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 1 

Executive Summary

1.1 Introduction

Air Quality Criteria for Ozone and Related Photochemical Oxidants evaluates the

latest scientific information useful in deriving criteria that form the scientific basis for U.S.

Environmental Protection Agency (EPA) decisions regarding the National Ambient Air

Quality Standards (NAAQS) for ozone (O ). This Executive Summary concisely summarizes 3

key conclusions from the document, which comprises nine chapters. Following this

Executive Summary is a brief Introduction (Chapter 2) containing information on the

legislative and regulatory background for review of the O NAAQS, as well as a brief 3

discussion of the issues presented and the format for their discussion in the document. 

Chapter 3 provides information on the chemistry, sources, emissions, measurement, and

transport of O and related photochemical oxidants and their precursors, whereas Chapter 4 3

covers environmental concentrations, patterns, and exposure estimates of O and oxidants. 3

Chapter 5 deals with environmental effects, and Chapters 6, 7, and 8 discuss animal

toxicological studies, human health effects, and extrapolation of animal toxicological data to

humans, respectively. The last chapter, Chapter 9, provides an integrative, interpretative

characterization of health effects associated with exposure to O . The following sections 3

conform to the chapter organization of the criteria document.

1.2 Legislative and Regulatory Background

The photochemical oxidants found in ambient air in the highest concentrations are

O and nitrogen dioxide (NO ). Other oxidants, such as hydrogen peroxide (H O ) and 3 2 2 2

peroxyacyl nitrates, also have been observed, but in lower and less certain concentrations. In

1971, EPA promulgated NAAQS to protect the public health and welfare from adverse effects

of photochemical oxidants, at that time, defined on the basis of commercially available

measurement methodology. After 1971, however, O -specific commercial analytical methods 3

became available, as did information on the concentrations and effects of the related nonO photochemical oxidants. As a result, the chemical designation of the standards was 3

changed in 1979 from photochemical oxidants to O . 3

The EPA is required under Sections 108 and 109 of the Clean Air Act to evaluate

periodically the air quality criteria that reflect the latest scientific information relevant to

review of the O NAAQS. These air quality criteria are useful for indicating the kind and 3
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extent of all identifiable effects on public health or welfare that may be expected from the

presence of O and related photochemical oxidants in ambient air. The last O criteria 3 3

document was released in 1986, and a supplement was released in 1992. These documents

were the basis for a March 1993 decision by EPA that revision of the existing 1-h NAAQS for

O was not appropriate at that time. That decision, however, did not take into consideration 3

more recent scientific information that has been published since the last literature review in

early 1989. The purpose of this revised criteria document, therefore, is to summarize the

pertinent information contained in the previous O criteria document and to critically evaluate 3

and assess the more recent scientific data associated with exposure to O and, to a lesser 3

extent, to H O and the peroxyacyl nitrates, particularly peroxyacetyl nitrate (PAN). This 2 2

document will be used by EPA's Office of Air Quality Planning and Standards to provide a

staff paper assessing the most significant scientific information and presenting staff

recommendations on whether revisions to the O NAAQS are appropriate. 3

1.3 Tropospheric Ozone and Its Precursors

Introduction

Ozone is found in the stratosphere, the "free" troposphere, and the planetary

boundary layer (PBL) of the earth's atmosphere. In the PBL, background O occurs as the 3

result of (1) the intrusions of stratospheric O into the "free" troposphere and downward 3

transport into the PBL, and (2) photochemical reactions of methane (CH ), carbon monoxide 4

(CO), and nitrogen oxides (NO ). These processes contribute to the background O near the x 3

surface. The major source of O in the PBL is the photochemical process involving 3

anthropogenic and biogenic emissions of NO with the many classes of volatile organic x

compounds (VOCs).

The topics considered in this section of the document include: tropospheric

O chemistry; meteorological influences on O formation and transport; precursor VOC and 3 3

NO emissions, ambient concentrations of VOCs and NO , and source apportionment and x x

reconciliation of measured VOC ambient concentrations with emission inventories; O air 3

quality models; and analytical methods for oxidants and precursors.

Tropospheric Ozone Chemistry

Ozone occurs in the stratosphere as the result of chemical reactions initiated by

short-wavelength radiation from the sun. In the "free" troposphere, O occurs as the result of 3

incursions from the stratosphere; upward venting from the PBL (the layer next to the surface

of the earth) through certain cloud processes; and photochemical formation from precursors,

notably CH , CO, and NO . 4 x

The photochemical production of O and other oxidants found at the surface of the 3

earth (in the PBL, troposphere, or ambient air [used interchangeably in this summary]) is the

result of atmospheric physical processes and complex, nonlinear chemical processes

involving two classes of precursor pollutants: (1) reactive anthropogenic and biogenic VOCs

and (2) NO . The only significant initiator of the photochemical production of O in the x 3

polluted troposphere is the photolysis of NO , yielding nitric oxide (NO) and a ground-state 2

oxygen atom that reacts with molecular oxygen to form O . The O thus formed reacts with 3 3

NO, yielding oxygen and NO . These cyclic reactions attain equilibrium in the absence of 2
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VOCs. However, in the presence of VOCs, which are abundant in polluted ambient air, the

equilibrium is upset, resulting in a net increase in O . Methane is the chief VOC found in the 3

free troposphere and in most "clean" areas of the PBL. The VOCs found in polluted ambient

air are much more complex and more reactive than CH , but, as with CH , their atmospheric 4 4

oxidative degradation is initiated through attack on the VOCs by hydroxyl (OH) radicals. As

in the CH oxidation cycle, the conversion of NO to NO during the oxidation of VOCs is 4 2

accompanied by the production of O and the efficient regeneration of the OH radical. The 3

O3, PAN, and higher homologues formed in polluted atmospheres increase with the NO2/NO

concentration ratio. 

At night, in the absence of photolysis of reactants, the simultaneous presence of

O and NO results in the formation of the nitrate (NO ) radical. Reactions with NO radicals 3 2 3 3

appear to constitute major sinks for alkenes, cresols, and several other compounds, although

the chemistry is not well characterized.

Most inorganic gas-phase processes (i.e., the nitrogen cycle and its

interrelationships with O production) are well understood. The chemistry of the VOCs in 3

ambient air is not as well understood. It is well known, however, that the chemical loss

processes of gas-phase VOCs include reaction with OH and NO radicals and O , and 3 3

photolysis. Reaction with the OH radical is the only important atmospheric reaction (loss

process) for alkanes, aromatic hydrocarbons, and the higher aldehydes and ketones that lack

>C=C< bonds; and the only atmospheric reaction of alcohols and ethers. Photolysis is the

major loss process for formaldehyde and acetone. Reactions with OH and NO radicals and 3

with O3 are all important loss processes for alkenes and for carbonyls containing >C=C<

bonds.

Uncertainties in the atmospheric chemistry of the VOCs can affect quantification

of the NO-to-NO conversion and of O yields, and can present difficulties in representation 2 3

of chemical mechanisms, products, and product yields in O air quality models. Major 3

uncertainties in understanding the atmospheric chemistry of the VOCs with NO in both x

urban and rural atmospheres include chemistry of alkyl nitrate formation, mechanisms and

products of >C4 n-alkanes and branched alkanes, mechanisms and products of alkeneO reactions, and mechanisms and products of aromatic hydrocarbons. 3

It should be noted that the atmospheric chemical processes involved in the

photooxidation of certain higher molecular weight VOCs and in the formation of O also can 3

lead to the formation of particulate-phase organic compounds. The OH radicals produced not

only can oxidize VOCs to particulate-phase organic compounds but also can react with NO2

and sulfur dioxide (SO ) to form nitric acid (HNO ) and sulfuric acid (H SO ), respectively, 2 3 2 4

portions of which become incorporated into aerosols as particulate nitrate and sulfate.

Meteorological Influences on Ozone Formation and Transport

The surface energy (radiation) budget of the earth strongly influences the dynamics

of the PBL. The redistribution of energy through the PBL creates thermodynamic conditions

that influence vertical mixing. Growing evidence indicates that the strict use of mixing

heights in modeling is an oversimplification of the complex processes by which pollutants are

redistributed within urban areas, and that it is necessary to treat the turbulent structure of the

atmosphere directly and acknowledge the vertical variations in mixing. Energy balances

therefore require study so that more realistic simulations can be made of the structure of the

PBL.
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Day-to-day variability in O concentrations depends heavily on day-to-day 3

variations in meteorological conditions, including temperature, solar radiation, and the degree

of mixing that occurs between release of a pollutant or its precursors and their arrival at a

receptor; the occurrence of inversion layers (layers in which temperature increases with height

above ground level); and the transport of O left overnight in layers aloft and subsequent 3

downward mixing of that O to the surface. 3

The transport of O and its precursors beyond the urban scale ( 50 km) to 3

neighboring rural and urban areas has been well documented. Episodes of high

O concentrations in urban areas are often associated with high concentrations of O in the 3 3

surroundings. Areas of O accumulation usually are characterized by synoptic-scale 3

subsidence of air in the free troposphere, resulting in development of an elevated inversion

layer; relatively low wind speeds associated with the weak horizontal pressure gradient

around a surface high pressure system; a lack of cloudiness; and high temperatures.

Ultraviolet (UV) radiation from the sun plays a key role in initiating the

photochemical processes leading to O formation and affects individual photolytic reaction 3

steps. Still, there is little empirical evidence in the literature linking day-to-day variations in

observed UV radiation levels to variations in O levels. An association, however, between 3

tropospheric O3 concentrations and temperature has been demonstrated. Empirical data from

four urban areas, for example, show an apparent upper bound on O concentrations that 3

increases with temperature. A similar qualitative relationship exists at a number of rural

locations.

The relationship between wind speed and O buildup varies from one part of the 3

country to another. 

Statistical techniques (e.g., regression techniques) can be used to help identify real

trends in O concentrations, both intra- and interannual, by normalizing meteorological 3

variability.

Precursors

Volatile Organic Compound Emissions

Hundreds of VOCs, usually containing from 2 to 12 carbon atoms, are emitted by

evaporative and combustion processes from a large number of source types. Total U.S.

anthropogenic VOC emissions in 1991 were estimated at 21.0 Tg; the two largest source

categories were (1) industrial processes (10.0 Tg) and (2) transportation (7.9 Tg). Emissions

of VOCs from highway vehicles accounted for almost 75% of the transportation-related

emissions; studies have shown that the majority of these VOC emissions come from about

20% of the automobiles in service, many, perhaps most, of which are older cars that are

poorly maintained. The accuracy of VOC emission estimates is difficult to determine for both

stationary and mobile sources. 

Vegetation emits significant quantities of VOCs into the atmosphere, chiefly

monoterpenes and isoprene, but also oxygenated VOCs, according to recent studies. The

most recent biogenic VOC emissions estimate for the United States showed annual emissions

of 29.1 Tg/year. 

Although the biogenic VOC emission estimates exceed the anthropogenic

estimates, the biogenic emissions are more diffusely distributed than the anthropogenic

emissions, which tend to be concentrated in population centers. However, the large
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uncertainties in both biogenic and anthropogenic VOC emission inventories prevent

establishing the relative contributions of these two categories.

Nitrogen Oxides Emissions

Anthropogenic NO is associated with combustion processes. The primary x

pollutant emitted is NO, formed at high combustion temperatures from nitrogen and oxygen

in the air and from nitrogen in the combustion fuel. Emissions of NO in 1991 in the United x

States totaled 21.39 Tg. The two largest single NO emission sources are electric power x

generating plants and highway vehicles. Because a large proportion of anthropogenic NOx

emissions come from distinct point sources, published annual estimates are thought to be

much more reliable than VOC estimates.

Natural NO sources include stratospheric intrusion, oceans, lightning, soil, and x

wildfires. Lightning and soil emissions are the only two significant natural sources of NO in x

the United States. It is estimated that combined natural sources contribute about 2.2 Tg of

NO to the troposphere over the continental United States; however, uncertainties in natural x

NO emission inventories are much greater than those for anthropogenic NO emissions. x x

Concentrations of Volatile Organic Compounds in Ambient Air

The VOCs most frequently analyzed in ambient air are the nonmethane

hydrocarbons (NMHCs). Morning (6:00 to 9:00 a.m.) concentrations most often have been

measured because of the use of morning data in the Empirical Kinetic Modeling Approach

(EKMA) and in air quality simulation models.

Concurrent measurements of anthropogenic and biogenic NMHCs have shown that

biogenic NMHCs usually constituted much less than 10% of the total NMHCs. For example,

average isoprene concentrations ranged from 0.001 to 0.020 ppm carbon (C) and terpenes

from 0.001 to 0.030 ppm C. 

Concentrations of Nitrogen Oxides in Ambient Air

Measurements of NO made in 22 and 19 U.S. cities in 1984 and 1985, x

respectively, showed median 6:00-to-9:00 a.m. NO concentrations ranging from 0.02 to x

0.08 ppm in most of these cities. Nonurban NO concentrations, reported as average seasonal x

or annual NO , range from <0.005 to 0.015 ppm. x

Ratios of Concentrations of Nonmethane Organic Compounds to Nitrogen Oxides

Ratios of 6:00-to-9:00 a.m. nonmethane organic compounds (NMOC) to NO are x

higher in southeastern and southwestern U.S. cities than in northeastern and midwestern U.S.

cities, according to data from EPA's multicity studies conducted in 1984 and 1985. Rural

NMOC/NO ratios tend to be higher than urban ratios. The NMOC/NO ratios trended x x

downward to well below 10 in the South Coast Air Basin and in cities in the eastern United

States during the 1980s. Based on these low ratios, hydrocarbon control should be more

effective than NO control within a number of cities. Morning (6:00-to-9:00 a.m.) x

NMOC/NO ratios are used in the EKMA type of trajectory model. The correlation of x

NMOC/NO ratios with maximum 1-h O concentrations, however, was weak in a recent x 3

analysis.
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Source Apportionment and Reconciliation

Source apportionment (regarded as synonymous with receptor modeling) refers to

determining the quantitative contributions of various sources of VOCs to ambient air pollutant

concentrations. Source reconciliation refers to the comparison of measured ambient VOC

concentrations with emissions inventory estimates of VOC source emission rates for the

purpose of validating the inventories.

Recent findings have shown that vehicle exhaust was the dominant contributor to

ambient VOCs in seven of eight U.S. cities studied. Whole gasoline contributions were

estimated to be equal to vehicle exhaust in one study and to 20% of vehicle exhaust in a

second study.

Estimates of biogenic VOCs at a downtown site in Atlanta, GA, in 1990 indicated

a lower limit of 2% (24-h average) for the biogenic percentage of total ambient VOCs at that

location (isoprene was used as the biogenic indicator species). The percentage varies during

the 24-h period because of the diurnal (e.g., temperature, light intensity) dependence of

isoprene concentrations. 

Source reconciliation data have shown disparities between emission inventory

estimates and receptor-estimated contributions. For biogenics, emission estimates are greater

than receptor-estimated contributions. The reverse has been true for natural gas contributions

estimated for Los Angeles, CA; Columbus, OH; and Atlanta; and for refinery emissions in

Chicago, IL. 

Ozone Air Quality Models

Models and Their Components

Photochemical air quality models are used to predict how O concentrations 3

change in response to prescribed changes in source emissions of NO and VOCs. These x

models operate on sets of input data that characterize the emissions, topography, and

meteorology of a region and produce outputs that describe air quality in that region. 

Two kinds of photochemical models are recommended in guidelines issued by

EPA: (1) the use of EKMA is accepted under certain circumstances, and (2) the grid-based

Urban Airshed Model (UAM) is recommended for modeling O over urban areas. The 3

1990 Clean Air Act Amendments mandate the use of three-dimensional (grid-based) air

quality models such as UAM in developing state implementation plans for areas designated as

"extreme", "severe", "serious", or "multistate moderate". General descriptions of EKMA and

grid-based models were given in the 1986 EPA criteria document for O . 3

The EKMA-based method for determining O control strategies has limitations, the 3

most serious of which is that predicted emissions reductions are critically dependent on the

initial NMHC/NO ratio used in the calculations. This ratio cannot be determined with any x

certainty and is expected to be quite variable in time and space in an urban area. 

Spatial and temporal characteristics of VOC and NO emissions are major inputs to x

a grid-based photochemical air quality model. Greater accuracy in emissions inventories is

needed for biogenics and for both mobile and stationary source components. Grid-based air

quality models also require as input the three-dimensional wind field for the photochemical

episode being simulated. 

A chemical kinetic mechanism, representing the important chemical reactions that

occur in the atmosphere, is used in an air quality model to estimate the net rate of formation of

each pollutant simulated as a function of time. 
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Dry deposition is an important removal process for O on both urban and regional 3

scales and is included in all urban- and regional-scale models. Wet deposition is generally not

included in urban-scale photochemical models, because O episodes do not occur during 3

periods of significant clouds or rain.

Concentration fields of all species computed by the model must be specified at the

beginning of the simulation ("initial conditions"). These initial conditions are determined

mainly with ambient measurements, either from routinely collected data or from special

studies; but interpolation can be used to distribute the surface ambient measurements.

Use of Ozone Air Quality Models

Photochemical air quality models are used for control strategy evaluation by first

demonstrating that a past episode or episodes can be simulated adequately. The hydrocarbon

or NO emissions or both are reduced in the model inputs, and the effects of these reductions x

on O in the region are assessed. The adequacy of control strategies based on grid-based 3

models depends, in part, on the nature of input data for simulations and model validation, on

input emissions inventory data, and on the mismatch between the spacial output of the model

and the current form of the NAAQS for O . Uncertainties in models obviously can affect 3

their outputs. Uncertainties exist in all components of grid-based O air quality models: 3

emissions, meteorological modules, chemical mechanisms, deposition rates, and

determination of initial conditions.

Grid-based models that have been widely used to evaluate control strategies for

O or acid deposition, or both, are the UAM, the California Institute of Technology/Carnegie 3

Institute of Technology model, the Regional Oxidant Model, the Acid Deposition and Oxidant

Model, and the Regional Acid Deposition Model. The UAM (Version IV) is the grid model

approved nationwide for control strategy development at this time.

Despite the many uncertainties in photochemical air quality modeling, including

emission inventories, these models are essential for regulatory analysis and solving the

O problem. Grid-based O air quality modeling is superior to the available alternatives for 3 3

O control planning, but the chances of its incorrect use must be minimized. 3

Analytical Methods for Oxidants and Their Precursors

Oxidants

Current methods used to measure O are chemiluminescence (CL); UV absorption 3

spectrometry; and newly developed spectroscopic and chemical approaches, including

chemical approaches applied to passive sampling devices (PSDs) for O . 3

The CL method has been designated as the reference method by EPA. Detection

limits of 0.005 ppm and a response time of <30 s are typical of currently available

commercial instruments. A positive interference from atmospheric water vapor was reported

in the 1970s and recently has been confirmed. Proper calibration can minimize this source of

error.

Commercial UV photometers for measuring O have detection limits of about 3

0.005 ppm and a response time of <1 min. Because the measurement is absolute, UV

photometry is also used to calibrate O methods. A potential disadvantage of UV photometry 3

is that atmospheric constituents that absorb 254-nm radiation, the wavelength at which O is 3

measured, will cause a positive interference in O measurements. Interferences have been 3

reported in two recent studies, but assessment of the potential importance of such
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interferences (e.g., toluene, styrene, cresols, nitrocresols) is hindered by lack of absorption

spectra data in the 250-nm range and by lack of aerometric data for the potentially interfering

species. There also can be some interference from water, possibly from the condensation of

moisture in sampling lines.

Calibration of O3 measurement methods (other than PSDs) is done by UV

spectrometry or by gas-phase titration (GPT) of O with NO. Ultraviolet photometry is the 3

reference calibration method approved by EPA. Ozone is unstable and must be generated in

situ at time of use to produce calibration mixtures. 

Peroxyacetyl nitrate and the higher peroxyacyl nitrates normally are measured by

gas chromatography (GC) using an electron capture detector. Detection limits have been

extended to 1 to 5 ppt. The preparation of reliable calibration standards is difficult because

PAN is unstable, but several methods are available.

Volatile Organic Compounds

The method recommended by EPA for total NMOC measurement involves the

cryogenic preconcentration of NMOCs and the measurement of the revolatilized NMOCs

using flame ionization detection (FID). The primary technique for speciated NMOC/NMHC

measurements is cryogenic preconcentration followed by GC-FID. Systems for sampling and

analysis of VOCs have been developed that require no liquid cryogen for operation.

Stainless steel canisters have become the containers of choice for collection of

whole-air samples for NMHC/NMOC data. Calibration procedures for NMOC

instrumentation require the generation, by static or dynamic systems, of dilute mixtures at

concentrations expected to occur in ambient air. 

Preferred methods for measuring carbonyl species (aldehydes and ketones) in

ambient air are spectroscopic methods; on-line colorimetric methods; and, the most common

method currently in use for measuring gas-phase carbonyl compounds in ambient air, the

high-performance liquid chromatography method, which employs 2,4-dinitrophenylhydrazine

derivatization in a silica gel cartridge. Use of an O scrubber has been recommended to 3

prevent interference by O in this method in ambient air. 3

Oxides of Nitrogen

Nitric oxide and NO comprise the NO compounds involved as precursors to 2 x

O and other photochemical oxidants. 3

The most common method of NO measurement is the gas-phase CL reaction with

O , which is essentially specific for NO. Commercial NO monitors have detection limits of a 3

few parts per billion by volume (ppbv) in ambient air but may not have sensitivity sufficient

for surface measurements in rural or remote areas or for airborne measurements. Direct

spectroscopic methods for NO exist that have very high sensitivity and selectivity for NO, but

their complexity, size, and cost restrict these methods to research applications. No PSDs exist

for measurement of NO. 

Chemiluminescence analyzers are the tools of choice for NO measurement, even 2

though they do not measure NO directly. Minimum detection levels for NO have been 2 2

reported to be 5 to 13 ppb, but more recent evaluations have indicated detection limits of

0.5 to 1 ppbv. Reduction of NO to NO is required for measurement. These analyzers 2

actually measure NO (NO + PAN + HNO + other reactive nitrogen species); however, for y x 3

most urban atmospheres, NO is the predominant species measured diurnally. x
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Several spectroscopic approaches to NO detection have been developed but share 2

the drawbacks of spectroscopic NO methods. Passive samplers for NO exist but are still in 2

the developmental stage for ambient air monitoring.

Calibration of methods for NO measurement is done using standard cylinders of

NO in nitrogen. Calibration of methods for NO measurement include use of cylinders of 2

NO in nitrogen or air, use of permeation tubes, and GPT. 2

1.4 Environmental Concentrations, Patterns, and

Exposure Estimates

Ozone is measured at concentrations above the minimum detectable level at all

monitoring locations in the world. In this section, hourly average concentration and exposure

information is summarized for urban, rural forested, and rural agricultural areas in the United

States.

Because O from urban area emissions is transported to rural downwind locations, 3

elevated O concentrations can occur at considerable distances from urban centers. Urban 3

O concentration values are often depressed because of titration by NO. Because of the 3

absence of chemical scavenging, O tends to persist longer in nonurban areas than in urban 3

areas, and nonurban exposures may be higher than those in urban locations.

Trends

Ozone hourly average concentrations have been recorded for many years by the

state and local air pollution agencies who report their data to EPA. The 10-year (1983 to

1992) composite average trend for the second highest daily maximum hourly average

concentration during the O season shows that the 1992 composite average for the trend sites 3

was 21% lower than the 1983 average. The 1992 value was the lowest composite average of

the 10-year period and was significantly less than each of the previous nine years, 1983 to

1991. The relatively high O concentrations in 1983 and 1988 likely were attributable, in 3

part, to hot, dry, stagnant conditions in some areas of the country, which were especially

conducive to O formation. 3

From 1991 to 1992, the composite mean of the second highest daily maximum

1-h O concentrations decreased 7%, and the composite average of the number of estimated 3

exceedances of the O standard decreased by 23%. Nationwide VOC emissions decreased 3

3% from 1991 to 1992. The composite average of the second daily maximum concentrations

decreased in 8 of the 10 EPA regions from 1991 to 1992, and remained unchanged in

Region VII. Except for Region VII, the 1992 regional composite means were lower than the

corresponding 1990 levels. Although meteorological conditions in the east during 1993 were

more conducive to O than those in 1992, the composite mean level for 1993 was the second 3

lowest composite average of the decade, 1984 to 1993.

Surface Concentrations

Published data provide evidence showing the occurrence at some sites of multihour

periods within a day of O at levels of potential health effects. Although most of these 3

analyses were made using monitoring data collected from sites in or near nonattainment areas,

in one analysis of five sites (two in New York state, two in rural California, and one in rural
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Oklahoma), none of which was in or near a nonattainment area, O concentrations showed 3

only moderate peaks but showed multihour levels above 0.1 ppm.

A small amount of the O concentration measured at a monitoring site is produced 3

by sources distant to the photochemical reactions occurring on an urban or regional scale. 

Typical sources include stratospheric intrusions into the troposphere, photochemical

production by the CH /CO/NO cycle in the troposphere, and transport of very distant 4 x

anthropogenic or biogenic VOCs and NO . The specific concentrations of this "background" x

O vary with averaging times ranging from the daily 1-h maximum to daily, monthly, 3

seasonal, or annual values. The background concentrations also vary with geographical

region and with elevation of the monitoring site.

On the basis of O data from isolated monitoring sites, EPA has indicated that a 3

reasonable estimate of O background concentration near sea level in the United States is 3

from 0.020 to 0.035 ppm for an annual average, 0.025 to 0.045 ppm for an 8-h daily summer

seasonal average, and from 0.03 to 0.05 ppm for the average summertime 1-h daily

maximum. This estimate includes a 0.005 to 0.015 ppm O contribution from stratospheric 3

intrusions into the troposphere.

Diurnal Variations

Diurnal patterns of O may be expected to vary with location, depending on the 3

balance among the many factors affecting O formation, transport, and destruction. Although 3

they vary with locality, diurnal patterns of O3 typically show a rise in concentration from low

levels, or levels near minimum detectable amounts, to an early afternoon peak. The diurnal

pattern of concentrations can be ascribed to three simultaneous processes: (1) downward

transport of O from layers aloft, (2) destruction of O through contact with surfaces and 3 3

through reaction with NO at ground level, and (3) in situ photochemical production of O . 3

Seasonal Patterns

Seasonal variations in O concentrations in urban areas usually show the pattern of 3

high O in the late spring or in the summer and low levels in the winter; however, weather 3

conditions in a given year may be more favorable for the formation of O and other oxidants 3

than during the prior or following year. 

Average O concentrations tend to be higher in the second versus the third quarter 3

of the year for many isolated rural sites. This observation has been attributed to either

stratospheric intrusions or an increasing frequency of slow-moving, high-pressure systems

that promote the formation of O . However, for several clean rural sites, the highest 3

exposures have occurred in the third quarter rather than in the second. For rural O sites in 3

the southeastern United States, the daily maximum 1-h average concentration was found to

peak during the summer months. 

Spatial Variations

Concentrations of O vary with altitude and with latitude. There appears to be no 3

consistent conclusion concerning the relationship between O exposure and elevation. 3
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Indoor Ozone

Until the early 1970s, very little was known about the O concentrations 3

experienced inside buildings; to date, the database on this subject is not large, and a wide

range of indoor/outdoor O concentration relationships can be found in the literature (reported 3

indoor/outdoor values for O are highly variable). Indoor/outdoor O concentration ratios 3 3

generally fall in the range from 0.1 to 0.7 and indoor concentrations of O almost invariably 3

will be less than outdoors.

Estimating Exposure

 Both fixed-site monitoring information and human exposure models are used to

estimate risks associated with O exposure. Because, for most cases, it is not possible to 3

estimate population exposure solely from fixed-station data, several human exposure models

have been developed. These models also contain submodels depicting the sources and

concentrations likely to be found in each microenvironment, including indoor, outdoor, and

in-transit settings. Two distinct types of O exposure models exist: (1) those that focus 3

narrowly on predicting indoor O levels and (2) those that focus on predicting O exposures 3 3

on a community-wide basis. These latter models and their distinguishing features are:

1. pNEM/O based on the National Air Quality Standards Exposure Model 3

(NEM) series of models

• Uses mass-balance approach and seasonal considerations for I/O ratio

estimation.

• Variables affecting indoor exposure obtained by Monte Carlo sampling

from empirical distributions of measured data.

2. Systems Applications International (SAI)/NEM

• More districts and microenvironments and more detailed mass-balance

model than pNEM/O . 3

• Human activity data outdated and inflexible.

3. Regional Human Exposure Model (REHEX)

• More detailed geographic resolution than NEM.

• Uses California-specific activity data and emphasizes in-transit and outdoor

microenvironments.

4. Event probability exposure model (EPEM)

• Estimates probability that a randomly selected person will experience a

particular exposure regime.

• Lacks multiday continuity.

Few data are available for individuals using personal exposure monitors. Results

from a pilot study demonstrated that fixed-site ambient measurements may not adequately

represent individual exposures. Models based on time-weighted indoor and outdoor

concentrations explained only 40% of the variability in personal exposures. 

 

Peroxyacyl Nitrates

Peroxyacetyl nitrate and peroxypropionyl nitrate (PPN) are the most abundant of

the non-O oxidants in ambient air in the United States, other than the inorganic nitrogenous 3

oxidants such as NO , and possibly HNO . Most of the available data on concentrations of 2 3

PAN and PPN in ambient air are from urban areas. The levels to be found in nonurban areas
1-12

will be highly dependent on the transport of PAN and PPN or their precursors from urban

areas, because the concentrations of the NO precursors to these compounds are considerably x

lower in nonurban areas than in urban areas.

Co-occurrence

Studies of the joint occurrence of gaseous NO /O and SO /O at rural sites have 2 3 2 3

concluded that the periods of co-occurrence represent a small portion of the potential

plant-growing period. For human ambient exposure considerations, in most cases, the

simultaneous co-occurrence of NO /O and SO /O was infrequent. Some researchers have 2 3 2 3

reported the joint occurrence of O , nitrogen, and sulfur in forested areas, combining 3

cumulative exposures of O with data on dry deposition of sulfur and nitrogen. One study 3

reported that several forest landscapes with the highest dry deposition loadings of sulfur and

nitrogen tended to experience the highest average O concentrations and largest cumulative 3

exposure. Although the authors concluded that the joint concentrations of multiple pollutants

in forest landscapes were important, nothing was mentioned about the hourly co-occurrences

of O and SO or O and NO . Acid sulfates, which are usually composed of H SO , 3 2 3 2 2 4

ammonium bisulfate, and ammonium sulfate, have been measured at a number of locations in

North America. The potential for O and acidic sulfate aerosols to co-occur at some locations 3

in some form (i.e., simultaneously, sequentially, or complex-sequentially) is real and requires

further characterization. For human ambient exposures, the simultaneous co-occurrence of

NO and O was infrequent. 2 3

In one study, the relationship between O and hydrogen ions in precipitation was 3

explored using data from sites that monitored both O and wet deposition simultaneously and 3

within one minute latitude and longitude of each other. It was reported that individual sites

experienced years in which both hydrogen ion deposition and total O exposure were at least 3

moderately high. With data compiled from all sites, it was found that relatively acidic

precipitation occurred together with relatively high O levels approximately 20% of the time, 3

and highly acidic precipitation occurred together with a high O level approximately 6% of 3

the time. Sites most subject to relatively high levels of both hydrogen ions and O were 3

located in the eastern part of the United States, often in mountainous areas.

The co-occurrence of O and acidic cloudwater in high-elevation forests has been 3

characterized. The frequent O -only and pH-only single-pollutant episodes, as well as the 3

simultaneous and sequential co-occurrences of O and acidic cloudwater, have been reported. 3

Both simultaneous and sequential co-occurrences were observed a few times each month

above cloud base. 

1.5 Environmental Effects of Ozone and Related

Photochemical Oxidants

Ozone is the gaseous pollutant most injurious to agricultural crops, trees, and

native vegetation. Exposure of vegetation to O can inhibit photosynthesis, alter carbon 3

(carbohydrate) allocation, and interfere with mycorrhizal formation in tree roots. Disruption

of the important physiological processes of photosynthesis and carbon allocation can suppress

the growth of crops, trees, shrubs, and herbaceous vegetation by decreasing their capacity to

form the carbon (energy) compounds needed for growth and maintenance and their ability to
1-13

absorb the water and mineral nutrients that they require from the soil. In addition, loss of

vigor impairs the ability of trees and crops to reproduce and increases their susceptibility to

insects and pathogens. The following section summarizes key environmental effects

associated with O exposure. 3

Effects on Agroecosystems

Methodologies Used in Vegetation Research

Most of the knowledge concerning the effects of O on vegetation comes from the 3

exposure-response studies of important agricultural crop plants and some selected forest and

urban tree species, mostly as seedlings. A variety of methodologies have been used, ranging

from field exposures without chambers to open-top chambers and to exposures conducted in

chambers under highly controlled conditions. In general, the more controlled conditions are

most appropriate for investigating specific responses and for providing the scientific basis for

interpreting and extrapolating results. The greatest body of knowledge is from OTC studies.

Mode of Action

Leaves are important regulators of plant stress and function. Stress resulting from

exposure to O produces a leaf-mediated response. Effects expressed within cells in the leaf 3

(i.e., inhibition of photosynthesis) affect a plant's carbon (energy) budget. Plant processes are

impaired only by the O that enters the plant through the stomata (opening in the leaves). An 3

effect will occur only if sufficient O reaches sensitive sites within the leaf cells. The uptake 3

and movement of O to sensitive cellular sites within a leaf are subject to various biochemical 3

and physiological controls. Leaf injury will not be detected if the rate of uptake is small

enough for the plant to detoxify or metabolize O and its derivatives, or the plant is able to 3

repair or compensate for the impact at a rate equal to or greater than the rate of uptake. 

Impairment of leaf cellular processes is the basis for all other plant effects. The diurnal

pattern of stomatal opening plays a critical role in O uptake, particularly at the canopy level. 3

 Visible injury is usually the first observable indication of cellular response; injury

can occur, however, with no visible effects. Early senescence of leaves or needles is also a

result of cellular response. Impairment of cellular processes inhibits the rate of

photosynthesis, reduces carbon (sugars, carbohydrate) production, and alters carbon

allocation, causing a shift in growth pattern that favors shoots over roots. The reduced

allocation of carbon to leaf repair and new leaf formation limits the availability of carbon for

reproduction; stem and root growth; and, particularly, the formation of the mycorrhizae on

roots necessary for nutrient and water uptake. Reduction of plant vigor by O can result in 3

mortality, particularly when plant susceptibility to insects and pathogens is increased.

Factors That Modify Plant Response

Plant response to O exposure is influenced by a variety of biological, chemical, 3

and physical factors. When determining the impact of O exposure on plants, both the 3

influence of environmental factors on plant response and the effects of O on that response 3

must be considered. Biological factors within plants that affect their response to stresses

include, genetic composition, stage of development, and the diurnal pattern of stomatal

opening. Genotype significantly influences plant sensitivity to O . Individuals, varieties, and 3

cultivars of a species are known to differ greatly in their responses to a given O exposure. 3
1-14

Genotype also influences the ability of plants to compete with one another for space,

nutrients, light, and water.

The magnitude of response of a particular species, variety, or cultivar depends on a

number of environmental factors. The plant's present and past environmental milieu, which

includes the temporal exposure pattern and stage of development, dictates the plant response. 

The corollary is also true: exposure to O can modify plant response to other environmental 3

variables. Available light, temperature, atmospheric turbulence and moisture, in both the

atmosphere and soil; soil nutrition; and exposure to and interaction with other pollutants such

as agricultural chemical sprays also influence the magnitude of plant response. 

Drought can reduce visible injury and the adverse effects of O on growth and 3

yield of crops. However, in the case of crops, drought, per se, much more adversely affects

yield than the effects of O . Ozone, on the other hand, tends to reduce the water-use 3

efficiency of well-watered crops. In some plants, O exposure reduces cold/winter hardiness. 3

Although exposure to O tends to reduce attacks by obligate pathogens, susceptibility of 3

plants to facultative pests and pathogens increases. 

Effects-Based Air Quality Exposure Indices

Environmental scientists for many years have attempted to characterize and

mathematically represent plant exposures to O . A variety of averaging times have been used. 3

Although most studies have characterized exposure by using mean concentrations, such as

seasonal, monthly, weekly, daily, or peak hourly means, other studies have used cumulative

measures (e.g., the number of hours above selected concentrations). None of these statistics

completely characterizes the relationships among O concentration, exposure duration, 3

interval between exposures, and plant response. 

The use of a mean concentration with long averaging times implies that all

concentrations of O are equally effective in causing plant responses and minimizes the 3

contributions of the peak concentrations to the response. Ozone effects are cumulative;

therefore, exposure duration should be included in any index if it is to be biologically

relevant. Present evidence suggests that cumulative effects of episodic exposures to either

peak or mid-range concentrations, or both, can play an important role in producing growth

responses. The key to plant response is timing because peak and mid-range concentrations do

not occur at the same time. Potentially, the greatest effect of O on plants will occur when 3

stomatal conductance is greatest. When peaks occur at the time of greatest stomatal

conductance, the effect of mid-range concentrations will not be observable. Atmospheric

conductivity also strongly influences plant response because O must be in contact with the 3

leaf surface if it is to be taken up by a plant. Effects on vegetation appear when the amount of

pollutant entering exceeds the ability of the plant to repair or compensate for the impact. 

Increasing uptake of O will inhibit photosynthesis and result in increased reductions in 3

biomass production. 

An index of ambient exposures that relates well to plant response should

incorporate, directly or indirectly, environmental influences (e.g., temperature, humidity, soilmoisture status) and exposure dynamics. Peak indices (e.g., second highest daily maximum)

imply that a single high-concentration exposure (1- or 8-h concentration) during the course of

a 70- to 120-day growing season is related to eventual yield or growth reductions. On the

other hand, mean indices (e.g., 7-h seasonal mean) imply that duration of the exposure is not

important, and that all concentrations have equal effect on plants. Neither of these indices
1-15

relates ambient O concentrations to biological effects on plants because these indices do not 3

consider the duration of exposure. An index that cumulates all hourly concentration during

the season and gives greater weight to higher concentrations appears to be a more appropriate

index for relating ambient exposures to growth or yield effects.

No experimental studies have been designed specifically to evaluate the adequacy

of the various peak-weighted indices that have been proposed. In retrospective analyses in

which O is the primary source of variation in response, year-to-year variations in plant 3

response are minimized by peak-weighted, cumulative exposure indices. However, a number

of different forms of peak-weighted, cumulative indices have been examined for their ability

to properly order yield responses from the large number of studies of the National Crop Loss

Assessment Network (NCLAN) program. These exposure indices (i.e., SUM00, SUM06,

SIGMOID, W126) all performed equally well, and it is not possible to distinguish among

them on the basis of statistical fits of the data. The biological basis for these indices has not

been determined.

Exposure Response of Plant Species

The emphasis of experimental studies usually has been on the more economically

important crop plants and tree species, as seedlings. Crop species usually are monocultures

that are fertilized and, in many cases, watered. Therefore, because crop plants are usually

grown under optimal conditions, their sensitivity to O exposures can vary from that of native 3

trees, shrubs, and herbaceous vegetation. 

The concept of limiting values was used in both the 1978 and 1986 criteria

documents to summarize visible foliar injury. Limiting values are defined as concentrations

and durations of exposure below which visible injury does not occur. The limit for visible

injury indicating reduced plant performance was an O exposure of 0.05 ppm for several 3

hours per day for more than 16 days. When the exposure period was decreased to 10 days,

the O concentration required to cause injury was increased to 0.10 ppm. A short, 6-day 3

exposure further increased the concentration to 0.30 ppm. These exposure and concentration

periods apply for those crops where appearance or aesthetic value (e.g, spinach, cabbage,

lettuce) is considered important. Limiting values for foliar injury to trees and shrubs range

from 0.06 to 0.10 ppm for 4 h. 

The following assertions can be made based on information from the 1986 criteria

document, its 1992 supplement, and literature published since 1986. Ambient

O concentrations in several regions of the country are high enough to impair growth and 3

yield of sensitive plant species. This clearly is indicated by comparison of data obtained from

crop yield in charcoal-filtered and unfiltered (ambient) exposures. These elevated levels are

further supported by data from studies using chemical protectants. These response data make

possible the extrapolation to plants not studied experimentally. Both approaches mentioned

above indicate that effects occur with only a few exposures above 0.08 ppm. Data from

regression studies conducted to develop an exposure-response function for estimating yield

loss indicated that at least 50% of the species and cultivars tested could be predicted to exhibit

a 10% yield loss at 7-h seasonal mean O concentrations of 0.05 ppm or less. 3

Effects on Natural Ecosystems

The responses of the San Bernardino mixed forest of Southern California to 50 or

more years of chronic ozone exposures based on many studies, present a classic example of
1-16

ecosystem response to severe stress. Data from an inventory conducted from 1968 through

1972 indicated that for 5 mo of each year, trees were exposed to O concentrations greater 3

than 0.08 ppm for more than 1,300 h. Concentrations rarely decreased below 0.05 ppm at

night near the crest of the mountain slope, approximately 5,500 ft. In addition, during the

years 1973 to 1978, average 24-h O concentrations ranged from a background of 0.03 to 3

0.04 ppm in the eastern part of the San Bernardino Mountains to a maximum of 0.10 to

0.12 ppm in the western part during May through September.

Plants accumulate, store, and use the energy in carbon compounds (sugars)

produced during photosynthesis to build their structures and to maintain the physiological

processes necessary for life. The patterns of carbon allocation to roots, stems, and leaves

directly influence growth. The strategy for carbon allocation changes during the life of a

plant, as well as with environmental conditions. Mature trees have a higher ratio of

respiration to photosynthetic tissue. Impairment of photosynthesis shifts carbon allocation

from growth and maintenance to repair; increased respiration can result in resource

imbalances. The significant changes observed in the San Bernardino forest ecosystem were a

possible outcome of the combined influences of O on carbon, water, and nutrient allocation. 3

The biochemical changes within the leaves of ponderosa and Jeffrey pine in the

San Bernardino forest, expressed as visible foliar injury, premature needle senescence,

reduced photosynthesis, and reduced carbohydrate production and allocation, resulted in

reduced tree vigor, growth, and reproduction. Reduced vigor increases susceptibility of trees

to insect pests and fungal pathogens. Premature needle senescence alters microorganismal

succession on confer needles and changes the detritus-forming process and associated nutrient

cycling. 

Altered carbon allocation is important in the formation of mycorrhizae (fungus

roots), which are an extremely important but unheralded component of all ecosystems; the

majority of all plants depend on them because they are integral in the uptake of mineral

nutrients and water from the soil. Carbon-containing exudates from the roots are necessary

for the formation of mycorrhizae. Reduced carbon allocation to plant roots affects

mycorrhizal formation and impacts plant growth. Exposure to ozone, therefore, affects plant

growth both above and below ground. 

Small changes in photosynthesis or carbon allocation can alter profoundly the

structure of a forest. Ecosystem responses to stress begin with the response of the most

sensitive individuals of a population. Stresses, whose primary effects occur at the molecular

level (within the leaves), must be propagated progressively through more integrated levels of

organ physiology (e.g., leaf, branch, root) to whole plant physiology, then to populations

within the stand (community), and finally to the landscape level to produce ecosystem effects. 

Only a small fraction of stresses at the molecular level become disturbances at the tree, stand,

or landscape level. The time required for a stress to be propagated from one level to the next

(it can take years) determines how soon the effects of the stress can be observed or measured. 

The primary effect of O on ponderosa and Jeffrey pine, two of the more 3

susceptible members of the San Bernardino forest community, was that the trees were no

longer able to compete effectively for essential nutrients, water, light, and space. Decline in

the sensitive trees, a consequence of altered competitive conditions, permitted the enhanced

growth of more tolerant species. Removal of the ecosystem dominants at the population level

changed its structure and altered the processes of energy flow and nutrient cycling, returning

the ecosystem to a less complex stage. 
1-17

The San Bernardino Mountains continue to experience exposure to O ; however, 3

there has been a gradual decline in concentrations and length of exposure. Ozone

concentrations of 0.06 ppm or higher of varying durations capable of causing injury to trees in

forest ecosystems have been observed during the past 5 years in the Sierra Nevada Mountains

and the Appalachian Mountains from Georgia to Maine. Visible injury to forest trees and

other vegetation in these areas has been observed. 

Injury to sensitive trees from exposure to ozone concentrations 0.06 ppm or greater

in the Sierra Nevada Mountains and the Appalachian Mountains has never had the impact on

these ecosystems that it did on the San Bernardino forest. Forest stands differ greatly in age,

species composition, stability, and capacity to recover from disturbance. In addition, the

position in the stand or community of the most sensitive species is extremely important. 

Ponderosa and Jeffrey pine were the dominant species in the San Bernardino forest. Removal

of populations of these trees altered both ecosystem structure and function. Both the Sierra

Nevada Mountains and the Appalachian Mountains are biologically more diverse. Removal

of sensitive individuals of eastern white pine and black cherry has not visibly altered the

forest ecosystems along the Appalachian Mountains, possibly because of the absence of

population changes in these species. Decline and dieback of trees on Mt. Mitchell, NC, and

Camel's Hump, VT, cannot be related solely to O injury. 3

Effects on Agriculture, Forestry, and Ecosystems: Economics

A number of economic assessments of the effects of O on agriculture have been 3

performed over the last decade. All use NCLAN response data to predict crop yield changes. 

Although these studies employ somewhat different economic assessment methodologies, each

shows national-level economic losses to major crops in excess of $1 billion (1990 dollars)

from exposure to ambient concentrations of O . These studies also evaluate the sensitivity of 3

the economic estimates to uncertainties in data, including the NCLAN response data. The

economic assessment models used could be adapted to future O -crop yield response findings, 3

if available.

The plant science literature shows that O adversely influences physiological 3

performance of both urban and native tree species; the limited economic literature also

demonstrates that changes in growth have economic consequences. However, the natural

science and economic literature on the topic are not yet mature enough to conclude

unambiguously that ambient O is imposing economic costs. The economic effects of O on 3 3

ecosystems have not yet been addressed in the published literature. There is, however, an

emerging interest in applying economic concepts and methods to the management of

ecosystems.

Effects on Materials

Over four decades of research show that O damages certain materials such as 3

elastomers, textile fibers, and dyes. The amount of damage to actual in-use materials and the

economic consequences of that damage are poorly characterized.

Natural rubber and synthetic polymers of butadiene, isoprene, and styrene, used in

products like automobile tires and protective outdoor electrical coverings, account for most of

the elastomer production in the United States. The action of O on these compounds is well 3

known, and concentration-response relationships have been established and corroborated by

several studies. These relationships, however, must be correlated with adequate exposure
1-18

information based on product use. For these and other economically important materials,

protective measures have been formulated to reduce the rate of oxidative damage. When

antioxidants and other protective measures are incorporated in elastomer production, the

O -induced damage is reduced considerably, although the extent of reduction differs widely 3

according to the material and the type and amount of protective measures used.

Both the type of dye and the material in which it is incorporated are important

factors in the resistance of a fabric to O . Some dyed fabrics, such as royal blue and red 3

rayon-acetate and plum cotton are resistant to O . On the other hand, anthraquinone dyes on 3

nylon fibers are sensitive to fading by O . Field studies and laboratory work show a positive 3

association between O levels and dye fading of nylon materials. At present, the available 3

research is insufficient to quantify the amount of damaged materials attributable to O alone. 3

The degradation of fibers from exposure to O is poorly characterized. In general, 3

most synthetic fibers, such as modacrylic and polyester, are relatively resistant, whereas

cotton, nylon, and acrylic fibers have greater but varying sensitivities to O . Ozone reduces 3

the breaking strength of these fibers, and the degree of strength reduction depends on the

amount of moisture present. The limited research in this area indicates that O in ambient air 3

may have a minimal effect on textile fibers, but additional research is needed to verify this

conclusion.

A number of artists' pigments and dyes are sensitive to O and other oxidants; in 3

particular, many organic pigments are subject to fading or other color changes when exposed

to O . Although most, but not all, modern fine arts paints are more O resistant, many older 3 3

works of art are at risk of permanent damage due to O -induced fading. 3

A great deal of work remains to be done to develop quantitative estimates of the

economic damage to materials from photochemical oxidants. Most of the available studies

are outdated in terms of O concentrations, technologies, and supply-demand relationships. 3

Additionally, little is known about the physical damage functions, so cost estimates have been

simplified to the point of not properly recognizing many of the scientific complexities of the

impact of O . 3

1.6 Toxicological Effects of Ozone and Related

Photochemical Oxidants

Respiratory Tract Effects of Ozone

Biochemical Effects

Knowledge of molecular targets provides a basis for understanding mechanisms of

effects and strengthening animal-to-human extrapolations. Ozone reacts with polyunsaturated

fatty acids and sulfhydryl, amino, and some electron-rich compounds. These elements are

shared across species. Several types of reactions are involved, and free radicals may be

created. Based on this knowledge, it has been hypothesized that the O molecule is unlikely 3

to penetrate the liquid linings of the respiratory tract (RT) to reach the tissue, raising the

possibility that reaction products exert effects.

In acute and short-term exposure studies, a variety of lung lipid changes occur,

including an increase in arachidonic acid, the further metabolism of which produces a variety

of biologically active mediators that can affect host defenses, lung function, the immune

system, and other functions.
1-19

The level of lung antioxidant metabolism increases after O exposure, probably as 3

a result of the increase in the number of Type 2 cells, which are rich in antioxidant enzymes.

Collagen (the structural protein involved in fibrosis) increases in O -exposed lungs 3

in a manner that has been correlated to structural changes (e.g., increased thickness of the

tissue between the air and blood after prolonged exposure). Some studies found that the

increased collagen persists after exposure ceases.

Generally, O enhances lung xenobiotic metabolism after both short- and long- 3

term exposure, possibly as a result of morphological changes (increased numbers of

nonciliated bronchiolar epithelial cells). The impact of this change is dependent on the

xenobiotics involved; for example, the metabolism of benzo[a]pyrene to active metabolites

was enhanced by O . 3

Lung Inflammation and Permeability Changes

Elevated concentrations of O disrupt the barrier function of the lung, resulting in 3

the entry of compounds from the airspaces into the blood and the entry of serum components

(e.g., protein) and white blood cells (especially polymorphonuclear leukocytes [PMNs]) into

the airspaces and lung tissue. This latter impact reflects the initial stage of inflammation. 

These cells can release biologically active mediators that are capable of a number of actions,

including damage to other cells in the lung. In lung tissue, this inflammation also can

increase the thickness of the air-blood barrier. 

Increases in permeability and inflammation have been observed at levels as low as

0.1 ppm O (2 h/day, 6 days; rabbits). After acute exposures, the influence of the time of 3

exposure (from two to several hours) increases as the concentration of O increases. 3

Long-term exposure effects are discussed under lung morphology.

The impacts of these changes are not fully understood. At higher

O concentrations (e.g., 0.7 ppm, 28 days), the diffusion of oxygen into the blood decreases, 3

possibly because the air-blood barrier is thicker; cellular death may result from the enzymes

released by the inflammatory cells; and host defense functions may be altered by mediators.

Effects on Host Defense Mechanisms

Exposure to elevated concentrations of ozone results in alterations of all defense

mechanisms of the RT, including mucociliary and alveolobronchiolar clearance, functional

and biochemical activity of the alveolar macrophage (AM), and immunologic competence. 

These effects can cause susceptibility to bacterial respiratory infections.

Mucociliary clearance, which removes particles and cellular debris from the

conducting airways, is slowed by acute, but not repeated exposures to O . Ciliated epithelial 3

cells that move the mucous blanket are altered or destroyed by acute and chronic exposures. 

Neonatal sheep exposed to O do not have normal development of the mucociliary system. 3

Such effects could prolong the retention of unwanted substances (e.g., inhaled particles) in the

lungs, allowing them to exert their toxicity for a longer period of time. 

Alveolar clearance mechanisms, which center on the functioning of AMs, are

altered by O . Short-term exposure to levels as low as 0.1 ppm O (2 h/day, 1 to 4 days; 3 3

rabbits) accelerates clearance, but longer exposures do not. Even so, after a 6-week exposure

of rats to an urban pattern of O , the retention of asbestos fibers in a region protected by 3

alveolar clearance is prolonged.
1-20

Alveolar macrophages engulf and kill microbes, as well as clear the deeper regions

of the lungs of nonviable particles; AMs also participate in immunological responses, but little

is known about the effects of O3 on this function. Acute exposures of rabbits to levels as low

as 0.1 ppm O decrease the ability of AMs to ingest particles. This effect is displayed in 3

decreases in the ability of the lung to kill bacteria after acute exposure of mice to levels as low

as 0.4 ppm O . 3

Both the pulmonary and systemic immune system are affected by O , but in a 3

poorly understood way. It appears that the part of the immune system dependent on T-cell

function is more affected than is the part dependent on B-cell function.

Dysfunction of host defense systems results in enhanced susceptibility to bacterial

lung infections. For example, acute exposure to O concentrations as low as 0.08 ppm for 3

3 h can overcome the ability of mice to resist infection with streptococcal bacteria, resulting in

mortality. However, more prolonged exposures (weeks, months) do not cause greater effects

on infectivity.

Effects on antiviral defenses are more complex and less well understood. Only

high concentrations (1.0 ppm O , 3 h/day, 5 days; mice) increase viral-induced mortality. 3

Apparently, O does not impact antiviral clearance mechanisms. Although O does not affect 3 3

acute lung injury from influenza virus infection, it does enhance later phases of the course of

an infection (i.e., postinfluenzal alveolitis). 

Morphological Effects

Elevated concentrations of O cause similar types of alterations in lung structure in 3

all laboratory animal species studied, from rats to monkeys. In the lungs, the most affected

cells are the ciliated epithelial cells of the airways and Type 1 epithelial cells of the

gas-exchange region. In the nasal cavity, ciliated cells are also affected.

The centriacinar region (CAR; the junction of the conducting airways and

gas-exchange regions) is the primary target, possibly because this area receives the greatest

dose of O . The ciliated cells can be killed and replaced by nonciliated cells (i.e., cells not 3

capable of clearance functions that also have increased ability to metabolize some foreign

compounds). Mucous-secreting cells are affected, but to a lesser degree. Type 1 cells, across

which gas exchange occurs, can be killed; they are replaced by Type 2 cells, which are thicker

and produce more lipids. An inflammatory response also occurs in the tissue. The tissue is

thickened further in later stages when collagen (a structural protein increased in fibrosis) and

other elements accumulate. Although fibrotic changes have been observed in the CAR, they

have not been distributed throughout the whole lung.

The distal airway is remodeled; more specifically, bronchiolar epithelium replaces

the cells present in alveolar ducts. Concurrent inflammation may play a role. This effect has

been observed at 0.25 ppm O (8 h/day, 18 mo) in monkeys; at a higher concentration, this 3

remodeling persists after exposure stops.

The progression of effects during and after a chronic exposure is complex. Over

the first few days of exposure, inflammation peaks and then drops considerably, plateauing

for the remainder of exposure, after which it largely disappears. Epithelial hyperplasia

increases rapidly over the first few days and rises slowly or plateaus thereafter; when

exposure ends, it begins to return toward normal. In contrast, fibrotic changes in the tissue

between the air and blood increase very slowly over months of exposure, and, after exposure

ceases, the changes sometimes persist or increase.
1-21

The pattern of exposure can make a major difference in effects. Monkeys exposed

to 0.25 ppm O (8 h/day) every other month of an 18-mo period had equivalent changes in 3

lung structure, more fibrotic changes, and more of certain types of pulmonary function

changes than did monkeys exposed every day over the 18 mo. From this work and rat

studies, it appears that natural seasonal patterns may be of more concern than more

continuous exposures. Thus, long-term animal studies with uninterrupted exposures may

underestimate some of the effects of O . 3

The morphologic lesions described in early publications on laboratory animals

exposed to O do not meet the current criteria for emphysema of the type seen in human 3

lungs.

Effects on Pulmonary Function

Pulmonary function changes in animals resemble those observed in humans after

acute exposure.

During acute exposure, the most commonly observed alterations are increased

frequency of breathing and decreased tidal volume (i.e., rapid, shallow breathing). This has

been reported at exposures as low as 0.2 ppm O for 3 h (rats). Typically, higher 3

concentrations (around 1 ppm) are required to affect breathing mechanics (compliance and

resistance). Extended characterizations of pulmonary function show types of changes

generally seen in humans. For example, there are decreased lung volumes at levels 0.5 ppm

O (a few hours; rats). 3

When rats are exposed to O for 2 h/day for 5 days, the pattern of attenuation of 3

pulmonary function responses is similar to that observed in humans. Other biochemical

indicators of lung injury did not return to control values by Day 5, and morphological changes

increased in severity over the period of exposure. Thus, attenuation did not result in

protection against all the effects of O . 3

Long-term exposures have provided mixed results on pulmonary function,

including no or minimal effects, restrictive effects, and obstructive effects. When changes

occurred and postexposure examinations were performed, pulmonary function recovered.

Genotoxicity and Carcinogenicity of Ozone

The chemical reactivities of O give it the potential to be a genotoxic agent. 3

In vitro studies are difficult to interpret because the culture systems used allowed

the potential formation of artifacts, and high or very high concentrations of O often were 3

used. Generally, in these studies, O causes DNA strand breaks, sometimes is weakly 3

mutagenic, and causes cellular transformation and chromosomal breakage. The latter finding

has been investigated in vivo, with mixed results in animals.

The few earlier long-term carcinogenic studies in laboratory animals, with or

without coexposure to known carcinogens, are either negative or ambiguous.

The National Toxicology Program (NTP) completed chronic rat and mouse cancer

bioassays using commonly accepted experimental approaches and designs. Both male and

female rats and mice were studied. Animals were exposed for 2 years (6 h/day, 5 days/week)

to 0.12, 0.5, and 1.0 ppm O or for a lifetime to the same levels (except 0.12 ppm). Following 3

their standard procedures for determination of weight-of-evidence for carcinogenicity, the

NTP reported "no evidence" in rats, "equivocal evidence" in male mice, and "some evidence"

in female mice. The increases in adenomas and carcinomas were observed only in the lungs. 
1-22

There was no concentration response. One of the reasons for the designation of "some

evidence" in female mice was that when the 2-year and lifetime exposure studies were

combined, there was a statistically significant increase in total tumors at 1.0 ppm. Lung

tumors from control and O -exposed mice also were examined for the presence of mutated 3

Ha-ras oncogenes. Although the types of mutations found were similar in both groups, a

higher incidence of mutations was found in lung tumors from the O -exposed mice. At the 3

present time, however, there is inadequate information to provide mechanistic support for the

finding in mice. Thus, the potential for animal carcinogenicity is uncertain.

In a companion NTP study, male rats were treated with a tobacco carcinogen and

exposed for 2 years to 0.5 ppm O . Ozone did not affect the response and therefore had no 3

tumor promoting activity.

Systemic Effects of Ozone

Ozone causes a variety of effects on tissues and organs distant from the lung. 

Because O itself is not thought to penetrate the lung, these systemic effects are either 3

secondary to lung alterations or result from reaction products of O . Effects have been 3

observed on clinical chemistry, white blood cells, red blood cells, the circulatory system, the

liver, endocrine organs, and the central nervous system. Most of these effects cannot be

interpreted adequately at this time and have not been investigated in humans, but it is of

interest to note that O exposures causing effects on the RT of animals cause a wide array of 3

effects on other organs also. 

Several behavior changes occur in response to O3. For example, 0.12 ppm O 3

(6 h, rats) decreases wheel-running activity, and 0.5 ppm (1 min) causes mice to avoid

exposure. These effects are not fully understood, but they may be related to lung irritation or

decreased ability to exercise.

Although cardiovascular effects, such as slowed heart rate and decreased blood

pressure, occur in O -exposed rats, some observed interactions with thermoregulation prevent 3

qualitative extrapolation of these effects to humans at this time.

Developmental toxicity studies in pregnant rats summarized in the 1986 O criteria 3

document showed that levels up to about 2.0 ppm O did not cause birth defects. Rat pups 3

from females exposed to 1.0 ppm O during certain periods of gestation weighed less or had 3

delays in development of behaviors (e.g., righting, eye opening). No "classical" reproductive

assays with O were found. 3

Other studies have indicated that O can affect some endocrine organs 3

(i.e., pituitary-thyroid-adrenal axis, parathyroid gland). It appears that the liver has less ability

to detoxify drugs after O exposure, but assays of liver enzymes involved in xenobiotic 3

metabolism are inconsistent.

Interactions of Ozone with Other Co-occurring Pollutants

Animal studies of the effects of O3 in combination with other air pollutants show

that antagonism, additivity, and synergism can result, depending on the animal species,

exposure regimen, and health endpoint. Thus, these studies clearly demonstrate the major

complexities and potential importance of interactions but do not provide a scientific basis for

predicting the results of interactions under untested ambient exposure scenarios.
1-23

1.7 Human Health Effects of Ozone and Related

Photochemical Oxidants

This section summarizes key effects associated with exposure to O , the major 3

component of photochemical oxidant air pollution that is clearly of most concern to the health

of the human population. Another, often co-occurring photochemical oxidant component of

"smog" is PAN, but this compound has been demonstrated to be primarily responsible for

induction of smog-related eye irritation (stinging of eyes). Limited pulmonary function

studies have shown no effects of PAN at concentrations below 0.13 to 0.30 ppm, which are

much higher than the generally encountered ambient air levels in most cities.

Controlled Human Studies of Acute Ozone Effects

Effects on Lung Function

Controlled studies in healthy adult subjects have demonstrated O -induced 3

decrements in pulmonary function, characterized by alterations in lung volumes and flow and

airway resistance and responsiveness. Respiratory symptoms, such as cough and pain on

deep inspiration, are associated with these changes in lung function.

Ozone-induced decreases in lung volume, specifically forced vital capacity (FVC)

and forced expiratory volume in 1 s (FEV ), largely can be attributed to decreases in 1

inspiratory capacity (the ability to take a deep breath), although at higher exposure

concentrations, there is clearly an additional component that is not volume dependent. Lung

volumes recover to a large extent within 2 to 6 h; normal baseline function typically is

reestablished within 24 h, but not fully with more severe exposures.

Ozone causes increased airway resistance and may cause reductions in expiratory

flow and the FEV /FVC ratio. 1

Ozone causes an increase in airway responsiveness to nonallergenic stimuli

(e.g., histamine, methacholine) in healthy and asthmatic subjects. There is no clear evidence

of a relationship between O -induced lung volume changes and changes in airway 3

responsiveness.

Inflammation and Host Defense Effects

Controlled studies in healthy adult subjects also indicate that O causes an 3

inflammatory response in the lungs characterized by elevated levels of PMNs, increased

epithelial permeability, and elevated levels of biologically active substances

(e.g., prostaglandins, proinflammatory mediators, cytokines).

Inflammatory responses to O can be detected within 1 h after a single 3

1-h exposure with exercise to concentrations 0.3 ppm; the increased levels of some

inflammatory cells and mediators persist for at least 18 h. The temporal response profile is

not defined adequately, although it is clear that the time course of response varies for different

mediators and cells.

Lung function and respiratory symptom responses to O do not seem to be 3

correlated with airway inflammation.

Ozone also causes inflammatory responses in the nose, marked by increased

numbers of PMNs and protein levels suggestive of increased permeability.
V

V

V

V

1-24

Alveolar macrophages removed from the lungs of human subjects after 6.6 h of

exposure to 0.08 and 0.10 ppm O have a decreased ability to ingest microorganisms, 3

indicating some impairment of host defense capability.

Ozone Exposure-Response Relationships

Functional, symptomatic, and inflammatory responses to O increase with 3

increasing exposure dose of O . The major determinants of the exposure dose are 3

O concentration (C), exposure duration (T), and the amount of ventilation ( ). 3 E

Exercise increases response to O by increasing (greater mass delivered), tidal 3 E

volume, inspiratory flow (greater percentage delivery), and the intrapulmonary

O concentration. 3

Repeated daily exposures to relatively high levels of O doses (C × T × ) 3 E

causing substantial reductions in FEV ( 20% decrement) typically cause exacerbation of the 1

lung function and respiratory symptom responses on the second exposure day. However,

attenuation of these responses occurs with continued exposures for a few days. Most

inflammatory responses also attenuate; for example, the PMN influx is absent after five

consecutive exposures.

Multihour exposures (e.g., for up to 7 h) to O3 concentrations as low as 0.08 ppm

cause small but statistically significant decrements in lung function, increases in respiratory

symptoms, and increases in PMNs and protein levels. Ozone C is a more important factor

than exercise or T in predicting responses to multihour low-level O exposure. There is E 3

clear evidence of a response plateau in terms of lung volume response to prolonged

O exposure. This evidence suggests that for a given combination of exercise and 3

O concentration (i.e., dose rate), there is a response plateau; continued exposure 3

(i.e, increased T) at that dose rate will not increase response. Therefore, quantitative

extrapolation of responses to longer exposure durations is not valid.

Mechanisms of Acute Pulmonary Responses

The mechanisms leading to the observed pulmonary responses induced by O are 3

beginning to be better understood. The available descriptive data suggest a number of

mechanisms leading to the alterations in lung function and respiratory symptoms, including

O delivery to the tissue (i.e., the inhaled concentration, breathing pattern, airway geometry; 3

O reactions with the airway lining fluid and epithelial cell membranes; local tissue responses, 3

including injury and inflammation; and stimulation of neural afferents (bronchial C-fibers)

and the resulting reflex responses and symptoms. The cyclooxygenase inhibitors block

production of prostaglandin E and interleukin-6 as well as reduce lung volume responses; 2

however, these drugs do not reduce inflammation and levels of cell damage markers such as

lactate dehydrogenase. 

Effects on Exercise Performance

Maximal oxygen uptake, a measure of peak exercise performance capacity, is

reduced in healthy young adults if preceded by O exposures sufficient to cause marked 3

changes in lung function (i.e., decreases of at least 20%) and increased subjective symptoms

of respiratory discomfort. Limitations in exercise performance may be related to increased

symptoms, especially those related to breathing discomfort.
V

1-25

Factors Modifying Responsiveness to Ozone

Many variables have the potential for influencing responsiveness to O ; however, 3

most are addressed inadequately in the available clinical data to make definitive conclusions.

Active smokers are less responsive to O exposure, which may reverse following 3

smoking cessation, but these results should be interpreted with caution. 

The possibility of age-related differences in response to O has been explored, 3

although young adults historically have provided the subject population for controlled human

studies. Children and adolescents have lung volume responses to O similar to those of young 3

adults, but lack respiratory symptoms. Pulmonary function responsiveness in adults appears

to decrease with age, whereas symptom rates remain similar to young adults. Group mean

lung function responses of adults over 50 years of age are less than those of children,

adolescents, and young adults.

The available data have not demonstrated conclusively that men and women

respond differently to O . Likewise, pulmonary function responses of women have been 3

compared during different phases of the menstrual cycle, but the results are conflicting. 

If gender differences exist for lung function responsiveness to O , they are not based on 3

hormonal changes, differences in lung volume, or the ratio of FVC to . E

There is no compelling evidence, to date, suggesting that any ethnic or racial

groups have a different distribution of responsiveness to O . 3

Seasonal and ambient factors may vary responsiveness to O , but further research 3

is needed to determine how they affect individual subjects. Individual sensitivity to O may 3

vary throughout the year, related to seasonal variations in ambient O concentrations. 3

The specific inhalation route appears to be of minor importance in exercising

adults. Exposure to O by oral breathing (i.e., mouthpiece) yields results similar to exposure 3

by oronasal breathing (i.e., chamber exposures).

Population Groups at Risk from Ozone Exposure

Population groups that have demonstrated increased responsiveness to ambient

concentrations of O consist of exercising healthy and asthmatic individuals, including 3

children, adolescents, and adults.

Available evidence from controlled human studies on subjects with preexisting

disease suggests that mild asthmatics have similar lung volume responses, but greater airway

resistance responses to O than nonasthmatics; and that moderate asthmatics may have, in 3

addition, greater lung volume responses than nonasthmatics.

Of all the other population groups studied, those with preexisting limitations in

pulmonary function and exercise capacity (e.g., chronic obstructive pulmonary disease,

chronic bronchitis, ischemic heart disease) would be of primary concern in evaluating the

health effects of O . Unfortunately, limitations of subject selection, standardized methods of 3

subject characterization, and range of exposure hamper the ability to make definitive

conclusions regarding the relative responsiveness of most chronic disease subjects.

Effects of Ozone Mixed with Other Pollutants

No significant enhancement of respiratory effects has been demonstrated

consistently for simultaneous exposures of O mixed with SO , NO , H SO , HNO , 3 2 2 2 4 3

particulate aerosols, or combinations of these pollutants. It is fairly well established that

simultaneous exposure of healthy adults and asthmatics to mixtures of O and other pollutants 3
1-26

for short periods of time (<2 h) induces pulmonary function responses not significantly

different from those following O alone when studies are conducted at the same 3

O concentration. Exposure to PAN has been reported to induce greater pulmonary function 3

responses than exposure to O alone, but at PAN concentrations (>0.27 ppm) much higher 3

than ambient levels. Unfortunately, only a limited number of pollutant combinations and

exposure protocols have been investigated, and subject groups are small and are

representative of only small portions of the general population. Thus, much is unknown

about the relationships between O and the complex mix of pollutants found in the ambient 3

air.

Prior exposure to O in asthmatics may cause an increase in response to other 3

pollutant gases, especially SO . Likewise, prior exposure to other pollutants can enhance 2

responses to O exposure. 3

Controlled Human Studies of Ambient Air Exposures

Mobile laboratory studies of lung function and respiratory symptoms in a local

subject population exposed to ambient photochemical oxidant pollution provide quantitative

information on exposure-response relationships for O3. A series of these studies from

Los Angeles has demonstrated pulmonary function decrements at mean ambient

O concentrations of 0.14 ppm in exercising healthy adolescents and increased respiratory 3

symptoms and pulmonary function decrements at 0.15 ppm in heavily exercising athletes and

at 0.17 ppm in lightly exercising healthy and asthmatic subjects. Comparison of the observed

effects in exercising athletes with controlled chamber studies at comparable O concentrations 3

showed no significant differences in lung function and symptoms, suggesting that coexisting

ambient pollutants have a minimal contribution to the measured responses under typical

summer ambient conditions in Southern California.

Field and Epidemiology Studies of Ambient Air Exposures

Individual-level field studies and aggregate-level time-series studies have

addressed the acute effects of O on lung function decrements and increased morbidity and 3

mortality in human populations exposed to real-world conditions of O exposure. 3

Camp and exercise studies of lung function provide quantitative information on

exposure-response relationships linking lung function declines with O exposure occurring in 3

ambient air. Combined statistical analysis of six recent camp studies in children yields an

average relationship between decrements in FEV and previous-hour O concentration of 1 3

0.50 mL/ppb. Two key studies of lung function measurements before and after well-defined

outdoor exercise events in adults have yielded exposure-response slopes of 0.40 and

1.35 mL/ppb. The magnitude of pulmonary function declines with O exposure is consistent 3

with the results of controlled human studies.

Daily life studies support a consistent relationship between O exposure and acute 3

respiratory morbidity in the population. Respiratory symptoms (or exacerbation of asthma)

and decrements in peak expiratory flow rate are associated with increasing ambient O , 3

particularly in asthmatic children; however, concurrent temperature, particles, acidity

(hydrogen ions), aeroallergens, and asthma severity or medication status also may contribute

as independent or modifying factors. Aggregate results show greater responses in asthmatic

individuals than in nonasthmatics, indicating that asthmatics constitute a sensitive group in

epidemiologic studies of oxidant air pollution.
1-27

Summertime daily hospital admissions for respiratory causes in various locations

of eastern North America have consistently shown a relationship with ambient levels of O , 3

accounting for approximately one to three excess respiratory hospital admissions per hundred

parts per billion O per million persons. This association has been shown to remain even after 3

statistically controlling for the possible confounding effects of temperature and copollutants

(e.g., hydrogen ions, sulfate, and particles less than 10 m), as well as when considering only

concentrations below 0.12 ppm O . 3

Many of the time-series epidemiology studies looking for associations between

O exposure and daily human mortality have been difficult to interpret because of 3

methodological or statistical weaknesses, including the a failure to account for other pollutant

and environmental effects. One of the two most useful new studies on O -mortality found a 3

small but statistically significant association in Los Angeles when peak 1-h maximum

O concentrations reached concentrations greater than 0.2 ppm during the study period. 3

A second study in regions with lower ( 0.15 ppm) maximum 1-h O concentrations 3

(St. Louis, MO, and Kingston-Harriman, TN) did not detect a significant O association with 3

mortality.

Only suggestive epidemiologic evidence exists for health effects of chronic

ambient O exposure in the population. All of the available studies of chronic respiratory 3

system effects in exposed children and adults are limited by a simplistic assignment of

exposure or by their inability to isolate potential effects related to O from those of other 3

pollutants, especially particles.

1.8 Extrapolation of Animal Toxicological Data to Humans

There have been significant advances in O dosimetry since 1986 that better enable 3

quantitative extrapolation with marked reductions in uncertainty. Experiments and models

describing the uptake efficiency and delivered dose of O in the RT of animals and humans 3

are beginning to present a clearer picture than has existed previously. 

The total RT uptake efficiency of rats at rest is approximately 50%. Within the RT

of the rat, 50% of the O taken up by the RT is removed in the head, 7% in the larynx/trachea, 3

and 43% in the lungs. 

In humans at rest, the total RT uptake efficiency is between 80 and 95%. Total RT

uptake efficiency falls as flow increases. As tidal volume increases, uptake efficiency

increases and flow dependence lessens. Pulmonary function response data and O uptake 3

efficiency data in humans generally indicate that the mode of breathing (oral versus nasal

versus oronasal) has little effect on upper RT or on total RT uptake efficiency, although one

study suggests that the nose has a higher uptake efficiency than the mouth. 

When all of the animal and human in vivo O uptake efficiency data are compared, 3

there is a good degree of consistency across data sets. This agreement raises the level of

confidence with which these data sets can be used to support dosimetric model formulations. 

Several mathematical dosimetry models have been developed since 1986. 

Generally, the models predict that net O dose to lung lining fluid plus tissue gradually 3

decreases distally from the trachea toward the end of the tracheobronchial region and then

rapidly decreases in the pulmonary region. 
1-28

When the dose of O to lung tissue is computed theoretically, it is found to be very 3

low in the trachea; to increase to a maximum in the terminal bronchioles of the first

generation pulmonary region; and then to decrease rapidly, moving further into the pulmonary

region. The increased tidal volume and flow, associated with exercise in humans, shifts

O dose further into the periphery of the lung and causes a disproportionate increase in distal 3

lung dose.

Predictions of delivered dose have been used to investigate both acute and chronic

O responses in the context of intra- and interspecies comparisons. In the case of intraspecies 3

comparisons, for example, the distribution of predicted O tissue dose to a ventilatory unit in a 3

rat as a function of distance from the bronchoalveolar duct junction is very consistent with the

distribution of alveolar wall thickening. In the case of interspecies comparisons (using the

delivered O dose to the proximal alveolar regions), although the functional responses 3

(e.g., rapid, shallow breathing) differ markedly between rats and humans, there is similarity of

acute dose-response patterns in inflammation (influx of cells and protein) among species, with

humans and guinea pigs more responsive than rats and rabbits, and similarity of chronic

dose-response patterns for increased alveolar interstitial thickness in the CAR of the lung,

with monkeys being more responsive than humans and rats less responsive. In other words,

the quantitative relationship between animal and human responses is dependent on the animal

species and the endpoint.

In summary, there is an emerging consistency among a variety of O dosimetry 3

data sets and between the experimental data and theoretical predictions of O dose. The 3

convergence of experimental data with theoretical predictions lends a degree of confidence to

the use of theoretical models to predict total and regional O dose. The use of O dosimetry 3 3

data and models is beginning to provide a useful extrapolation of effects between animals and

humans. The data and models have thus far helped demonstrate that humans may be more

responsive to O than rats, but less responsive than monkeys with respect to acute and chronic 3

inflammatory responses. However, the monkey, with its similarity to the human in distal

airway structure, provides chronic effects data that may best reflect the degree to which a

comparably exposed human would respond. These findings, therefore, suggest that long-term

exposure to O could impart a chronic effect in humans. 3

1.9 Integrative Summary of Ozone Health Effects

This section summarizes the primary conclusions derived from an integration of

the known effects of O provided by animal toxicological, human clinical, and 3

epidemiological studies.

1. What are the effects of short-term (<8-h) exposures to ozone?

Recent epidemiology studies addressing the effects of short-term ambient exposure

to O in the population have yielded significant associations with a wide range of health 3

outcomes, including lung function decrements, aggravation of preexisting respiratory disease,

increases in daily hospital admissions and emergency department visits for respiratory causes,

and increased mortality. Results from lung function epidemiology studies generally are

consistent with the experimental studies in laboratory animals and humans. 
1-29

Short-term O exposure of laboratory animals and humans causes changes in 3

pulmonary function, including tachypnea (rapid, shallow breathing), decreased lung volumes

and flows, and increased airway responsiveness to nonspecific stimuli. Increased airway

resistance occurs in both humans and laboratory animals, but typically at higher exposure

levels than other functional endpoints. In addition, adult human subjects experience

O -induced symptoms of airway irritation such as cough or pain on deep inspiration. The 3

changes in pulmonary function and respiratory symptoms occur as a function of exposure

concentration, duration, and level of exercise. Adult human subjects with mild asthma have

responses in lung volume and airway responsiveness to bronchoconstrictor drugs that are

qualitatively similar to those of nonasthmatics. Respiratory symptoms are also similar, but

wheezing is a prevalent symptom in O -exposed asthmatics in addition to the other 3

demonstrated symptoms of airway irritation. Airway resistance, however, increases relatively

more in asthmatics from an already higher baseline. Recovery from the effects of O on 3

pulmonary function and symptoms is usually complete within 24 h of the end of exposure,

although other responses may persist somewhat longer.

• An association between daily mortality and O concentration for areas with 3

high O levels (e.g., Los Angeles) has been suggested, although the magnitude 3

of such an effect is unclear.

• Increased O levels are associated with increased hospital admissions and 3

emergency department visits for respiratory causes. Analyses from data in the

northeastern United States suggest that O air pollution is associated with a 3

substantial portion (on the order of 10 to 20%) of all summertime respiratory

hospital visits and admissions.

• Pulmonary function in children at summer camps in southern Ontario, Canada,

in the northeastern United States, and in Southern California is associated with

O concentration. Meta-analysis indicates that a 0.5-mL decrease in FEV is 3 1

associated with a 1-ppb increase in O concentration. For preadolescent 3

children exposed to 120 ppb (0.12 ppm) ambient O , this amounts to an 3

average decrement of 2.4 to 3.0% in FEV . Similar responses are reported for 1

children and adolescents exposed to O in ambient air or O in purified air for 3 3

1 to 2 h while exercising.

• Pulmonary function decrements generally are observed in healthy subjects (8 to

45 years of age) after 1 to 3 h of exposure as a function of the level of exercise

performed and the O concentration inhaled during the exposure. Group mean 3

data from numerous controlled human exposure and field studies indicate that,

in general, statistically significant pulmonary function decrements beyond the

range of normal measurement variability (e.g., 3 to 5% for FEV ) occur 1

(1) at >0.50 ppm O when at rest, 3

(2) at >0.37 ppm O with light exercise (slow walking), 3 

(3) at >0.30 ppm O with moderate exercise (brisk walking), 3

(4) at >0.18 ppm O with heavy exercise (easy jogging), and 3

(5) at >0.16 ppm O with very heavy exercise (running). 3 

Smaller group mean changes (e.g., <5%) in FEV have been observed at lower 1

O concentrations than those listed above. For example, FEV decrements 3 1

have been shown to occur with very heavy exercise in healthy adults at 0.15 to

0.16 ppm O , and such effects may occur in healthy young adults at levels as 3
1-30

low as 0.12 ppm. Also, pulmonary function decrements have been observed in

children and adolescents at concentrations of 0.12 and 0.14 ppm O with heavy 3

exercise. Some individuals within a study may experience FEV decrements in 1

excess of 15% under these exposure conditions, even when the group mean

decrement is less than 5%.

• For exposures of healthy subjects performing moderate exercise during longer

duration exposures (6 to 8 h), 5% group mean decrements in FEV were 1

observed at

(1) 0.08 ppm O after 5.6 h, 3

(2) 0.10 ppm O after 4.6 h, and 3

(3) 0.12 ppm O after 3 h. 3

For these same subjects, 10% group mean FEV decrements were observed at 1

0.12 ppm O after 5.6 and 6.6 h. As in the shorter duration studies, some 3

individuals experience changes larger than those represented by the group

mean changes. 

• An increase in the incidence of cough has been reported at O concentrations 3

as low as 0.12 ppm in healthy adults during 1 to 3 h of exposure with very

heavy exercise. Other respiratory symptoms, such as pain on deep inspiration,

shortness of breath, and lower respiratory scores (a combination of several

symptoms), have been observed at 0.16 to 0.18 ppm O with heavy and very 3

heavy exercise. Respiratory symptoms also have been observed following

exposure to 0.08, 0.10, and 0.12 ppm O for 6.6 h with moderate levels of 3

exercise.

• Increases in nonspecific airway responsiveness in healthy adults have been

observed after 1 to 3 h of exposure to 0.40 but not 0.20 ppm O at rest and 3

have been observed at concentrations as low as 0.18 but not to 0.12 ppm

O during exposure with very heavy exercise. Increases in nonspecific airway 3

responsiveness during 6.6-h exposures with moderate levels of exercise have

been observed at 0.08, 0.10, and 0.12 ppm O . 3

Short-term O exposure of laboratory animals and humans disrupts the barrier 3

function of the lung epithelium, permitting materials in the airspaces to enter lung tissue,

allowing cells and serum proteins to enter the airspaces (inflammation), and setting off a

cascade of responses.

• Increased levels of PMNs and protein in lung lavage fluid have been observed

following exposure of healthy adults to 0.20, 0.30, and 0.40 ppm with very

heavy exercise and have not been studied at lower concentrations for 1- to 3-h

exposures. Increases in lung lavage protein and PMNs also have been

observed at 0.08 and 0.10 ppm O during 6.6-h exposures with moderate 3

exercise; lower concentrations have not been tested.

Short-term O exposure of laboratory animals and humans impairs AM clearance 3

of viable and nonviable particles from the lungs and decreases the effectiveness of host

defenses against bacterial lung infections in animals and perhaps in humans. The ability of

AMs to engulf microorganisms is decreased in humans exposed to 0.08 and 0.10 ppm O for 3

6.6 h with moderate exercise. 
1-31

2. What are the effects of repeated, short-term exposures to ozone?

During repeated short-term exposures, some of the O -induced responses are 3

partially or completely attenuated. Over a 5-day exposure, pulmonary function changes are

typically greatest on the second day, but return to control levels by the fifth day of exposure. 

Most of the inflammatory markers (e.g., PMN influx) also attenuate by the fifth day of

exposure, but markers of cell damage (e.g., lactate dehydrogenase enzyme activity) do not

attenuate but continue to increase. Attenuation of lung function decrements is reversed

following 7 to 10 days without O . Some inflammatory markers also are reversed during this 3

time period, but others still show attenuation even after 20 days without O . The mechanisms 3

and impacts involved in attenuation are not known, although animal studies show that the

underlying cell damage continues throughout the attenuation process. In addition, attenuation

may alter the normal distribution of O within the lung, allowing more O to reach sensitive 3 3

regions, possibly affecting normal lung defenses (e.g., PMN influx in response to inhaled

microorganisms). 

3. What are the effects of long-term exposures to ozone?

Available data indicate that exposure to O for months and years causes structural 3

changes in several regions of the RT, but effects may be of the greatest importance in the

CAR (where the alveoli and conducting airways meet); this region typically is affected in

most chronic airway diseases of the human lung. This information on O effects in the distal 3

lung is extrapolated from animal toxicological studies because, to date, comparable data are

not available from humans. The apparent lack of reversal of effects during periods of clean

air exposure raises concern that seasonal exposures may have a cumulative impact over many

years. The role of adaptive processes in this response is unknown but may be critically

dependent on the temporal frequency or profile of exposure. Furthermore, the interspecies

diversity in apparent sensitivity to the chronic effects of O is notable, with the rat 3

representing the lower limit of response, and the monkey the upper limit. Epidemiological

studies attempting to associate chronic health effects in humans with long-term O exposure 3

provide only suggestive evidence that such a linkage exists.

Long-term exposure of one strain of female mice to high O levels (1 ppm) caused 3

a small, but statistically significant increase in lung tumors. There was no concentrationresponse relationship, and rats were not affected. Genotoxicity data are either negative or

weak. Given the nature of the database, the effects in one strain of mice cannot yet be

extrapolated qualitatively to humans. Ozone (0.5 ppm) did not show tumor-promoting

activity in a chronic rat study.

4. What are the effects of binary pollutant mixtures containing ozone?

Combined data from laboratory animal and controlled human exposure studies of

O support the hypothesis that coexposure to pollutants, each at low-effect levels, may result 3

in effects of significance. The data from human studies of O in combination with NO , SO , 3 2 2

H SO , HNO , or CO show no more than an additive response on lung spirometry or 2 4 3

respiratory symptoms. The larger number of laboratory animal studies with O in mixture 3

with NO and H SO show that effects can be additive, synergistic, or even antagonistic, 2 2 4

depending on the exposure regimen and the endpoint studied. This issue of exposure to

copollutants remains poorly understood, especially with regard to potential chronic effects.
1-32

5. What population groups are at risk as a result of exposure to ozone?

Identification of population groups that may show increased sensitivity to O is 3

based on their biological responses to O , preexisting lung disease (e.g., asthma), activity 3

patterns, personal exposure history, and personal factors (e.g., age, nutritional status). 

The predominant information on the health effects of O3 noted above comes from

clinical and field studies on healthy, nonsmoking, exercising subjects, 8 to 45 years of age. 

These studies demonstrate that, among this group, there is a large variation in sensitivity and

responsiveness to O , with at least a 10-fold difference between the most and least responsive 3

individuals. Individual sensitivity to O also may vary throughout the year, related to 3

seasonal variations in ambient O exposure. The specific factors that contribute to this large 3

intersubject variability, however, remain undefined. Although differences in response may be

due to the dosimetry of O in the RT, available data show little difference on O deposition in 3 3

the lungs for inhalation through the nose or mouth.

Daily life studies reporting an exacerbation of asthma and decrease in peak

expiratory flow rates, particularly in asthmatic children, appear to support the controlled

studies; however, those studies may be confounded by temperature, particle or aeroallergen

exposure, and asthma severity of the subjects or their medication use. In addition, field

studies of summertime daily hospital admissions for respiratory causes show a consistent

relationship between asthma and ambient levels of O in various locations in the northeastern 3

United States, even after controlling for independent contributing factors. Controlled studies

on mild asthmatics suggest that they have similar lung volume responses but greater airway

resistance changes to O than nonasthmatics. Furthermore, limited data from studies of 3

moderate asthmatics suggest that this group may have greater lung volume responses than

nonasthmatics. 

Other population groups with preexisting limitations in pulmonary function and

exercise capacity (e.g., chronic obstructive pulmonary disease, chronic bronchitis, ischemic

heart disease) would be of primary concern in evaluating the health effects of O . 3

Unfortunately, not enough is known about the responses of these individuals to make

definitive conclusions regarding their relative responsiveness to O . Indeed, functional effects 3

in these individuals with reduced lung function may have greater clinical significance than

comparable changes in healthy individuals.

Currently available data follow on personal factors or personal exposure history

known or suspected of influencing responses to O . 3

• Human studies have identified a decrease in pulmonary function

responsiveness to O with increasing age, although symptom rates remain 3

similar. Toxicological studies are not easily interpreted but suggest that young

animals are not more responsive than adults.

• Available toxicological and human data have not demonstrated conclusively

that males and females respond differently to O . If gender differences exist 3

for lung function responsiveness to O , they are not based on differences in 3

baseline pulmonary function.

• Data are not adequate to determine whether any ethnic or racial group has a

different distribution of responsiveness to O . In particular, the responses of 3

nonwhite asthmatics have not been investigated.
1-33

• Information derived from O exposure of smokers is limited. The general trend 3

is that smokers are less responsive than nonsmokers. This reduced

responsiveness may wane after smoking cessation. 

• Although nutritional status (e.g., vitamin E deficiency) makes laboratory rats

more susceptible to O3-induced effects, it is not clear if vitamin E

supplementation has an effect in human populations. Such supplementation

has no or minimal effect in animals. The role of such antioxidant vitamins in

O responsiveness, especially their deficiency, has not been well studied. 3

Based on information presented in this document, the population groups that have

demonstrated increased responsiveness to ambient concentrations of O consist of exercising, 3

healthy and asthmatic individuals, including children, adolescents, and adults.