Particulate Matter

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. 

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.  

Health Effects of PM

This section provides a summary of the health effects associated with
exposure to ambient concentrations of PM.  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). 

The ISA concludes that ambient concentrations of PM are associated with
a number of adverse health effects.  The 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.

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.  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.

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.  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.

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. 
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.

Effects Associated with Ultrafine Particles

The ISA concludes that the evidence is suggestive of a causal
relationship between short-term exposures to UFPs and cardiovascular
effects, including changes in heart rhythm and vasomotor function (the
ability of blood vessels to expand and contract).  

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.

Ozone

Background

Ground-level ozone pollution is typically formed through reactions
involving VOCs and NOX in the lower atmosphere in the presence of
sunlight.  These pollutants, often referred to as ozone precursors, are
emitted by many types of pollution sources such as highway and nonroad
motor vehicles and engines, power plants, chemical plants, refineries,
makers of consumer and commercial products, industrial facilities, and
smaller area sources. 

The science of ozone formation, transport, and accumulation is complex. 
Ground-level ozone is produced and destroyed in a cyclical set of
chemical reactions, many of which are sensitive to temperature and
sunlight.  When ambient temperatures and sunlight levels remain high for
several days and the air is relatively stagnant, ozone and its
precursors can build up and result in more ozone than typically occurs
on a single high-temperature day.  Ozone and its precursors can be
transported hundreds of miles downwind of precursor emissions, resulting
in elevated ozone levels even in areas with low VOC or NOX emissions. 

The highest levels of ozone are produced when both VOC and NOX emissions
are present in significant quantities on clear summer days.  Relatively
small amounts of NOX enable ozone to form rapidly when VOC levels are
relatively high, but ozone production is quickly limited by removal of
the NOX.  Under these conditions NOX reductions are highly effective in
reducing ozone while VOC reductions have little effect.  Such conditions
are called “NOX-limited.”  Because the contribution of VOC emissions
from biogenic (natural) sources to local ambient ozone concentrations
can be significant, even some areas where man-made VOC emissions are
relatively low can be NOX-limited.

Ozone concentrations in an area also can be lowered by the reaction of
nitric oxide (NO) with ozone, forming nitrogen dioxide (NO2); as the air
moves downwind and the cycle continues, the NO2 forms additional ozone. 
The importance of this reaction depends, in part, on the relative
concentrations of NOX, VOC, and ozone, all of which change with time and
location.  When NOX levels are relatively high and VOC levels relatively
low, NOX forms inorganic nitrates (i.e., particles) but relatively
little ozone.  Such conditions are called “VOC-limited.”  Under
these conditions, VOC reductions are effective in reducing ozone, but
NOX reductions can actually increase local ozone under certain
circumstances.  Even in VOC-limited urban areas, NOX reductions are not
expected to increase ozone levels if the NOX reductions are sufficiently
large.  Rural areas are usually NOX-limited, due to the relatively large
amounts of biogenic VOC emissions in such areas.  Urban areas can be
either VOC- or NOX-limited, or a mixture of both, in which ozone levels
exhibit moderate sensitivity to changes in either pollutant.

Health Effects of Ozone

Exposure to ambient ozone contributes to a wide range of adverse health
effects.  These health effects are well documented and are critically
assessed in the EPA ozone air quality criteria document (ozone AQCD) and
EPA staff paper.,  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.  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.

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.,
, , , ,   Repeated exposure to ozone can increase susceptibility to
respiratory infection and lung inflammation and can aggravate
preexisting respiratory diseases, such as asthma., , , ,  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., , , 

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.  For example, summer camp studies have reported statistically
significant reductions in lung function in children who are active
outdoors., , , , , , ,   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., , , 

Nitrogen Oxides and Sulfur Oxides

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 and
NO2 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   REF _Ref305681120 \w \h  6.1.1.2 .  NOx along
with VOCs are the two major precursors of ozone.  The health effects of
ozone are covered in Section   REF _Ref305681131 \w \h  6.1.2.2 .

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.  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 in
moderate to heavy levels of exercise, with more limited evidence of
respiratory effects among exercising asthmatics exposed to
concentrations as low as 0.2-0.3 ppm.  A clear concentration-response
relationship has been demonstrated in these studies following exposures
to SO2 at concentrations between 0.2 and 1.0 ppm, both in terms of
increasing severity of respiratory symptoms and decrements in lung
function, as well as the percentage of asthmatics adversely affected. 

In epidemiologic studies, respiratory effects have been observed in
areas where the mean 24-hour SO2 levels range from 1 to 30 ppb, with
maximum 1 to 24-hour average SO2 values ranging from 12 to 75 ppb. 
Important new multicity studies and several other studies have found an
association between 24-hour average ambient SO2 concentrations and
respiratory symptoms in children, particularly those with asthma. 
Generally consistent associations also have been observed between
ambient SO2 concentrations and emergency department visits and
hospitalizations for all respiratory causes, particularly among children
and older adults (≥ 65 years), and for asthma.  A limited subset of
epidemiologic studies has examined potential confounding by copollutants
using multipollutant regression models.  These analyses indicate that
although copollutant adjustment has varying degrees of influence on the
SO2 effect estimates, the effect of SO2 on respiratory health outcomes
appears to be generally robust and independent of the effects of gaseous
and particulate copollutants, suggesting that the observed effects of
SO2 on respiratory endpoints occur independent of the effects of other
ambient air pollutants. 

Consistent associations between short-term exposure to SO2 and mortality
have been observed in epidemiologic studies, with larger effect
estimates reported for respiratory mortality than for cardiovascular
mortality.  While this finding is consistent with the demonstrated
effects of SO2 on respiratory morbidity, uncertainty remains with
respect to the interpretation of these associations due to potential
confounding by various copollutants.  The U.S. EPA has therefore
concluded that the overall evidence is suggestive of a causal
relationship between short-term 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.

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.  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 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.

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.  The
ISA concludes that ambient concentrations of CO are associated with a
number of adverse health effects.  This section provides a summary of
the health effects associated with exposure to ambient concentrations of
CO.  

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.  

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.

Health Effects of Air Toxics

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.,,  EPA states in its IRIS database that data
indicate a causal relationship between benzene exposure and acute
lymphocytic leukemia and suggest a relationship between benzene exposure
and chronic non-lymphocytic leukemia and chronic lymphocytic leukemia. 
EPA’s IRIS documentation for benzene also lists a range of 2.2 x 10-6
to 7.8 x 10-6 as the unit risk estimate (URE) for benzene.,  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.,  

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.,  The most sensitive noncancer effect
observed in humans, based on current data, is the depression of the
absolute lymphocyte count in blood.,  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.,,,  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/m3 for 1-14 days exposure.,

1,3-Butadiene

EPA has characterized 1,3-butadiene as carcinogenic to humans by
inhalation.,  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.,,  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 × 10-5 per µg/m3.  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.  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).

Ethanol

EPA is planning to develop an assessment of the health effects of
exposure to ethanol, a compound which is not currently listed on EPA’s
IRIS database.  Extensive health effects data are available for
ingestion of ethanol, while data on inhalation exposure effects are
sparse.  In developing the assessment, EPA is evaluating pharmacokinetic
models as a means of extrapolating across species (animal to human) and
across exposure routes (oral to inhalation) to better characterize the
health hazards and dose-response relationships for low levels of ethanol
exposure in the environment.

The IARC has classified “alcoholic beverages” as carcinogenic to
humans based on sufficient evidence that malignant tumors of the oral
cavity, pharynx, larynx, esophagus, and liver are causally related to
the consumption of alcoholic beverages.  The U.S. DHHS in the 12th
Report on Carcinogens also identified “alcoholic beverages” as a
known human carcinogen (they have not evaluated the cancer risks
specifically from exposure to ethanol), with evidence for cancer of the
mouth, pharynx, larynx, esophagus, liver and breast.  There are no
studies reporting carcinogenic effects from inhalation of ethanol.  EPA
is currently evaluating the available human and animal cancer data to
identify which cancer type(s) are the most relevant to a human health
assessment of low-level oral and inhalation exposure to ethanol.

Noncancer health effects data are available from animal studies as well
as epidemiology studies.  The epidemiologic data were obtained from
studies of alcoholic beverage consumption.  Reported effects included
neurological impairment, developmental effects, cardiovascular effects,
immune system depression, and effects on the liver, pancreas and
reproductive system.  There is evidence that children prenatally exposed
via maternal ingestion of alcoholic beverages during pregnancy are at
increased risk of hyperactivity and attention deficits, impaired motor
coordination, a lack of regulation of social behavior or poor
psychosocial functioning, and deficits in cognition, mathematical
ability, verbal fluency, and spatial memory.,,,,,,,  Genetic factors
influencing ethanol metabolism may render certain subpopulations more
susceptible to the health effects of ethanol exposure by altering
internal levels of ethanol and/or its metabolites (e.g., acetaldehyde).

Formaldehyde

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

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.,,  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.  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.  Finally, a study of embalmers reported formaldehyde exposures
to be associated with an increased risk of myeloid leukemia but not
brain cancer. 

	Health effects of formaldehyde in addition to cancer were reviewed by
the Agency for Toxics Substances and Disease Registry in 1999 and
supplemented in 2010, and by the World Health Organization.  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.  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 2011
(http://www.nap.edu/catalog.php?record_id=13142).  The EPA is currently
revising the draft assessment in response to this review.

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.  The URE in IRIS for
acetaldehyde is 2.2 × 10-6 per µg/m3.  Acetaldehyde is reasonably
anticipated to be a human carcinogen by the U.S. DHHS in the 12th Report
on Carcinogens and is classified as possibly carcinogenic to humans
(Group 2B) by the IARC.,  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.  In short-term (4
week) rat studies, degeneration of olfactory epithelium was observed at
various concentration levels of acetaldehyde exposure.,  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.  The
agency is currently conducting a reassessment of the health hazards from
inhalation exposure to acetaldehyde.  

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

Lesions to the lungs and upper respiratory tract of rats, rabbits, and
hamsters have been observed after subchronic exposure to acrolein.  The
Agency has developed an RfC for acrolein of 0.02 µg/m3 and an RfD of
0.5 µg/kg-day.  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.  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.  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 hyper-responsiveness.  
Based on animal data (more pronounced respiratory irritancy in mice with
allergic airway disease in comparison to non-diseased mice) 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/m3 for 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/m3 and 0.7 µg/m3, respectively.  

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.

Effects of ground-level smog causing intense eye irritation have been
attributed to photochemical oxidants, including PAN.  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.

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.  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. 
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.  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)., 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.

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.  Chronic (long term)
exposure of workers and rodents to naphthalene has been reported to
cause cataracts and retinal damage.  EPA released an external review
draft of a reassessment of the inhalation carcinogenicity of naphthalene
based on a number of recent animal carcinogenicity studies.  The draft
reassessment completed external peer review.  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.  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.  

Naphthalene also causes a number of chronic non-cancer effects in
animals, including abnormal cell changes and growth in respiratory and
nasal tissues.  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.  The ATSDR MRL for acute exposure to
naphthalene is 0.6 mg/kg/day.

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.

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.,  Other systematic
reviews of relevant literature are cited were appropriate.

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.    REF _Ref305501234 \h
 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.

Figure 6-  SEQ Figure \* ARABIC  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.  Although the
ATUS does not indicate their mode of travel, the majority of trips
undertaken nationally is by motor vehicle.  As such, daily travel
activity brings nearly all residents into a high-exposure
microenvironment for part of the day. 

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 traffic-related exposure and cardiovascular mortality.  

Cardiovascular effects

Cardiac physiology

Exposure to traffic-associated pollutants has been associated with
changes in cardiac physiology, including cardiac function.  One common
measure of cardiac function is heart rate variability (HRV), an
indicator of the heart’s ability to respond to variations in stress,
reflecting the nervous system’s ability to regulate the heart. 
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.,

Heart attack and atherosclerosis

The HEI panel concluded that epidemiologic evidence of the association
between traffic-related 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.

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.

Respiratory effects

Asthma

Pediatric asthma and asthma symptoms are the effects that have been
evaluated by the largest number of studies in the epidemiologic
literature on the topic.  In general, studies consistently show effects
of residential or school exposure to traffic and asthma symptoms, and
the effects are frequently statistically significant.  Studies have
employed both short-term and long-term exposure metrics, and a range of
different respiratory measures.  HEI Special Report 17 (HEI Panel on the
Health Effects of Traffic-Related Air Pollution, 2010) concluded that
there is sufficient evidence for a causal association between exposure
to traffic-related air pollution and exacerbation of asthma symptoms in
children.  

While there is general consistency in studies examining asthma incidence
in children, the available studies employ different definitions of
asthma (e.g., self-reported vs. hospital records), methods of exposure
assessment, and population age ranges.  As such, the overall evidence,
while supportive of an association between traffic exposure and new
onset asthma, are less consistent than for asthma symptoms. The HEI
report determined that there is “sufficient” or “suggestive”
evidence of a causal relationship between exposure to traffic-related
air pollution and incident (new onset) asthma in children (HEI Panel on
the Health Effects of Traffic-Related Air Pollution, 2010).  A recent
meta-analysis of studies on incident asthma and air pollution in
general, based on studies dominated by traffic-linked exposure metrics,
also concluded that available evidence that exposures is consistent with
a effect of exposure on asthma incidence (Anderson et al., 2011).  The
study reported excess main risk estimates for different pollutants
ranging from 7-16 percent per 10 g/m3 of 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.

Chronic obstructive pulmonary disease (COPD)

The HEI panel reviewed available studies examining COPD in the context
of traffic-associated 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.

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.,,,  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.

Lung function

There are numerous measurements of breathing (spirometry) that indicate
the presence or degree of airway disease, such as asthma and chronic
obstructive pulmonary disease (COPD).  Forced vital capacity (FVC) is
measured when a patient maximally fills their lungs and then blows their
hardest in completely exhaling.  The peak expiratory flow (PEF) is the
maximum air flow achievable during exhalation.  The forced expiratory
volume in the first second of exhalation is referred to as FEV1.  FEV1
and PEF reflect the function of the large airways.  FVC and FEV1, along
with their ratio (FVC/FEV1) are used to classify airway obstruction in
asthma and COPD.  Measurements of air flow at various times during
forced exhalation, such as 25 percent, 50 percent, and 75 percent, are
also used.  The flow at 75 percent of forced exhalation (FEF75) reflects
the status of small airways, which asthma and COPD affect. 

The HEI panel concluded that the available literature suggests that
long-term exposure to traffic-related air pollution is associated with
reduced lung function in adolescents and young adults and that lung
function is lower in populations in areas with high traffic-related air
pollutant levels.  However, the panel noted the difficulty of
disentangling traffic-specific exposures from urban air pollution in
general.  The studies reviewed that were more specifically oriented
toward traffic were not consistent in their findings.  As a result, the
panel found that the evidence linking lung function and traffic exposure
is “inadequate and insufficient” to infer a causal relationship. 

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.  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.

Cancer

Childhood cancer

A number of studies examining various types of childhood cancer have
been published with mixed results.  The HEI panel concluded that the
available epidemiologic evidence was “inadequate and insufficient”
to infer a causal relationship between traffic-related air pollution and
childhood cancer.  An earlier review article on the topic noted that
studies reporting positive effects tended to be small, while those with
null effects tended to be larger, suggesting the potential for
publication bias in the available literature. 

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.

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.

Environmental Effects of Criteria and Toxic Pollutants

Visibility Degradation

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

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.    REF _Ref300739587 \h  Figure 6-2 
shows the location of the 156 Mandatory Class I Federal areas. 

Figure   STYLEREF 1 \s  6 -  SEQ Figure \* ARABIC \s 1  2  Mandatory
Class I Federal Areas in the U.S.

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   REF _Ref300739587 \h  Figure 6-2 ).  This long-term
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 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.

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.,,

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.  

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 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.  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.  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.

In the U.S. numerous forests now show severe symptoms of nitrogen
saturation.  These forests include:  the northern hardwoods and mixed
conifer forests in the Adirondack and Catskill Mountains of  New York;
the red spruce forests at Whitetop Mountain, Virginia, and Great Smoky
Mountains National Park, North Carolina; mixed hardwood watersheds at
Fernow Experimental Forest in West Virginia; American beech forests in
Great Smoky Mountains National Park, Tennessee;  mixed conifer forests
and chaparral watersheds in southern California and the southwestern
Sierra Nevada in Central California; the alpine tundra/subalpine conifer
forests of the Colorado Front Range; and red alder forests in the
Cascade Mountains in Washington.

Excess nutrient inputs into aquatic ecosystems (i.e. streams, rivers,
lakes, estuaries or oceans) either from direct atmospheric deposition,
surface runoff, or leaching from nitrogen saturated soils into ground or
surface waters can contribute to conditions of severe water oxygen
depletion; eutrophication and algae blooms; altered fish distributions,
catches, and physiological states; loss of biodiversity; habitat
degradation; and increases in the incidence of disease.

Atmospheric deposition of nitrogen is a significant source of total
nitrogen to many estuaries in the United States.  The amount of nitrogen
entering estuaries that is ultimately attributable to atmospheric
deposition is not well-defined.  On an annual basis, atmospheric
nitrogen deposition may contribute significantly to the total nitrogen
load, depending on the size and location of the watershed.  In addition,
episodic nitrogen inputs, which may be ecologically important, may play
a more important role than indicated by the annual average
concentrations.  Estuaries in the U.S. that suffer from nitrogen
enrichment often experience a condition known as eutrophication. 
Symptoms of eutrophication include changes in the dominant species of
phytoplankton, low levels of oxygen in the water column, fish and
shellfish kills, outbreaks of toxic alga, and other population changes
which can cascade throughout the food web.  In addition, increased
phytoplankton growth in the water column and on surfaces can attenuate
light causing declines in submerged aquatic vegetation, which serves as
an important habitat for many estuarine fish and shellfish species.

Severe and persistent eutrophication often directly impacts human
activities.  For example, losses in the nation’s fishery resources may
be directly caused by fish kills associated with low dissolved oxygen
and toxic blooms.  Declines in tourism occur when low dissolved oxygen
causes noxious smells and floating mats of algal blooms create
unfavorable aesthetic conditions.  Risks to human health increase when
the toxins from algal blooms accumulate in 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.

Deposition of Heavy Metals

Heavy metals, including cadmium, copper, lead, chromium, mercury, nickel
and zinc, have the greatest potential for impacting forest growth. 
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 re-enter the soil
and further cycle in the environment.  Although there has been no direct
evidence of a physiological association between tree injury and heavy
metal exposures, heavy metals have been implicated because of
similarities between metal deposition patterns and forest decline.  This
hypothesized relationship/correlation was further explored in high
elevation forests in the northeastern U.S.  These studies measured
levels of a group of intracellular compounds found in plants that bind
with metals and are produced by plants as a response to sublethal
concentrations of heavy metals.  These studies indicated a systematic
and significant increase in concentrations of these compounds associated
with the extent of tree injury.  These data strongly imply that metal
stress causes tree injury and contributes to forest decline in the
northeastern United States.  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.,  Upon entering the soil environment, PM
pollutants can alter ecological processes of energy flow and nutrient
cycling, inhibit nutrient uptake, change ecosystem structure, and affect
ecosystem biodiversity.  Many of the most important effects occur in the
soil.  The soil environment is one of the most dynamic sites of
biological interaction in nature. It is inhabited by microbial
communities of bacteria, fungi, and actinomycetes.  These organisms are
essential participants in the nutrient cycles that make elements
available for plant uptake.  Changes in the soil environment that
influence the role of the bacteria and fungi in nutrient cycling
determine plant and ultimately ecosystem response. 

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.,  Over fifty percent of the mercury in
the Chesapeake Bay has been attributed to atmospheric deposition. 
Overall, the National Science and Technology Council identifies
atmospheric deposition as the primary source of mercury to aquatic
systems.  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.,  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.  Plant
uptake of platinum has been observed at these locations.

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. 
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.  

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.,  Analyses of PAH deposition in Chesapeake and Galveston Bay
indicate that dry deposition and gas exchange from the atmosphere to the
surface water predominate.,  Sediment concentrations of PAHs are high
enough in some segments of Tampa Bay to pose an environmental health
threat.  EPA funded a study to better characterize the sources and
loading rates for PAHs into Tampa Bay.  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. 
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.  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.  

Cousins et al. estimate that more than ninety percent of semi-volatile
organic compound (SVOC) emissions in the United Kingdom deposit on soil.
 An analysis of PAH concentrations near a Czechoslovakian roadway
indicated that concentrations were thirty times greater than background.

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.

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.  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.”  Like carbon dioxide (CO2) and other gaseous
substances, ozone enters plant tissues primarily through apertures
(stomata) in leaves in a process called “uptake.”  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.,  If enough tissue becomes
damaged from these effects, a plant’s capacity to fix carbon to form
carbohydrates, which are the primary form of energy used by plants, is
reduced, 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.,

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),,  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.

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/tree-response 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 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. 

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.,  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.,

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.  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.,,  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.  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. 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.”  In addition, economic studies have shown reduced economic
benefits as a result of predicted reductions in crop yields associated
with observed ozone levels.,,

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.  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.

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.,  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   REF _Ref300740364 \h  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.,

 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 Mid-Atlantic 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 ozone-sensitive 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. 

Figure   STYLEREF 1 \s  6 -  SEQ Figure \* ARABIC \s 1  3  Ozone Injury
to Forest Plants in U.S. by EPA Regions, 2002ab

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.  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.  

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.

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

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.,,  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. 

References

 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.

  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.

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

 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.  

  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.

  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.  

  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.

 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.

 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.

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

 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.

 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.

Chapter 1 

*** E.O. 12866 Review – Revised Version 3/25/13 – Do Not Cite,
Quote, or Release During Review ***

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