&EPA
United States
Environmental Protection
Agency
Environmental Criteria and
Assessment Office
Research Triangle Park, NC 27711
EPA-600/8-84-020aF
August 1986
Research and Development
Air Quality
Criteria for
Ozone and Other
Photochemical
Oxidants
Volume I of V
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EPA-600/8-84-020aF
August 1986
Air Quality Criteria
for Ozone and Other
Photochemical Oxidants
Volume I of V
Environmental Criteria and Assessment Office
Office of Health and Environmental Assessment
Office of Research and Development
U.S. Environmental Protection Agency
Research Triangle Park, N.C. 27711
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DISCLAIMER
This document has been reviewed in accordance with U.S. Environmental
Protection Agency policy and approved for publication. Mention of trade names
or commerical products does not constitute endorsement or recommendation for
use.
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ABSTRACT
Scientific information is presented and evaluated relative to the health
and welfare effects associated with exposure to ozone and other photochemical
oxidants. Although it is not intended as a complete and detailed literature
review, the document covers pertinent literature through early 1986.
Data on health and welfare effects are emphasized, but additional infor-
mation is provided for understanding the nature of the oxidant pollution pro-
blem and for evaluating the reliability of effects data as well as their
relevance to potential exposures to ozone and other oxidants at concentrations
occurring in ambient air. Information is presented on the following exposure-
related topics: nature, source, measurement, and concentrations of precursors
to ozone and other photochemical oxidants; the formation of ozone and other
photochemical oxidants and their transport once formed; the properties, chem-
istry, and measurement of ozone and other photochemical oxidants; and the
concentrations of ozone and other photochemical oxidants that are typically
found in ambient air.
The specific areas addressed by chapters on health and welfare effects
are the toxicological appraisal of effects of ozone and other oxidants; effects
observed in controlled human exposures; effects observed in field and epidemio-
logical studies; effects on vegetation seen in field and controlled exposures;
effects on natural and agroecosystems; and effects on nonbiological materials
observed in field and chamber studies.
m
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AIR QUALITY CRITERIA FOR OZONE
AND OTHER PHOTOCHEMICAL OXIDANTS
Page
VOLUME I
Chapter 1. Summary and Conclusions 1-1
VOLUME II
Chapter 2. Introduction 2-1
Chapter 3. Properties, Chemistry, and Transport of Ozone and
Other Photochemical Oxidants and Their Precursors 3-1
Chapter 4. Sampling and Measurement of Ozone and Other
Photochemical Oxidants and Their Precursors 4-1
Chapter 5. Concentrations of Ozone and Other Photochemical
Oxidants in Ambient Air 5-1
VOLUME III
Chapter 6. Effects of Ozone and Other Photochemical Oxidants
on Vegetation 6-1
Chapter 7. Effects of Ozone on Natural Ecosystems and Their
Components 7-1
Chapter 8. Effects of Ozone and Other Photochemical Oxidants
on Nonbiological Materials 8-1
VOLUME IV
Chapter 9. Toxicological Effects of Ozone and Other
Photochemical Oxidants 9-1
VOLUME V '
Chapter 10. Controlled Human Studies of the Effects of Ozone
and Other Photochemical Oxidants 10-1
Chapter 11. Field and Epidemic!ogical Studies of the Effects
of Ozone and Other Photochemical Oxidants 11-1
Chapter 12. Evaluation of Health Effects Data for Ozone and
Other Photochemical Oxidants 12-1
iv
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TABLE OF CONTENTS
LIST OF TABLES 1x
LIST OF FIGURES . x
LIST OF ABBREVIATIONS xi
AUTHORS, CONTRIBUTORS, AND REVIEWERS , xv
1. SUMMARY AND CONCLUSIONS 1-1
1.1 INTRODUCTION 1-1
1.2 PROPERTIES, CHEMISTRY, AND TRANSPORT OF OZONE AND OTHER
PHOTOCHEMICAL OXIDANTS AND THEIR PRECURSORS 1-3
1.2.1 Descriptions and Properties of Ozone and Other
Photochemical Oxidants 1-3
1.2.2 Nature of Precursors to Ozone and Other
Photochemical Oxidants 1-4
1.2.3 Atmospheric Reactions of Ozone and Other Oxidants
Including Their Role in Aerosol Formation 1-5
1.2.3.1 Formation and Transformation of Ozone
and Other Photochemical Oxidants 1-6
1.2.3.2 Atmospheric Chemical Processes
Involving Ozone 1-7.
1.2i3.3 Atmospheric Reactions of PAN, H202,
and HCOOH ...-.' 1-8
1.2.4 Meteorological and Climatological Processes 1-9
1.2.4.1 Atmospheric Mixing . .... 1-9
1.2.4.2 Wind Speed and Direction 1-11
1.2.4.3 Effects of Sunlight and Temperature 1-11
1.2.4.4 Transport of Ozone and Other Oxidants
and Their Precursors 1-12
, 1.2.4.5 Stratospheric-Tropospheric Ozone
Exchange '...;...... 1-13
1.2.4.6 Stratospheric Ozone at Ground Level ....' 1-14
1.2,4.7 Background Ozone from Photochemical
Reactions 1-15
1.2.5 Sources, Emissions, and Concentrations of
Precursors to Ozone and Other Photochemical
Oxidants 1-16
1.2.5.1 Sources and Emissions of Precursors 1-16
1.2.5.2 Representative Concentrations in
Ambient Air 1-17
1.2,6 Source-Receptor (Oxidant-Precursor) Models 1-19
1.2.6.1 Trajectory Models 1-19
1.2.6.2 Fixed-Grid Models 1-20
1.2.6.3 Box Models 1-20
1.2.6.4 Validation and Sensitivity Analyses
for Dynamic Model s 1-20
1.3 SAMPLING AND MEASUREMENT OF OZONE AND OTHER
PHOTOCHEMICAL OXIDANTS AND THEIR PRECURSORS 1-21
1.3.1 Sampling and Measurement of Ozone and Other
Photochemical Oxidants 1-21
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TABLE OF CONTENTS
(continued)
1.3.1.1 Quality Assurance and Sampling
1.3.1.2 Measurement Methods for Total Oxidants
and Ozone 1-22
1.3.1.3 Calibration Methods 1-24
1.3.1.4 Relationships of Total Oxidants and
Ozone Measurements 1-26
1.3.1.5 Methods for Sampling and Analysis of
Peroxyacetyl Nitrate and Its
Homo! ogues 1-27
1.3.1.6 Methods for Sampling and Analysis of
Hydrogen Peroxide 1-30
1.3.2 Measurement of Precursors to Ozone and Other
Photochemical Oxidants 1-32
1.3.2.1 Nonmethane Organic Compounds . ...... 1-32
1.3.2.2 Nitrogen Oxides 1-34
1.4 CONCENTRATIONS OF OZONE AND OTHER PHOTOCHEMICAL
OXIDANTS IN AMBIENT AIR 1-36
1.4.1 Ozone Concentrations in Urban Areas 1-36
1.4.2 Trends in Nationwide Ozone Concentrations 1-39
1,4.3 Ozone Concentrations in Nonurban Areas 1-41
1.4.4 Diurnal and Seasonal Patterns in Ozone
Concentrations 1-44
1.4.5 Spatial Patterns in Ozone Concentrations 1-46
1.4.5.1 Urban-Nonurban Differences in Ozone
Concentrations 1-46
1.4.5.2 Geographic, Vertical, and Altitudinal
Variations in Ozone Concentrations 1-47
1.4.5.3 Other Spatial Variations in Ozone
Concentrations 1-50
1.4.6 Concentrations and Patterns of Other
Photochemical Oxidants 1-51
1.4.6.1 Concentrations 1-51
1.4.6.2 Patterns 1-52
1.4.7 Relationship Between Ozone and Other
Photochemical Oxidants . 1-53
1.5 EFFECTS OF OZONE AND OTHER PHOTOCHEMICAL OXIDANTS
ON VEGETATION 1-55
1.5.1 Limiting Values of Plant Response to Ozone 1-56
1.5.2 Methods for Determining Ozone Yield Losses 1-58
1.5.3 Estimates of Ozone-Induced Yield Loss 1-60
1.5.3.1 Yield Loss: Determination by
Regression Analysis 1-61
1.5.3.2 Yield Loss: Determination from
Discrete Treatment 1-67
1.5.3.3 Yield Loss: Determination with
Chemical Protectants 1-67
1.5.3.4 Yield Loss: Determination from
Ambient Exposures 1-69
1.5.3.5 Yield Loss Summary ... 1-69
vi
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TABLE OF CONTENTS
(continued)
1.5.4 Effects on Crop Quality . 1-72
1.5.5 Statistics Used to Characterize Ozone Exposures .. 1-72
1.5.6 Relationship Between Yield Loss and Foliar
.Injury 1-75
1.5.7 Physiological Basis of Yield Reductions 1-75
1.5.8 Factors Affecting Plant Response to Ozone 1-76
1.5.8.1 Environmental Conditions 1-77
1.5.8.2 Interaction with Plant Diseases 1-78
1.5.8.3 Interaction of Ozone with Other
Air Pollutants 1-78
1.5.9 Economic Assessment of Effects of Ozone on
Agriculture 1-79
1.5.10 Effects of Peroxyacetyl Nitrate on'Vegetation 1-90
1.5.10.1 Factors Affecting Plant Response to
PAN 1-90
1.5.10.2 Limiting Values of Plant Response .. 1-90
1.5.10.3 Effects of PAN on Plant Yield 1-91
1.6 EFFECTS OF OZONE ON NATURAL ECOSYSTEMS AND THEIR
COMPONENTS 1-91
1.6.1 Responses of Ecosystems to Ozone Stress 1-91
1.6.2 Effects of Ozone on Producers 1-92
1.6.3 Effects of Ozone on Other Ecosystem Components
and on Ecosystem Interactions 1-94
1.6.4 Effects of Ozone on Specific Ecosystems 1-95
1.6.5 Economic Valuation of Ecosystems 1-97
1.7 EFFECTS OF OZONE AND OTHER PHOTOCHEMICAL OXIDANTS ON
NONBIOLOGICAL MATERIALS 1-98
1.8 TOXICOLOGICAL EFFECTS OF OZONE AND OTHER PHOTOCHEMICAL
OXIDANTS 1-103
1.8.1 Introduction 1-103
1.8.2 Regional Dosimetry in the Respiratory Tract 1-104
1.8.3 Effects of Ozone on the Respiratory Tract 1-107
1.8.3.1 Morphological Effects 1-107
1.8.3.2 Pulmonary Function 1-110
1.8.3.3 Biochemical Effects 1-116
1.8.3.4 Host Defense Mechanisms 1-122
1.8.3.5 Tolerance 1-127
1.8.4 Extrapulmonary Effects of Ozone 1-131
1.8.4.1 Central Nervous System and Behavioral
Effects 1-131
1.8.4.2 Cardiovascular Effects 1-132
1.8.4.3 Hematological and Serum Chemistry
Effects 1-132
1.8.4.4 Cytogenetic and Teratogenic Effects .... 1-134
1.8.4.5 Other Extrapulmonary Effects 1-135
1.8.5 Interaction of Ozone with Other Pollutants 1-136
1.8.6 Effects of Other Photochemical Oxidants 1-136
1.9 CONTROLLED HUMAN STUDIES OF THE EFFECTS OF OZONE AND
OTHER PHOTOCHEMICAL OXIDANTS 1-142
vii
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TABLE OF CONTENTS
(continued)
1.10 FIELD AND EPIDEMIOLOGICAL STUDIES OF THE EFFECTS OF
OZONE AND OTHER PHOTOCHEMICAL OXIDANTS 1-151
1.11 EVALUATION OF HEALTH EFFECTS DATA FOR OZONE AND OTHER
PHOTOCHEMICAL OXIDANTS 1-156
1.11.1 Health Effects in the General Human Population ... 1-156
1.11.2 Health Effects in Individuals with
Preexisting Disease 1-162
1.11.3 Extrapolation of Effects Observed in Animals
to Human Populations 1-163
1.11.4 Health Effects of Other Photochemical Oxidants
and Pol 1utant Mixtures 1-164
1.11.5 Identification of Potentially At-Risk
Groups 1-164
1.12 REFERENCES 1-166
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LIST OF TABLES
Table Page
1-1 Summary of ozone monitoring techniques 1-23
1-2 Ozone calibration techniques 1-25
1-3 Second-highest ozone concentrations among daily maximum 1-hr
values in 1983 in Standard Metropolitan Statistical Areas
with populations >1 million, given by census divisions and
regions 1-37
1-4 Ozone concentrations for short-term exposures that
produce 5 or 20 percent injury to vegetation grown under
sensitive conditions 1-57
1-5 Summary of ozone concentrations predicted to cause 10 percent
and 30 percent yield losses and summary of yield losses
predicted to occur at 7-hr seasonal mean ozone concentrations
of 0.04 and 0.06 ppm 1-64
1-6 Ozone concentrations at which significant yield losses have
been noted for a variety of plant species exposed under
various experimental conditions 1-68
1-7 Effects of ozone on crop yield as determined by the use
of chemical protectants 1-70
1-8 Effects of ambient oxidants on yield of selected crops 1-71
1-9 Summary of estimates of regional economic consequences of
ozone pollution 1-81
1-10 Summary of estimates of national economic consequences of
ozone pollution 1-73
1-11 Summary table: Morphological effects of ozone in
experimental animals 1-112
1-12 Summary table: Effects on pulmonary function of short-term
exposures to ozone in experimental animals 1-116
1-13 Summary table: Effects on pulmonary function of long-term
exposures to ozone in experimental animals 1-118
1-14 Summary table: Biochemical changes in experimental animals
exposed to ozone 1-124
1-15 Summary table: Effects of ozone on host defense mechanisms
in experimental animals 1-129
1-16 Summary table: Extrapulmonary effects of ozone in
experimental animals 1-138
1-17 Summary table: Interaction of ozone with other pollutants
in experimental animals 1-140
1-18 Summary table: Controlled human exposure to ozone 1-143
1-19 Summary table: Acute effects of ozone and other photochemical
oxidants in field studies with a mobile laboratory 1-152
IX
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LIST OF FIGURES
Figure Page
1-1 National trend in composite average of the second highest
value among daily maximum 1-hour ozone concentrations at
selected groups of sites, 1975 through 1983 1-40
1-2 Distributions of the three highest 1-hour ozone concentrations
at valid sites (906 station-years) aggregated for 3 years
(1979, 1980, and 1981) and the highest ozone concentrations
at NAPBN sites aggregated for those years (24 station-years) 1-42
1-3 Relationship between ozone concentration, exposure duration,
and a reduction in plant growth or yield 1-59
1-4 Examples of the effects of ozone on the yield of soybean
and wheat cultivars , 1-62
1-5 Examples of the effects of ozone on the yield of cotton,
tomato, and turnip 1-63
1-6 Number and percentage of 37 crop species or cultivars
predicted to show a 10 percent yield loss at various ranges
of 7-hr seasonal mean ozone concentrations 1-66
1-7 Summary of morphological effects in experimental animals
exposed to ozone 1-111
1-8 Summary of effects of short-term ozone exposures on pulmonary
function in experimental animals . 1-115
1-9 Summary of effects of long-term ozone exposures on pulmonary
function in experimental animals 1-117
1-10 Summary of biochemical changes in experimental animals
exposed to ozone 1-123
1-11 Summary of effects of ozone .on host defense mechanisms in
experimental animals 1-128
1-12 Summary of extrapulmonary effects of ozone in experimental
animals 1-137
1-13 Summary of effects in experimental animals exposed to
ozone combined with other pollutants 1-139
1-14 Group mean decrements in 1-sec forced expiratory volume
during 2-hr ozone exposures with different levels of
intermittent exercise 1-158
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LIST OF ABBREVIATIONS
AChE
avg
BAKI
Be (7Be)
C
°C
CA
CC
CHSC(0)02
cm
CNS
CO
C02
COLD
cone., concn.
CV
dbh
DN'PH
ECD
EDU
EKMA
FEF
Fe2(S04)3
FEV
FEV1
FID
fR
FTIR
FVC
G-6-PD
GC
GPT
acetylcholinesterase
average
boric acid buffered potassium iodide
beryllium (radioactive isotope of beryllium)
carbon, concentration
degrees Celsius
chromotropic acid
closing capacity
acetylperoxy radical
centimeter
central nervous system
carbon monoxide
carbon dioxide
chronic obstructive lung disease
concentration
closing volume
diameter at breast height
2,4-di nitropheny1hydrazi ne
electron-capture detector
ethylenediurea
Empirical Kinetic Modeling Approach
forced expiratory flow
ferric sulfate
forced expiratory volume
forced expiratory volume in 1 sec
flame ionization detector
respiratory frequency
Fourier-transform infrared
forced vital capacity
glucose-6-phosphate dehydrogenase
gas chromatography
gas-phase titration
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LIST OF ABBREVIATIONS
(continued)
GSH
HC
HCOOH
HN03
HN04
HO
H02
MONO
HPLC
HPPA
hr
hr/day
HRP
H2Q2
H2S04
I
I"
1C
I/O
IR
KI03
km
LAAPCD
LCV
LDH
L/min
M
ra
MBTH
mi
glutathione
hydrocarbon(s)
formic acid
nitric acid
peroxynitric acid
hydroxy
hydroperoxy
nitrous acid
high-pressure liquid chromatograpy; high-performance
liquid chrotnatography
3-(p_-hydroxyphenyl)propionic acid
hour(s)
hours per day ;
horseradish peroxidase
hydrogen peroxide
sulfuric acid
impact
iodide ion
inspiratory capacity
ratio of indoor to outdoor ozone concentrations '
infrared
potassium iodate
ki 1ometer .
Los Angeles Air Pollution Control District ';
leuco crystal violet
lactate deyhydrogenase
liters per minute , ..-. . ••.
molar .••. • : .
meter(s). . , •; -.., >••-
3-methyl-2-benzothiazolinone hydrazone
mile(s) •
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LIST OF ABBREVIATIONS
(continued)
NADPH nicotinamide adenine dinucleotide phosphate
NAPBN National Air Pollution Background Network
NBKI neutral buffered potassium iodide
(NH4)2S04 ammonium sulfate
NF National Forest
nm nanometer(s)
NMHC nonmethane hydrocarbons
NMOC nonmethane organic compounds
NO nitric oxide
NQ2 nitrogen dioxide
N03 nitrogen trioxide
NO nitrogen oxides
r\
AN2 nitrogen washout
NPSH non-protein sulfhydryls
NR natural rubber
N20 nitrous oxide
OH hydroxyl group (or radical)
Q2 oxygen
03 ozone
OZIPP . , Ozone Isopleth Plotting Package
PAN peroxyacetyl nitrate
PA02 alveolar partial pressure of oxygen
M »
PBzN peroxybenzoyl nitrate
PEFR peak expiratory flow rate
pH negative log of H ion concentration
PPN peroxypropionyl nitrate
ppb parts per billion
ppm parts per million
rad radiation absorbed dose
RBC red blood cell
RV residual volume
xm
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LIST OF ABBREVIATIONS
(continued)
Sa02
SAROAD
SBR
sec
SGaw
SNAAQS
S02
S04
S0x
SRaw
SRM
SURE
T
TF
Tg/yr
TGS-ANSA
TLC
TSH
pg/m3
pm/hr
UV
voc
ZnS04
arterial oxygen saturation
Storage and Retrieval of Aerometric Data
styrene-butadiene rubber
second(s)
specific airway conductance
Secondary National Ambient Air Quality Standards
sulfur dioxide
sulfate
sulfur oxide(s)
specific airway resistance
Standard Reference Material
Sulfate Regional Experiment Sites
time, temperature
tropopause-folding events
teragrams per year
triethanolamine, guaiacol(£-methoxyphenol), sodium
metabisulfite; and 8-anilino-l-naphthalene sulfonic acid
total lung capacity
thyroid stimulating hormone
microgram(s) per cubic meter
micrometer(s) per hour
ultraviolet
tidal volume
minute ventilation; expired volume per minute
volatile organic compounds
zinc sulfate
xiv
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.AUTHORS, CONTRIBUTORS, AND REVIEWERS
Authors:
Dr. Richard M. Adams
Department of Agricultural and Resource Economics
Oregon State University
Con/all is, OR 97331
Dr. Donald E. Gardner
Northrop Services, Inc.
Environmental Sciences
P. 0. Box 12313
Research Triangle Park, NC 27709
Dr. J. H. B. Garner
Environmental Criteria and Assessment Office
MD-52
U.S. Environmental Protection Agency
Research Triangle Park, NC 27711
Dr. Judith A. Graham
Health Effects Research Laboratory
MD-51
U.S. Environmental Protection Agency
Research Triangle Park, NC 27711
Dr. Milan J. Hazucha
School of Medicine
Center for Environmental Health and Medical Sciences
University of North Carolina
Chapel Hill, NC 27514
Dr. Jimmie A. Hodgeson
U.S. Environmental Protection Agency
Environmental Monitoring Systems Laboratory
26 West St. Clair
Cincinnati, OH 45268
Mr. Michael W. Holdren
BatteHe, Columbus Laboratories
505 King Avenue
Columbus, OH 43201
Dr. Donald H. Horstman
Health Effects Research Laboratory
MD-58
U.S. Environmental Protection Agency
Research Triangle Park, NC 27711
xv
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Authors (continued):
Mr. James M. Kawecki
TRC Environmental Consultants, Inc.
2001 Wisconsin Avenue, N.W.
Suite 261
Washington, DC 20007
Dr. Jan G. Laarman
Department of Forestry
North Carolina State University
Raleigh, NC 27607
Mr. Thomas B. McMullen
Environmental Criteria and Assessment Office
MD-52
U.S. Environmental Protection Agency
Research Triangle Park, NC 27711
Dr. Daniel B. Menzel
Laboratory of Environmental Toxicology and Pharmacology
Duke University Medical Center
P. 0. Box 3813
Durham, NC 27710
Mr. James A. Raub
Environmental Criteria and Assessment Office
MD-52
U.S. Environmental Protection Agency
Research Triangle Park, NC 27711
Ms. Beverly E. Tilton
Environmental Criteria and Assessment Office
MD-52
U.S. Environmental Protection Agency
Research Triangle Park, NC 27711
Dr. David T. Tingey
Environmental Research Laboratory
200 SW 35th Street
Corvallis, OR 97330
Dr. Halvor Westberg
Director, Laboratory for Atmospheric Research,
and Professor, Civil and Environmental Engineering
Washington State University
Pullman, WA 99164-2730
Dr. Arthur M. Winer
Assistant Director
Statewide Air Pollution Research Center
University of California
Riverside, CA 92521
xv i
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Contributing Authors:
Dr. Robert Frank
Department of Environmental Health Sciences .
Johns Hopkins School of Hygiene and Public Health
615 N. Wolfe Street
Baltimore, MD 21205
Dr. Michael D. Lebowitz
Department of Internal Medicine
College of Medicine
University of Arizona
Tucson, AZ 85724
Dr. Harold G. Richter
Office of Air Quality Planning and Standards
Monitoring and Data Analysis Division
MD-14
U.S. Environmental Protection Agency
Research Triangle Park, NC 27711
Mr. Elmer Robinson
Director, Mauna Loa Observatory
National Oceanic and Atmospheric Administration
(NOAA/CMCO)
P.O. Box 275
Hilo, HI 96720
xv
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SCIENCE ADVISORY
CLEAN AIR SCIENTIFIC ADVISORY COMMITTEE
The substance of this document was reviewed by the Clean Air Scientific
Advisory Committee of the Science Advisory Board in public sessions.
SUBCOMMITTEE ON OZONE
Chai rman
Dr. Morton Lippmann
Professor
Department of Environmental Medicine
New York University Medical Center
Tuxedo, New York 10987
Members
Dr. Mary 0. Amdur
Senior Research Scientist
Energy Laboratory
Massachusetts Institute of Technology
Cambridge, Massachusetts 02139
Dr. Eileen G. Brennan
Professor
Department of Plant Pathology
Martin Hall, Room 213, Lipman Drive
Cook College-NJAES
Rutgers University
New Brunswick, New Jersey 08903
Dr. Edward D. Crandall
Professor of Medicine
School of Medicine
Cornell University
New York, New York 10021
Dr. James D. Crapo
Associate Professor of Medicine
Chief, Division of Allergy, Critical
Care and Respiratory Medicine
Duke University Medical Center
Durham, North Carolina 27710
Dr. Robert Frank
Professor of Environmental Health
Sciences
Johns Hopkins School of Hygiene
and Public Health
615 N. Wolfe Street
Baltimore, Maryland 21205
Professor A. Myrick Freeman II
Department of Economics
Bowdoin College
Brunswick, Maine 04011
Dr. Ronald J. Hall
Senior Research Scientist and Leader
Aquatic and Terrestrial Ecosystems
Section
Ontario Ministry of the Environment
Dorset Research Center
Dorset, Ontario
Canada POA1EO
Dr. Jay S. Jacobson
Plant Physiologist
Boyce Thompson Institute
Tower Road
Ithaca, New York 14853
xvm
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Dr. Warren B. Johnson
Director, Atmospheric Science Center
SRI International
333 Ravenswood Avenue
Menlo Park, California 94025
Dr. Jane Q. Koenig
Research Associate Professor
Department of Environmental Health
University of Washington
Seattle, Washington 98195
Dr. Paul Kotin
Adjunct Professor of Pathology
University of Colorado Medical School
4505 S. Yosemite, #339
Denver, Colorado 80237
Dr. Timothy Larson
Associate Professor
Environmental Engineering and
Science Program
Department of Civil Engineering
University of Washington
Seattle, Washington 98195
Professor M. Granger Morgan
Head, Department of Engineering
and Public Policy
Carnegie-Mellon University
Pittsburgh, Pennsylvania 15253
Dr. D. Warner North
Principal
Decision Focus Inc., Los Altos
Office Center, Suite 200
4984 El Camino Real
Los Altos, California 94022
Dr. Robert D. Rowe
Vice President, Environmental and
Resource Economics
Energy and Resources Consultants, Inc.
207 Canyon Boulevard
Boulder, Colorado 80302
Dr. George Taylor
Environmental Sciences Division
P.O. Box X
Oak Ridge National Laboratory
Oak Ridge, Tennessee 37831
Dr. Michael Treshow
Professor
Department of Biology
University of Utah
Salt Lake City, Utah 84112
Dr. Mark J. Utell
Co-Director, Pulmonary Disease Unit
Associate Professor of Medicine and
Toxicology in Radiation Biology
and Biophysics
University of Rochester Medical
Center
Rochester, New York 14642
Dr. James H. Ware
Associate Professor
Harvard School of Public Health
Department of Biostatisties
677 Huntington Avenue
Boston, Massachusetts 02115
Dr. Jerry Wesolowski
Air and Industrial Hygiene Laboratory
California Department of Health
2151 Berkeley Way
Berkeley, California 94704
Dr. James L. Whittenberger
Director, University of California
Southern Occupational Health Center
Professor and Chair, Department of
Community and Environmental Medicine
California College of Medicine
University of California - Irvine
19772 MacArthur Boulevard
Irvine, California 92717
Dr. George T. Wolff
Senior Staff Research Scientist
General Motors Research Labs
Environmental Science Department
Warren, Michigan 48090
xix
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PROJECT TEAM FOR DEVELOPMENT
OF
Air Quality Criteria for Ozone and Other Photochemical Oxidants
Ms. Beverly E. Til ton, Project Manager
and Coordinator for Chapters 1 through 5, Volumes I and II
Environmental Criteria and Assessment Office (MD-52)
U.S. Environmental Protection Agency
Research Triangle Park, NC 27711
Mr. Norman E. Chi Ids
Environmental Criteria and Assessment Office (MD-52)
U.S. Environmental Protection Agency
Research Triangle Park, NC 27711
Dr. J.H.B. Garner
Coordinator for Chapters 7 and 8, Volume III
Environmental Criteria and Assessment Office (MD-52)
U.S. Environmental Protection Agency
Research Triangle Park, NC 27711
Mr. Thomas B. McMullen
Environmental Criteria and Assessment Office (MD-52)
U.S. Environmental Protection Agency
Research Triangle Park, NC 27711
Mr. James A. Raub
Coordinator for Chapters 9 through 12, Volumes IV and V
Environmental Criteria and Assessment Office (MD-52)
U.S. Environmental Protection Agency
Research Triangle Park, NC 27711
Dr. David T. Tingey
Coordinator for Chapter 6, Volume III
Environmental Research Laboratory
U.S. Environmental Protection Agency
200 S.W. 35th Street
Corvallis, OR 97330
xx
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1. SUMMARY AND CONCLUSIONS
1.1 INTRODUCTION
This document is a revision of Air Quality Criteria for Ozone and Other
Photochemical Oxidants, published in 1978 (U.S. Environmental Protection
Agency, 1978). Its purpose is to provide the scientific basis for the deriva-
tion of National Ambient Air Quality Standards (NAAQS) by consolidating'and
assessing knowledge regarding the origin and distribution of. ozone and other
photochemical oxidants and the effects of these pollutants on humans, experi-
mental animals, vegetation, terrestrial ecosystems, and nonbiological materials.
Because the indirect contributions of the photochemical oxidants to visibility
degradation, climatic changes, and acidic deposition cannot, "at present be
quantified, these atmospheric effects and phenomena are not addressed in this
document. They have been addressed, however, in other, recent air quality
criteria documents (U.S. Environmental Protection Agency, 1982a,b).
Research has established that photochemical oxidants in ambient air
consist mainly of ozone, peroxyacetyl nitrate,,and nitrogen dioxide, and of.
considerably lesser amounts of other peroxyacyl nitrates, hydrogen peroxide,
alkyl hydroperoxides, nitric and nitrous acids, and formic acid. Other••oxida'nts
suspected to occur in ambient air but only in trace amounts include peracids
and ozonides. Only data on ozone, peroxyacyl nitrates,..hydrogen peroxide, and
formic acid are examined in this document. Coverage has been limited to these
photochemical oxidants on the basis of available information on effects,
ambient air concentrations, or both. Of these'oxidants, only ozone and, peroxy-
acetyl nitrate have been studied at concentrations having relevance for potential
exposures of human populations or of vegetation, ecosystems, or nonbiological
materials. Although by definition a photochemical oxidant, nitrogen dioxide
is not included among the oxidants discussed in this document. Separate
criteria documents are issued for oxides of nitrogen, and the second document
in that series, completed in 1982, presented information on nitrosamines and
inorganic nitrogen acids, as well as the oxides of nitrogen (U.S. Environmental
Protection Agency, 1982a).
This document presents a review and evaluation of relevant literature on
ozone and other photochemical oxidants published through early 1986. The
document is not intended as a complete literature review, however; but is
1-1
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intended, rather, to present current data of probable consequence for the
derivation of national ambient air quality standards for protecting public
health and welfare. The scientific information selected for review and comment
in the text generally came from the more recent literature, with emphasis on
studies conducted at or near pollutant concentrations found in ambient air.
Generally, only published material that has undergone scientific peer review
has been included. In the interest of admitting new and important information,
however, some material not published in the open literature but meeting other
standards of scientific reporting may have been included. In addition, the
studies reviewed in the health- and we!fare-related chapters met other selection
criteria, including the appropriate use and satisfaction of statistical tests.
In the early chapters of this document, an overview is presented of the
nature, origins, and distribution in ambient air of those organic and inorganic
compounds that serve as precursors to ozone and other photochemical oxidants.
The currently available measurement techniques for these precursors are briefly
evaluated, inasmuch as the assessment of the occurrence of the precursors
depends upm. their accurate measurement. Similarly, an overview is presented
of the chemical and physical processes in the atmosphere by which precursors
give rise to the production of ozone and other photochemical oxidants. In
addition, the properties of ozone and other photochemical oxidants are presented
as background for understanding information presented in the chapters on
health and welfare effects. Likewise, techniques for the measurement of
ozone, total oxidants, and individual oxidant species other than ozone are
evaluated, since the significance of aerometric and exposure data on these
pollutants is dependent upon the accuracy and specificity of the analytical
techniques used. Typical concentrations of the respective oxidants are pre-
sented to permit assessment of potential exposures of human populations and
other receptors.
Remaining chapters of the document contain the actual air quality criteria;
that is, quantitative and qualitative information that describes the nature of
the health and welfare effects attributable to ozone and other photochemical
oxidants and the concentrations at which these pollutants are thought to
produce the observed effects.
Neither techniques nor strategies for the abatement of photochemical
oxidants are reviewed in this document. Technology for controlling the emissions
of nitrogen oxides and of volatile organic compounds is discussed in documents
1-2
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issued by the Office of Air Quality Planning and Standards (OAQPS) of the U.S.
Environmental Protection Agency (e.g., U.S. Environmental Protection Agency,
1978b, 1983). Likewise, research findings and issues germane to the scientific
basis for control strategies are addressed in numerous documents issued by
OAQPS and by the Office of Research and Development.
In addition, certain issues of direct relevance to standard-setting are
not explicitly addressed in this document, but are addressed instead in
documentation prepared by OAQPS as part of its regulatory analyses. Such
analyses include: (1) discussion of what constitutes an "adverse effect,"
that is, the effect or effects the NAAQS are intended to protect against; (2)
assessment of risk; and (3) discussion of factors to be considered in providing
an adequate margin of safety. While scientific data contribute significantly
to decisions regarding these three issues, their resolution cannot be achieved
solely on the basis of experimentally acquired information. Final decisions
on items (1) and (3) are made by the Administrator of the U.S. Environmental
Protection Agency.
The legislative basis for the development and issuance of the air quality
criteria and related information presented in this document is found in
Sections 108 and 109 of the Clean Air Act (U.S.C., 1982).
1.2 PROPERTIES, CHEMISTRY, AND TRANSPORT OF OZONE AND OTHER PHOTOCHEMICAL
OXIDANTS AND THEIR PRECURSORS
1.2.1 Descriptions and Properties of Ozone and OtherPhotochemical Oxidants
' Ozone (03) and other photochemical oxidants occurring at low concentra-
tions in ambient air, such as peroxyacetyl nitrate (PAN), hydrogen peroxide
(HpOp), and formic acid (HCOOH), are characterized chiefly by their ability to
remove electrons from, or. to share electrons with, other molecules or ions
(i.e., oxidation). The capability of a chemical species for oxidizing or
reducing other chemical species is termed "redox potential" (positive or
negative standard potential) and is expressed in volts. A reactive allotrope
of oxygen that is only about one-tenth as soluble as oxygen in water, ozone
has a standard potential of +2.07 volts in aqueous systems for the redox pair,
03/H20 (Weast, 1977). Hydrogen peroxide, which is highly soluble in water and
other polar solvents, has a standard potential of +1.776 in the redox pair,
H2Q2/H2Q (Weast, 1977). No standard potential for peroxyacetyl nitrate in
neutral or buffered aqueous systems, such as those that occur in biological
1-3
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systems, appears in the literature. In acidic solution (pH 5 to 6), PAN
hydrolyzes fairly rapidly (Lee et al., 1983; Holdren et al., 1984); in alkaline
solution it decomposes with the production of nitrite ion and molecular oxygen
(Stephens, 1967; Nicksic et al., 1967). An important property.of PAN, especi-
ally in the laboratory, is its thermal instability. Its explosiveness dic-
tates its synthesis for experimental and calibration purposes by experienced
personnel only.
Formic acid is formed as a stable product in photochemical air pollution.
It has the structure of both an acid and an aldehyde and in concentrated form
is a pungent-smelling, highly corrosive liquid.
The toxic effects of oxidants are attributable to their oxidizing abili-
ty. Their oxidizing properties also form the basis of several measurement
techniques for 03 and PAN. The calibration of ozone and PAN measurements,
however, is achieved via their spectra in the ultraviolet and infrared regions,
respectively. All three pollutants of most concern in this document (0-,, PAN,
and HpOp) must be generated j_n situ for the calibration of measurement tech-
niques. For ozone and H?®?' generation of calibration gases is reasonably
straightforward.
1.2.2 Nature of Precursors to Ozone and Other Photochemical Oxidants
Photochemical oxidants are products of atmospheric reactions involving
volatile organic compounds (VOC) and oxides of nitrogen (NO ), as well as
/\
hydroxyl (OH) and other radicals, oxygen, and sunlight (see, e.g., Demerjian
et al., 1974; National Research Council, 1977; U.S. Environmental Protection
Agency, 1978; Atkinson, 1985). The oxidants are largely secondary pollutants
formed in the atmosphere from their precursors by processes that are a complex,
non-linear function of precursor emissions and meteorological factors.
The properties of organic compounds that are most relevant to their role
as precursors to ozone and other oxidants are their volatility, which governs
their emissions into the atmosphere; and their chemical reactivity, which
determines their lifetime in the atmosphere. Although vapor-phase hydrocar-
bons (compounds of carbon and hydrogen only) are the predominant organic
compounds in ambient air that serve as precursors to photochemical oxidants,
other volatile organic compounds are also photochemically reactive in those
atmospheric processes that give rise to oxidants. In particular, halogenated
organics (e.g., haloalkenes) that participate in photochemical reactions are
1-4
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present in ambient air, although at lower concentrations than the hydrocarbons.
They are oxidized through the same initial step involved in the oxidation of
the hydrocarbons; that is, attack by hydroxyl radicals. Alkenes, haloalkenes,
and aliphatic aldehydes are, as classes, among the most reactive organic.
compounds found in ambient air (e.g., Altshuller and Bufalini, 1971; Darnall
et al., 1976; Pitts et a!., 1977; U.S. Environmental Protection Agency, 1978,
and references therein). Alkenes and haloalkenes are unique among VOC in
ambient air in that they are susceptible both to attack by OH radicals (OH)
and by ozone (Niki et al., 1983). Methane, halomethanes, and certain haloe-
thenes are of negligible reactivity in ambient air and have been classed as
unreactive by the U.S. Environmental Protection Agency (1980a,b). Since methane
is considered only negligibly reactive in ambient air, the volatile organic
compounds of importance as oxidant precursors are usually referred to as
nonmethane hydrocarbons (NMHC) or, more properly, as nonmethane organic
compounds (NMOC).
The oxides of nitrogen that are important as precursors to ozone and
other photochemical oxidants are nitrogen dioxide (N0?) and nitric oxide (NO).
Nitrogen dioxide is itself an oxidant that produces deleterious effects, which
are the subject of a separate criteria document (U.S. Environmental Protection
Agency, 1982). Nitrogen dioxide is an important precursor (1) because its
photolysis in ambient air leads to the formation of oxygen atoms that combine
with molecular oxygen to form ozone; and (2) because it reacts with acetyl-
peroxy radicals to form peroxyacetyl nitrate, a phytotoxicant and a lachryma-
tor. Although ubiquitous, nitrous oxide (N-O) is unimportant in the production
of oxidants in ambient air because it is virtually inert in the troposphere.
1.2,3 Atmospheric Reactio.ns of Ozone and Other Oxidants Including Their Role
in Aerosol Formation
The chemistry of the polluted atmosphere is exceedingly complex, but an
understanding of the basic phenomena is not difficult to acquire. Three
processes occur: the emission of precursors to ozone from predominantly
manmade sources; photochemical reactions that take place during the disper-
sion and transport of these precursors; and scavenging processes that reduce
the concentrations of both 0^ and precursors along the trajectory.
The specific atmospheric reactions of ozone and of other photochemical
oxidants such as peroxyacetyl nitrate and hydrogen peroxide are becoming
1-5
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increasingly well-character!zed. The reactions of these species result in
products and processes that may have significant environmental and health- and
welfare-related implications, including effects on biological systems, nonbio-
logical materials, and such phenomena as visibility degradation and acidifica-
tion of cloud and rain water.
1.2.3.1 Formationand Transformation of Ozone and OtherPhotochemical Oxi-
dants. In the troposphere, ozone is formed through the dissociation of N02 by
sunlight to yield an oxygen atom, which then reacts with molecular oxygen (0?)
to produce an 0, molecule. If it is present, NO can react rapidly with 03 to
form NOp and an 02 molecule. In the absence of competing reactions, a steady-
state or equilibrium concentration of 03 is soon established between 03, N02,
and NO (National Research Council, 1977). The injection of organic compounds
into the atmosphere upsets the equilibrium and allows the ozone to accumulate
at higher than steady-state concentrations. The length of the induction
period before the accumulation of 03 begins depends heavily on the initial
N0/N09 and NMOC/NOV ratios (National Research Council, 1977).
£m /\
The major role played by organic compounds in smog reactions is attribut-
able to the hydroxyl radical (OH), since it reacts with essentially all organic
compounds (e.g., Atkinson, 1985; Herron and Huie, 1977, 1978; Dodge and Arnts,
1979; Niki et a!., 1981). Aldehydes, which are constituents of automobile
exhaust as well as decomposition products of most atmospheric photochemical
reactions involving hydrocarbons, and nitrous acid (HONO), are important
sources of OH radicals, as is 0~ itself. Other free radicals, such as hydro-
and alkylperoxy radicals and the nitrate (NO,,) radical play important roles in
photochemical air pollution.
The presence of organic compounds, oxides of nitrogen, and sunlight does
not mean that the photochemical reactions will continue indefinitely. Termi-
nation reactions gradually remove NOp from the reaction mixtures, such that
the photochemical cycles slowly come to an end unless fresh NO and N02 emis-
sions are injected into the atmosphere. Compounds containing nitrogen, such
as PAN, nitric acid (HN03), and peroxynitric acid (HWL), as well as organic
and inorganic nitrates, are formed in these termination reactions.
Recent studies on the photooxidation of organic compounds under simulated
atmospheric conditions have been reasonably successful. The rate constants
for the reaction of OH radicals with a large number of organic compounds have
been measured (e.g., Atkinson et a!., 1979; Atkinson et al., 1985). The
1-6
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mechanisms of the reactions of paraffinic compounds are fairly well under-
stood, as are those of olefinic compounds, at least for the smaller compounds.
Photooxidation reactions of the aromatic compounds, however, are poorly under-
stood.
In the presence of NO , natural hydrocarbons (i.e., those organic com-
x\
pounds emitted from vegetation) can also undergo photooxidation reactions to
yield 0~, although most naturally emitted hydrocarbons are olefins and are
scavengers as well as producers of 0~ (e.g., Lloyd et al., 1983;.Atkinson
et al., 1979; Kamens" et al., 1982; Killus and Whitten, 1984; Atkinson and
Carter, 1984),
1.2.3.2 Atmospheric Chemical Processes Involving Ozone. Ozone can react with
organic compounds in the boundary layer of the troposphere (Atkinson and
Carter, 1984), It is important to recognize, however, that organics undergo
competing reactions with OH radicals in the daytime (Atkinson et al., 1979;
Atkinson, 1985) and, in certain cases, with N03 radicals during the night
(Japar and Niki, 1975; Carter et al. , 1981a; Atkinson et al., 1984a,b,c,d,e;
Winer et al., 1984), as well as photolysis, in the case of aldehydes and other
oxygenated organics. Only for organics whose ozone reaction rate constants
*5™f *3 "1 "1
are greater than -10 cm molecule sec can consumption by ozone be
considered to be atmospherically important (Atkinson and Carter, 1984),
Ozone reacts rapidly with the acyclic mono-, di-, and tri-alkenes and
— 1 Q
with cyclic alkenes. The rate constants for these reactions range from ~10
—14. 3 —i "~"i
to ~10 cm molecule sec (Atkinson and Carter, 1984), corresponding to
atmospheric lifetimes ranging from a few minutes for the more reactive cyclic
alkenes, such as the monoterpenes, to several days. In polluted atmospheres,
a significant portion of .the consumption of the more reactive alkenes will
occur via reaction with ozone rather than with OH radicals, especially in the
afternoons during photochemical oxidant episodes. Reactions between ozone and
alkenes can result in aerosol formation (National Research Council, 1977;
Schuetzle and Rasmussen, 1978), with alkenes of higher carbon numbers the
chief contributors.
Because of their respective rate constants, neither alkanes (Atkinson and
Carter, 1984) nor alkynes (Atkinson and Aschmann, 1984) are expected to react
with ozone in the atmosphere, since competing reactions with OH radicals have
higher rate constants (Atkinson et al., 1979; Atkinson, 1985).
1-7
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The aromatics react with ozone, but quite slowly (Atkinson and Carter,
1984), such that their reactions with ozone are expected to be unimportant in
the atmosphere. Cresols are more reactive toward ozone than the aromatic
hydrocarbons (Atkinson and Carter, 1984), but their reactions with OH radicals
(Atkinson, 1985) or NCL radicals (Carter et a!., 1981a; Atkinson et al.,
1984d) predominate.
For oxygen-containing organic compounds, especially those without carbon-
carbon double bonds, reactions with ozone are slow. For carbonyls and ethers
(other than ketene) that contain unsaturated carbon-carbon bonds, however,
much faster reactions are observed (Atkinson and Carter, 1984).
Certain reactions of ozone other than its reactions with organic com-
pounds are important in the atmosphere. Ozone reacts rapidly with NO to form
NOg, and subsequently with N02 to produce the nitrate (NO,) radical and an
oxygen molecule. Photolysis of ozone can be a significant pathway for the
formation of OH radicals, particularly in polluted atmospheres when ozone
concentrations are at their peak.
Ozone may play a role in the oxidation of S02 to H2$Q4, both indirectly
in the gas phase (via formation of OH radicals and Criegee biradicals) and
directly in aqueous droplets.
1.2.3.3 Atmospheric Reactions of PAN, Hg02, and HCOQH. Because PAN is in
equilibrium with acetyl peroxy radicals and N02, any process that leads to the
removal of either of these species will lead to the decomposition of PAN. One
such process is the reaction of NO with acetyl peroxy radicals. This can
lead, however, to the formation of OH radicals. Thus, PAN remaining overnight
from an episode on the previous day can react with NO emitted from morning
traffic to produce OH radicals (Cox and Roffey, 1977; Carter et al., 1981b)
that will enhance smog formation on that day (e.g., Tuazon et al., 1981). In
the absence of significant NO concentrations, and in regions of moderate to
lower temperatures, PAN will persist-, in the atmosphere (Wallington et al.,
1984; Aikin et al., 1983) and contribute to the long-range transport of NO .
! " X
Although hydrogen,( peroxide;formed in Ithe^gas phase from the reactions of
hydroperoxyl radicals plays ,a role-in HOY chemistry in the,; troposphere, and
* - ' y^
especially in the stratosphere (Crutzen ;and Fishman, 1977; Cox and Burrows,
1979), its major importance arises from its high solubility in water. The
latter ensures that a .large fraction of gaseous H202 will be.taken up in
aqueous droplets. Over the past decade, evidence has accumulated that H,
1-8
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dissolved in cloud, fog, and rainwater may play an important, and, in acidic
droplets (i.e., pH £5), even a dominant role in the oxidation of S02 to H^SO^
(e.g., Hoffman and Edwards, 1975; Martin and Damschen, 1981; Chameides and
Davis, 1982; Calvert and Stockwell, 1983, 1984; Schwartz, 1984). Hydrogen
peroxide may also play a role in the oxidation of N0? dissolved in aqueous
droplets, although relevant data are limited and additional research is required
(see, e.g., Gertler et a!., 1984). Substantial uncertainties remain concerning
the quantitative role of H^O^ in acidification of aqueous particles and droplets
(Richards et al., 1983).
Because it can be scavenged rapidly into water droplets, formic acid can
potentially function as an oxidant in cloud water and rain water. Thus, HCOOH
is an example of a compound that is a non-oxidant or weak oxidant in the gas
phase but that is transformed upon incorporation in aqueous solutions into an
effective oxidizer of S(IV). Although much uncertainty remains concerning the
quantitative role of HCOOH and the higher organic acids, they, potentially play
a minor but still significant role in the acidification of rain.
1.2.4 Meteorological and Climatological Processes
Meteorological and climatological processes are important in determining
the extent to which precursors to ozone and other photochemical oxidants can
accumulate, and thereby the concentrations of ozone and other oxidants that
can result. The meteorological factors most important in the formation and
transport of ozone and other photochemical oxidants in the lower troposphere
are: (1) degree of atmospheric stability; (2) wind speed and direction;
(3) intensity and wavelength of sunlight; and (4) synoptic weather conditions.
These factors are in turn dependent upon or interrelated with geographic,
seasonal, and other climatological factors.
Incursions of ozone from the stratosphere are an additional source of the
ozone found in the lower troposphere. The physical and meteorological mechanisms
by which ozone is brought into the troposphere from the stratosphere are
important in determining the resulting ground-level concentrations, ground-level
locations impacted,- and the seasonality of incursions of stratospheric ozonek
1.2.4.1 Atmospheric Mixing. The concentration of a pollutant in ambient air
depends significantly on the degree of atmospheric mixing that occurs from the
time the pollutant or its precursors are emitted and the arrival of the pollu-
tant at the receptor. The rate at which atmospheric mixing proceeds and the
1-9
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extent of the final dilution depends on the amount of turbulent mixing that
occurs and on wind speed and direction. Atmospheric stability is one of the
chief determinants of turbulent mixing since pollutants do not spread rapidly
within stable layers nor do they mix upward through stable layers to higher
altitudes.
Temperature inversions, in which the temperature increases with increasing
altitude, represent the most stable atmospheric conditions. Surface inver-
sions (base at ground level) and elevated inversions (the entire layer is
above the surface) are both common (Hosier, 1961; Holzworth, 1964) and both
can occur simultaneously at the same location. Surface inversions show a
diurnal pattern, forming at night in the absence of solar radiation but break-
ing up by about mid-morning as the result of surface heating by the sun (Hosier,
1961; Slade, 1968). Elevated inversions can persist throughout the day and
pollutants can be trapped between the ground surface and the base of the
inversion. The persistence of elevated inversions is a major meteorological
.factor contributing to high pollutant concentrations and photochemical smog
conditions along the California coast (Hosier, 1961; Holzworth, 1964; Robinson,
1952). In coastal areas generally, such as the New England coast (Hosier,
1961) and along the Great Lakes (Lyons and Olsson, 1972), increased atmospheric
stability (and diminished mixing) occurs in summer and fall as the result of
the temperature differential between the water and the land mass.
The depth of the layer in which turbulent mixing can occur (i.e., the
"mixing height") shows geographical dependence. Summer morning mixing heights
are usually >300 m in the United States except for the Great Basin (part of
Oregon, Idaho, Utah, Arizona, and most of Nevada), where the mixing height is
~200 m (Holzworth, 1972). By mid-morning, mixing heights increase markedly
such that only a few coastal areas have mixing heights <1000 m.
Summer afternoon mixing heights are generally an indication of the poten-
tial for recurring photochemical oxidant problems. Photochemical smog problems
in the United States are somewhat unexpected since the lowest afternoon mixing
height is ~600 m (Holzworth, 1972). Elevated inversions having bases <500 m
(i.e., low-level inversions) occur in the United States, however, with the
following frequencies: 90 percent on the California coast; >20 percent on the
Atlantic coast (New Jersey to Maine); >5 percent along the Great Lakes; and 5
to 10 percent from Louisiana to Arkansas and eastward to about Atlanta, Georgia.
For most areas of the United States, though, the persistence through the
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afternoon of low-level stable layers is a rare event, occurring on <1 day in
20 (Holzworth and Fisher, 1979).
1.2.4.2 Wind Speed and Direction. For areas in which mixing heights are not
restrictive, wind speed and, in some cases, wind direction are major determi-
nants of pollution potential. Since strong winds dilute precursors to ozone
and other photochemical oxidants, a location may have good ventilation despite
the occurrence of persistent inversions (e.g., San Francisco). Conversely,
light winds can result in high oxidant levels even if the mixing layer is
deep.
The frequency of weak winds, then, is important in oxidant formation. In
industrialized, inland areas east of the Mississippi River, surface inversions
in the morning coupled with wind speeds £2.5 m/sec ( mi/hr) occur with a
frequency :>50 percent (Holzworth and Fisher, 1979). These surface inversions
break up by afternoon, however, permitting dispersion.
The effects of wind speed and direction include the amount of dilution
occurring in the source areas, as well as along the.trajectory followed by an
urban or source-area plume. Regions having steady prevailing winds, such that
a given air parcel can pass over a number of significant source areas, can
develop significant levels of pollutants even in the absence of weather patterns
that lead to the stagnation type of air pollution episodes. The Northeast
states are highly susceptible to pollutant plume transport effects, although
some notable stagnation episodes have also affected this area (e.g., Lynn
et al.j 1964). Along the Pacific Coast, especially along the coast of
California, coastal winds and a persistent low inversion layer contribute to
major pollutant buildups in urban source areas and downwind along the urban
plume trajectory (Robinson, 1952; Neiburger et a!., 1961).
1.2.4.3 Effects of Sunlight and Temperature. The effects of sunlight on
photochemical oxidant formation, aside from the role of solar radiation in
meteorological processes, are related to its intensity and its spectral dis-
tribution. Intensity varies diurnally, seasonally, and with latitude, but the
effect of latitude is strong only in the winter. Experimental studies have
verified the effects on oxidant formation of Tight intensity (Peterson, 1976;
Demerjian et a!., 1980) and its diurnal variations (Jeffries et a!., 1975;
1976), as well as on the overall photooxidation process (Jaffee et a!., 1974;
Winer et al., 1979).
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A correlation between high oxidant concentrations and warm, above-normal
temperatures has been demonstrated generally (Bach, 1975; Wolff and Lioy,
1978) and for specific locations, e.g., St. Louis (Shreffler and Evans, 1982).
Coincident meteorology appears to be the cause of the observed correlation.
Certain synoptic weather conditions are favorable both for the occurrence of
higher temperatures and for the formation of ozone and other oxidants, so that
temperature is often used to forecast the potential for high oxidant concen-
trations (e.g., Wolff and Lioy, 1978; Shreffler and Evans, 1982). Data from
smog chamber studies show an effect of temperature on ozone formation (e.g.,
Carter et al., 1979; Countess et al., 1981), but the effect is thought to
result from the volatilization and reaction of chamber wall contaminants as
the temperature is increased.
1.2.4.4 Transport of Ozone and Other Oxidants and Their Precursors. The
levels of ozone and other oxidants that will occur at a given receptor site
downwind of a precursor source area depend upon many interrelated factors,
which include but are not restricted to: (1) the concentrations of respective
precursors leaving the source area; (2) induction time; (3) turbulent mixing;
(4) wind speed and wind direction; (5) scavenging during transport; (6) in-
jection of new emissions from source areas in the trajectory of the air mass;
and (7) local and synoptic weather conditions.
Ozone and other photochemical oxidants can be transported hundreds of
miles from the place of origin of their precursors, as documented by the
numerous studies on transport phenomena that were described in the 1978 cri-
teria document for ozone and other photochemical oxidants (U.S. Environmental
Protection Agency, 1978). In that document, transport phenomena were classi-
fied into three categories, depending upon transport distance: urban-scale,
mesoscale, and synoptic-scale. In urban-scale transport, maximum concentra-
tions of 0, are produced about 20 miles or so (and about 2 to 3 hours) down-
wind from the major pollutant source areas. In mesoscale transport, 0, has
been observed up to 200 miles downwind from the sources of its precursors.
Synoptic-scale transport is associated with large-scale, high-pressure air
masses that may extend over and persist for many hundreds of miles.
Urban-scale transport has been identified as a significant, characteris-
tic feature of the oxidant problem in the Los Angeles Basin (Tiao et al.,
1975), as well as in San Franciso, New York, Houston, Phoenix, and St. Louis
(e.g., Altshuller, 1975; Coffey and Stasiuk, 1975; Shreffler and Evans, 1982;
1-12
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Wolff et al., 1977a). Simple advection of a photochemically reactive air
mass, local wind patterns, and diurnal wind cycles appear to be the main
factors involved in urban-scale transport,
Mesoscale transport is in many respects an extension of urban-scale
transport and is characterized by the development of urban plumes. Bell
documented cases in 1959 in which precursors from the Los Angeles Basin and
the resultant oxidant plume were transported over the coastal Pacific Ocean,
producing elevated oxidant concentrations in San Diego County the next day
(Bell, 1960). Similar scales of transport have been reported by Cleveland
et al. (1976a,b) for the New York-Connecticut area; by Wolff and coworkers and
others (Wolff et al., 1977a,b; Wolff and Lioy, 1978; Clark and Clarke, 1982;
Clarke et al., 1982; Vaughan et al., 1982) for the Washington, DC-Boston
corridor; and by Westberg and coworkers for the Chicago-Great Lakes area
(Sexton and Westberg, 1980; Westberg et al., 1981). These and other studies
have demonstrated that ozone-oxidant plumes from major urban areas can extend
downwind about 100 to 200 miles and can have widths of tens of miles (Sexton,
1982), frequently up to half the length of the plume.
Synoptic-scale transport is characterized by the general and widespread
occurrence of elevated oxidants and precursors on a regional or air-mass scale
as the result of certain favorable weather conditions, notably, slow-moving,
well-developed high-pressure, or anti-cyclonic, systems characterized by weak
winds and limited vertical mixing (Korshover, 1967; 1975). The size of the
region that can be affected has been described by Wolff and coworkers, who
reported the occurrence of haze and elevated ozone levels in an area extending
from the Midwest to the Gulf Coast (Wolff et al., 1982) and the occurrence of
elevated ozone concentrations extending in a virtual "ozone river" from the
Gulf Coast to New England that affected anywhere from a few hundred square
miles to a thousand square miles during a 1-week period in July 1977 (Wolff
and Lioy, 1980).
1.2.4.5 Stratospheric-Tropospheric Ozone Exchange. The fact that ozone is
formed in the stratosphere, mixed downward, and incorporated into the tropo-
sphere, where it forms a more or less uniformly mixed background concentra-
tion, has been known in various degrees of detail for many years (Junge,
1963). It is widely accepted that the long-term average tropospheric back-
ground concentration of about 30 ppb to 50 ppb results primarily, though not
1-13
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exclusively, from the transfer of stratospheric ozone into the upper tropo-
sphere, followed by subsequent dispersion throughout the troposphere (e.g.,
Kelly et al., 1982).
The exchange of ozone between the stratosphere and the troposphere in the
middle latitudes occurs to a major extent in events called "tropopause folds"
(TF) (Reiter, 1963; Reiter and Mahlman, 1965; Danielsen, 1968; Reiter, 1975;
Danielsen and Mohnen, 1977; Danielsen, 1980), in which the polar jet stream
plays a major role. From recent studies, Johnson and Viezee (1981) proposed
four types or mechanisms of TF injection and concluded that two of these, both
of which are consistent with theory, could cause substantial effects in terms
of high ozone concentrations at ground level. They concluded, in addition,
that all low-pressure trough systems may induce TF events and cause the trans-
tropopause movement of ozone-rich air into the troposphere (Johnson and Viezee,
1981).
1.2.4.6 Stratospheric Ozone at Ground Level. From a detailed review of
studies on background tropospheric ozone, Viezee and Singh (1982) concluded
that the stratosphere is a major but not the sole source of background ozone
in the unpolluted troposphere, a conclusion reached by other investigators as
well (e.g., Kelly et al., 1982). The stratospheric ozone reservoir shows a
strong seasonal cycle that is reflected at ground-level. At some stations
that monitor background ozone levels, average spring background levels may be
as high as 80 ppb, with average fall levels ranging from 20 to 40 ppb (e.g.,
Singh et al., 1977; Mohnen, 1977; U.S. Environmental Protection Agency, 1978).
Viezee and Singh (1982) and Viezee et al. (1983) concluded that relatively
high ozone concentrations can occur for short periods of time (minutes to a
few hours) over local areas as a result of stratospheric intrusions.
A number of investigators have attempted to quantify the amount of the
surface ozone that can be attributed to stratospheric sources. The method
most commonly used is based on the assumption that beryllium-7 ( Be) is a
unique tracer for air parcels of stratospheric origin. Calculated correlations
between surface ozone and Be show, however, that their relationship is highly
variable (e.g., Kelly et al., 1982; Ferman and Monson^ 1978; Johnson and
Viezee, 1981; Husain et al., 1977). Singh et al. (1980) and Viezee and Singh
(1982) have pointed out problems with using this technique to quantify the
contribution of stratospheric ozone to surface ozone. Singh et al. (1980)
1-14
-------
concluded that "the experimental technique involving a Be/0- ratio to esti-
mate the daily stratospheric component of ground-level Q~ is unverified and
considered to be inadequate for air quality applications" (p. 1009). This
group of investigators have suggested, however, that Be may be used, under
appropriate meteorological conditions, as a qualitative tracer for air masses
of stratospheric origin (Johnson and Viezee, 1981; Viezee et a!., 1979).
Other methods used to attempt to quantify the stratospheric component of
surface ozone include aircraft observations of TF events coupled with calcula-
tions of downward ozone flux, and examination of surface ozone data records.
From such data, Viezee et al. (1983) concluded that direct ground-level contri-
butions from stratospheric ozone are infrequent (<1 percent of the time),
short-lived, and associated with ozone concentrations <0.1 ppm.
Notwithstanding difficulties with quantifying its contribution to surface
ozone, however, stratospheric ozone is clearly present in atmospheric surface
layers, and the meteorological mechanisms responsible have been described by a
number of investigators (e.g., Danielsen, 1968; Wolff et al., 1979; Johnson
and Viezee, 1981).
1.2.4.7 Background Ozone from Photochemical Reactions. Whereas stratospheric
ozone is thought by many investigators to be the dominant contributor to
background levels of ozone, as discussed above, other investigators have
concluded that as much as two-thirds of the annual average background concen-
trations may result from photochemical reactions. For example, Altshuller
(1986), in a recent review article, has concluded that photochemically generated
ozone should equal or exceed the stratospheric contribution at lower-elevation
remote locations; and that photochemically generated ozone from manmade emissions
probably constitutes most of the ozone measured at more polluted rural locations
during the warmest months of the year. His conclusions were based, in part,
on an analysis of global circulation (e.g., Levy et al., 1985) and photochemical
modeling studies (e.g., Fishman and Seiler, 1983; Fishman and Carney, 1984;
Fishman et al., 1985; Dignon and Hameed, 1985). In these modeling studies,
the photochemical contribution to background ozone levels was estimated to
range from ~15 ppb (long-term) to ~80 ppb (summertime), depending on the level
of NO emissions assumed.
/\
Studies on the role of NO in nonurban ozone photochemistry have shown
/\
that ozone formation at many of the locations is not NO -limited, but depends
s\
on VOC reactions, as well (e.g., Martinez and Singh, 1979; Kelly et al., 1984;
1-15
-------
Liu et al., 1984). Background NO concentrations at most remote, clean "loca-
J\
tions range from <0.05 ppb upward. Mean concentrations of NO at nonurban
J\
locations in the United States east of the Rocky Mountains range from ~1 ppb
to 10 ppb (Altshuller, 1986; see also Section 1,2.5.2.4 and Chapter 3). These
background concentrations of NO are higher than previously thought (see,
e.g., Singh et al., 1980; Kelly et al., 1984, regarding global models and
assumed reservoirs of NO-)-
X
The contributions of biogenic VOC to background ozone, although a matter
of controversy in recent years, appear not to be significant under most atmos-
pheric conditions, since ambient air concentrations of biogenic VOC are quite
low, even at rural sites (Altshuller, 1983).
Thus, photochemistry and stratospheric intrusions are both regarded as
contributing to background ozone concentrations, but the apportionment of
background to respective sources remains a matter of investigation.
1.2.5 Sources, Emissions, and Concentrations of Precursors to Ozone and Other
Photochemical Oxidants
As noted earlier, photochemical production of ozone depends both on the
presence of precursors, volatile organic compounds (VOCs) and nitrogen oxides
(NOV), emitted by manmade and by natural sources; and on suitable conditions
-------
biogenic emissions of organic compounds in the United States are highly
inferential but data suggest that the yearly rate is the same order of magni-
tude as manmade emissions. Most of the biogenic emissions actually occur
during the growing season, however, and the kinds of compounds emitted are
different from those arising from manmade sources.
Emissions of manmade NO in the United States were estimated at 19.4 Tg/yr
x>.
for 1983. Retrospective estimates show that manmade NO emissions rose from
x>.
about 6.8 Tg/yr in 1940 to about 18.1 Tg/yr in 1970 (U.S. Environmental Protec-
tion Agency, 1986). Annual emissions of manmade NO were some 12 percent
s\
higher in 1983 than in 1970, but the rate leveled off in the late 1970s and
exhibited a small decline from about 1980 through 1982 (U.S. Environmental
Protection Agency, 1984). The increase over the period 1970 through 1983 had
two main causes: (1) increased fuel combustion in stationary sources such as
power plants; and (2) increased fuel combustion in highway motor vehicles, as
the result of the increase in vehicle miles driven. Total vehicle miles
driven increased by 42 percent over the 14 years in question.
Estimated biogenic NO emissions are based on uncertain extrapolations
x>.
from very limited studies, but appear to be about an order of magnitude less
than manmade NO emissions.
s\
1.2.5.2 Representative Concentrations in Ambient Air.
1.2.5.2.1 Hydrocarbons in urban areas. Most of the available ambient air
data on the concentrations of nonmethane hydrocarbons (NMHC) in urban areas
have been obtained during the 6:00 to 9:00 a.m. period. Since hydrocarbon
emissions are at their peak during that period of the day, and since atmospheric
dispersion is limited that early in the morning, NMHC concentrations measured
then generally reflect maximum diurnal levels. Representative data for urban
areas show mean NMHC concentrations between 0.4 and 0.9 ppm.
The hydrocarbon composition of urban atmospheres is dominated by species
in the C~ to C-.Q molecular-weight range. The paraffinic hydrocarbons (alkanes)
are most prominent, followed by aromatics and alkenes. Based on speciation
data obtained in a number of urban areas, alkanes generally constitute 50 to
60 percent of the hydrocarbon burden in ambient air, aromatics 20 to 30 percent,
with alkenes and acetylene making up the remaining 5 to 15 percent (Sexton and
Westberg, 1984).
1.2.5.2.2 Hydrocarbons in nonurban areas. Rural nonmethane hydrocarbon
concentrations are usually one to two orders of magnitude lower than those
1-17
-------
measured in urban areas (Ferman, 1981; Sexton and Westberg, 1984). In samples
from sites carefully selected to guarantee their rural character, total NMHC
concentrations ranged from 0.006 to 0.150 ppm C (e.g., Cronn, 1982; Seila,
1981; Holdren et al.} 1979). Concentrations of individual species seldom
exceeded 0.010 ppm C. The bulk of species present in rural areas are alkanes;
ethane, propane, ri-butane, iso-pentane, and ri-pentane are most abundant.
Ethylene and propene are sometimes present at £0.001 ppm C, and toluene is
usually present at ~0.001 ppm C. Monoterpene concentrations are usually
£0.020 ppm C. During the summer months, isoprene concentrations as high as
0.150 ppm C have been measured (Ferman, 1981), The maximum concentrations of
isoprene usually encountered, however, are in the range of 0,030 to 0.040
ppm C.
1.2.5.2.3 Nitrogen oxides in urban areas. Concentrations of NO , like hydro-
Ps,
carbon concentrations, tend to peak in urban areas during the early morning,
when atmospheric dispersion is limited and automobile traffic is dense. Most
NO is emitted as nitric oxide (NO), but the NO is rapidly converted to N0~,
/> £-
initially by thermal oxidation and subsequently by ozone and peroxy radicals
produced in atmospheric photochemical reactions. The relative concentrations
of NO versus N02 fluctuate day-to-day, depending on diurnal and day-to-day
fluctuations in ozone levels and photochemical activity.
Urban NO concentrations during the 6:00 to 9:00 a.m. period in 10 cities
/tr\
ranged from 0.05 to 0.15 ppm in studies done in the last 5 to 7 years (e.g.,
Westberg and Lamb, 1983; Richter, 1983; Eaton et a!., 1979), although concen-
trations two to three times higher occur in cities such as Los Angeles.
Concurrent NMHC measurements for these 10 cities showed that NMHC/NO ratios
J\
ranged from 5 to 16.
1.2.5,2.4 Nitrogen oxides in nonurban areas. Concentrations of NO in clean
'::::::::::::::--": "•'• '"" -»----^-"'==-::- •:--::.•::-..:. ^
remote environments are usually <0.5 ppb (Logan, 1983). In exceptionally
clean air, NO concentrations as low as 0.015 ppb have been recorded (Bellinger
*>
et al., 1982). Concentrations of NO at nonurban sites in the northeastern
y\
United States appear to be higher than NO concentrations in the west by a
J\
factor of ten (Mueller and Hidy, 1983). From the limited amount of data
available, NO concentrations in unpopulated nonurban areas in the west average
f{
<1 ppb; but in nonurban northeastern areas average NO can exceed 10 ppb.
*"""
1-18
-------
1.2,6 Source-Receptor (Oxldant-Precursor) Models
In order to apply knowledge of the atmospheric chemistry of precursors,
and of ozone and other photochemical oxidants, during their dispersion and
transport, models describing these phenomena have been developed in a variety
of forms over the past 15 years. Most of these models relate the rates of
precursor emissions from mobile and stationary sources, or precursor atmos-
pheric concentrations, to the resulting ambient concentrations of secondary
pollutants that impact receptors at downwind sites. For this reason they have
been described as source-receptor models.
Current air quality, or source-receptor, models can be classified as
either statistical or computational-dynamic. Statistical models are generally
based on a statistical analysis of historical air quality data, and are not
explicitly concerned with atmospheric chemistry or meteorology. An example of
statistical models is the linear rollback concept.
Computational, or dynamic, models attempt to describe mathematically the
atmospheric chemical and physical processes influencing air pollution forma-
tion and impacts. Examples of computational models include trajectory and
fixed-grid airshed models. Two phenomenologically different approaches have
been employed in dynamic models with respect to the coordinate systems chosen.
A coordinate system fixed with respect to the earth is termed Eulerian, while
in Lagrangian models the reference frame moves with the air parcel whose
behavior is being simulated. ,
1.2.6.1 Trajectory Models. In trajectory models, a moving-coordinate system
describes pollutant transport as influenced by local meteorological conditions.
Trajectory models provide dynamic descriptions of atmospheric source-receptor
relationships that are simpler and less expensive to derive than those obtained
from fixed-cell models. , . ,
The simplest form of, trajectory model is the empirical kinetic modeling
approach (EKMA), This approach was developed from earlier efforts (Dimitriades,
1972) to use smog chamber data.to develop graphical relationships between
morning NMOC and NO levels and afternoon-maximum concentrations of ozone. In
}\
applying EKMA, the Ozone Isopleth Plotting Package (OZIPP) (Whitten and Hogos
1978) is used to generate ozone isop!eths at .various levels of sophistication
corresponding to "standard" EKMA, "city-specific" EKMA, or the simplified
trajectory model (F.R., 1979). The EKMA isopleths generated are used to
determine the relative degree of control of precursor emissions needed to
achieve a given percentage reduction in ozone.
1-19
-------
The use of EKMA in ozone abatement programs is relatively widespread. It
is therefore worth noting the general control implications of EKMA isopleths.
For areas with high levels of morning precursor emissions and meteorology
conducive to oxidant formation, such as Los Angeles, for example, EKMA isopleths
predict that (1) at high NMOC/NO concentration ratios, reductions in NO will
s\ /\
decrease ozone formation; (2) at moderate NMOC/NO ratios, reductions in NMOC
/\
and NO will decrease ozone formation; and (3) at very low NMOC/NO ratios,
/\ /\
increases in NO will inhibit ozone formation. These predictions cannot be
/\
assumed to apply to all urban areas, or even to all high-oxidant urban areas,
since the shape of the EKMA isopleths is a function of numerous factors, many
of which are location-specific. For discussions of the specific assumptions
employed in EKMA and the underlying chemistry and meteorology, the primary
literature should be consulted (e.g., Dimitriades, 1970, 1972, 1977a,b; Dodge,
1977a,b; Whitten and Hogo, 1977; U.S. Environmental Protection Agency, 1977,
1978; Whitten, 1983). Likewise, the primary literature should be consulted
for additional data and discussions on the respective effects on ozone forma-
tion of controlling NMHC and NO (e.g., Liu and Grisinger, 1981; Chock et al.,
/\
1981; Kelly, 1985; Kelly et al., 1986; Glasson and Tuesday, 1970; Dimitriades,
1970, 1972, 1977a,b).
1.2.6.2 Fixed-Grid Models. Fixed-grid models, also called regional airshed
models, are based on two- or three-dimensional arrays of grid cells and are
the most sophisticated source-receptor models presently available. Such
models are computationally complex and require the most extensive set of input
data; but they also provide the most realistic treatment of the various processes
involved in photochemical air pollution formation.
1.2.6.3 Box Models. Box models (Hanna, 1973; Demerjian and Schere, 1979;
Derwent and Hov, 1980) are the simplest of dynamic, models. They treat the
atmosphere as a single cell, bounded by the mixing layer, having an area on
the order of 100 square miles.
1.2.6.4 Validation and Sensitivity Analyses for Dynamic Models. All present-
ly available source-receptor models require a degree of simplifying assump-
tions to deal with practical limitations imposed by existing computer capabil-
ities, time and cost constraints, or lack of knowledge concerning inputs such
as boundary conditions, emissions, or detailed reaction mechanisms. The
reliability and applicability of any particular model therefore depends upon
its specific limitations, data requirements, and degree of validation against
1-20
-------
experimental data from ambient air measurements or environmental chamber runs.
Reliability and applicability also depend on the quality of the chemical
kinetics mechanisms used to define the Oo~HC~NO relationship,
O j\
Attempts are made to validate model predictions by comparing them with
real observations; and operating parameters are often varied to determine the
sensitivity of the model to respective parameter changes (Gelinas and Vajk,
1979). In addition, the extent of agreement between the results from two
simulations can be tested. In this way, completely different models may be
compared, or an internal component, such as the chemical kinetics mechanisms,
may be substituted and the model run again to ascertain the effect of such
substitutions.
1.3 SAMPLING AND MEASUREMENT OF OZONE AND OTHER PHOTOCHEMICAL OXIDANTS AND
THEIR PRECURSORS
1.3.1 Sampling and Measurement of Ozone and Other Photochemical Oxidants
The analysis of ozone and other, related atmospheric oxidants includes
requirements for representative sampling, specific and sensitive measurement
methodologies, methods for the generation of standard samples, absolute methods
for the calibration of these standards, and procedures by which to provide
quality assurance for the whole measurement process. In this summary, recom-
mended procedures are given for all of these analytical steps. Because of the
large existing data base that employed measurements for "total oxidants,"
non-specific iodometric techniques are discussed and compared to current
specific OT measurements,
1.3.1.1 Quality Assurance and Sampling. A quality assurance program is
employed by the U.S. Environmental Protection Agency for assessing the accuracy
and precision of monitoring data and for maintaining and improving the quality
of ambient air data. Procedures and operational details have been prescribed
in each of the following areas: selection of analytical methods and instrumen-
tation (i.e., reference and equivalent methods); method specifications for
gaseous standards; methods for primary and secondary (transfer standards) :
calibration; instrumental zero and span check requirements, including frequency
of checks, multiple-point calibration procedures, and preventive and remedial
maintenance requirements; procedures for air pollution episode monitoring;
methods for recording and validating data; and information on documenting
1-21
-------
quality control (U.S. Environmental Protection Agency, 1977b). Design criteria
for DO monitoring stations, to help ensure the quality of aerometric data,
have been established (U.S. Environmental Protection Agency, 1977a; National
Research Council, 1977).
1.3.1.2 Measurement Methods for Total Qxidants and Ozone. Techniques for the
continuous monitoring of total oxidants and 03 in ambient air are summarized
in Table 1-1. The earliest methods used for routinely monitoring oxidants in
the atmosphere were based on iodometry. When atmospheric oxidants are absorbed
in potassium iodide (KI) reagent, the iodine produced,
03 + 3I~ + H20 •» I3~ + 02 + 20H~ (1-1)
is measured by ultraviolet absorption in colorimetric instruments and by
amperometric means in electrochemical instruments. The term "total oxidants"
is of historical significance only and should not be construed to mean that
such measurements yield a sum of the concentrations of the oxidants in the
atmosphere. The various oxidants in the atmosphere react to yield iodine at
different rates and with different stoichiometries. Only ozone reacts immedi-
ately to give a quantitative yield of iodine. As discussed below, the total
oxidants measurement correlates fairly well with the specific measurement of
ozone, except during periods when significant nitrogen dioxide (NOp) and
sulfur dioxide (SOp) interferences are present. The major problem with the
total oxidants measurement was the effect of these interferences. Total
oxidants instruments have now been replaced by specific ozone monitors in all
aerometric networks and in most research laboratories. Biases among and
between total oxidants and ozone methods are still important, however, for
evaluating existing data on health and welfare effects levels where concentra-
tions were measured by total oxidants methods.
The reference method promulgated by EPA for compliance monitoring for
ozone is the chemiluminescence technique based on the gas-phase ozone-ethylene
reaction (F.R., 1971). The technique is specific for ozone, the response is a
linear function of concentration, detection limits of 0.001 to 0.005 ppm are
readily obtained, and response times are 30 seconds or less. Prescribed
methods of testing and prescribed performance specifications that a commercial
analyzer must meet in order to be designated as a reference method or an
equivalent method have been published by EPA (F.R., 1975). An analyzer may be
1-22
-------
TABLE 1-1. SUMMARY OF OZONE MONITORING TECHNIQUES
l\5
CO
Principle
Continuous
colon metric
Continuous
electrochemical
Chemi 1 uai neseence
Chemi 1 umi nescence
Ultraviolet
photometry
Reagent
10(20)% KI
buffered at
pH = 6.8
2% KI
buffered at
pH =6.8
Ethyl ene,
gas-phase
Rhodamine-B
None
Response
Total
oxidants
Total
oxidants
Q3~specific
03-specific
03-specific
Minimum
detection limit
0.010 ppn
0.010 ppm
0.005 ppin
0.001 ppat
0.005 ppra
Response
tine, 90% FSa
3 to 5 minutes
1 minute
< 30 seconds
< 1 minute
30 seconds
Major
interferences
N02(+20%, 10%KI)
S02(-100%)
N02(+6%)
so2(-ioo%)
Noneb
None
Species that
absorb at 254 nn
References
Littman and Benoliel (1953)
Tokiwa et al. (1972)
Brewer and Mil ford (1960)
Mast and Saunders (1962)
Tokiwa et al. (1972)
Nederbragt'et al. (1965)
Stevens and Hodgeson (1973)
Regener (1960, 1964)
Hodgeson et al. (1970)
Bowman and Horak (1972)
aFS = full response.
A signal enhancement of 3 to 12% has been reported for measurement of 03 in humid versus dry air (California Air Resources Board, 1976).
No significant interferences have been reported in routine ambient air monitoring. If abnormally high concentrations of species that
absorb at 254 ran (e.g., aromatic hydrocarbons and mercury vapor) are present, some positive response may be expected. If high aerosol
concentrations are sampled, inlet filters must be used to avoid a positive response. •
-------
designated as a reference method if it is based on the same principle as the
reference method and meets performance specifications. An acceptable equiva-
lent method must meet the prescribed performance specifications and also show
a consistent relation with the reference method.
The designated equivalent methods are based on either the gas-solid
chemiluminescence procedure or the ultraviolet absorption procedure (Table
1-1). The first designated equivalent method was based on ultraviolet absorp-
tion by ozone of the mercury 254 nm emission line. The measurement is in
principle an absolute one, in that the ozone concentration can be computed
from the measured transmission signal since the absorption coefficient and
pathlength are accurately known. In the gas-solid chemiluminescence analyzer,
the reaction between ozone and Rhodamine-B adsorbed on activated silica pro-
duces chemiluminescence, the intensity of which is directly proportional to
ozone concentration.
1.3.1.3 Calibration Methods. All the analyzers discussed above must be
calibrated periodically with ozonized air streams, in which the ozone concen-
tration has been determined by some absolute technique. This includes the
ultraviolet (UV) absorption analyzer, which, when used for continuous ambient
monitoring, may experience ozone losses in the inlet system because of contami-
nation.
A primary ozone calibration system consists of a clean air source, ozone
generator, sampling manifold, and means for measuring absolute ozone concentra-
tion. The ozone generator most often used is a photolytic source employing a
mercury lamp that irradiates a quartz tube through which clean air flows at a
controlled rate (Hodgeson et a!., 1972). Once the output of the generator has
been calibrated by a primary reference method, it may be used to calibrate 0,,
transfer standards, which are portable generators, instruments, or other
devices used to calibrate analyzers in the field. Reference calibration
procedures that have been used for total oxidants and ozqne-specific analyzers
in the United States are summarized in Table 1-2.
The original reference calibration procedure promulgated by EPA was the
1 percent neutral buffered potassium iodide (NBKI) method (F.R., 1971). This
technique was employed in most of the United States, with the exception of
California. The California Air Resources Board (CARB) (1976) and the Los
Angeles Air Pollution Control District (LAAPCD) employed different versions of
iodometric techniques. Procedural details of the calibration methods are
1-24
-------
TABLE 1-2. OZONE CALIBRATION TECHNIQUES
Method
1% NBKI
2% NBKIC
11 Unbuffered
KI
UV photometry
\—>
i
en Gas-phase
titration (GPT)
1% BAKI
Reagent
1% KI,
phosphate buffer
pH = 6.8
2% KI
phosphate buffer
pH = 6.8
1% KI
pH = 7
None
Nitric oxide
standard reference
gas
1% KI,
boric acid buffer
pH = 5
Primary standard3
Reagent grade
arsenious oxide
Reagent grade
potassium biiodate
03 absorptivity at
Hg 254 nm emission
line
Nitric oxide SRM
(50 ppm in N2)
from NBS
Standard KI03g
solutions
Organization
and dates
EPA
1971-1976
CARB
until 1975
LAAPCD
until 1975
All
1979-present
EPA, States
1973-present
EPA
1975-1979
Bias,
Purpose [03]./[03]
Primary reference 1.12 ± 0.05
procedure
Primary reference 1. 20 ± 0.05
procedure
Primary reference 0,96
procedure
Primary reference
procedure
Alternative reference 1,030 ± 0,015
procedure (1973-1979)
Transfer standard (1979-present)
Alternative reference 1.00 ± 0.05
procedure
In the case of the iodometric methods, the primary standard is the reagent used to prepare or standardize iodine solutions,
The uncertainty limits represent the range of values obtained in several independent studies.
cPre-humidified air used for the ozone source.
Only one study available (DeMore et al., 1976).
p
UV photometry used as reference method by CARB since 1975. This technique used as an interim, alternative reference procedure by
EPA from 1976 to 1979.'
This is the value reported in the latest definitive study (Fried and Hodgeson, 1982). Previous studies reported biases ranging from
0 to 10 percent (Burton et al., 1976; Paur and McElroy, 1979),
^This procedure also recommended a standard I3 solution absorptivity to be used instead of the preparation of standard iodine solutions.
-------
summarized In Table 1-2. A number of studies conducted between 1974 and 1978
revealed several deficiencies with KI methods, the most notable of which were
poor precision or inter!aboratory comparability and a positive bias of NBKI
measurements relative to simultaneous absolute UV absorption measurements.
The positive bias observed is peculiar to the use of phosphate buffer in the
NBKI techniques. The bias was not observed in the unbuffered LAAPCD method
(which nevertheless suffered from poor precision), nor in the 1 percent EPA KI
method without phosphate buffer (Hodgeson et a!., 1977), nor in a KI procedure
that used boric acid as buffer (Flamm, 1977). A summary of results of these
prior studies was presented in the previous criteria document (U.S. Environ-
mental Protection Agency, 1978) and in a review by Burton et al. (1976).
Correction factors for converting NBKI calibration data to a UV photometry
basis are given in Table 1-2 and discussed in Chapter 4 (Section 4.2.4.2,1).
Subsequently, EPA evaluated four alternative reference calibration proce-
dures and selected UV photometry on the basis of superior accuracy and precision
and simplicity of use (Rehme et al., 1981). In 1979 regulations were amended
to specify UV photometry as the reference calibration procedure (F.R., 1979).
Laboratory photometers used as reference systems for absolute CU measurements
have been described by DeMore and Patapoff (1976) and Bass et al. (1977).
These laboratory photometers contain long path cells (1 to 5 m) and
employ sophisticated digital techniques for making effective double beam
measurements of small absorbancies at low ozone concentrations, A primary
standard UV photometer is one that meets the requirements and specifications
given in the revised ozone calibration procedures (F.R., 1979e). Since these
are currently available in only a few laboratories, EPA has allowed the use of
transfer standards, which are devices or methods that can be calibrated against
a primary standard and transferred to another location for calibration of 0,
analyzers. Examples of transfer standards that have been used are commercial
O.j photometers, calibrated generators, and gas-phase titration (GPT) apparatus.
Guidelines on transfer standards have been published by EPA (McElroy, 1979).
1.3.1.4- Relationships of Total Oxidants and Ozone Measurements. The temporal
and quantitative relationship, between simultaneous total oxidants and ozone
measurements has been examined in this chapter because of the existence of a
data base obtained by total oxidants measurements. Such a comparison is com-
plicated by the relative scarcity of simultaneous data, the presence of both
positive (N02) and negative (S02) interferences in total oxidants measurements
1-26
-------
of ambient air, and the change in the basis of calibration. In particular,
the presence of NOp and SCL interferences prevent the establishment of a
definite quantitative relationship between ozone and oxidants data. Neverthe-
less, some interesting conclusions can be drawn and are summarized below.
Studies concluded in the early to mid-1970s were reviewed in the previous
criteria document (U.S. Environmental Protection Agency, 1978). Averaged data
showed fairly good qualitative and quantitative agreement between diurnal
variations of total oxidants and ozone. For example, uncorrected monthly
averaged data from Los Angeles and St. Louis showed distinct morning and
evening peaks resulting from NOp interference (Stevens et al. , 1972a,b). The
most recent comparison in the literature involved simultaneous ozone and total
oxidant measurements in the Los Angeles basin by the California Air Resources
Board (1978) in 1974, 1976, and 1978. The maximum hourly data pairs were
correlated (Chock et al., 1982) and yielded the following regression equation
for 1978, in which a large number (927) of data pairs were available:
Oxidant (ppm) = 0.870 03 + 0.005
(Correlation coefficient = 0.92) (1-2)
The oxidant data were uncorrected for NOp and SQp interferences, and on individ-
ual days maximum oxidant averages were both higher than and lower than ozone
averages.
In summary, specific ozone measurements agree fairly well with total oxi-
dants corrected for NOp and SQp interferences, and in such corrected total
oxidants measurements ozone is the dominant contributor. Indeed, it is diffi-
cult to discern the presence of other oxidants in corrected total oxidant
data. Without corrections there can be major temporal discrepancies between
ozone and oxidants data, which are primarily a result of oxidizing and reducing
interferences with KI measurements. As a result of these interferences, on
any given day the total oxidant values may be higher than or lower than simul- :
taneous ozone data. The measurement of ozone is a more reliable indicator
than total oxidant measurements of oxidant a-ir quality.
1.3.1.5 Methods for Sampling and Analysis of Peroxyacetyl Nitrate and Its
Homologues. Only two analytical techniques have been used to obtain signifi-
cant data on ambient peroxyacetyl nitrate (PAN) concentrations. These are gas
chromatography with electron capture detection (GC-ECD) and long-path Fourier
1-27
-------
transform infrared (FTIR) spectrometry. Atmospheric data on PAN concentrations
have been obtained predominantly by GC-ECD because of its relative simplicity
and superior sensitivity. These techniques have been described in this chapter
along with attendant methods of sampling, PAN generation, absolute analysis,
and calibration.
By far the most widely used technique for the quantitative determination
of ppb concentrations of PAN and its homologues is GC-ECD (Darley et a!.,
1963; Stephens, 1969), With carbowax or SE30 as a stationary phase, PAN,
peroxypropionyl nitrate (PPN), peroxybenzoyl nitrate (PBzN), and other homo-
logues (e.g., peroxybutyryl nitrate) are readily separated from components such
as air, water, and other atmospheric compounds, as well as ethyl nitrate,
methyl nitrate, and other contaminants that are present in synthetic mixtures.
Electron-capture detection provides sensitivities in the ppb and sub-ppb
ranges. Typically, manual air samples are collected in 50- to 200-ml ungreased
glass syringes and purged through the gas-sampling valve. Continuous analyses
are performed by pumping ambient air through a gas sampling loop of an auto-
matic valve, which periodically injects the sample onto the column. Samples
collected from the atmosphere should be analyzed as soon as possible because
PAN and its homologues undergo thermal decomposition in the gas phase and at
the surface of containers. The recent work of Singh and Salas (1983a,b) on
the measurement of PAN in the free (unpolluted) troposphere (see Chapter 5) is
illustrative of current capabilities for measuring low concentrations. A
minimum detection limit of 0.010 ppb was obtained.
The literature contains conflicting reports on the effects of variable
relative humidity on PAN measurements by GC-ECD. If a moisture effect is
suspected in a PAN analysis, the bulk of this evidence suggests that humidifi-
cation of PAN calibration samples (to a range approximating the humidity of
the samples being analyzed) would be advisable.
Conventional long-path infrared spectroscopy and Fourier-transform in-
frared spectroscopy (FTIR) have been used to detect and measure atmospheric
PAN, Sensitivity is enhanced by the use of FTIR. The most frequently used IR
bands have been assigned and the absorptivities reported in the literature
(Stephens, 1964; Bruckmann and Wiliner, 1983; Holdren and Spicer, 1984) permit
the quantitative analysis of PAN without calibration standards. The absorptiv-
w. ~\
ity of the 990 cm band of PBzN, a higher homologue of PAN, has been reported
by Stephens (1969). Tuazon et al. (1978) describes an FTIR system operable at
1-28
-------
pathlengths up to 2 km for ambient measurements of PAN and other trace constit-
uents. This system employed an eight-mirror multiple reflection cell with a
22.5-m base path. Tuazon et al. (1981b) reported maximum PAN concentrations
ranging from 6 to 37 ppb over a 5-day smog episode in Claremont, CA. Hanst et
al. (1982) made measurements with a 1260-m folded optical path system during a
2-day smog episode in Los Angeles in 1980. An upper limit of 1 ppb of PBzN
was placed, and the maximum PAN concentration observed was 15 ppb. The large
internal surface area of the White cells may serve to promote the decomposition
or irreversible adsorption of reactive trace species such as PAN. High volume
sampling rates and inert internal surface materials are used to minimize these
effects.
Because of the thermal instability of dilute PAN samples and the explosive
nature of liquefied PAN, calibration samples are not commercially available !
and must be prepared. Earlier methods used to synthesize PAN have been summa-
rized by Stephens (1969). The photolysis of alkyl nitrites in oxygen was the
most commonly used procedure and,may still be used by some investigators. As
described by Stephens et al. (1965), the liquefied crude mixture obtained at
the outlet of the photolysis chamber is purified by preparative-scale GC.
[CAUTION: Both the liquid crude mixture and the purified PAN samples are
violently explosive and should be handled behind explosion shields using
plastic full-face protection, gloves, and a leather coat at all times.] The
pure PAN is usually diluted to about 1000 ppm in cylinders pressurized with
nitrogen to 100 psig and stored at reduced temperatures, <15°C.
Gay et al. (1976) have used the photolysis of chlorine: aldehyde: nitrogen
dioxide mixtures in air or oxygen for the preparation of PAN and a number of
its homologues at high yields. This procedure has been utilized in a portable
PAN generator that can be used for the calibration of GC-ECD instruments in
the field (Grosjean, 1983; Grosjean et al., 1984).
Several investigators have recently reported on a condensed-phase synthesis
of PAN by nitration of peracetic acid in a hydrocarbon solvent. High yields
are produced of a pure product free of other alkyl nitrates (Hendry and Kenley,
1977; Kravetz et al. , 1980; Nielsen et al., 1982; Holdren and Spicer, 1984).
After the nitration is complete, the hydrocarbon fraction containing PAN
concentrations of 2 to 4 mg/ml can be stored at -20°C for periods longer than
a year (Holdren and Spicer, 1984).
1-29
-------
The most direct method for absolute analysis of these PAN samples is by
infrared absorption, using the IR absorptivities mentioned earlier. Conven-
tional IR instruments and 10-cm gas cells can analyze gas standards of concen-
trations >35 ppm.. Liquid microcells can be used for the analysis of PAN in
isooctane solutions. The alkaline hydrolysis of PAN to acetate ion and nitrite
ion in quantitative yield (Nicksic et al,, 1967) provides a means independent
of infrared for the quantitative analysis of PAN. Following hydrolysis,
nitrite ion may be quantitatively analyzed by the Saltzman colorimetric proce-
dure (Stephens, 1969). The favored procedures now use ion chromatography to
analyze for nitrite (Nielsen et a!., 1982) or acetate (Grosjean, 1983; Grosjean
et al., 1984) ions. Another calibration procedure has been proposed that is
based on the thermal decomposition of PAN in the presence of excess and measured
NO concentrations (Lonneman et al., 1982). The acetylperoxy radical, CH~C(Q)Og,
>J £.
and its decomposition products rapidly oxidize nitric oxide (NO) to NGp with a
stpichiometry that has been experimentally determined.
1.3.1,6 Methodsfor Sampling and Analysis of Hydrogen Peroxide. Hydrogen
peroxide (H202) is significant in photochemical smog as a chain terminator; as
an index of the hydroperoxyl radical (H02) concentration (Bufalini and Brubaker,
1969; Demerjian et al., 1974); and as a reactant in the aqueous-phase oxidation
-?
of S02 to SO, and in the acidification of rain (Penkett et al., 1979; Dasgupta,
1980a,b; Martin and Damschen, 1981; Overton and Durham, 1982).
One of the major problems, however, in assessing the role of atmospheric
HpOp has been a lack of adequate measurement methodology. Earlier measurements
(Gay and Bufalini, 1972a,b; Bufalini et al., 1972; Kok et al., 1978a,b) reporting
HLOg concentrations of 0.01 to 0.18 ppm are now believed to be far too high,
and to be the result of artifact H?02 formation from reactions of absorbed Q~
(Zika and Saltzman, 1982; Heikes et al., 1982; Heikes, 1984). Maximum tropo-
spheric H202 concentrations predicted by modeling calculations (Chameides and
Tan, 1981; Logan et al., 1981) and observed in recent field studies (Das et
al., 1983) are on the order of 1 ppb. .:.
Almost all of the methods used for the measurement of atmospheric hLOg
have used aqueous traps for sampling. Atmospheric 03, however, which is also
absorbed at concentrations much higher than H202, reacts in the bulk aqueous
phase and at surfaces to produce H202 and thus is a serious interference (Zika
and Saltzman, 1982; Heikes et al., 1982; Heikes, 1984). The removal of absorbed
03 by purging immediately after sample collection may remove or substantially
1-30 , , ,.
-------
reduce this interference (Zika and Saltzman, 1982; Das et al., 1983). Another
problem identified with aqueous sampling is that other atmospheric species (in
particular, SCL) may interfere with the generation of H_Q2 in aqueous traps •
and also react with collected HJD? to reduce the apparent concentration measured
(Heikes et al., 1982).
Of the techniques that have been used for the measurement of aqueous and
gas-phase HpO?, only the chemiluminescence and enzyme-catalyzed methods are
summarized below. The other techniques are now believed to have inadequate
sensitivity and specificity for the atmospheric concentrations actually present.
In addition, the use of a tunable diode infrared laser source should eliminate
the problem associated with nearby water bands, and this method is currently
under investigation for atmospheric measurements (unpublished work in progress,
Schiff, 1985). . V:
Hydrogen peroxide in the atmosphere may be detected at low concentrations
by the chemiluminescence obtained from copper(ID-catalyzed oxidaton of luminol
(5-amino-2,3-dihydro-l,4-phthalazinedione) by H^O^ (Armstrong and Humphreys,
1965; Kok et al., 1978a,b). This technique as initially employed suffered the
interferences from 0- and SQ2 discussed above for aqueous traps. Das et al.
(1982) employed a static version of the method of Kok et al. (1978a) to measure
H202 concentrations in the 0.01 to 1 ppb range. In addition, samples were
purged with argon immediately after collection to eliminate, reportedly, the
03 interference. Recently, a modified chemiluminescence method has been
reported which used hemin, a blood component, as a catalyst for the luminol-
based H^O* oxidation (Yoshizumi et al., 1984).
The most promising chemical approach employs the catalytic activity of
the enzyme horseradish peroxidase (HRP) on the oxidation of organic substrates
by H^Op. The production or decay of the fluorescence intensity of the substrate
or reaction product is measured as it is oxidized by HpO,,, catalyzed by HRP.
Some of the more widely used substrates have been scopoletin (6-methoxy~7-
hydroxyl,2-benzopyrone) (Andreae, 1955; Perschke and Broda, 1961); 3-(p_-hydroxy-
phenyl)propionic acid (HPPA) (Zaitsu and Okhura, 1980);, and leuco crystal
violet (LCV) (Mottola et al., 1970).
The decrease in the fluorescence intensity of scopoletin is measured as a
function of HuQ9 concentration. Detection limits have been reported to be
-.11
quite low (10 M). The chief disadvantage of this approach is that the
concentration- of H?0? must be within a narrow range to obtain an accurately
1-31
-------
measureable decrease in fluorescence. Oxidation of LCV produces intensely
5 ~1
colored crystal violet, which has a molar absorption coefficient of 10 M
••II
on at the analytical wavelength, 596 nm. The detection limit reported was
**8
10 M in 5 cm cells. Two quite similar hydrogen donor substrates have been
used. Zaitsu and Okhura (1980) employed 3-(p_-hydroxyphenyl) propionic acid
and more recently the p_-hydroxyphenyl acetic acid homologue is being used
(Kunen et a!., 1983; Dasgupta and Hwang, 1985). The measurement of the fluores-
cence intensity of the product dimer provides a quite sensitive means for the
assay of H202.
As with 03, ti?®? calibration standards are not commercially available and
are usually prepared at the time of use. The most convenient method for pre-
paring aqueous samples containing micromolar concentrations of H?0p is simply
the serial dilution of commercial grade 30 percent HyQ- (Fisher Analytical
Reagent). Techniques for the convenient generation of gas-phase standards are
not available. A technique often used for generating ppm concentrations of
H«0« in air involves the injection of micro!iter quantities of 30 percent HpO«
solution into a metered stream of air that flows into a Teflon bag. Aqueous
and gas-phase samples are then standardized by conventional iodometric proce-
dures (Allen et al., 1952; Cohen et al., 1967).
1.3.2 Measurement of Precursors to Ozone and Other Photochemical Oxidants
1.3.2.1 Nonmethane Organic Compounds. Numerous analytical methods have been
employed to determine nonmethane organic compounds (NMOC) in ambient air.
Measurement methods for the organic species may be grouped according to three
major classifications: nonmethane hydrocarbons, aldehydes, and other oxygenated
compounds.
Nonmethane hydrocarbons have been determined primarily by methods that
employ a flame ionization detector (FID) as the sensing element. Early methods
for the measurement of total nonmethane hydrocarbons did not provide for
speciation of the complex mixture of organics in ambient air. These methods,
still in use for analysis of total nonmethane organic compounds, are essentially
organic carbon analyzers, since the response of the FID detector is essentially
proportional to the number of carbon atoms present in the organic'molecule
(Sevcik, 1975). Carbon atoms bound, however, to oxygen, nitrogen^ or halogens
give reduced relative responses (Dietz, 1967). The FID detectqfiihas been
utilized both as a stand-alone continuous detection system (non-speciation
1-32
-------
analyzers have indicated an overall poor performance of the commercial instru-
ments when calibration or ambient mixtures containing nonmethane organic
compounds (NMOC) concentrations less than 1 ppm C were analyzed (e.g., Reckner,
1974; McElroy and Thompson, 1975; Sexton et a!., 1982). The major problems
associated with the non-speciation analyzers have been summarized in a recent
technical assistance document published by the U.S. Environmental Protection
Agency (1981). The document also presents ways to reduce some of the existing
problems.
Because of the above deficiencies, other approaches to the measurement of
nonmethane hydrocarbons are currently under development. The use of gas
chromatography coupled to an FID ^system circumvents many of the problems
associated with continuous non-speciation analyzers. This method, however,
requires sample preconcentration because the organic components are present at
part-per-billion (ppb) levels or lower in ambient air. The two main preconcen-
tration techniques in present use are cryogenic collection and the use of
solid adsorbents (McClenny et al., 1984; Jayanty et a!., 1982; Westberg
et al., 1980; Ogle et al., 1982). The preferred preconcentration method for
obtaining speciated data is cryogenic collection. Speciation methods involving
cryogenic preconcentration have also been compared with continuous nonspeciation
analyzers (e.g., Richter, 1983). Results indicate poor correlation between
methods at ambient concentrations below 1 part-per-million carbon (ppm C).
Aldehydes, which are both primary and secondary pollutants in ambient
air, are detected by total NMOC and NMHC speciation methods but can not be
quantitatively determined by those methods. Primary measurement techniques
for aldehydes include the chromotropic acid (CA) method for formaldehyde
(Altshuller and McPherson, 1963; Johnson et al.s 1981), the 3-methyl~2-benzo-
thiazolene (MBTH) technique for total aldehydes (e.g., Sawicki et al., 1961;
Mauser and Cummins, 1964), Fourier-transform infrared (FTIR) spectroscopy
(e.g., Hanst et al., 1982; Tuazon et al., 1978, 1980, 1981a), and high-perfor-
mance liquid chromatography employing 2»4-dim"trophenyl-hydrazine derivatiza-
tion (HPLC-ONPH) for aldehyde speciation (e.g., Lipari and Swarin, 1982; Kuntz
et al.5 1980). The CA and MBTH methods utilize wet chemical procedures and
spectrophotometric detection. Interferences from other compounds have been
reported for both techniques. The FTIR method offers good specificity and
direct jji situ analysis of ambient air. These advantages are offset, however,
by the relatively high cost and lack of portability of the instrumentation.
1-33
-------
On the other hand, the HPLC-DNPH method not only offers good specificity but
can also be easily transported to field sites. A few intercomparison studies
of the above methods have been conducted and considerable differences in
measured concentrations were found. The data base is still quite limited at
present, however, and further intercomparisons are needed.
Literature reports describing the vapor-phase organic composition of
ambient air indicate that the major fraction of material consists of ^substi-
tuted hydrocarbons and aldehydes. With the exception of formic acid, other
oxygenated species are seldom reported. The lack of oxygenated hydrocarbon
data is somewhat surprising since significant quantities of these species are
emitted into the atmosphere by solvent-related industries and since at least
some oxygenated species appear to be emitted by vegetation. In addition to
direct emissions, it is also expected that photochemical reactions of hydro-
carbons with oxides of nitrogen, CU, and hydroxyl radicals will produce signi-
ficant quantities of oxygenated species. Difficulties during sample collection
and analysis may account for the apparent lack of data. Attempts have been
made to decrease adsorption by deactivating the reactive surface or by modifying
the compound of interest (Qsman et al.} 1979; Westberg et a!., 1980). Additional
research efforts should focus on this area.
1,3.2.2 Nitrogen Oxides. Aside from the essentially unreactive nitrous oxide
(NgO), only two oxides of nitrogen occur in ambient air at appreciable concen-
trations: nitric oxide (NO) and nitrogen dioxide (N02). Both compounds,
together designated as NO , participate in the cyclic reactions in the atmosphere
/\
that lead to the formation of ozone. Other minor reactive oxides of nitrogen
in ambient air include peroxyacyl nitrates, nitrogen trioxide, dinitrogen
pentoxide, and peroxynitric acid.
The preferred means (Federal Reference Method) of measuring NO and N02 is
the chemiluminescence method (F.R., 1976). The measurement principle is the
gas-phase chemiluminescent reaction of 0- and NO (Fontijn et a!., 1970).
While NO is determined directly in this fashion, NO^ is detected indirectly by
first reducing or thermally decomposing the gas quantitatively to NO with a
converter. The reaction of NO and 0_ forms excited N0« molecules that release
light energy that is proportional to the NO concentration. Although the NO
chemiluminescence is interference-free, other nitrogen compounds do interfere
when directed through the NQ2 converter. The magnitude of these interferences
is dependent upon the type of converter used (Winer et a!., 1974; Joshi and
1-34
-------
Bufalini, 1978). The detection limit of commercial chemiluminescence instru-
ments for N02 measurement is 2.5 ug/m3 (0.002 ppm) (Katz, 1976).
Development of an instrument based on the chemiluminescent reaction of
NQ2 with luminol (5-amino-2, 3-dihydro-l, 4-phthalazine dione) has been reported
by Maeda et al. (1980). Wendel et al. (1983) have reported modifications of
this luminol-based method in which better response time and less interference
from Q3 have been achieved.
Other acceptable methods for measuring ambient N0? levels, including two
methods designated as equivalent methods, are the Lyshkow-modified Griess-
Saltzman method, the instrumental colorimetric Griess-Saltzman method, the
triethanolamine method, the sodium arsenite method, and the TGS-ANSA method
[TGS-ANSA = triethanolamine, guaiacol (o-methoxyphenol), sodium metabisulfite,
and 8-anilino-1-naphthalene sulfonic acid]. The sodium arsenite method and
the TGS-ANSA method were designated as equivalent methods in 1977. For descrip-
tions of these methods, the reader is referred to the EPA criteria document
for nitrogen oxides (U.S. Environmental Protection Agency, 1982). While some
of these methods measure the species of interest directly, others require
oxidation, reduction, or thermal decomposition of the sample, or separation •
from interferences, before measurement. None of these other techniques,
however, is widely used to monitor air quality.
Careful adherence to specified calibration procedures is essential for
obtaining accurate NO measurements. The U.S. Environmental Protection Agency
A
(1975) has issued a technical assistance document that describes in detail the
two acceptable calibration methods for NO : (1) the use of standard reference
"
materials (SRMs) and (2) gas-phase titration (GPT) of NO with Og. The SRM for
NO is a cylinder of compressed NO in N-; the mixture is both accurate and
stable (Hughes, 1975). The SRM for N0£ is the N02 permeation tube (O'Keeffe
and Ortman, 1966; Scaringelli et al., 1970). The gas-phase titration, described
by Rehme et al. (1974), is based upon the bimolecular reaction of NO with Oo t
to form N02 plus 02- The U.S. Environmental Protection Agency (1975) recommends
the combined use of GPT plus SRM procedures, using one technique to check the
other.
1-35
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1.4 CONCENTRATIONS OF OZONE AND OTHER PHOTOCHEMICAL OXIDANTS IN AMBIENT AIR
In the context of this document, the concentrations of ozone and other
photochemical oxidants found in ambient air are important for: (1) assessing
potential exposures of human and other receptors; (2) determining the range of
ambient air concentrations of ozone and other photochemical oxidants relative
to demonstrated "effects levels" (Chapters 6-12); (3) determining indoor-
outdoor gradients for exposure analyses; (4) assessing whether the concentra-
tions of oxidants other than ozone, singly, collectively, or in combination
with ozone, are cause for concern; and (5) evaluating the adequacy of ozone as
a control surrogate for other photochemical oxidants, if concentrations of the
other oxidants are cause for concern given the effects and the "effects levels"
of those oxidants.
1-4.1 Ozone Concentrations in Urban Areas
In Table 1-3, 1983 ozone concentrations for Standard Metropolitan Stat-
istical Areas (SMSAs) having populations >^ 1 million are given by geographic
area, demarcated according to United States Census divisions and regions (U.S.
Department of Commerce, 1982). The second-highest concentrations among daily
maximum 1-hour values measured in 1983 in the 38 SMSAs having populations of
at least 1 million ranged from 0.10 ppm in the Ft. Lauderdale, Florida; Phila-
delphia, Pennsylvania; and Seattle, Washington, areas to 0.37 ppm in the Los
Angeles-Long Beach, California, area. The second-highest value among daily
maximum 1-hour ozone concentrations for 35 of the 38 SMSAs in Table 1-3
equaled or exceeded 0.12 ppm. The data clearly show, as well, that the
highest 1-hour ozone concentrations in the United States occurred in the
northeast (New England and Middle Atlantic states), in the Gulf Coast area
(West South Central states), and on the west coast (Pacific states). Second-
highest daily maximum 1-hour concentrations in 1983 in the SMSAs within each
of these three areas averaged 0.16, 0.17, and 0.21 ppm, respectively. It
should be emphasized that these three areas of the United States are subject
to those meteorological and climatological factors that are conducive to local
oxidant formation, or transport, or both. It should also be emphasized that 9
of the 16 SMSAs in the country with populations > 2 million are located in
these areas.
1-36
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TABLE 1-3. SECOND-HIGHEST OZONE CONCENTRATIONS AMONG DAILY MAXIMUM 1-hr
VALUES IN 1983 IN STANDARD METROPOLITAN STATISTICAL AREAS WITH POPULATIONS
> 1 MILLION, GIVEN BY CENSUS DIVISIONS AND REGIONS3
Division
and region
SMSA
population,
SMSA millions
Second-highest
1983 03
concn. , ppm
Northeast
New England
Boston, MA
.Middle Atlantic Buffalo, NY
Nassau-Suffolk, NY
Newark, NJ
New York, NY/NJ
Philadelphia, PA/NJ
Pittsburgh, PA
South
South Atlantic
South
West South
Central
North Central
East North
Central
West North
Central
Atlanta, GA
Baltimore, MD
Ft. Lauderdale-Hollywood,
Miami, FL
Tampa-St. Petersburg, FL
Washington, DC/MD/VA
Dallas-Ft. Worth, TX
Houston, TX
New Orleans, LA
San Antonio, TX
Chicago, IL
Detroit, MI
Cleveland, OH
Cincinnati, OH/KY/IN
Milwaukee, WI
Indianapolis, IN
Columbus, OH
FL
St. Louis, MO/IL
Minneapolis-St. Paul
Kansas City, MO/KS
>2
1 to <2
>2
1 to <2
>2
>2
>2
>2
>2
1 to <2
1 to <2
1 to <2
>2
>2
>2
1 to <2
1 to <2
MN/WI
>2
>2
1 to <2
1 to <2
1 to <2
1 to <2
1 to <2
>2
>2
1 to <2
0.18
0.12
0.17
0.25
0.19
0.10
0.14
0.17
0.19
0.10
0.12
0.14
0.17
0.16
0.28
0.12'
0.12
0.17
0.17
0.15
0.15
0.18
0.14
0.12
0.18
0.13
0.13
1-37
-------
TABLE 1-3 (cont'd). SECOND-HIGHEST OZONE CONCENTRATIONS AMONG DAILY MAXIMUM
1-hr VALUES IN 1983 IN STANDARD METROPOLITAN STATISTICAL AREAS
WITH > 1 MILLION, GIVEN BY CENSUS DIVISIONS AND REGIONS9
Division
and region
West
Mountai n
Pacific
SMSA
Denver-Boulder, CO
Phoenix, AZ
Los Angeles- Long Beach, CA
San Francisco-Oakland, CA
Anaheim-Santa Ana-
Garden Grove, CA
San Diego, CA
Seattle- Everett, WA
Riverside-San Bernardino-
Ontario, CA
San Jose, CA
Portland, OR/WA
Sacramento, CA
SMSA Second-highest
population, 1983 Og
millions concn. , ppm
1 to <2
1 to <2
>2
>2
1 to <2
1 to <2
1 to <2
I to <2
1 to <2
1 to <2
1 to <2
0.14
0.16
0.37
0.17
0.28
0.2Q
0.10
0.34
0.16
0.12
0.15
Standard Metropolitan Statistical Areas and geographic divisions and regions
as defined by Statistical Abstract of the United States (U.S. Department of
Commerce, 1982).
Source; U.S. Environmental Protection Agency (1984).
Emissions of manmade oxidant precursors are usually correlated with
population, especially emissions from area source categories such as transpor-
tation and residential heating (Chapter 3). Accordingly, when grouped by
population, the 80 largest SMSAs had the following median values for their
collective second-highest daily maximum 1-hour ozone concentrations in 1983:
populations > 2 million, 0.17 ppm 03; populations of 1 to 2 million, 0.14 ppm
03; and populations of 0.5 to 1 million, 0.13 ppm 03- As noted above, however,
coincident meteorology favorable for oxidant formation undoubtedly contributes
to the apparent correlation between population and ozone levels.
Among all stations reporting valid ozone data ( 0.28 ppm.
1-38
-------
A pattern of concern in assessing responses to ozone in human populations
and in vegetation is the occurrence of repeated or prolonged multiday periods
when the ozone concentrations in ambient air are in the .range of those known
to elicit responses (see Chapters 10 and 12). In addition, the number of days
of respite between such multiple-day periods of high ozone is of possible
consequence. Data show that repeated, consecutive-day exposures to or respites
from specified concentrations are location-specific. At a site in Dallas,
Texas, for example, daily maximum 1-hour concentrations were >_ 0.06 ppm for '
2 to 7 days in a row 37 times in a 3-year period (1979 :throtigh 1981). A -con-
centration of >0.18 ppm was recorded at that site on only 2 single days,
however, and no multiple-day recurrences of that concentration or greater were
recorded over the 3-year period. At a site in Pasadena, California, daily
maximum 1-hour concentrations >0.18 ppm recurred on 2 to 7 consecutive days
33 times in that same 3-year period (1979 through 1981) and occurred, as well,
on 21 separate days. These and other data demonstrate the occurrence in some
urban areas of multiple-day potential exposures to relatively high concentra-
tions of ozone. .
1.4.2. Trends in Nationwide Ozone Concentrations . V
Trends in ozone concentrations nationwide are important for estimating ;
potential exposures in the future of human populations and other receptors,'as
well as for examining the effectiveness of abatement programs. The,determina^
tion of nationwide trends requires the application of statistical tests to
comparable, representative, multiyear aerometric data. The derivation of
recent trends in ozone concentrations and the interpretation of.those, trends ;
is complicated by two potentially significant factors that have affected', ;
aerometric data since 1979: (1) the promulgation by EPA. in 1979 of a new ,,K
calibration procedure for ozone monitoring (see Chapter 4);. ,and (2) the intro-
duction by EPA of a quality-assurance, program that has resulted in .improved-.,;
data-quality audits. The effects of these factors on ozone concentration ,,
measurements are superimposed on the effects on concentrations of any changes -
in meteorology or in precursor emission rates that may have occurred over the <
same time span. . . • , ,,
The nationwide trends in ozone concentrations for a 9-year period, 1975 •
through 1983:, are shown in Figure 1-1 (U. S. Environmental Protection Agency,--\
1984). The data given are trends as gauged by the composite, average of the
1-39.
-------
0.18
0,17
E
ex
Q.
o
<
oc
t-
z
Ul
o
z
o
o
U!
z
o
M
o
0.16
0.15
0.14
,0.13
0.12
CA (27 Stations) _
f
D (MAMS STATIONS (62)
T 95% CONFIDENCE
1 INTERVALS
CALL STATIONS (176)
"95% CONFIDENCE
X INTERVALS
A CALIFORNIA STATIONS (27)
V ALL STATIONS EXCEPT
CALIFORNIA (149)
I I I I
I
1975 1976 1977 1978 1979 1980 1981 1982 1983
YEAR
Figure 1 -1. National trend in composite average of the second highest value
among daily maximum 1 -hour ozone concentrations at selected groups of
sites, 1975 through 1983.
Source: U.S. Environmental Protection Agency (1984).
1-40
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second-highest value among daily maximum 1-hour ozone concentrations. Data
from four subsets of monitoring stations, most of them urban stations, are
given: (1) California stations only; (2) all stations except those in
California; (3) all stations including those in California; and (4) all
National Air Monitoring Stations (NAMS), which report data directly to EPA.
Only stations reporting > 75 percent of possible hourly values in the respective
years are represented in the data.
For the entire 9-year period, 1975 through 1983, all subsets of monitoring
stations show a decline in the composite second-highest daily maximum 1-hour
ozone concentration. Between 1979, when the new, more accurate calibration
procedure was promulgated, and 1982, a small decline of 9 to 10 percent in
nationwide ozone concentrations occurred. From 1982 to 1983, however, concen-
trations increased by about 10 percent in California, by about 14 percent
outside California, and by about 12 percent nationwide (all states). Recently
published data for 1984 from a somewhat smaller number of sites (163) (U.S.
Environmental Protection Agency, 1986) show a decrease in nationwide ozone
concentrations from 1983 to 1984, with 1984 levels approximating those recorded
in 1981. The portion of the apparent decline in ozone nationwide from 1975
through 1984 that is attributable to the calibration change of 1979 cannot be
determined by simply applying a correction factor to all data, since not all
monitoring stations began using the UV calibration procedure in the same year.
Figure 1-2 shows the nationwide frequency distributions of the first-,
second-, and third-highest 1-hour 0, concentrations at predominantly urban
stations aggregated for 1979, 1980, and 1981, as well as the highest 1-hour 03
concentration at site of the National Air Pollution Background Network (NAPBN)
aggregated for the same 3 years. As shown by Figure 1-2, 50 percent of the
second-highest 1-hour values from non-NAPBN sites in this 3-year period were
0.12 ppm or less and about 10 percent were equal to or greater than 0.20 ppm.
At the NAPBN sites, the collective 3-year distribution (1979 through 1981) is
such that about 6 percent of the values are less than 0.10 ppm and fewer than
20 percent are higher than 0.12 ppm.
1.4.3. Ozone Concentrations in Nonurban Areas
Few nonurban areas have been routinely monitored for ozone concentrations.
Consequently, the aerometric data base for nonurban areas is considerably less
substantial than for urban areas. Data are available, however, from two
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no
O
I
z
O
z
O
O
UJ
z
O
N
O
3359
0.45
0.40
0.35
0.30
0,25
0,20
0.15
0.10
0.05
99.9 99,8
99 98 95 SO
60 70 60 50 40 30 20
TO
1 0.5 0,2 0.1 0.05 0.01
Mill
HIGHEST
2nd-HIGHEST
3rd-HIGHEST
HIGHEST, NAPBN SITES
1 I 1 1 1 i
1 I I I I I II I
_LL
0.01 0.05 0.1 0.2 0.5 1 2 5 10 20 30 40 SO 60 TO 80 90 95 98 99 99.8 99.9 99.99
STATIONS WITH PEAK 1-hour CONCENTRATIONS < SELECTED VALUE, percent
Figure 1 -2. Distributions of the three highest 1 -hour ozone concentrations at valid sites (906
station-years) aggregated for 3 years (1979,1980, and 1981) and the highest ozone
concentrations at NAPBN sites aggregated for those years (24 station-years).
Source; U.S. Environmental Protection Agency (1980,1981,1982).
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special-purpose networks, the National Air Pollution Background Network (NAPBN)
and the Sulfate Regional Experiment network (SURE). Data on maximum 1-hour
concentrations and arithmetic mean 1-hour concentrations reveal that maximum
1-hour concentrations at nonurban sites classified as rural (SURE study,
Martinez and Singh, 1979; NAPBN studies, Evans et al., 1983) can sometimes ,
exceed the concentrations observed at sites classified as suburban (SURE ,
study, Martinez and Singh, 1979). For example, maximum 1-hour ozone concentra-
tions measured in 1980 at Kisatchie National Forest (NF), Louisiana; Custer
NF5 Montana; and Green Mt. NF, Vermont, were 0.105, 0.070, and 0.115 ppm, "
respectively. Arithmetic mean 1-hour concentrations for 1980 were 0.028, "
0.037, and 0.032 ppm at the respective sites. For four nonurban (rural) sites
in the SURE study, maximum 1-hour ozone concentrations were 0.106, 0.107,
0.117, and 0.153; and mean 1-hour concentrations ranged from 0.021 to 0.035
ppm. At the five nonurban (suburban) sites of the SURE study, maximum concen-
trations were 0.077, 0.099, 0.099, 0.080, and 0.118 ppm, respectively. The
mean 1-hour concentrations at those sites were 0.023, 0.030, 0.025, 0.020,.and
0.025 ppm, respectively.
Ranges of concentrations and the maximum 1-hour concentrations at some of
the NAPBN and SURE sites show the probable influence of ozone transported from
urban areas. In one documented case, for example, a 1-hour peak ozone concen-
tration of 0.125 ppm at an NAPBN site in Mark Twain National Forest, Missouri,
was measured during passage of an air mass whose trajectory was calculated to
have included Detroit, Cincinnati, and Louisville in the preceding hours
(Evans et al., 1983). •
The second-highest concentration among all the daily maximum 1-hour
concentrations measured at the NAPBN sites appear to be about one-half the
corresponding concentrations measured at urban sites in the same years. No
trend in concentrations at these NAPBN sites is discernible in the data record
for 1979 through 1983.
These data corroborate the conclusion given in the 1978 criteria document
(U.S. Environmental Protection Agency, 1978) regarding urban-nonurban and
urban-suburban gradients; i.e., nonurban areas may sometimes sustain higher
peak ozone concentrations than those found in urban areas.
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1.4.4. Diurnal and Seasonal Patterns in Ozone Concentrations
Since the photochemical reactions of precursors that result in ozone for-
mation are driven by sunlight, as well as by emissions, the patterns of ozone
occurrence in ambient air depend on daily and seasonal variations in sunlight
intensity. The typical diurnal pattern of ozone in ambient air has a minimum
ozone level around sunrise (near zero in most urban areas), increasing through
the morning to a peak concentration in early afternoon, and decreasing toward
minimal levels again in the evening. The 1978 criteria document ascribed the
daily ozone pattern to three simultaneous processes: (1) downward transport of
ozone from layers aloft; (2) destruction of ozone through contact with surfaces
and through reaction with nitric oxide (NO) at ground level; and (3) in situ
photochemical production of ozone (U.S. Environmental Protection Agency, 1978;
Coffey et al., 1977; Mohnen, 1977; Reiter, 1977). Obviously, meteorology is
a controlling factor; if strong winds disperse the precursors or heavy clouds
intercept the sunlight, high ozone levels will not develop. Another important
variation on the basic diurnal pattern appears in some localities as a secondary
peak in addition to the early afternoon peak. This secondary peak may occur
any time from midafternoon to the middle of the night and is attributed to
ozone transported from upwind areas where high ozone levels have occurred
earlier in the day. Secondary peak concentrations can be higher than concen-
trations resulting from the photochemical reactions of locally emitted precursors
(Martinez and Singh, 1979). At one nonurban site in Massachusetts (August
1977), for example, primary peak concentrations of about 0.11, 0.14, and 0.14
occurred at noon, from noon to about 4:00 p.m., and at noon, respectively, on
3 successive days of high ozone levels (Martinez and Singh, 1979). Secondary
peaks at the same site for those same 3 days were about 0.150, 0.157, and
0.130 ppm at about 6:00 p.m., 8:00 p.m., and 8:00 p.m., respectively (Martinez
and Singh, 1979).
Because weather patterns, ambient temperatures, and the intensity and
wavelengths of sunlight all play important roles in oxidant formation, strong
seasonal as well as diurnal patterns exist. The highest ozone levels generally
occur in the spring and summer (second and third quarters), when sunlight
reaching the lower troposphere is most intense and stagnant meteorological
conditions augment the potential for ozone formation and accumulation. Average
summer afternoon levels can be from 150 to 250 percent of the average winter
afternoon levels. Minor variations in the smog season occur with location,
1-44
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however. In addition, it is possible for the maximum and second-highest
1-hour ozone concentration to occur outside the two quarters of highest average
ozone concentrations. Exceptions to seasonal patterns are potentially important
considerations with regard to the protection of crops from ozone damage,
especially since respective crops have different growing seasons in terms of
length, time of year, and areas of the country in which they are grown.
In addition to the seasonal meteorological conditions that obtain in the
lower troposphere, stratospheric-tropospheric exchange mechanisms exist that
produce relatively frequent but sporadic, short-term incursions into the
troposphere of stratospheric ozone (see Chapter 3). Such incursions show a
seasonal pattern, usually occurring in late winter or early spring.
Percentile distributions, by season of the year, of concentration data
from all eight NAPBN sites show that the arithmetic mean 1-hour concentration
(averaged over a minimum of 3 years of data at each site, for 1977 through
1983) was higher in the second quarter of the year (April, May, June) at seven
of the eight stations; and was only negligibly lower than the third-quarter
value at the eighth station. The maximum 1-hour concentrations at respective
sites, aggregated over 3 to 6 years, depending on the data record for each
site, ranged from 0.050 ppm at Custer NF, MT (in the fourth quarter) to
0.155 ppm at Mark Twain NF, MO (in the third quarter). The second-highest
1-hour concentration among maximum daily 1-hour values ranged from 0.050 ppm
at Custer NF, MT (in the fourth quarter), to 0.150 ppm at Mark Twain NF, MO
(in the third quarter). The data also show that 99 percent of the 1-hour
concentrations measured were well below 0.12 ppm, even in the second quarter
of the year, when incursions of stratospheric ozone are expected to affect,
at least to some degree,, the concentrations measured at these stations.
Excursions above 0.12 ppm were recorded in 1979 and 1980 at NAPBN sites; but
none were recorded in 1981 (Evans et al., 1983; Lefohn, 1984).
Because of the diurnal patterns of ozone, averaging across longer-term
periods such as a month, a season, or longer masks the occurrence of peak
concentrations (see, e.g., Lefohn and Benedict, 1985). This is an obvious and
familiar statistical phenomenon. It is pointed out, however, because it has
direct relevance to the protection of public health and welfare. Averaging
times must correspond to, or be related in a consistent manner to, the pattern
of exposure that elicits untoward responses.
1-45
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1.4.5 SpatialPatterns In Ozone Concentrations
In addition to temporal variations, both macro- and microscale spatial
variations in ozone concentrations occur that have relevance ranging from
important to inconsequential for exposure assessments. Differences in concen-
trations or patterns of occurrence, or both, are known to exist, for example,
between urban and nonurban areas, between indoor and outdoor air, within large
metropolises, and between lower and higher elevations. The more important
variations are summarized below.
1,4.5.1 Urban-Nonurban Differences in Ozone Concentrations. Ozone concentra-
tions differ between urban and rural, between urban and remote, and even
between rural and remote sites, as discussed in part in the preceding section
on temporal variations. The variations with area and type of site are varia-
tions 1th in the timing and the magnitude of the peak concentrations, and, in
the case of transported ozone, are related to the temporal variations between
urban and nonurban areas discussed above. Data from urban, suburban, rural,
uid remote sites (see, e.g., SAROAD, 1985a-f; Martinez and Singh, 1979; Lefohn,
1984 Evans, 1985; respectively) corroborate the conclusion drawn in the 1978
c'iuHa document (U.S. Environmental Protection Agency, 1978) that ozone
;_,nce»,trations can sometimes be higher in some suburban or even rural areas
jowwtnd of urban plumes than in the urban areas themselves; and, furthermore,
, at higher concentrations can persist longer in rural and remote areas,
iarge1:1 because of the absence of nitric oxide (NO) for chemical scavenging.
onurban areas downwind of urban plumes, peak concentrations can
air, as the result of transport, at virtually any hour of the day or night,
df -ending upon many factors, such as the strength of the emission source,
i'ldvtvion time, scavenging, and wind speed (travel time) and other meteorological
far-.IKS. The dependence of the timing of peak exposures upon these transport-
related factors is well-known and is illustrated here by two studies. Evans
-------
ozone concentrations occurred at distances of 10 to 70 km (ca. 6 to 43 mi)
north-northeast of the urban center. Consequently, it would be possible for
peak ozone concentrations to occur in the late afternoon or early evening in
nonurban areas downwind of Detroit. Kelly et al. (1986) also found that
concentrations diminished again beyond 70 km (ca. 43 mi) downwind of the urban
center. Thus, as illustrated by these and similar data, beyond the distance
traversed in the time required for maximum ozone formation in an urban plume,
ozone concentrations will decrease (unless fresh emissions are injected into
the plume) as the rate of ozone formation decreases, the plume disperses,
surface deposition or other scavenging occurs, and meteorological conditions
intervene.
It is not surprising, therefore, that in rural areas lying beyond the
point of maximum ozone formation, for a given set of conditions, peak concen-
trations are lower and average diurnal profiles are flatter than in urban and
near-urban areas (see, e.g., SAROAD, 1985b-f, for rural and remote sites). In
remote areas beyond the influence of urban plumes, average peak concentrations
will be still lower and average diurnal profiles still flatter (see e.g.,
Evans, 1985). Exceptions to these generalizations occur, of course, because
of the complex interactions of topography, meteorology, and photochemistry.
Such temporal and spatial differences between ozone concentrations in
urban versus nonurban areas are important considerations for accurately assessing
actual and potential exposures for human populations and for vegetation in
nonurban areas, especially since the aerometric data for nonurban areas are
far from abundant.
1.4.5.2 Geographic, Vertical, and AltitudinalVariations in OzoneConcentrations.
Although of interest and concern when estimating global ozone budgets, demon-
strated variations in ozone concentrations with latitude and the lesser variations
with longitude have little practical significance for assessing exposure
within the contiguous United States. The effects on ozone concentrations of
latitude and longitude within the contiguous states are minor, and are reflected
in the aerometric data bases. Of more importance, ozone concentrations are
known to increase with increasing height above the surface of the earth.
Conversely, they may be viewed as decreasing with proximity to the surface of
the earth, since the earth's surface acts as a sink for ozone (see, e.g.,
Seller and Fishman, 1981; Gal bally and Roy, 1980; Oltmans, 1981, cited in
Logan et al., 1981). The most pertinent vertical and altitudinal gradients in
1-47
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ozone concentrations are: (1) increases in concentration with height above
the surface of the ground (regardless of altitude); (2) increases in concentra-
tion with altitude; and (3) variations in concentrations with elevation in
mountainous areas, attributable to transport and overnight conservation of
ozone aloft, nocturnal inversions, trapping inversions, upslope flows, and
other, often location-specific interactions between topography, meteorology,
and photochemistry.
The importance of monitoring concentrations at the proper height above
the surface of the ground has been known for a long time, and EPA guidance on
the placement of monitoring instruments (see Chapter 4) is predicated on the
existence of a vertical gradient as ozone is depleted by reaction with ground-
level emissions of NO or by deposition on reactive surfaces such as vegetation.
Data illustrative of the near-surface gradient were reported by Pratt et al.
(1983), who measured ozone concentrations at two separate heights (3 and 9 or
6 and 9 meters) above the ground at three rural, vegetated sites. Although
the maximum mean difference between 3 and 9 meters was 3 ppb, this difference
was similar to the mean difference between sites at the same height. Given
the height of some vegetation canopies, especially forests, even such small
differences over a spread of 6 meters should probably be taken into considera-
tion when interpreting reported dose-response functions.
The gradual increase in ozone concentrations with altitude has been
documented by a number of workers (see e.g., Viezee et al., 1979; Seller and
Fishman, 1981; Oltmans, 1981, as cited in Logan et al., 1981). There is a
general increase in concentration with altitude, but as described by Seiler
and Fishman (1981) and Oltmans (1981; cited in Logan et al., 1981), for example,
two relatively pronounced gradients exist, one between the surface of the
earth and 2 km (ca. 1 mi) and one even more pronounced between 8 and 12 km
(ca. 5 and 7.5 mi).
Increases in concentrations with altitude could potentially be of some
consequence for passengers and airline personnel on high-altitude flights in
the absence of adequate ventilation-filtration systems (see Chapter 11).
Variations with height above the surface and with elevation, in mountainous
areas, however, should be taken into account to ensure the accurate assessment
J»
of exposures and the accurate derivation of dose-response functions, especially
for forests and other vegetation.
1-48
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Variations in ozone concentrations with elevation, not always consistent
or predictable, have been reported by researchers investigating the effects of
ozone on the mixed-conifer forest ecosystem of the San Bernardino Mountains of
California. Measurements taken at four monitoring stations at four different
elevations showed that peak ozone concentrations occurred progressively later
in the day at progressively higher elevations (Miller et al., 1972). Ozone
concentrations >0.10 ppm occurred for average durations of 9, 13, 9, and
8 hr/day at the four respective stations, going from lower to higher elevations.
The occurrence for 13 hr/day of concentrations >0.10 ppm at the station at
817 m (2860 ft) was probably the result of contact of that zone of the mountain-
side with the inversion layer (U.S. Environmental Protection Agency, 1978).
Nighttime concentrations rarely decreased below 0.05 ppm at the mountain crest;
whereas at the lowest elevation, the basin station at 442 m (1459 ft), the
nighttime concentration usually decayed to near zero. Trapping inversions
were major contributors to the elevational gradients observed in this study,
which was conducted in the 1970s. Oxidant concentrations within the inversion
were found not to be uniform but to occur in multiple layers and strong vertical
gradients. The important result of the trapping of oxidants in the inversion
layers was the prolonged contact of high terrain with oxidants at night (U.S.
Environmental Protection Agency, 1978).
In a more recent report, Wolff et al. (1986) described measurements made
in July 1975 at three separate elevations at High Point Mountain in northeastern
New Jersey. The daily ozone maxima were similar at different elevations. At
night, however, ozone concentrations were nearly zero in the valley but increased
with elevation on the mountainside. Greater cumulative doses (number of hours
at >0.08 ppm) were sustained at the higher elevations, 300 and 500 m, respec-
tively (ca. 990 and 1650 ft, respectively). Wolff et al. (1986) related this
phenomenon to the depth of the nocturnal inversion layer and the contact with
the mountainside of ozone conserved aloft at night.
These concentration gradients with increased elevation are important for
accurately describing concentrations at which injury or damage to vegetation,
especially forests, may occur. Researchers investigating the effects of ozone
on forest ecosystems have seldom measured nighttime ozone concentrations
because the stomates of most species are thought to be closed at night, thus
preventing the internal flux of ozone that produces injury or damage (see
Chapter 6). If stomates remain even partially open at night, however, the
1-49
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possible occurrence of nighttime peak concentrations of ozone, the occurrence
of multiple peaks in a 24-hour period, or the persistence of elevated concen-
trations that do not decay to near zero overnight should not be overlooked.
Furthermore, the lack of NO for nighttime scavenging in nonurban areas and the
persistence of ozone overnight at higher elevations will result in the presence
of relatively higher concentrations in such areas at sunrise when the stomates
open and photosynthesis begins. This possibility requires that exposure
assessments, in the absence of sufficient aerometric data for forests and
other vegetated areas, take such factors into consideration.
1.4.5.3 Other Spatial Variations in Ozone Concentrations. Other spatial varia-
tions are important for exposure assessments for human populations. Indoor-
outdoor gradients in ozone concentrations are known to occur even in buildings
or vehicles ventilated by fresh air rather than air conditioning (e.g., Sabersky
et a!., 1973; Thompson et a!., 1973; Peterson and Sabersky, 1975). Ozone reacts
with surfaces inside buildings, so that decay may occur fairly rapidly, depending
upon the nature of interior surfaces and furnishings (e.g., Davies et a!., 1984;
Content et al., 1985). Ratios of indoor-to-outdoor (I/O) ozone concentrations
are quite variable, however, since cooling and ventilation systems, air infil-
tration or exchange rates, interior air circulation rates, and the composition
of interior surfaces all affect indoor ozone concentrations. Ratios (I/O,
expressed as percentages) in the literature thus vary from 100 percent in a
non-air-conditioned residence (Contant et al., 1985); to 80 ± 10 percent
(Sabersky et al., 1973) in an air-conditioned office building (but with 100 per-
cent outside air intake); to 10 to 25 percent in air-conditioned residences
(Berk et al., 1981); and to as low as near zero in air-conditioned residences
(Stock et al., 1983; Contant et al., 1985).
On a larger scale, within-city variations in ozone concentrations can
occur, even though ozone is a "regional" pollutant. Data show, for example,
relatively homogeneous ozone concentrations in New Haven, Connecticut (SAROAD,
1985a), a moderately large city that is downwind of a reasonably well-mixed
urban plume (Wolff et al., 1975; Cleveland et al.; 1976a,b). In a large
metropolis, however, appreciable gradients in ozone concentrations can exist
from one side of the city to the other, as demonstrated for New York City
(Smith, 1981), and for Detroit (Kelly et al., 1986). Such gradients should be
taken into consideration, where possible, in exposure assessments.
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1.4.6 Concentrations and Patterns of Other Photochemical Oxidants
1.4.6.1 Concentrations. No aerometric data are routinely obtained by Federal,
state, or local air pollution agencies for any photochemical oxidants other
than nitrogen dioxide and ozone. The concentrations presented in this document
for non-ozone oxidants were all obtained in special field investigations. The
limitations in the number of locations and areas of the country represented in
the information presented simply reflect the relative paucity of data in the
published literature.
The four non-ozone photochemical oxidants for which at least minimal
concentration data are available are formic acid, peroxyacetyl nitrate (PAN),
peroxypropionyl nitrate (PPN), and hydrogen peroxide (Hp02). Peroxybenzoyl
nitrate has not been clearly identified in ambient air in the United States.
The highest concentrations of PAN reported in the older literature, 1960
through the present, were those found in the Los Angeles area: 70 ppb (1960),
214 ppb (1965); and 68 ppb (1968) (Renzetti and Bryan, 1961; Mayrsohn and
Brooks, 1965; Lonneman et a!., 1976; respectively).
The highest concentrations of PAN measured and reported in urban areas in
the past 5 years were 42 ppb at Riverside, California, in 1980 (Temple and
Taylor, 1983) and 47 ppb at Claremont, California, also in 1980 (Grosjean
1981). These are clearly the maximum concentrations of PAN reported for
California and for the entire country in this period. Other maximum PAN
concentrations measured in the last decade in the Los Angeles Basin have been
in the range of 11 to 37 ppb. Average concentrations of PAN in the Los Angeles
Basin in the past 5 years have ranged from 4 to 13 ppb (Tuazon et a!., 1981a;
Grosjean, 1983; respectively). The only published study covering urban PAN
concentrations outside California in the past 5 years is that of Lewis et al.
(1983) for New Brunswick, New Jersey, in which the average PAN concentration
was 0.5 ppb and the maximum was 11 ppb during September 1978 through May 1980.
Studies outside California from the early 1970s through 1978 showed average
PAN concentrations ranging from 0.4 ppb in Houston, Texas, in 1976 (Westberg
et al., 1978) to 6.3 ppb in St. Louis, Missouri, in 1973 (Lonneman et al. ,
1976). Maximum PAN concentrations outside California for the same period
ranged from 10 ppb in Dayton, Ohio, in 1974 (Spicer et al., 1976) to 25 ppb in
St. Louis (Lonneman et al., 1976).
The highest PPN concentration reported in studies over the period 1963
through the present was 6 ppb in Riverside, California (Darley et al., 1963).
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The next highest reported PPN concentration was 5 ppb at St. Louis, Missouri,
in 1973 (Lonneman et al., 1976). Among more recent data, maximum PPN concentra-
tions at respective sites ranged from 0.07 ppb in Pittsburgh, Pennsylvania, in
1981 (Singh et al., 1982) to 3.1 ppb at Staten Island, New York (Singh et al.,
1982). California concentrations fell within this range. Average PPN concentra-
tions at the respective sites for the more recent data ranged from 0.05 ppb at
Denver and Pittsburgh to 0.7 ppb mt Los Angeles in 1979 (Singh et a!., 1981).
Altshuller (1983) has succinctly summarized the nonurban concentrations
of PAN and PPN by pointing out that they overlap the lower end of the range of
urban concentrations at sites outside California. At remote locations, PAN
and PPN concentrations are lower than even the lowest of the urban concentra-
tions by a factor of 3 to 4.
The concentrations of H^O/j reported in the literature to date must be
regarded as inaccurate since ozone is now thought to be an interference in all
methods used to date except FTIR (Chapter 4). Measurements by FTIR, the most
specific and accurate method now available, have not demonstrated unambiguously
the presence of HpO? in ambient air, even in the high-oxidant atmosphere of the
Los Angeles area. (The limit of detection for a 1-ktn-pathlength FTIR system
is around 0.04 ppm.)
Recent data indicate the presence in urban atmospheres of only trace
amounts of formic acid: < 15 ppb, measured by FTIR (Tuazon et al., 1981b).
Estimates in the earlier literature (1950s) of 600 to 700 ppb of formic acid
in smoggy atmospheres were erroneous because of faulty measurement methodology
(Hanst et al., 1982).
1.4.6.2 Patterns. The patterns of formic acid (HCOOH), PAN, PPN, and ^9
may be summarized fairly succinctly. Qualitatively, diurnal patterns are
similar to those of ozone, with peak concentrations of each of these occurring
in close proximity to the time of the ozone peak. The correspondence in time
of day is not exact, but is close. As demonstrated by the work of Tuazon
et al. (1981b), ozone concentrations return to baseline levels somewhat faster
than the concentrations of PAN, HCOOH, or I-LQo (PPN was not measured).
Seasonally, winter concentrations (third and fourth quarters) of PAN are
lower than summer concentrations (second and third quarters). The percentage
of PAN concentrations (PAN/03 x 100) relative to ozone, however, is higher in
winter than in summer. Data are not readily available on the seasonal patterns
of the other non-ozone oxidants.
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Indoor-outdoor data on PAN are limited to one report (Thompson et a"!.,
1973), which confirms the pattern to be expected from the known chemistry of
PAN; that is, it persists longer indoors than ozone. Data are lacking on
indoor-outdoor ratios for the other non-ozone oxidants.
1.4.7 RelationshipBetweenOzone and Other Photochemical Qxidants
The relationship between ozone concentrations and the concentrations of
PAN, PPN, tiyOy, and HCOOH is important only if these non-ozone oxidants are
shown to produce potentially adverse health or welfare effects, singly, in
combination with each other, or in various combinations with ozone at concentra-
tions correponding to those found in ambient air. If only ozone is shown to
produce adverse health or welfare effects in the concentration ranges of
concern, then only ozone must be controlled. If any or all of these other
four oxidants are shown to produce potentially adverse health or welfare
effects, at or near levels found in ambient air, then such oxidants will also
have to be controlled. Since ozone and all four of the other oxidants arise
from reactions among the same organic and inorganic precursors, an obvious
question is whether the control of ozone will also result in the control of
the other four oxidants.
Control!ed-exposure studies on these non-ozone oxidants have employed
concentrations much higher than those found in ambient air (see Chapters 9 and
10). Because PAN may have contributed, however, to the eye irritation symptoms
reported in earlier epidemiological studies, and because PAN is the most
abundant of these non-ozone oxidants, the relationship between ozone and PAN
concentrations in ambient air remains of interest.
The patterns of PAN and ozone concentrations are not quantitatively
similar but do show qualitative similarities for most locations at which both
pollutants have been measured in the same study. That a quantitative, monotonic
relationship between ozone and PAN is lacking, however, is shown by the range
of PAN-to-ozone ratios, expressed as percentages, between locations and at the
same location, as reported in the review of Altshuller (1983).
Certain other information bears out the lack of a monotonic relationship
between .PAN,.and ozone. Not only are PAN-ozone relationships not consistent
between different urban areas, but they are not consistent in urban versus
nonurban areas, in summer versus winter, in indoor versus outdoor environments,
or even, as the data show, in location, timing, or magnitude of diurnal peak
1-53
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concentrations within the same city. Data obtained in Houston by Jorgen
et al. (1978), for example, show variations in peak concentrations of PAN and
in relationships to ozone concentrations of those peaks among three separate
monitoring sites. Temple and Taylor (1983) have showh that PAN concentrations
are a greater percentage of ozone concentrations in winter than in the remainder
of the year in California. Lonneman et al. (1976) demonstrated that PAN,
absolutely and as a percentage of ozone, is considerably lower in nonurban
than in urban areas. Thompson et al. (1973), in what is apparently the only
published report on indoor concentrations of PAN, showed that PAN persists
longer than ozone indoors. (This is to be expected from its enhanced stability
at cooler-than-ambient temperatures such as found in air-conditioned buildings.)
Tuazon et al. (1981b) demonstrated that PAN persists in ambient air longer
than ozone, its persistence paralleling that of nitric acid, at least in the
locality studied (Claremont, CA). Reactivity data presented in the 1978
criteria document for ozone and other photochemical oxidants indicated that
all precursors that give rise to PAN also give rise to ozone. The data also
showed, however, that not all precursors giving rise to ozone also give rise
to PAN, and that not all that give rise to both are equally reactive toward
both, with some precursors preferentially giving rise, on the basis of units
of product per unit of reactant, to more of one product than the other (U.S.
Environmental Protection Agency, 1978).
In the review cited earlier, Altshuller (1983) examined the relationships
between ozone and a variety of other smog components, including PAN, PPN,
HgOg, HCOOH, aldehydes, aerosols, and nitric acid. He concluded that "the
ambient air measurements indicate that ozone may serve directionally, but
cannot be expected to serve quantitatively, as a surrogate for the other
products" (Altshuller, 1983). It must be emphasized that the issue Altshuller
examined was whether ozone could serve as an abatement surrogate for all
photochemical products, not just the subset of non-ozone oxidants of concern
in this document. Nevertheless, a review of the data presented indicates that
his conclusion is applicable to the non-ozone oxidants examined in this docu-
ment.
1-54
-------
1.5 EFFECTS OF OZONE AND OTHER PHOTOCHEMICAL OXIDANTS ON VEGETATION
Foliar injury on vegetation is one of the earliest and most obvious
manifestations of CL injury. The effects of 03 are not limited to visible
injury, however. Impacts can range from reduced plant growth and decreased
yield, to changes in crop quality and alterations in susceptibility to abiotic
and biotic stresses. The plant foliage is the primary site of 03 effects,
although significant secondary effects, including reduced growth (both roots
and foliage) and yield, can occur.
Ozone exerts a phytotoxic effect only if a sufficient amount reaches the
sensitive cellular sites within the leaf. The 03 diffuses from the ambient
air into the leaf through the stomata, which can exert some control on 0-
uptake, to the active sites within the leaf. Ozone injury will not occur if
(1) the rate of 0» uptake is low enough that the plant can detoxify or metab-
olize 0-, or its metabolites; or (2) the plant is able to repair or compensate
for the effects (Tingey and Taylor, 1982). This is analogous to the plant
response to SO, (Thomas et al., 1950). Cellular disturbances that are not
repaired or compensated are ultimately expressed as visible injury to the leaf
or as secondary effects that can be expressed as reduced root growth, or
reduced yield of fruits or seeds, or both.
Plant growth and yield are the end products of a series of biochemical
and physiological processes related to uptake, assimilation, biosynthesis, and
translocation. Sunlight drives the processes that convert carbon dioxide into
the organic compounds (assimilation) necessary for plant growth and development.
In addition to nutrients supplied through photosynthesis, the plant must
extract from the soil the essential mineral nutrients and water for plant
growth. Plant organs convert these raw materials into a wide array of compounds
required for plant growth and yield. A disruption or reduction in the rates
of uptake, assimilation, or subsequent biochemical reactions will be reflected
in reduced plant growth and yield. Ozone would be expected to reduce plant
growth or yield if (1) it directly impacted the plant process that was limiting
plant growth; or (2) it impacted another step sufficiently so that it becomes
the step limiting plant growth (Tingey, 1977). Conversely, 0, will not limit
plant growth if the process impacted by u\ is not or does not become rate-
limiting. This implies that not all effects of 03 on plants are reflected in
growth or yield reductions. These conditions also suggest that there are
combinations of 03 concentration and exposure duration that the plant can
1-55
-------
experience that will not result in visible injury or reduced plant growth and
yield. Indeed, numerous studies have demonstrated combinations of concentration
and time that did not cause a significant effect on the plant growth or yield.
Ozone induces a diverse range of effects on plants and plant communities.
These effects are usually classified as either injury or damage. Injury
encompasses all plant reactions such as reversible changes in plant metabolism
(e.g., altered photosynthesis), leaf necrosis, altered plant quality, or
reduced growth that does not impair yield or the intended use of the plant
(Guderian, 1977). In contrast, damage or yield loss includes all effects that
reduce or impair the intended use or the value of the plant. Thus, for example,
visible foliar injury to ornamental plants, detrimental responses in native
species, and reductions in fruit and grain production are all considered
damage or yield loss. Although foliar injury is not always classified as damage,
its occurrence is an indication that phytotoxic concentrations of 0, are present.
The occurrence of injury indicates that additional studies should be conducted
in areas where vegetation shows foliar injury to assess the risk of 0- to
vegetation and to determine if the intended use or value of the plants is being
impaired.
1.5.1 Linn ting Values of Plant Response to Ozone
Several approaches have been used to estimate the 0« concentrations and
exposure durations that induce foliar injury. Most of these studies used
short-term exposures (less than 1 day) and measured visible injury as the
response variable. One method for estimating the 0~ concentrations and exposure
durations that would induce specific amounts of visible injury involves exposing
plants to a range of 03 concentrations and exposure durations, and then evalua-
ting the data by regression analysis (Heck and Tingey, 1971). The data obtained
by this method for several species are summarized in Table 1-4 to illustrate
the range of concentrations required to induce foliar injury (5% and 20%) on
sensitive, intermediate, and less sensitive species.
An alternative method for estimating the 03 concentrations and exposure
durations that induce foliar injury is the use of the limiting-value approach
(Jacobson, 1977). The limiting-value method, which was developed from a
review of the literature, identified the lowest concentration and exposure
duration reported to cause visible injury on various plant species. The
analysis was based on more than 100 studies of agricultural crops and 18
1-56
-------
TABLE 1-4. OZONE CONCENTRATIONS FOR SHORT-TERM
EXPOSURES THAT 5 OR 20 PERCENT INJURY TO VEGETATION
GROWN UNDER SENSITIVE CONDITIONS3
(ppm)
Ozone concentrations that may produce
Exposure
time, hr
0.
1.
2.
4.
8.
5
0
0
0
0
Sensitive plants
0.
(0.
0.
(0.
0.
(0.
0.
(0.
0.
35 -
45 -
15 -
20 -
09 -
12 -
04 -
10 -
02 -
0.50
0.60)
0.25
0.35)
0.15
0.25)
0.09
0.15)
0.04
Intermediate
0.
(0.
0.
(0.
0.
(0.
0.
(0.
0.
55 -
65 -
25 -
35 -
15 -
25 -
10 -
15 -
07 -
0.
0.
0.
0.
0.
0.
0.
0.
0.
plants
70
85)
40
55)
25
35)
15
30)
12
5% (20%)
injury:
Less
sensitive plants
£0.
SO.
W.
£0.
£0.
70 (0.
40 (0.
30 (0.
25 (0.
20 (0.
85)
55)
40)
35)
30)
aThe concentrations in parenthesis are for the 20% injury level.
Source: U.S. Environmental Protection Agency (1978).
studies of tree species. The analysis yielded the following range of concen-
trations and exposure durations that were likely to induce foliar injury (U.S.
Environmental Protection Agency, 1978):
1. Agricultural crops:
a. 0.20 to 0.41 ppm for 0.5 hr.
b. 0.10 to 0.25 ppm for 1.0 hr.
c. 0.04 to 0.09 ppm for 4.0 hr.
2. Trees and shrubs:
a. 0.20 to 0,51 ppm for 1.0 hr.
b. 0.10 to 0.25 ppm for 2.0 hr.
c. 0.06 to 0.17 ppm for 4.0 hr.
It should be emphasized that both methods described above can estimate
concentrations and exposure durations that might induce visible injury, but
that neither method can predict impacts of 0~ on crop yield or intended use.
The concept of limiting values also was used to estimate the 0~ concen-
trations and exposure durations that could potentially reduce plant growth and
yield (U.S. Environmental Protection Agency, 1978). The data were analyzed
1-57
-------
and plotted In a manner similar to the approach used by Jacobson (1977)
(Figure 1-3). In Figure 1-3 the line bounds mean 0~ concentrations and
exposure durations below which effects on plant growth and yield were not
detected. This graphical analysis used data from both greenhouse and field
studies and indicated that the lower limit for reduced plant performance was a
mean 0, concentration of 0.05 ppm for several hours per day for exposure
periods greater than 16 days. At 10 days the 0,, response threshold increased
to about 0.10 ppm, and to about 0.30 ppm at 6 days.
1.5.2 Methods for Determining Ozone Yield Losses
Diverse experimental procedures have been used to study the effects of 0~
on plants, ranging from studies done under highly controlled conditions, to
exposures in open-top chambers, and to field exposures without chambers. In
general, the more controlled conditions-are most appropriate for investigating
specific responses and for providing the scientific basis for interpreting and
extrapolating results. These systems are powerful tools for adding to an
understanding of the biological effects of air pollutants. To assess, however,
the impact of 0, on plant yield and to provide data for economic assessments,
deviations from the typical environment in which the plant is grown should be
minimized. For field crops, this implies that the studies should be conducted
in the field, but for crops that are typically grown in glass houses, the
studies should be conducted under glass-house conditions.
To improve estimates of yield loss in the field, the National Crop Loss
Assessment Network (NCLAN) was initiated by EPA in 1980 to estimate the magnitude
of crop losses caused by 03 (Heck et a!., 1982). The primary objectives of
NCLAN were:
1. To define the relationships between yields of major agricultural
crops and 03 exposure as required to provide data necessary for
economic assessments and the development of National Ambient
Air Quality Standards;
2. To assess the national economic consequences resulting from the
exposure of major agricultural crops to Oq>
3. To advance understandng of the cause and effect relationships that
determine crop responses to pollutant exposures.
1-58
-------
1.0
-I I III
1 I II I 1111
E
a
a.
O
oc
Ul
o
O
O
ui
O
N
O
0.1
0.01
— \ 44* 19OM8 *45
DO17 Q31
1514 30 59
40* 010 ft) •
\ 1213 41 AC
\ 26 *8
CD 29
21Q 11D
146 •48-52
,10
VO24
\
39
7QD20
• 42430D9
5« 33
54 «• 55, 56
3*
58
\
57
53
EXPOSURE, hr/day
A < 1.99
D 2 TO 3.99
O 4 TO 5.99
• > 6
NOS. = REF. NOS. ON TABLE 11-4
ll
I I I II I I I I
8 10
20 40 60 80 100
EXPOSURE PERIOD, days
200
400
Figure 1-3. Relationship between ozone concentration,
exposure duration, and reduction in plant growth or yield (see
Table 6-18; also U.S. EPA, 1978).
Source: U.S. Environmental Protection Agency (1978).
1-59
-------
In the NCLAN studies, the cultural conditions used approximated typical
agronomic practices, and open-top field exposure chambers were used to minimize
perturbations to the plant environment during the exposure. The studies have
attempted to use a range of realistic CL concentrations and sufficient repli-
cation to permit the development of exposure-response models. In the NCLAN
studies, plants were exposed to a range of 0, concentrations. Chambers were
supplied with either charcoal-filtered air (control), ambient air, or ambient
air supplemented with 0- to provide concentrations three or four levels greater
than ambient. Consequently, the CL exposures were coupled to the ambient 03
level; days with the highest ambient CL were also the same days when the
highest concentrations occurred in a specific treatment in a chamber. As the
ambient 03 varied from day-to-day, the base to which additional 03 was added
also varied. This coupling of the 0., exposures to the ambient environment
means that high 03 concentrations occurred in the chambers when the environ-
mental and air chemistry conditions, in the ambient air, were conducive for
producing elevated ambient 0» levels. The plant response data have been
analyzed using regression approaches. The exposures were typically character-
ized by a 7-hr (9:00 a.m. to 4:00 p.m.) seasonal mean 03 concentration. This
is the time period when 0,, was added to the exposure chambers.
»5
1.5.3 Estimates ofOzone-Induced Yield Loss
Yield loss is defined as an impairment or decrease in the intended use of
the plant. Included in the concept of yield loss are reductions in aesthetic
values, the occurrence of foliar injury (changes in plant appearance), and
losses in terms of weight, number, or size of the plant part that is harvested.
Yield loss may also include changes in physical appearance, chemical composi-
tion, or ability to withstand storage; which collectively are traits called
crop quality. Losses in aesthetic values.are difficult to quantify. For
example, because of its aesthetic value, the loss of or adverse effect on a
specimen plant in a landscape planting may result in a greater economic loss
than that incurred by the same impact on a plant of the same species growing
as a part of natural plant community. Foliar injury symptoms may decrease the
value of ornamental plants with or without concomitant growth reductions.
Similarly, foliar injury on crops in which the foliage is the marketable plant
part (e.g., spinach, lettuce, cabbage) can substantially reduce marketability
and thus can constitute yield loss. Attainment of the limiting values for
1-60
-------
ozone previously discussed in this section should be sufficient to prevent
foliar injury and thereby reduce this type of yield loss. Most studies of the
relationship between yield loss and ozone concentration have focused on yields
as measured by weight of the marketable plant organ, and that kind of yield
loss will be the primary focus of this section.
Studies have been conducted, frequently using open-top field exposure
chambers, to estimate the impact of 0,, on the yield of various crop species.
These studies can be grouped into two types, depending on the experimental
design and statistical methods used to analyze the data: (1) studies that
developed predictive equations relating 03 exposure to plant response, and (2)
studies that compared discrete treatment levels to a control. The advantage
of the regression approach is that exposure-response models can be used to
interpolate results between treatment levels.
When the regression approach was used to estimate yield loss, 03 was
added to either charcoal-filtered or ambient air to create a range of 03
concentrations. In summarizing the data, 0~-induced yield loss was derived
from a comparison of the performance of the plants in charcoal-filtered air,
although other reference concentrations have been used. Various regression
techniques have been used to derive exposure-response functions. The use of
regression approaches permits the estimation of the CL impact on plant yield
over the range of concentrations, not just at the treatment means as is the
case with analysis of variance methods.
1.5.3.1 Yield Loss: Determination by Regression Analysis. Examples of the
relationship between 0- concentration and plant yield are shown in Figures
1-4 and 1-5. These cultivars and species were selected because they also
illustrated the type of year-to-year variation in plant response to ozone that
may occur. The derived regression equations can be used to determine the
concentrations that would be predicted to cause a specific yield loss or to
estimate the predicted yield loss that would result from a specifc 0, concen-
tration. Both approaches have been used to summarize the data on crop responses
to 03 using the Weibull function (Raw!ings and Cure, 1985). As an example of
response, the 0~ concentrations that would be predicted to cause a 10 or 30
percent yield loss have been estimated (Table 1-5). A brief review of the
data in this table indicates that for some species mean yield reductions of 10
percent were predicted when the 7-hr seasonal mean 0, concentration exceeded
0.04 to 0.05 ppm. Concentrations of 0.028 to 0.033 ppm were predicted to
1-61
-------
6000
6OOO
19
.C
O
LU
2 3000
w
2000
1000
W
SOYBEAN (DAVIS)
RALEIGH. 1981 AND 1 982
1981(Ol
y " ES93'1'
I
_L
°-872
_L
J_
0 0.02 0.04 0.06 0.08 0.1 0.12 0.14
Oa CONCENTRATION, ppm
6000
BOOO
S 4000
O
ui
O
ui
3000
2000
1000
WHEAT (ABE)
ARGONNE,
1982 AND 19S3
1§82(O)
1983 (A)
y = B873-«V°'10B»14'4
0 0.020.040.060.080,1 0.120.14
6000
6000
o
I
4000
3000
2000
1000
fBjV SOYBEAN (WILLIAMS)
* XBELTSVIIIF un ioni
BELTSVILLE, MD, 1981 AND 1982
1981JO)
1982(A|
I
I
0 0.020.040.060,080.1 0.120.14
O3 CONC5ENTRATION. ppm
6000
5000
* 4000
O
ui
3000
2000
1000
(D)
WHEAT (ARTHUR 71)
ARGONNE, 1982 AND 1883
1982(0)
y = 4B13-(0,/0.146)2-BB
- 1983 |A|
0 0.02 0.04 0.06 0.08 0.1 0,12 0.14
O3 CONCENTRATION, ppm Oa CONCENTRATION, ppm
Figure 1 -4. Examples of the effects of ozone on the yield of soybean and wheat
cultivars. The O3 concentrations are expressed as 7-hr seasonal mean concentrations.
The cultivars were selected as examples of O3 effects and of year-to-year variations in
plant response to O3.
Source: Soybean data from Hecket al. (1984); wheat data from Kress etal. C1985).
1-62
-------
6000
5500
5000
| 4500
3
!g 4000
Q
UJ
UJ
V) 3500
Q
z
£ 3000
3
2500
2000
1500
(A)
COTTON (SJ-2)
SHAFFER. CA. 1981 AND 1982
1981 (0)
= 6546-<°3/0.199)1-228
1982(A)
0 0.02 0.04 0.06 0.08 0.1 0.12 0.14 0.16 0.18 0.2
O3 CONCENTRATION, ppm
34
33
32
31
S 30
*a
o 29
> 28
I -
26
25
24
23
n TOMATO (MURIETTA)
TRACV. CA, 1981 AND 1982
1981 (O)
= 329-(0,/0.142)3-807
\
1982(A)
y = 32.3-l3-06
I t II I 111 I
0 0.02 0.04 0.06 0.08 0.1 0.12 0.14 0.16
O3 CONCENTRATION, ppm
a.
o>
UJ
SE
S
O
16
15
14
13
12
11
10
9
8
7
6
5
4
3
2
1
TURNIP (TOKYO CROSS)
RALEIGH, 1979 AND 1980
1980(A)
y=16.25-«V0.094)3-94
I
I
I
0 0.02 0.04 0.06 0.08 0.1 0.12 0.14 0.16
O3 CONCENTRATION, ppm
Figure 1 -5. Examples of the effects of ozone on the yield of cotton, tomato, and
turnip. The O3 concentrations are expressed as 7-hr seasonal mean concentrations.
The species were selected as examples of O3 effects and of year-to-year variations in
plant response to O3.
Source: Cotton and tomato data from Heck et al. (1984); turnip data from Heagle et
al. (1985).
1-63
-------
TABLE 1-5. SUMMARY OF OZONE CONCENTRATIONS PREDICTED TO CAUSE
10 PERCENT AND 30 PERCENT YIELD LOSSES AND SUMMARY OF YIELD LOSSES PREDICTED
TO OCCUR AT 7-hr SEASONAL MEAN OZONE CONCENTRATIONS OF 0.40 and 0.06 ppm
03 concentrations, ppm,
Species
Legume crops
Soybean, Corsoy
Soybean , Davis (81)
Soybean, Davis (CA-82)
Soybean, Davis (PA-82)
Soybean, Essex
Soybean, Forrest
Soybean, Williams
Soybean, Hodgson
Bean, Kidney
Peanut, NC-6
Grain crops
Wheat, Abe
Wheat, Arthur 71
Wheat, Roland
Wheat, Von a
Wheat, Blueboy II
Wheat, Coker 47-27
Wheat, Holly
Wheat, Oasis
Corn, PAG 397
Corn, Pioneer 3780
Corn, Coker 16
Sorghum, DeKalb-28
Barley, Poco
Fiber crops
Cotton, Acala SJ-2 (81)
Cotton, Acala SJ-2 (82)
Cotton, Stone vi lie
Horticultural crops
Tomato, Murrieta (81)
Tomato, Murrieta (82)
Lettuce, Empire
Spinach, America
Spinach, Hybrid
Spinach, Viroflay
Spinach, Winter Bloom
Turnip, Just Right
Turnip, Pur Top W. G.
Turnip, Shogoin
Turnip, Tokyo Cross
predicted to
yield losses
10%
0.048
0.038
0.048
0.059
0.048
0.076
0.039
0.032
0.033
0.046
0.059
0.056
0.039
0.028
0.088
0.064
0.099
0.093
0.095
0.075
0.133
0.108
0.121
0.044
0.032
0.047
0.079
0.040
0.053
0.046
0.043
0.048
0. 049
0.043
0.040
0.036
0.053
aThe yield losses are derived from Weibull
cause
of;
30%
0.082
0.071
0.081
0.081
0.099
0.118
0.093
0.066
0.063
0.073
0.095
0.094
0.067'
0.041
0.127
0.107
0.127
0.135
0.126
0.111
0.175
0.186
0.161
0.096
0.055
0.075
0.108
0.059
0.075
0.082
0.082
0.080
0.080
0.064
0.064
0.060
0.072
equations
Percent yield
to occur at
losses predicted
7-hr seasonal
mean Os concentration of;
0.04 ppm
6.4
11.5
6.4
2.0
7.2
1.7
10.4
15.4
14.9
6.4
3.3
4.1
10.3
28.8
0.5
2.2
0.0
0.4
0.3
1.4
0.0
0.0
0.0
8.3
16.1 '
4.6
0.8
10.3
0.0
6.8
2.6
6.0
5.8
7.7
10.1
13.0
3.3
and are based
0.06 ppm
16.6
24.1
16.5
10.4
14.3
5.3
18.1
18.4
28
19.4
10.4
11.7
24.5
51.2
2.8
8.4
0.9
2.4
1.5
5.1
0.3
2.7
0.5
16.2
35.1
16.2
3.7
,31.2
16.8
17.2
9.2
16.7
16.5
24.9
26.5
29.7
15.6
on the control
yields in charcoal-filtered air.
Source: Derived from Heck
et al. (1984).
1-64
-------
cause a 10 percent yield loss in Vona wheat, kidney bean, and Hodgson soybean.
At a 7-hr seasonal mean 0, concentration of 0,04 ppm, mean yield reductions
ranged from zero percent in sorghum, barley, and a corn cultivar to a high of
28.8 percent in Vona wheat.
A histogram of the 7-hr seasonal mean 03 concentrations predicted to
cause a 10 percent yield loss (Table 1-5) is given in Figure 1-6 to help
illustrate the range of concentrations and their relative frequency of occur-
rence. The data in Figure 1-6 are based on 37 species or cultivar yield-
response functions developed from studies in open-top field exposure chambers.
Approximately 57 percent of the species or cultivars were predicted to exhibit
10 percent yield reductions at 7-hr seasonal mean concentrations below 0.05
ppm. Thirty-five percent of plant types were predicted to display a 10 percent
yield loss at 7-hr mean concentrations between 0.04 and 0.05 ppm. Seven-hr
seasonal mean concentrations in excess of 0.08 ppm were required to cause a 10
percent yield loss in almost 19 percent of the species or cultivars. The data
indicate that approximately.il percent of the species or cultivars would
display a 10 percent loss at 7-hr seasonal mean concentrations below 0.035
ppnij suggesting that these plant types are very sensitive to On-induced yield
losses.
A review of the data in Table 1-5 indicates that the grain crops were
apparently generally less sensitive than the other crops to 0~. Mean yield
reductions at 0.04 ppm were predicted to be less than 5 percent for all the
species and cultivars tested except for the Roland and Vona wheat cultivars.
The data also demonstrate that sensitivity differences within a species may be
as large as differences between species. For example, at 0.04 ppm 0^, estimated
yield losses ranged from 2 to 15 percent in soybean and from 0 to 28 percent
in wheat. In addition to differences in sensitivity among species and cultivars,
the data in Figures 1-4 and 1-5 illustrate year-to-year variations in plant
response to 0,.
Several exposure-response models, ranging from simple linear to complex
nonlinear models, have been used to describe the relationship between plant
yield and 0- exposure. When exposure-response models are used, it is important
for the fitted equations not to show systematic deviation from the data points
2
and for the coefficient of determination (R ) to be high. Although linear
regression equations have been used to estimate yield loss, there appear to be
systematic deviations from the data for some species and cultivars even though
1-65
-------
9-
| o
3
e
OT 7 —
CC
1 6~
CC S~
o
05
O
UJ
to 3—
0.
o
CC 2-
0 *
1-
0-
10.8%
©
10.8%
©
13.S%
©
21,6%
(4)
13.5%
©
10.8%
©
18,9%
0
I
<0.035 | 0.040-0.044 | 0.050-0.059 I _>0.080
0.035-0.039 0.045-0:049 0.060-0.079
7-hr SEASONAL MEAN
OZONE CONCENTRATIONS, ppm
Figure 1 -6. Number and percentage of 37 crop species or cuitivars predicted to
show a 10 percent yield loss at various ranges of 7-hr seasonal mean ozone
concentrations. Concentration ranges and 10% yield loss data are derived from
Table 1 -5. Data represent 12 separate crop species; circled numbers represent
separate species for each concentration range.
1-66
-------
o
the equations have moderate-to-high coefficients of determination (R ). ,
Plateau-linear or polynomial equations appear to fit the data better. More
recently, a Weibull model has been used to estimate percentage yield loss
(Heck et al.s 1983). The Weibull model yields a curvilinear response line
that seems to provide a reasonable fit to the data. Based on available data,
it is recommended that curvilinear exposure-response functions be used to
describe and analyze plant response to 0~.
1.5.3.2 Yield Loss: Determination from Discrete Treatments. In addition to
the use of regression approaches in some studies, various other approaches
have been used to investigate the effects of 03 on crop yield. These studies
were designed to test whether specific 0., treatments were different from the
control rather than to develop exposure-response equations. In general, these
data were analyzed using analysis of variance. To summarize the data from
studies that used discrete treatments, the lowest 0, concentration that signi-
ficantly reduced yield was determined from analyses done by the authors (Table
1-6). The lowest concentration reported to reduce yield was frequently the
lowest concentration used in the study; hence it was not always possible to
estimate a no-effect exposure concentration. In general, the data indicate
that Q~ concentrations of 0.10 ppm (frequently the lowest concentration used
in the studies) for a few hours per day for several days to several weeks
generally caused significant yield reductions. Although it appears from this
analysis that a higher CL concentration was required to cause an effect than
was estimated from the regression studies, it should be noted that the concen-
trations derived from the regression studies were based on a 10 percent yield
loss,.while in studies using analysis of variance (Table 1-6) the 0.10 ppm
concentration frequently induced mean yield" losses of 10 to 50 percent.
1.5.3.3 Yield Loss: Determination with Chemical Protectants. Chemical
protectants (antioxidants) have been used to estimate the impact of ambient 0,,
on crop yield. In these studies, some plots were treated with the chemical
and others were not. Yield loss was determined by comparing the yield in the
plots treated with the chemical to the yield in untreated plots. When chemical
protectants are used, care must be used in interpreting the data because the
chemical itself may alter plant growth. The chemical may not be effective
against all concentrations of all pollutants in the study area, which would
result in an underestimation of yield loss. With an understanding of these
limitations, however, researchers have concluded that chemical protectants are
1-67
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TABLE 1-6. OZONE CONCENTRATIONS AT WHICH SIGNIFICANT YIELD LOSSES HAVE BEEN NOTED FOR
A VARIETY OF PLANT SPECIES EXPOSED UNDER VARIOUS EXPERIMENTAL CONDITIONS
Plant species
Alfalfa
Alfalfa
Pasture grass
Ladino clover
Soybean
Sweet corn
Sweet corn
Wheat
Radish
Beet
Potato
Pepper
Cotton
Carnation
Coleus
Begonia
Ponderosa pine
Western white
pine
Loblolly pine
Pitch pine
Poplar
Hybrid poplar
Hybrid poplar
Red maple
American
sycamore
Sweetgum
White ash
Green ash
Willow oak
Sugar maple
Exposure duration
7 hr/day, 70 days
2 hr/day, 21 day
4 hr/day, 5 days/wk, 5 wk
6 hr/day, 5 days
6 hr/day, 133 days
6 hr/day, 64 days
3 hr/day, 3 days/wk, 8 wk
4 hr/day, 7 day
3 hr
2 hr/day, 38 days
3 hr/day, every 2 wk,
120 days
3 hr/day, 3 days/wk, 11 wk
6 hr/day, 2 days/wk, 13 wk
24 hr/day, 12 days
2 hr
4 hr/day, once every 6 days
for a total of 4 times
6 hr/day, 126 days
6 hr/days, 126 days
6 hr/day, 28 days
6 hr/day, 28 days
12 hr/day, 5 mo
12 hr/day, 102 days
8 hr/day, 5 day/wk, 6 wk
8 hr/day, 6 wk
6 hr/day, 28 days
6 hr/day, 28 days
6 hr/day, 28 days
6 hr/day, 28 days
6 hr/day, 28 days
6 hr/day, 28 days
Yield reduction,
% of control
51, top dry wt
16, top dry.wt
20, top dry wt
20, shoot dry wt
55, seed wt/plant
45, seed wt/plant
13, ear fresh wt
30, seed yield
33, root dry wt
40, storage root dry wt
25, tuber wt
19, fruit dry wt
62, fiber dry wt
74, no. of flower buds
20, flower no.
55, flower wt
21, stem dry wt
9, stem dry wt
18, height growth
13, height growth
+1333, leaf abscission
58, height growth
50, shoot dry wt
37, height growth
9, height growth
29, height growth
17, total dry wt
24, height growth
19, height growth
12, height growth
Q3 concentration,
pptn
0.10
0.10
0.09
0.10
0.10
0.10
0.20
0.20
0.25
0.20
0.20
0.12
0.25
0.05-0.09
0.20
0.25
0.10
0.10
0.05
0.10
0.041
0,15
0.15
0.25
0.05
0.10
0.15
0.10
0.15
0.15
Reference
Neely et al. (1977)
Hoffman et al. (1975)
Horsman et al. (1980)
Blum et al. (1982)
Heagle et al. (1974)
Heagle et al. (1972)
Oshima (1973)
Shannon and Mulchi (1974)
Adedipe and Ormrod (1974)
Ogata and Haas (1973)
Pell et al. (1980)
Bennett et al. (1979)
Oshinia et al. (1979)
Feder and Campbell (1968)
Adedipe et al. (1972)
Reinert and Nelson (1979)
Wilhour and Neely (1977)
Wilhour and Neely (1977)
Wilhour and Neely (1977)
Wilhour and Neely (1977)
Wilhour and Neely (1977)
Patton (1981)
Patton (1981)
Dochinger and Townsend (1979)
Kress and Skelly (1982)
Kress and Skelly (1982)
Kress and Skelly (1982)
Kress and Skelly (1982)
Kress and Skelly (1982)
Kress and Skelly (1982)
-------
an objective method of assessing the effects of CU on crop yield, especially
in conjunction with other methods. Results of several studies with chemical
protectants showed decreased crop yield from exposure to ambient oxidants
(Table 1-7), Crop yields were reduced 18 to 41 perecent when the ambient
oxidant concentration exceeded 0.08 ppm for 5 to 18 days over the growing
season of the crop.
1.5.3.4 Yield Loss: Determination fromAmbient Exposures. A number of
research studies have demonstrated that ambient 0- concentrations in a number
of locations in the United States are sufficently high to impair plant yield.
Of studies to determine the impact of ambient oxidants (primarily 0~) on plant
yield, most have compared the yield differences between plants grown in ambient
air and those grown in charcoal-filtered air. Early research documented that
ambient oxidants reduced the yield and quality of citrus, grape, tobacco,
cotton, and potato (U.S. Environmental Protection Agency, 1978). Subsequent
studies substantiated the impacts of ambient oxidants on plant yield (Table
1-8). Over several years, bean yields varied from a 5 percent increase to a
22 percent decrease in response to 0, concentrations in excess of 0.06 ppm
(Heggestad and Bennett, 1981).
Studies conducted on eastern white pine in the southern Appalachian
mountains showed that ambient 0~ may have reduced the, radial growth of sensitive
individuals as much as 30 to 50 percent annually over the,last 15 to 20 years
(Mann et al. , 1980). Field studies in the San Bernardino National Forest
showed that during the last 30 years ambient 0- may have reduced height growth
of ponderosa pine by as much as 25 percent, radial growth by 37 percent,, and
the total wood volume produced by 84 percent (Miller et al., 1982). Calcula-
tions of biomass in these studies were based, however, on apparent reductions
in radial growth without standardization of radial growth data with respect to
tree age. "
1.5.3.5 YieldLoss Summary. Several general conclusions can be drawn from
the various approaches used to estimate crop yield loss. The data from the
comparisons of crop yield in charcoal-filtered and unfiltered air (ambient
exposures) clearly show that ambient levels of G~ are sufficiently elevated in
several parts of the country to impair the growth and yield of plants. The
data from the chemical protectant studies support and extend this conclusion
to other plant species. Both approaches indicate that the effects occur at
low mean concentrations, with only a few 0, occurrences greater than 0.08 ppm.
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TABLE 1-7. EFFECTS OF OZONE ON CROP YIELD .
AS DETERMINED BY THE USE OF CHEMICAL PROTECTANTSC
I
o
Species
Beans (green)
\ield reduction,
% of control
03 exposure,
ppm
Reference
41
Onion
Tomato
Bean (dry)
Tobacco
Potato
38
30
24
18
36
>0.08 for total
of 27 hr over
3.5 months
>0.08 on 5 days out
of 48
>0.08 on 15 days '
over 3 months
>0.08 on 11 days
(total of 34 hr)
over 3 months
>0.08 on 14 days
during the summer
>0.08 ppm on 18 days
(total of 68 hr)
over 3 months
Potato
25
Manning et al. (1974)
Wukasch and Hofstra (1977b)
Legassicke and Ormrod (1981)
Temple and Bisessar (1979)
Bisessar and Palmer (1984)
Bisessar (1982)
Clarke et al. (1983)
aAll the species were treated with the antioxidant, EDU, except the bean study by
Manning et al. (1974) which used the systemic fungicide, benomyl.
Yield reduction was determined by comparing the yields of plants treated with
chemical protectants (control) to those that were not treated.
GThis study was run over 2 years when the 03 doses were 65 and 110 ppm-hr,
respectively, but the yield loss was similar both years.
-------
TABLE 1-8, EFFECTS Of AMBIENT OXIDANTS ON YIELD OF SELECTED CROPS
Plant species
Tomato
(Fireball 861 VR)
Bean
(Tendergreen)
os
concentration,
ppm
0.035
(0.017-0.072)
0.041
(0.017-0.090)
Exposure duration
99 day average
(6:00 a.m. - 9:00 p.m.)
43 day average
(6:00 a.m. - 9:00 p.m.)
Yield, %
reduction
from control
33, fruit fresh
wt
26, pod fresh wt
Location
of study
New York
Reference
MacLean and
Schneider (1976)
Snap bean (3 cultivars:
Astro, BBL 274, BBL
290)
Soybean (4 cultivars:
Cutler, York, Clark,
Dare)
Forbs, grasses, sedges
Sweet corn
(Bonanza)
0.042
>0.05
0.052
0.051
0.035
>0.08
3 mo average
(9:00 a.m. - 8:00 p.m.)
31% of hr between
8:00 a.m. - 10:00 p.m.
from late June to mid-
September over three
summers; 5% of the time
the concentration was
>0.08 ppm
1979, 8 hr/day average
(10:00 a.m. - 6:00 p.m.),
April-September
1980, 8 hr/day average
(10:00 a.m. - 6:00 p.m.),
April-September
1981, 8 hr/day average
(10:00 a.m. - 6:00 p.m.),
April-September
58% of hr (6:00 a.m.
9:00 p.m.),
1 July-6 September
1, pod wt
20, seed wt
32, total above-
ground biomas
20, total above-
ground biomass
21, total above-
ground biomass
9, ear fresh wt
Mary!and
Maryland
Virginia
Virginia
California
Heggestad and
Bennett (1981)
Howell et al.
(1979); Howe11
and Rose (1980)
Duchelle et al.
(1983)
Thompson et al.
(1976a)
(Monarch Advance)
>0.08
28, ear fresh wt
-------
Growth and yield data from the previous criteria document (U.S. Environmental
Protection Agency, 1978), shown in Figure 1-3, indicate that effects on
growth and yield of several plant species occurred when the mean 0, concentra-
tion (for 4 to 6 hr/day) exceeded 0.05 ppm for at least 2 wk. The data from
the regression studies, conducted to develop exposure-response functions for
estimating yield loss, indicated that at least 50 percent of the species/culti-
vars tested were predicted to display a 10 percent yield loss at 7-hr seasonal
mean 0- concentrations of 0.05 ppm or less. Most of the data from the discrete
treatment studies did not use levels low enough to support these values directly.
The magnitude of yield losses reported at 0.10 ppm, however, indicate that
maintenance of a substantially lower concentration than 0.10 ppm is needed to
prevent 0- effects, although a specific value cannot be derived from the
discrete treatment studies.
1.5.4 Effectson CropQua!ity
Based on results of the few studies that have been conducted, 03 can
reduce crop quality in addition to reducing the total yield of the crop.
Quality is a general term that includes many features of the crop, such as
nutritional composition, appearance, taste, and ability to withstand storage
and shipment. Examples of On-induced alterations in quality are decreased oil
in soybean seeds (Howell and Rose, 1980; Kress and Miller, 1983); decreased
p-carotene, vitamin C, and carbohydrates in alfalfa (Thompson et al., 1976b;
Neely et al.s 1977); and increased reducing sugars that are associated with
undesirable darkening when potatoes are used to make potato chips (Pell et
al., 1980).
1.5.5 Statistics Used to Characterize Ozone Exposures
The characterization and representation of plant exposures to 0~ has
been, and continues to be a major problem. Research has not yet clearly
identified which components of the pollutant exposure cause the plant response.
Most studies have characterized the exposure by the use of mean 0,, concentra-
tions, although various averaging times have been used. Some studies have
also used cumulative 03 dose. The difficulty of selecting an appropriate
statistic to characterize plant exposure has been summarized by Heagle and
Heck (1980). Ambient and experimental 03 exposures have been presented as
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seasonal, monthly, weekly, or daily means; peak hourly means; number of hours
above a selected concentration; or the number of hours above selected concen-
tration intervals. None of these statistics adequately characterize the
relationships among 03 concentration, exposure duration, interval between
exposures, and plant response. The use of a mean concentration (with long
averaging times) (1) implies that all concentrations of 0-> are equally effec-
tive in causing plant responses and (2) minimizes the contributions of the
peak concentrations to the resonse. The mean treats low-level, long-term
exposures the same as high-concentration, short-term ones. Thus, the use of a
long-term mean concentration ignores the importance of peak concentrations; to
ignore the peaks is inconsistent with the literature.
The total ozone dose (concentration multiplied by time) has been used to
describe plant exposure; however, it suffers from the same problem as the
mean. The total dose is simply the summation of the ppm-hr over the study
period, which also treats all concentrations as being equally effective.
Several investigators have attempted to give greater importance to peak 0~
concentrations. For example, Oshima et al. (1977a,b) and Lefohn and Benedict
(1982) have summed only the ppm-hr of exposure greater than some preselected
value. Larsen et al. (1983) have introduced the concept of "impact" to describe
the effects of 03 and SCL on soybeans. The "impact (I)" is calculated similarly
to total dose, except the concentration is raised to an exponent greater than
W
one (I = C X T); this method of calculation effectively gives greater weight
to the higher concentrations. More recently, Larsen and Heck (1984) have
suggested the term "effective mean" to describe an approach in which greater
importance is given to higher concentrations. The "effective mean" is defined
as the average hourly impact raised to an exponent and divided by the duration.
Several lines of evidence suggest that higher concentrations should be
regarded as having the greater influence in determining the impact of 0~ on
vegetation. Studies have shown that plants can tolerate some combinations of
exposure duration and concentration without exhibiting foliar injury or effects
on growth or yield, illustrating that not all concentrations are equally
effective in causing a response. From the toxicological perspective, it is
the peaks or concentrations above some level that are most likely to have an
impact. Effects occur on vegetation when the amount of pollutant that the
plant has absorbed exceeds the ability of the organism to repair or compensate
for the impact.
1-73
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Studies with beans and tobacco (Heck et al., 1966) showed that a dose
(concentration times time) distributed over a short period induced more injury
than did the same dose distributed over a longer period. Tobacco studies
showed that the 0~ concentration was substantially more important than exposure
duration in causing foliar injury (Tonneijck, 1984). In beans, foliar injury
2
occurred when the internal 0~ flux exceeded 115 umoles/m in 1 hr (Bennett,
1979). A single 3-hr exposure, however, at approximately half the concentration
(0.27 compared with 0.49 ppm) required a 64 percent greater internal flux of
Og to produce the same amount of foliar injury as the 1-hr exposure required.
More recently, Amiro et al. (1984) showed that higher concentrations were more
important than low concentrations in causing injury. Their study also suggested
the existence of a biochemical injury threshold (i.e., the 0~ uptake rates
that plants can experience without incurring visible foliar injury). The
greater importance of concentration compared to exposure duration has also
been reported by other authors (e.g., Heck and Tingey, 1971; Henderson and
Reinert, 1979; Reinert and Nelson, 1979).
Studies with soybean (Johnston and Heagle, 1982), tobacco (Heagle and
Heck, 1974), and bean (Runeckles and Rosen, 1977) showed that plants exposed
to a low level of 03 for a few days became more sensitive to subsequent 0-
exposures. In studies with tobacco, Mukammal (1965) showed that a high Oj
concentration on one day caused substantial injury, whereas an equal or higher
concentration on the second day caused only slight injury. Using stress
ethylene as an indicator of 03 effects, Stan and Schicker (1982) showed that a
series of successive short exposures was more injurious to plants than a
continuous exposure at the same 0, concentration for the same total exposure
period. Walmsley et al. (1980) continuously exposed radishes to 03 for several
weeks and found that the plants acquired some 0~ tolerance. The acquired
tolerance displayed two components: (1) the exposed plants developed new
leaves faster than the controls, and (2) there was a progressive decrease in
sensitivity of the new leaves to Og. The newer leaves also displayed a slower
rate of senescence. The observations by Elkiey and Ormrod (1981) that the 0,
uptake decreased during a 3-day study period may provide an explanation for
the results with radish.
Not only are concentration and time important but the dynamic nature of
the G*3 exposure is also important; i.e. whether the exposure is at a constant
or variable concentration. Musselman et al. (1983) recently showed that
constant concentrations of 03 caused the same types of plant responses as
1-74
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variable concentrations at equivalent doses. Constant concentrations, however,
had less effect on plant growth responses than variable concentrations at
similar doses. Exposures of radishes to ambient 0» in open-top exposure
**»
chambers showed that significant yield reductions occurred when the maximum CL
«5
concentration exceeded 0.06 ppm at least 10 percent of the days when the crop
was growing (Ashmore, 1984). Initial studies have compared the response of
alfalfa to daily peak and episodic 0~ exposure profiles that gave the equivalent
total 0~ dose over the growing season (Hogsett et al., 1985). Alfalfa yield
was reduced to a greater extent in the episodic than in the daily peak exposure.
This study also illustrates the problem with the 7-hr seasonal mean concentra-
tion; i.e., it does not properly account for the peak concentrations. The
plants that displayed the greater growth reduction (in the episodic exposure)
were exposed to a significantly lower 7-hr seasonal mean concentration.
Studies with SOg also showed that plants exposed to variable concentrations
exhibited a greater plant response than those exposed to a constant concentra-
tion (Mclaughlin et al., 1979; Male et al., 1983).
1-5.6 Relationship Between Yield Loss and Foliar Injury
Because plant growth and production depend on photosynthetically functional
leaves, various studies have been conducted to determine the association
between foliar injury and yield for species in which the foliage is not part
of the yield. Some research has demonstrated significant yield loss with
little or no foliar injury (e.g., Tingey et al., 1971; Tingey and Reinert,
1975; Kress and Skelly, 1982; Feder and Campbell, 1968; Adedipe et al., 1972).
Other studies showed that significant foliar injury was not always associated
with yield loss (Heagle et al., 1974; Oshima et al.s 1975). The relative
sensitivities of two potato cultivars were reversed when judged by foliar
injury versus yield reductions (Pell et al., 1980). In field corn, foliar
injury occurred at a lower 0» concentration than yield reductions; but as the
03 concentration increased, yield was reduced to a greater extent than foliar
injury was increased (Heagle et al., 1979a). In wheat, foliar injury was not
a good predictor of 0~-induced yield reductions (Heagle et al., 1979b).
1.5.7 Physiological Basis of Yield Reductions
As discussed earlier in this summary, plant growth is the summation of a
series of biochemical and physiological processes related to uptake, assimila-
tion, biosynthesis, and trans!ocation. An impairment in these processes may
lead to reduced plant yield if the process is limiting.
1-75
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For plant growth to occur, plants must assimilate C02 and convert it into
organic substances; an inhibition in carbon assimilation may be reflected in
plant growth or yield. In several species 03 (at 0.05 ppm and higher) inhibited
photosynthesis, as measured by gas-exchange (e.g., U.S. Environmental Protection
Agency, 1978; Coyne and Bingham, 1978; Black et al., 1982; Bennett and Hill,
1974; Yang et al., 1983). Biochemical studies showed that 03 (0.12 ppm for 2
hr) inhibited an enzyme that catalyzes the assimilation of CO,, (Pell and
Pearson, 1983).
Ozone, in addition to decreasing the total amount of COy that is assimi-
lated, alters that pattern by which the reduced amount of assimilate is parti-
tioned throughout the plant. There is generally less photosynthate translo-
cated to the roots and to the reproductive organs (e.g., Tingey et al., 1971;
Jacobson, 1982; Oshima et al., 1978, 1979; Bennett et al., 1979). This reduces
root size and marketable yield as well as rendering the plant more sensitive
to injury from environmental stresses. Another consequence of reduced root
growth and altered carbon allocation is an impairment of symbiotic nitrogen
fixation (U.S. Environmental Protection Agency, 1978; Ensing and Hofstra,
1982).
The reproductive capacity (flowering and seed set) is reduced by 0, in
ornamental plants, soybean, corn, wheat, and other plants (Adedipe et al.,
1972; Feder and Campbell, 1968; Heagle et al., 1972, 1974; Shannon and Mulchi,
1974). These data suggest that 0, impairs the fertilization process in plants.
This suggestion has been confirmed in tobacco and corn studies using low
concentrations of 0, (0.05 to 0.1.0 ppm) for a few hours (Feder, 1968; Mumford
et al., 1972).
Ozone both in the field and in chamber studies stimulates premature
senescence and leaf drop (Menser and Street, 1962; Heagle et al., 1974;
Heggestad, 1973; Pell et al., 1980; Hofstra et al., 1978). In part, the
Oo-induced yield reduction has been attributed.to premature senescence. The
premature leaf drop decreases the amount of photosynthate that a leaf can
contribute to plant growth.
1.5.8 Factors Affecting Plant Response to Ozone
Numerous factors influence the type and magnitude of plant response to
Oo. Most studies of the factors influencing plant response have been limited
to effects on foliar injury; however, some studies have measured yield and
1-76
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some have researched the physiological basis for the influences. The para-
meters studied include environmental factors, biological factors, and inter-
actions with other air pollutants.
1.5.8,1 Environmental Conditions. Environmental conditions before and during
plant exposure are more influential than post-exposure conditions in determining
the magnitude of the plant response. The influence of environmental factors
has been studied primarily under controlled conditions, but field observations
have substantiated the results. Most studies have evaluated the influence of
only a single environmental factor and have relied primarily upon foliar
injury as the plant response measure. Some generalizations of the influence
of environmental factors can be made:
1. Light conditions that are conducive to stomatal opening appear to
enhance 0., injury (U.S. Environmental Protection Agency, 1978).
Light is required to induce stomatal opening, which permits the
plant to absorb pollutants.
2. No consistent pattern relating plant response to temperature has
been observed (U.S. Environmental Protection Agency, 1978). Plants
do not appear to be as sensitive at extremely high or low tempera-
tures, however, as they are under more moderate conditions.
3. Plant, injury tends to increase with increasing relative humidity
(U.S. Environmental Protection Agency, 1978). The relative humidity
effect appears to be related to stomatal aperture, which tends to
increase with increasing relative humidity. McLaughlin and Taylor
(1981) demonstrated that plants absorb significantly more CL at high
humidity than at low humidity. It is generally accepted that plants
in the eastern United States are injured by lower concentrations of
CL .than their counterparts in California; this phenomenon has been
attributed to differences in humidity (U.S. Environmental Protection
. Agency, 1978).
4. As soil moisture decreases, plant water stress increases and there
is a reduction in plant sensitivity to 0-, (U.S. Environmental Protec-
tion Agency, 1978). The reduced 03 sensitivity is apparently related
to stomatal closure, which reduces 0~ uptake (U.S. Environmental
Protection Agency, 1978; Olszyk and Tibbitts, 1981; Tingey et a!.,
1982). Water stress does not confer a permanent tolerance to 0~;
1-77
-------
once the water stress has been alleviated, the plants regain their
sensitivity to 03 (Tingey et a!., 1982).
1.5.8.2 Interaction with Plant Diseases. Ozone can affect the development of
disease in plant populations. Laboratory evidence suggests that 0., (at ambient
concentrations or greater for 4 hr or more) inhibits infection by pathogens
and subsequent disease development (Laurence, 1981; Heagle, 1982; U.S. Environ-
mental Protection Agency, 1978). Increases, however, in diseases from "stress
pathogens" have been noted. For example, plants exposed to 03 were more
readily injured by Botrytis than plants not exposed to 03 (Manning et a!.,
1970a,b; Wukasch and Hofstra, 1977a,b; Bisessar, 1982). Both field and labora-
tory studies have confirmed that the roots and cut stumps of 03-injured ponderosa
and Jeffrey pines are more readily colonized by a root rot (Heterobasidion
annosus). The degree of infection was correlated with the foliar injury
(James et al., 1980; Miller et a!., 1982). Studies in the San Bernardino
National Forest showed that 0.,-injured trees were predisposed to attack by
bark beetles and that fewer bark beetles were required to kill an 0~-injured
tree (Miller et al., 1982).
1.5.8.3 Interaction of Ozone with Other Air Pollutants. The report of Menser
and Heggestad (1966) provided the initial impetus for studying the interaction
of 0- with SOp. They showed that Bel W-3 tobacco plants exposed to CU (0.03
ppm) or S02 (0.24 to 0.28 ppm) were uninjured but that substantial foliar
injury resulted when the plants were exposed to both gases simultaneously.
Subsequent studies have confirmed and extended the observation that combinations
of 03 and S02 may cause more visible injury than expected based on the injury
from the individual gases. This injury enhancement (synergism) is most common
at low concentrations of each gas and also when the amount of foliar injury
induced by each gas, individually, is small. At higher concentrations or when
extensive injury occurs, the effects of the individual gases tend to be less
than additive (antagonistic). In addition to foliar injury, the effects of
pollutant combinations have also been investigated in relation to other plant
effects, and these have been discussed in several reviews and numerous individual
reports (e.g., Reinert et al., 1975; Ormrod, 1982; Jacobson and Colavito, 1976;
Heagle and Johnston, 1979; Olszyk and Tibbitts, 1981; Flagler and Youngner,
1982; Foster et al., 1983; Heggestad and Bennett, 1981; Heagle et al., 1983a).
1-78
-------
Field studies have investigated the influence of SOp on plant response to
0-3 at ambient and higher concentrations in several plant species: soybean
(Heagle et al., 19835; Reich and Amundson, 1984), beans (Qshima, 1978; Heggestad
and Bennett, 1981), and potatoes (Foster et al., 1983). In these studies, G3
altered plant yield but SCL had no significant effect and did not interact
with On to reduce plant yield unless the SO,, exposure concentrations and
frequency of occurrence were much greater than the concentrations and fre-
quencies of occurrence typically found in the ambient air in the.United States.
The applicability of the yield results from pollutant combination studies
to ambient conditions is not known. An analysis of ambient air monitoring
data for instances of co-occurrence of 03 and SQp indicated that at sites
where the two pollutants were monitored, they both were present for ten or
fewer periods during, the growing season (Lefohn and Tingey, 1984). Co-
occurrence was defined as the simultaneous occurrence of hourly averaged
concentrations of 0.05 ppm or greater for both pollutants. At this time, it
appears that most of the studies of the effects on pollutant combinations (0_
and SOp) on plant yield have used a longer exposure duration and a higher
frequency of pollutant co-occurrence than are found in the ambient air.
Only a few studies have investigated the effects of 03 when combined with
pollutants other than SOp, and no clear trend is available. Preliminary
studies using three-pollutant mixtures (03, SOp, NQp) showed that the additions
of SOp and NOp (at low concentrations) caused a greater growth reduction than
DO alone.
1.5.9 Economic Assessment of Effects of Ozone on Agriculture
Evidence from the plant science literature clearly demonstrates that fl-
at ambient levels will reduce yields of some crops (see Chapter 6, Section
6.4.3.2.2). In view of the importance of U.S. agriculture to both domestic
and world consumption of food and fiber, such reductions in crop yields could
adversely affect human welfare. The plausibility of this premise has resulted
in numerous attempts to assess, in monetary terms, the losses from ambient 0~
or the benefits of Q~ control to agriculture. Many of these assessments have
been performed since publication of the 1978 03 criteria document (U.S.
Environmental Protection Agency, 1978). The utility of these post-1978
studies in regulatory decision-making can be evaluated in terms of how well
the requisite biological, aerometric, and economic inputs conform to specific
criteria, as discussed in Section 6.5 of Chapter 6.
1-79
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While a complete discussion of the criteria for evaluating economic
*s
assessments is not appropriate here, it is instructive to highlight certain
key issues. First, the evidence on crop response to 03 should reflect how
crop yields will respond under actual field conditions. Second, the air
quality data used to frame current or hypothetical effects of CL on crops
O
should represent the actual exposures sustained by crops in each production
area. Finally, the assessment methodology into which such data are entered
should (1) capture the economic behavior of producers and consumers as they
adjust to changes in crop yields and prices that may accompany changes in 0™
air quality; and (2) ideally, should accurately reflect institutional considera-
tions, such as regulatory programs, that may result in market distortions.
The assessments of 0^ damages to agriculture found in the literature
display a range of procedures for calculating economic losses, from simple
monetary calculation procedures to more complex economic assessment methodol-
ogies. The simple procedures calculate monetary effects by multiplying
predicted yield or production changes resulting from exposure to 0., by an
assumed constant crop price, thus failing to recognize possible crop price
changes arising from yield changes as well as not accounting for the processes
underlying economic response. Conversely, a rigorous economic assessment will
provide estimates of the benefits of air pollution control that account for
producer-consumer decision-making processes, associated market adjustments,
and perhaps some measure of distributional consequences between affected
parties. It is important to distinugish between those studies based on naive
or simple models and those based on correct procedures, since the naive proce-
dure may be badly biased, leading to potentially incorrect policy decisions.
Most of the post-1978 economic assessments focus on 03 effects in specific
regions, primarily California and the Corn Belt (Illinois, Indiana, Iowa,
Ohio, and Missouri). This regional emphasis may be attributed to the relative
abundance of data on crop response and air quality for selected regions, as
well as the national importance of these agricultural regions. Economic
estimates for selected regions are presented in Table 1-9. In addition to
reporting the monetary loss or benefit estimates derived from each assessment,
this table provides some evaluation of the adequacy of the plant science,
aerometric, and economic data, and assumptions used in each assessment.
Adequacy as defined here does not mean that the estimates are free of error;
rather, it implies that the estimates are based on the most defensible biologic,
1-80
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TABLE 1-9. SUMMARY OF ESTIMATES OF REGIONAL ECONOMIC CONSEQUENCES OF OZONE POLLUTION
Reference and
study region
Crops
Annual benefits
of control,
$ million
Evaluation of critical data and assumptions
Plant response data
Aeronetric data
Economic model data
Additional comments
Adams et al. 12 annual crops: $45 (in 1976
(1982);
Southern
California
beans, broccoli,
cantaloupes,
carrots, cauli-
flower, celery,
lettuce, onions,
potatoes, tomatoes
cotton, and sugar
beets.
dollars)
Inadequate; uses Larsen- Adequate; exposure
Heck (1976) foliar injury measured as cumu-
models converted to yield lative seasonal
losses. exposure in
excess of Cali-
fornia standard
(0.08 ppn), from
hourly data col-
lected for sites
closest to produc-
tion regions.
Adequate; a price endo-
genous mathematical
(quadratic) programming
model reflecting agro-
nomic, environmental,
and economic conditions
in 1976.
Economic effect measured as a
change in economic surplus (sum
of consumers and producers'
surpluses) between base case
(actual Oa levels in 1976)'
aftd economic surplus that
would be realize.d if all
regions were in compliance with
1971 photochemical oxidant
standard of 0.08 ppm.
i
00
Lueng et al. 9 crops: lemons, $103 (in 1975
(1982); oranges (Valencia dollars)
Southern and Navel), straw-
California berry, tomato,
alfalfa, avocado,
lettuce, and celery.
Inadequate; 03-yield
response functions
estimated from second-
ary data on crop yields.
Adequate for some
regions; exposure
measured in aver-
age monthly con-
centration in ppm
for 12 hr period
(7:00 a.m. to
7:00 p.m.). Data
from 61 Calfornia
Air Resources
Board monitoring
sites.
Adequate on demand side;
economic model is
composed of linear
supply and demand
curves for each crop
estimated with data
from 1958-1977, but
ignores producer-level
adjustments.
Economic effect is measured as
a change in economic surplus
between base case (1975) and a
clean air environnent reflecting
zero 03.
Howitt et al. 13 crops: alfalfa, From $35 (bene-
(1984a,b); barley, beans, fit of control
California celery, corn, to 0.04 ppm) to
cotton, grain sor- $117 (loss for
ghum, lettuce, Increase to
onions, potatoes, 0.08 ppm) (in
rice, tomatoes, 1978 dollars).
and wheat.
Adequate for some crops; Adequate; Califor- Adequate; economic model Economic effects measured as
most response functions nia Air Resources similar to Adams et al. changes in economic surplus
derived from NCLAN data Board data for (1982) but includes some across three 03 changes fron
through 1982. Surrogate monitoring sites perennial crops and re- 1978 actual levels. These
responses used for celery, closest to rural fleets 1978 economic and include changes in ambient 03
onions, rice and potatoes production areas. technical environment.
are questionable. Exposure measured
as the seasonal
7-hr average in
each production
area for compati-
bility with NCLAN
exposure.
to 0.04, 0.05, and 0.08 ppm
across all regions.
-------
TABLE 1-9 (cont'd). SUMMARY OF ESTIMATES OF REGIONAL ECONOMIC CONSEQUENCES OF OZONE POLLUTION
Reference and
study region
Rowe et al.
(1984);
San Joaquin
Valley In
California
h- '
1
00
J^J
Adaas and
MeCarl
(1985);
Corn Belt
Annual benefits
of control ,
Crops $ million
14 annual and $43 to $117
perennial crops: depending on
alfalfa, barley, degree of
beans, carrots, control,
corn, cotton, nteasured in
grain sorghum, 1978 dollars.
grass hay, grapes,
pasture, potatoes,
saf flower,
tomatoes and
wheat.
3 crops: corn, $668 (in 1980
soybeans, and dollars)
wheat.
Evaluation of critical data and assumptions3
Plant response data
Adequate for $om crops;
response functions based
on both experimental and
secondary data, Host
crops from NCLAN data.
Responses for the remain-
ing crops were based on
surrogate responses of
similar crops in the
data set.
Adequate; 03 yield
response information
from NCLAN "for 3 yr
(1980-1982). Yield
adjustments estimated
from Weibull response
models.
Aerometric data
Adequate; 4 expo-
sure levels were
tested. The aver-
age hourly concen-
tration was used
in roost functions
to predict changes.
All data were froi
California Air
Resources Board
monitoring sites in
predominantly rural
areas.
Adequate except
for linkage of
7-hr seasonal
mean to hourly
standards. Data
are interpolated
from SAROAD
monitoring sites
by Krigingb
procedure ,
measured as
1980 seasonal
7-hr average.
Regulatory
analysis assumes
that 03 is log-
normally
distributed.
Economic model data
Adequate; sane as in
Howitt et al.
(1984a,b).
Adequate; economic
estimates are generated
by a mathematical pro-
gramming model of U.S.
agriculture reflecting
1980 conditions. Farm-
level response is
portrayed by 12
individual "represen-
tative" farm models
to generate supply
adjustments used in
the national-level
model .
Additional comments
Economic effects Measured as the
change in economic surplus be-
tween the 1978 base case and three
increasingly stringent control
scenarios: (1) a 50% reduction in
in no. of hr >0.10 ppm; (2)
meeting the current standard of
0.10 ppm; and (3) meeting an 03
standard of 0.08 ppm.
Economic estimates represent
changes in economic surplus
(sum of consumers' and pro-
ducers' surpluses) between
current (1980) 03 levels and
increases and decreases in
ambient 03 levels. Reduction
to a uniform ambient level of
0,04 ppm across all regions
results in benefits of $668
million.
-------
TABLE 1-9 (cont'd). SUMMARY OF ESTIMATES OF REGIONAL ECONOMIC CONSEQUENCES OF OZONE POLLUTION
i
00
CO
Reference and
study region Crops
Hjelde et al, 3 crops: corn,
(1984); soybeans, and
Illinois wheat.
Page et al. 3 crops: corn,
(1982); soybeans and
Ohio River wheat.
Basin
Annual benefits
of control ,
$ million
Ranges from
$55 to $220
annually for
period 1976
to 1980.
$7.022 measured
as present
value of pro-
ducer losses
for period
1976 to 2000.
Annual i zed
losses are
approx. $270
in 1976
dollars.
Evaluation of critical data and assumptions3
Plant response data
Adequate when cross*
checked against NCLAN
data; responses are
estimated from secon-
dary (non-experimental)
data on actual farmer
yield, input, and 03
concentrations. Results
are translated into yield
effects and compared to
NCLAN data from Illinois.
Inadequate; crop losses
provided by Loucks and
Armentano (1982);
responses derived by
synthesis of existing
experimental data.
•*
Aerometric data
Adequate; same
Kn'ged data set as
used in Adams and
HcCarl (1985),
except only for
Illinois and
cover 5 yr
(1976-1980).
Exposure is mea-
sured as seasonal
7-hr average to
facilitate compa-
rison with NCLAN
response estimates.
Inadequate; dose
measured as cumu-
lative seasonal
exposure for a
7-hr period
(9:30 a.m. to
4:30 p.m.)
Monitoring sites
at only 4 loca-
tions were used
to characterize
the regional
exposure.
Economic model data
Adequate at producers
level; economic model
consists of a series
of annual relationships
on fanners' profits
These functions).
These functions are
adjusted to represent
changes in 03 (±25X)
for each year. Model
does not include consumer
(demand) effects.
Inadequate; the econo-
mic model consists of
• regional supply curves
for each crop. The
predicted changes in
production between
"clean air" case and
each scenario are used
to shift crop supply
curves. The analysis
ignores price changes
from shifts in supply.
Additional comments
The estimates represent increases
in farmers' profits that could
arise for a 25% reduction in 03
for each year (1976-1980). Years
with higher ambient levels have
highest potential increase in
profits for changes.
Losses are measured as differ-
ences in producer surplus across
the various scenarios. Since
prices are assumed fixed (In
real terms) over the period,
no consumer effects are
measured.
-------
TABLE 1-9 (cont'd). SUMMARY OF ESTIMATES OF REGIONAL ECONOMIC CONSEQUENCES OF OZONE POLLUTION
Reference and
study region
Crops
Annual benefits
of control,
$ Million
Evaluation of critical data and assumptions
Plant response data
Aerometric data
EconoMic node! data
Additional comments
Benson et al. 4 crops: alfalfa, $30.5 (measured Inadequate; but innova-
(1982); wheat, corn, and in 1980 dollars) tive crop loss models
Minnesota potatoes. Cultivar estimated using experi-
believed to be mental yield-03 data
,_, limited to one per from other researchers.
i crop. Crop loss modeling
2 includes both chronic
and espisodic response
and crop development
stage as factors in
yield response, by
regressing yield on 03
exposures for various
time windows, during the
growing season.
Adequate; air
quality data are
for state of
Minnesota for
1979 and 1980.
Exposure measured
several ways but
generally as a
daily exposure sta-
tistic reflecting
either sun of hourly
averages or the mean
hourly average.
Adequate on demand side;
The economic estimates
are derived from a
comprehensive economic
model calibrated to
1980 values.
The economic effect measured
in terms of short-run profit
changes for Minnesota producers.
If yields are assumed to change
only in Minnesota then losses to
Minnesota producers are $30.5
million. If yields change in
Minnesota and the rest of U.S.,
then producers gain $67 million
as a result of increases in crop
prices.
Adequacy as defined here does not mean that the estimates are free of error; rather, it implies that the estimates are based on the most defensible
biologic, aerometric, or economic information and models currently available.
Kriging is a spatial interpolation procedure that has been used to generate 03 concentration data for rural areas in which no monitoring sites have
been established. See Heck et al. (1983b).
-------
aerometric, or economic information and models currently available in the
literature. The estimates can then be ranked relative to the strength of
these data and assumptions. Of the eight regional studies reviewed, most have
adequate economic models, but only four are judged adequate across all input
categories. Further, most regional studies abstract from the interdependences
that exist between regions, which limits their utility in evaluating secondary
national ambient air quality standards (SNAAQS).
National-level studies can overcome this limitation of regional analyses
by accounting for economic linkages between groups and regions. A proper
accounting for these linkages, however, requires additional data and more
complex models, and frequently poses more difficult analytical problems.
Thus, detailed national assessments tend to be more costly to perform. As a
result, there are fewer assessments of pollution effects at the national than
at the regional level. Six national-level assessments performed since the
last criteria document was published in 1978 are reported in Table 1-10. Of
these, two used the simple "price times quantity" approach to quantify dollar
effects. Four used more defensible economic approaches. As with Table 1-9,
an evaluation of the adequacy of critical plant science, aerometric, and
economic data is presented, along with the estimates of benefits or damages.
As is evident from the evaluation, most of the national studies reviewed
here suffer from either plant science and aerometric data problems, incomplete
economic models, or both. As a result of these limitations, decision-makers
should be cautious in using these estimates to evaluate the efficiency of
alternative SNAAQS. Two of the studies, however, are judged to be much more
adequate in terms of the three critical areas of data inputs. Together, they
provide reasonably comprehensive estimates of the economic consequences of
changes in ambient air 03 levels on agriculture.
In the first of these studies, Kopp et al. (1984) measured the national
economic effects of changes in ambient air 0., levels on the production of
corn, soybeans, cotton, wheat, and peanuts. In addition to accounting for
price effects on producers and consumers, the assessment methodology used is
notable in that it placed emphasis on developing producer-level responses to
03~induced yield changes (from NCLAN data) in 200 production regions. The
results of the Kopp et al. (1984) study indicated that a reduction in 0., from
1978 regional ambient levels to a seasonal 7-hr average of approximately 0.04
ppm would result in a $1.2 billion net benefit in 1978 dollars. Conversely,
1-85
-------
TABU 1-10. SWHftRY OF ESTIMATES OF NATIONAL ECONOMIC CONSEQUENCES OF OZONE POLLUTION
Annual benefits
of control, Evaluation of critical data and assumptions3
Study Crops
Ryan at al. 16 crops: alfalfa,
(1981) beets, broccoli,
cabbage, corn
(sweet and field),
hay, Una beans,
oats, potatoes,
sorghun, soybeans,
spinach, tobacco,
tomatoes, and
wheat.
i — »
i Shriner et al. 4 crops: corn,
g (1982) soybeans, wheat,
and peanuts.
Multiple cultivars
of all crops but
peanuts.
Adams and 3 crops; corn,
Crocker (1984) soybeans, and
cotton. Two corn
cultivars, three
soybean, two
cotton.
$ billion Plant response data
$1.747 (in 1980 Inadequate; yield-response
dollars). information derived fron
a synthesis of 5 yield
studies in the literature
prior to 1980. Synthe-
sized response functions
estimated for both chronic
and acute exposures
for six crops. For
the remaining 10 crops
surrogates are used.
Yield changes are based
on reductions in Q3 to
meet 1980 Federal stan-
dard of 0.12 ppm In non-
compliance counties.
$3.0 (in 1978 Adequate; analysis uses
dollars). NCLAN response data for
1980. Functions esti-
mated in linear fora.
Yield changes reflect
difference between 1978
ambient 03 levels of
each county and assumed
background of 0.025 ppm
concentration.
$2.2 (in 1980 Adequate; analysis uses
dollars). NCLAN 03-yield data for
1980 and 1981. Functions
estimated in linear form.
Yield changes measured
between 1980 ambient
levels and an assumed 03
concentration of 0.04 ppm
across all production
regions.
Atroaetric data
Inadequate; dose
measured in sev-
eral ways to
correspond to
underlying
response function.
Qa data derived
froa National
Aeroisetric
Data Bank and
from Lawrence
Berkeley
Laboratory, for
period 1974-1976.
Unknown; exposure
may be measured as
highest 7-hr.
average, rather
than 7-hr NCLAN
average. Rural
ambient concen-
trations for 1978
tconoaic node! data
Inadequate; naive econo-
mic model. Monetary
impact calculated by
multiplying changes in
county production by
crop price in 1980.
Measures impact on
producers only.
Inadequate; same as Ryan
et al. (1981) except
uses 1978 crop prices.
Additional comments
Dollar estimate Is for the 531
counties exceeding the
Federal standard of 0.12 ppm.
This study is essentially an
updated version of Benedict
et al. (1971) reported in 1978
criteria document.
Dollar estimates are for all
counties producing the four
crops. As with Ryan et al.
(1981), estimates are for
for producer level effects
only.
estimated by Kriging0
procedure applied
to SAROAD data.
Adequate; 1980
ambient Q3 levels
estimated by
Kriging of SAROAD
monitoring sites,
translated into a
seasonal 7-hr
average.
Adequate on demand side;
inadequate on modeling
producer behavior; eco-
nomic model consists of
crop demand and supply
curves. Corresponding
price and quantity
adjustments result in
changes in economic
surplus. No producer
level responses
modeled; only measures
aggregate effects.
Economic estimate measured in
terms of changes in consumer
and producer surpluses associated
with the change in 03.
-------
TABLE 1-10 (cont'd). SUMMARY OF ESTIMATES OF NATIONAL ECONOMIC CONSEQUENCES OF OZONE POLLUTION
Study
Adams et al.
(1984a)
Annual benefits
of control ,
Crops $ billion
4 crops: corn, $2.4 (in 1980
soybeans, wheat, dollars).
and cotton. Two
cultivars for corn
and cotton, three
for soybeans and
and wheat.
Evaluation of critical data and assumptions3
Plant response data Aerometric data
Adequate; analysis uses Adequate; same
NCLAN 03-yield data for as Adams and
1980 through 1982. Yield Crocker (1984).
changes measured between
1980 ambient levels and
25% reduction in
Oa across all regions.
Functions estimated in
both linear and quadratic
form.
Economic model data
Inadequate producer
model; same as Adams
and Crocker (1984),
except that analysis
examines range of
economic estimates
reflecting variability
in yield predictions
resulting from sample
size and functional
form.
Additional comments
Same as Adams and Crocker (1984).
Linear functions result in higher
yield losses and hence higher
economic loss estimates.
Reported estimate ($2.4 billion)
is for quadratic response
function.
Kopp et al,
(1984)
i
co
5 crops; corn,
soybeans, wheat,
cotton, and
peanuts. Multiple
cultivars of each
crop except peanuts.
$1.2 (in 1978
dollars).
Adequate; analysis uses
NCLAN Os yield response
data for 1980 through
1982. Yield losses (for
estimates reported here)
measured as the differ-
ence between ambient 1978
03 and a level assumed to
represent compliance with
an 0.08 ppin standard.
Adequate; same as
Adams and Crocker
(1984) and Adams
et al. (1984b)
but for 1978
growing season.
Adequate; economic model
consists of producer-
level models, by crop,
for numerous production
regions. Predicted
yield changes are used
to generate supply
shifts for each region/
crop combined with crop
demand relationships
to estimate producer
and consumer surpluses.
In addition to measuring the
change in economic surplus for
various assumed 03 levels, the
analysis also includes an exam-
ination of the sensitivity of
the estimates to the nature of
the demand relationships used
in the model.
Adams et al.
(1984b)
6 crops: barley,
corn, soybeans,
cotton, wheat, and
sorghum. Multiple
cultivars used for
each crop except
barley and grain
sorghum; two for
cotton, three for
wheat, two for corn,
and nine for soybean.
$1.7 (in 1980
dollars).
Adequate; analysis uses
NCLAN 03 yield response
data for 1980 through
1983. Yield changes
reflect changes from
1980 ambient 03 of 10
and 40% reduction and
a 25% increase for each
response.
Adequate; same as
above but for 1980
and 1976 through
1980 periods.
Adequate; economic model
consists of two compo-
nents: a series of farm-
level models for each of
55 production regions
and a national model
of crop use and demand.
Yield changes are used
to generate regional
supply shifts used in
national model.
Consumer surplus estimated
for both domestic and foreign
markets; producer surplus
nationally and by region. The
analysis includes a range of
economic estimates reflecting.
changes in response and 03
data and assumptions.
Adequacy as defined here does not mean that the estimates are free of error; rather, it implies that the estimates are based on the most defensible
biologic, aerometric, or economic information and models currently available.
is a spatial interpolation procedure that has been used to generate 03 concentration data for rural areas in which no monitoring sites have
been established. See Heck et al. (1983b).
-------
an increase in 0^ to an assumed ambient concentration of 0.08 ppm (seasonal
7-hr average) across all regions produced a net loss of approximately $3.0
billion.
The second study, by Adams et al. (1984b), is a component of the NCLAN
program. The results were derived from an economic model of the U.S. agricul-
tural sector that includes individual farm models for 55 production regions
integrated with national supply-and-demand relationships for a range of crop
and livestock activities. Using NCLAN data, the analysis examined yield
changes for six major crops (corn, soybeans, wheat, cotton, grain, sorghum,
and barley) that together account for over 75 percent of U.S. crop acreage.
The estimated annual benefits (in 1980 dollars) from 0., adjustments are sub-
stantial, but make up a relatively small percentage of total agricultural
output (about 4 percent). Specifically, in this analysis, a 25 percent reduc-
tion in ozone from 1980 ambient levels resulted in benefits of $1.7 billion.
A 25 percent increase in ozone resulted in an annual loss (negative benefit)
of $2,363 billion. When adjusted for differences in years and crop coverages,
these estimates are quite close to the Kopp et al. (1984) benefit estimates.
While the estimates from both Kopp et al. (1984) and Adams et al. (1984b)
were derived from conceptually sound economic models and from the most defen-
sible plant science and aerometric data currently available, there are several
sources of uncertainty. These include the issue of exposure dynamics (7-hr
per day exposures from the NCLAN experiments versus longer exposure periods,
such as 12-hr exposures), and the lack of environmental interactions, particu-
larly Q~~moisture stress interactions, in many of the response experiments.
Also, the 0, data in both studies are based on a limited set of the monitoring
sites in the SAROAD system of EPA, mainly sites in urban and suburban areas.
While the spatial interpolation process used for obtaining 0- concentration
data (Kriging) results in a fairly close correspondence between predicted and
actual 03 levels at selected validation points, validation requires more
monitoring sites in rural areas. The economic models, with their large number
of variables, and parameters, and the underlying data used to derive these
values, contain potential sources of uncertainty, including the effects on bene-
fits estimates of market-distorting factors such as the Federal farm programs.
The inclusion of these possible improvements in future assessments is not
likely, however, with the possible exception of market-distorting factors, to
alter greatly the range of agricultural benefits provided in the Kopp et al.
(1984) and Adams et al. (1984b) studies, for several reasons. First, the
1-88
-------
current studies cover about 75 to 80 percent of U.S. agricultural crops (by
value). For inclusion of the other 20 percent to change the estimates signifi-
cantly would require that their sensitivities to 03 be much greater than for
the crops included to date. Second, model sensitivity analyses from existing
studies indicate that changes in key plant science parameters must be substan-
tial to translate into major changes in economic estimates. From experience
to date it seems unlikely that use of different dose measures or interaction
effects would result in changes of the magnitude already addressed in some of
the sensitivity analyses. Third, even if there are such changes, there are
likely to be countervailing responses; e.g., longer exposure periods may
predict greater yield losses but 03~water stress tends to dampen or reduce the
yield estimates. Finally, it should be noted that potential improvements in
economic estimates are policy-relevant only to the extent that they alter the
relationship between total benefits and total costs of that policy. Uncertain-
ties in other effects categories are probably greater.
In conclusion, the recent economic estimates of benefits to agriculture
of 03 control, particularly those estimates by Kopp et al. (1984) and Adams et
al. (1984b), meet the general criteria discussed in Section 6.5 and hence
provide the most defensible evidence given in the literature to date of the
general magnitude of such effects. Relative to estimates given in the 1978
criteria document (U.S. Environmental Protection Agency, 1978) and economic
information on most other 0- effects categories (non-agricultural), these two
studies, in combination with the underlying NCLAN data on yield effects,
provide the most comprehensive economic information to date on which to base
judgments regarding the economic efficiency of alternative SNAAQS. As noted
above, there are still gaps in plant science and aerometric data and a strong
need for meteorological modeling of 03 formation and transport processes for
use in formulating rural 03 scenarios. With regard to the economic data and
models used, the impact of factors that upset free-market equilibria needs
further analysis. Additionally, it must be emphasized that none of the studies
has accounted for the compliance costs of effecting changes in 0, concentra-
tions in ambient air. For a cost-benefit analysis to be complete, the annual-
ized estimated benefits to agriculture that would result from 0, control would
have to be combined with benefits accruing to other sectors and then compared
with the overall annualized compliance costs.
1-89
-------
1.5.10 Effects of Peroxyacetyl Nitrate on Vegetation
Peroxyacetyl nitrate (PAN) is a highly phytotoxic air pollutant that is
produced by photochemical reactions similar to those that produce 0,. Both CL
and PAN can coexist in the photochemical oxidant complex in ambient air. The
effects of PAN were a concern in southern California for almost 20 years
before the phytotoxicity of 0, under ambient conditions was identified. The
symptoms of photochemical oxidant injury that were originally described (prior
to 1960) were subsequently shown to be identical with the symptoms produced by
PAN. Following the identification of PAN as a phytotoxic air pollutant, PAN
injury (foliar symptoms) has been observed throughout California and in several
other states and foreign countries.
1.5.10.1 Factors Affecting PlantResponse to PAN. Herbaceous plants are
sensitive to PAN and cultivar differences in sensitivity have been observed in
field and controlled studies. Trees and other woody species, however, are
apparently resistant to visible foliar injury from PAN (Taylor, 1969; Davis,
1975, 1977).
Taylor et al. (1961) demonstrated that there is an absolute requirement
for light before, during, and after exposure or visible injury from PAN will
not develop. Field observations showed that crops growing under moisture
stress developed little or no injury during photochemical oxidant episodes
while, adjacent to them, recently irrigated crops were severly injured (Taylor,
1974).
Only a few studies have investigated the effects of PAN and 03 mixtures
on plants. When plants were exposed to both gases at their respective injury
thresholds, no interaction between the gases was found (Tonneijck, 1984). At
higher concentrations, the effects were less than additive. Studies with
petunia confirmed that 0^ tended to reduce PAN injury (Nouchi et al., 1984).
1.5.10.2 LimitingValues of Plant Response. The limiting-value method has
been used to estimate the lowest PAN concentration and exposure duration
reported to cause visible injury on various plant species (Jacobson, 1977).
The analysis yielded the following range of concentrations and exposure dura-
tions likely to induce foliar injury: (1) 200 ppb for 0.5 hr; (2) 100 ppb for
1.0 hr; and (3) 35 ppb for 4.0 hr.
Other studies, however, suggest that these values need to be lowered by
30 to 40 percent to reduce the likelihood of foliar injury (Tonneijck, 1984).
For example, foliar injury developed on petunia plants exposed at 5 ppb PAN
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for 7 hr (Fukuda and Terakado, 1974). Under field conditions, injury symptoms
may develop on sensitive species when PAN concentrations reach approximately
15 ppb for 4 hr (Taylor, 1969).
1.5.10.3 Effects of PAN on Plant Yield. Only a few limited studies have been
conducted to determine the effects of PAN on plant growth and yield. In
greenhouse studies, radish, oat, tomato, pinto bean, beet, and barley were
exposed to PAN concentrations of up to 40 ppb for 4 hr/day, twice/wk, from
germination to crop maturity (Taylor et a!., 1983). No significant effects on
yield were detected. This is supportive of field observations, in which
foliar injury from ambient PAN exposures was found but no evidence was seen of
reduced yield in these crops. In contrast, lettuce and Swiss chard exposed to
PAN concentrations of up to 40 ppb for 4 hr/day, twice/wk, from germination to
crop maturity showed yield losses up to 13 percent (lettuce) and 23 percent
(Swiss chard) without visible foliar injury symptoms (Taylor et a!., 1983).
The results indicate that PAN at concentrations below the foliar-injury
threshold can cause significant yield losses in sensitive cultivars of leafy
vegetable crops. In addition, photochemical oxidant events have caused foliar
injury on leafy vegetables (Middleton et a!., 1950) for which the foliage is
the marketable portion. After severe PAN damage, entire crops may be unmarket-
able or else extensive hand work may be required to remove the injured leaves
before the crop may be marketed.
A comparison of PAN concentrations likely to cause either visible injury
or reduced yield with measured ambient concentrations (see Chapter 5) indicates
that it is unlikely that ambient PAN will impair the intended use of plants in
the United States except in some areas of California and possibly in a few
other localized areas.
1.6 EFFECTS OF OZONE ON NATURAL ECOSYSTEMS AND THEIR COMPONENTS
1.6.1 Responses of Ecosystems to Ozone Stress
The responses to ozone of individual species and subspecies of herbaceous
and woody vegetation are well documented. They include (1) injury to foliage,
(2) reductions in growth, (3) losses in yield, (4) alterations in reproductive
capacity, and (5) alterations in susceptibility to pests and pathogens, espec-
ially "stress pathogens" (National Research Council, 1977; U.S. Environmental
Protection Agency, 1978; this document, Chapter 6). The responses elicited by
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ozone in individual species and subspecies of primary producers (green plants)
have potential consequences for natural ecosystems because effects that alter
the interdependence and interrelationships among individual components of
populations can, if the changes are severe enough, perturb ecosystems. Because,
however, of the numerous biotic and abiotic factors known to influence the
response of ecosystem components such as trees (see, e.g., Cowling, 1985;
Manion, 1985), it is difficult to relate natural ecosystem changes to ozone
specifically, and especially to ozone alone. Ozone can only be considered a
contributing factor.
Evidence indicates that any impact of ozone on ecosystems will depend on
the responses to ozone of the producer community. Producer species (trees and
other green plants) are of particular importance in maintaining the integrity
of an ecosystem, since producers are the source, via photosynthesis, of all
new organic matter (energy/food) added to an ecosystem. Any significant
alterations in producers, whether induced by ozone or other stresses, can
potentially affect the consumer and decomposer populations of the ecosystem,
and can set the stage for changes in community structure by influencing the
nature and direction of successional changes (Woodwell, 1970; Bormann, 1985),
with possibly irreversible consequences (see, e.g., Odum, 1985; Bormann,
1985).
1.6.2 Effects of Ozone on Producers
In forest ecosystems, tree populations are the producers. As such, they
determine the species composition, trophic relationships, and energy flow and
nutrient cycling of forest ecosystems (Ehrlich and Mooney, 1983). Ozone-
induced effects on the growth of trees has been clearly demonstrated in
controlled studies (see Chapter 6). For example, Kress and Skelly (1982)
showed the following reductions in growth in height in seedlings exposed to
ozone for 6 hr/day for 28 days: American sycamore, 9 percent (0.05 ppm 0-);
sweetgum, 29 percent (0.10 ppm 03); green ash, 24 percent (0.10 ppm); willow
oak, 19 percent (0.15 ppm 0~); and sugar maple, 25 percent (0.15 ppm). Similar
results have been obtained for other tree species by other investigators
(e.g., Dochinger and Townsend, 1979; Mooi, 1980; Patton, 1981; Kress et al.,
1982). Some species, however, have been shown to exhibit increased growth in
short-term ozone exposures (e.g., yellow poplar and white ash; Kress and
Skelly, 1982). Hogsett et al. (1985) found reductions in growth in height, in
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radial growth, and in root growth in slash pine seedlings exposed for up to
112 days to 7-hr seasonal mean concentrations of 0.104 ppm 0- (with a 1-hr
daily maximum of 0.126 ppm 0,) and 0.076 ppm 0_ (with a 1-hr daily maximum of
0.094 ppm 03).
Field studies on the Cumberland Plateau (near Oak Ridge, TN) have shown
reductions in growth in eastern white pine exposed to ambient air 0., concentra-
tions >0.08 ppm (1-hr) (Mann et a!., 1980), with 1-hr concentrations ranging
over the multi-year study from 0.12 ppm to 0.2 ppm (Mclaughlin et al., 1982).
It should be noted, however, that in the Mclaughlin et al. (1982) study trees
classified as ozone-tolerant sustained greater percentage reductions in radial
growth in the last 4 years (1976 to 1979) of the 1962 to 1979 period for which
growth was examined than the reductions observed in trees classified as ozone-
sensitive. In the Blue Ridge Mountains of Virginia, Benoit et al. (1982)
found reductions in radial growth of sensitive eastern white pine in a multi-
year study in which 1-hr 0~ concentrations were generally 0.05 to 0.07 ppm but
peaked at ^0.12 ppm on as many as 5 consecutive days at a time.
The concentrations of ozone reported for sites on the Cumberland Plateau
and in the Blue Ridge Mountains may not fully represent the actual exposures
at those sites, however, since measurements were made in the daytime only.
For species in which stomates remain open at night, such as eastern white
pine, the possible occurrence of peak ozone concentrations at night, from
transported urban plumes, is an important consideration for accurately
assessing concentration-response relationships.
Exposures of trees and other producers to ozone have been shown to reduce
photosynthesis (e.g., Miller et al., 1969; Botkin et al., 1972; Barnes, 1972;
Carlson, 1979;.Coyne and Bingham, 1981; Yang et al., 1983; Reich and Amundson,
1985) and to alter carbohydrate allocation, especially the partitioning of
photosynthate between roots and tops (e.g., Price and Treshow, 1972; Tingey
et al., 1976; Mclaughlin et al., 1982). Krause et al. (1984) have associated
growth reductions in ozone-exposed seedlings with foliar leaching. All three
of these effects have been postulated as mechanisms of the reduced growth seen
in ozone-exposed vegetation.
Responses to ozone are not uniform among plants of the same species and
the same approximate age. Differential responses have been attributed in part
to differences in genetic potential (e.g., Mann et al., 1980; Coyne and Bingham,
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1981; Benoit et al., 1982). In addition, the age of the plant and its develop-
mental stage at time of exposure influence its response to ozone (see Chapter 6).
Other factors, as well, influence the types and magnitude of plant responses
to ozone, including such macro- and microenvironmental factors as temperature,
relative humidity, soil moisture, light intensity, and soil fertility (see
Chapter 6).
Trees may respond rapidly to CL stress. Needles of sensitive eastern
white pine usually exhibit injury symptoms within a few days after exposure to
high CL concentrations. In other instances, responses are more subtle and may
not be observable for years because trees are perennials and must therefore
cope over time with the cumulative effects of multiple short- and long-term
stresses. Reductions in the growth of annual rings observed in ponderosa,
Jeffrey, and eastern white pine have been attributed to the exposure of the
trees to 0, over a period of 10 to 20 years (Miller and Elderman, 1977; Miller
et al., 1982; Mclaughlin et al., 1982; Benoit et al., 1982). Decline and
dieback of red spruce in the northeastern United States and reduced growth
rates of red spruce, balsam fir, and Fraser fir in central West Virginia and
western Virginia also have been attributed to stresses, to which air pollution
is a possible contributor, that began at least 20 years ago (Johnson and
Siccama, 1983; Adams et al., 1985).
1.6.3 Effects of Ozone on Other Ecosystem Components and on Ecosystem
Interactions
Evidence for the effects of ozone on other ecosystem components indicates
that most are indirect, occurring chiefly as a result of the direct effects of
ozone on trees and other producers. Significant alterations in producer
species can change the ability of a species to compete and thus can influence
the nature and direction of successional changes in the ecosystem. Likewise,
significant alterations in producers can result in changes in the consumer and
decomposer populations that depend on producers as their food source. Studies
in the San Bernardino Mountain ecosystems in the 1970s have provided some
evidence of successional shifts and of predisposition to infestation by pests
and pathogens as the result of oxidant-induced changes in ponderosa arid Jeffrey
pines (see Section 1.6.4 below).
Marked morphological deterioration of the common lichen species, Hypogymm'a
enteromorpha, was documented in areas of the San Bernardino Mountains having
high oxidant concentrations. A comparison of the species of lichens found
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growing on ponderosa and Jeffrey pine with collections from the early 1900's
indicated the presence of 50 fewer species (Sigal and Nash, 1983).
McCool et al. (1979) and Parmeter et al. (1962) reported decreases in
mycorrhizal infections and rootlets in ozone-stressed citrange (a citrus
hybrid) and ponderosa pine, respectively. Mahoney (1982), on the other hand,
found no evidence of impairment in the development of mycorrhizal associations
in loblolly pine seedlings exposed to ozone plus sulfur dioxide; however,
shoot dry weight was decreased by 12 percent.
The effects of ozone on mycorrhizae are of particular note here, since
mycorrhizae are essential for the optimal development of most plants because
of the functions they perform. Mycorrhizal fungi increase the solubility of
minerals, improve the uptake of nutrients for host plants, protect roots
against pathogens, produce plant growth hormones, and move carbohydrates from
one plant to another (Hacskaylo, 1972). Ozone may disrupt the association
between mycorrhizal fungi and plants, possibly by inhibiting photosynthesis
and reducing the amounts of sugars and carbohydrates available for transfer
from leaves of producers to the roots. Mycorrhizae are known to be sensitive
to alterations in carbon allocation to the roots in host,plants (Hacskaylo,
1973).
Because of the complex interactions among plants, pests, pathogens, and
other biotic and abiotic factors, Laurence and Weinstein (1981) have emphasized
the critical importance of examining pollutant-pathogen and poll ant-insect
interactions in determining the growth impact of a pollutant. Manion (1985)
has emphasized the necessity of taking non-pollutant stresses, both biotic and
abiotic, into account when attempting to attribute changes in forest ecosystems
to air pollutants.
1.6.4 Effects of Ozone on Specific Ecosystems
One of the most thoroughly studied ecosystems in the United States is the
mixed-conifer forest ecosystem in the San Bernardino Mountains of southern
California. Sensitive plant species there began showing injury in the early
1950's (Miller and Elderman, 1977) and the source of the injury was identified
as oxidants (ozone) in 1962 (Miller et al., 1963). In an inventory begun in
1968, Miller found that sensitive ponderosa and Jeffrey pines were being
selectively removed by oxidant air pollution. Mortality of 8 and 10 percent
was found in two respective populations of ponderosa pine studied between 1968
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and 1972. Monitoring in that period showed ozone concentrations >0.08 ppm for
£1300 hours, with concentrations rarely decreasing below 0.05 ppm at night
near the crest of the mountain slope (Miller, 1973).
In a subsequent interdisciplinary study (1973 through 1978), biotic and
abiotic components and ecosystem processes were examined. The ecosystem
components most directly affected were various tree species, the fungal micro-
flora of needles, and the foliose lichens on the bark of trees. In May through
September, 1973 through 1978, 24-hr-average ozone concentrations ranged from
about 0.03 to 0.04 ppm to about 0.10 to 0.12 ppm. (Monitoring was done by the
Mast meter through 1974 and by the UV method from 1975 through 1978). Foliar
injury on sensitive ponderosa and Jeffrey pine was observed when the 24-hr-
average ozone concentrations were 0.05 to 0.06 ppm (Miller et al., 1982).
Injury, decline, and death of these species were associated with the major
ecosystem changes observed (Miller et al., 1982).
Growth reductions attributable to oxidant air pollution were calculated
by McBride et al. (1975) for ponderosa pine saplings. Assuming 1910 to 1940
to be a period of low oxidant pollution and 1944 to 1974 a period of high
oxidant pollution, they used radial growth increments (dbh) to calculate an
oxidant-induced decrease in diameter of 40 percent. On the basis of the
3-year growth of saplings in. filtered and nonfiltered air in portable green-
houses, they calculated oxidant-induced reductions of 26 percent in height
growth (McBride et al., 1975). No standardized methods, for determining tree
ring widths were available at the time of this study.
Carbon flow and mineral nutrient cycling were influenced by the accumula-
tion of litter under stands with the most severe needle injury and by defolia-
tion, as well as by a reduction in the number of species and the population
density of the fungi that normally colonize living needles and later participate
in decomposition. The most likely result of heavy litter accumulation is a
reduction in pine seedling establishment and greater establishment and growth
of oxidant-tolerant understory species on some sites and oxidant-tolerant
trees on other sites (Miller et al., 1982).
Changes in the energy available to trees influenced the biotic interac-
tions, so that weakened ponderosa pines were more susceptible to attack by
predators such as bark beetles and to pathogens such as root rot fungi (Stark
and Cobb, 1969). Fewer western pine beetles were required to kill weakened
trees (Dahlsten and Rowney, 1980); and stressed pines became more susceptible
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to root rot fungi (James et al., 1980) and showed a decrease in mycorrhizal
rootlets and their replacement by saprophytic fungi (Parmeter et al., 1962).
Accelerated rates of mortality of ponderosa and Jeffrey pine in the
forest overstory, resulting from 03 injury, root rot, and pine beetle attack,
and in some cases, removal by fire, changed the basic structure of the forest
ecosystem (Phase IV; Bormann, 1985) by causing replacement, of the dominant
conifers with self-perpetuating, fire-adapted, 03-tolerant shrub and oak
species, which are considered less beneficial than the former pine forest and
which inhibit reestablishment of conifers (Miller et al., 1982).
Injury to vegetation in other ecosystems has also been reported. Duchelle
et al. (1983) found reductions in the growth and productivity of graminoid and
forb vegetation in the Shenandoah National Park, where 1-hr ozone concentra-
tions ranged from 0.08 to 0.10 ppm in the 3-year study period, with 1-hr
concentrations >0.06 ppm occurring for 1218, 790, and 390 hours in 1979, 1980,
and 1981, respectively. Treshow and Stewart (1973) fumigated species that
grow in the Salt Lake Valley and the Wasatch Mountains in Utah and found key,
dominant species to be ozone-sensitive. The National Park Service (1985) has
recently reported ozone-induced injury to vegetation in the Santa Monica
Mountains National Recreational Area, the Sequoia and Kings Canyon National
Parks, Indiana Dunes National Lakeshore, Great Smoky Mountains National Park,
and the Congaree Swamp National Monument. ,The impact of injury to vegetation
in these ecosystems has not been appraised.
It should be emphasized that the relative importance of a given species
in a given ecosystem must be considered in any assessment of the impact of
ozone (or other stresses) on an ecosystem. Ozone has not had the impact on
other ecosystems that it has had on the San Bernardino mixed-conifer forest
because the plant species injured do not have a role equal in importance to
the role of ponderosa and Jeffrey pines in the San Bernardino ecosystem.
1.6.5 Economic Valuation of Ecosystems
At the present time, economists and ecologists remain unable to devise a
mutually acceptable framework for estimating the economic value of ecosystems.
In addition, the credibility of any attempt to estimate at present the economic
value of ecosystems would be diminished by a lack of scientific data (1) on
the time-course of the manifestation of stress-induced effects on ecosystems,
(2) on the point at which ecosystems lose the capacity for self-repair, and
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(3) on the points at which they begin to lose their ability to provide,
respectively, priced and unpriced benefits to society. In addition, estimation
of the economic losses that might be associated with the specific effects of
ozone on ecosystems requires other data that are presently in short supply,
i.e., better and more aerometric data and better and more data on additional
variables, so that significant contributions from abiotic factors other than
ozone, as well as from biotic factors, can be credibly estimated.
1.7 EFFECTS OF OZONE AND OTHER PHOTOCHEMICAL OXIDANTS ON NONBIOLOGICAL MATERIALS
Over two decades of research show that ozone damages certain nonbiological
materials; the amount of damage to actual in-use materials, however, is poorly
characterized. Knowledge of indoor/outdoor ozone gradients, for example, has
expanded considerably in recent years, and this type of exposure information
has not been incorporated in materials damage studies. Moreover, virtually
all materials research on photochemical oxidants has focused on ozone. Theo-
retically, a number of the less abundant oxidants may equal or surpass ozone
in reactivity with certain materials, but this possibility has not been tested
empirically. In the absence of photochemical pollution, oxidative damage to
certain materials still occurs from atmospheric oxygen, but at a much reduced
rate and through different chemical mechanisms. Generally, ozone damages
elastomers by cracking along the line of physical stress, whereas oxygen
causes internal damage to the material.
The materials most studied in ozone research are elastomers and textile
fibers and dyes. 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 ozone on these compounds is well known, and
dose-response relationships have been established and corroborated by several
studies. These relationships, however, must be correlated with adequate expo-
sure 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 dose-cracking rate is reduced
considerably, although the extent of reduction differs widely according to the
material and the type and amount of protective measures used.
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The formation of cracks and the depth of cracking in elastomers are re-
lated to ozone dose and are influenced greatly by humidity and mechanical
stress. Dose is defined as the product of concentration and time of exposure.
The importance of ozone dose was demonstrated by Bradley and Haagen-Smit
(1951), who used a specially formulated ozone-sensitive natural rubber.
Samples exposed to ozone at a concentration of 20,000 ppm cracked almost
instantaneously, and those exposed to lower concentrations took a propor-
tionately longer time to crack. At concentrations of 0.02 to 0.46 ppm, and
under 100-percent strain, the cracking rate was directly proportional to the
time of exposure, from 3 to 65 min.
Similar findings were reported by Edwards and Storey (1959), who exposed
two SBR elastomers to ozone at a concentration of 0.25 ppm for 19 to 51 hr
under 100-percent strain. With ozone doses of 4.75 ppm-hr to 12.75 ppm-hr, a
proportional rate in cracking depth was observed, averaging 2.34 pm/hr for
cold SBR and 4.01 um/hr for hot SBR. When antiozonants were added to the com-
pounds, the reduction in cracking depth rate was proportional to the amount
added. Haynie et al. (1976) exposed samples of a tire sidewall to ozone at
concentrations of 0.08 and 0.5 ppm for 250 to 1000 hr under 10 and 20 per-
cent-strain. Under 20-percent strain, the mean cracking rate for 0.08 ppm was
1.94 um/hr. From these and other data, they estimated that at the ozone stan-
dard of the time (0.08 ppm, 1-hr average), and at the annual NO standard of
X
0.05 ppm, it would take 2.5 years for a crack to penetrate cord depth.
In addition to stress, factors affecting the cracking rate include atmos-
pheric pressure, humidity, sunlight, and other atmospheric pollutants. Veith
and Evans (1980) found a 16-percent difference in cracking rates reported from
laboratories located at various geographic elevations.
Ozone has been found to affect the adhesion of plies (rubber-layered
strips) in tire manufacturing. Exposure to ozone concentrations of 0.05 to
0.15 ppm for a few hours significantly decreased adhesion in an NR/SBR blend,
causing a 30-percent decrease at the highest ozone level. This adhesion prob-
lem worsened at higher relative humidities. When fast-blooming waxes and
antiozonants or other antioxidants were added, only the combination of protec-
tive measures allowed good adhesion and afforded protection from ozone and
sunlight attack. Wenghoefer (1974) showed that ozone (up to 0.15 ppm), espe-
cially in combination with high relative humidity (up to 90 percent), caused
greater adhesion losses than did heat and N02 with or without high relative
humidity.
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The effects of ozone on dyes have been known for nearly three decades.
In 1955, Salvin and Walker exposed certain red and blue anthraquinone dyes to
a 0.1 ppm concentration of ozone and noted fading, which until that time was
thought to be caused by N02> Subsequent work by Schmitt (I960, 1962) confirmed
the fading action of ozone and the importance of relative humidity in the
absorption and reaction of ozone in vulnerable dyes. The acceleration in
fading of certain dyes by high relative humidity was noted later by Beloin
(1972, 1973) at an ozone concentration of 0.05 ppm and relative humidity of 90
percent. Kamath et al. (1982) also found that a slight rise in relative humidity
(85 to 90 percent) caused a 20-percent dye loss in nylon fibers.
Both the type of dye and the material in which it is incorporated are
important factors in a fabric's resistance to ozone. Haynie et al. (1976) and
Upham et al. (1976) found no effects from ozone concentrations of 0.1 to 0.5
ppm for 250 to 1000 hr under high and low relative humidity (90 vs. 50 percent)
on royal blue rayon-acetate, red rayon-acetate, or plum cotton. On the other
hand, Haylock and Rush (1976, 1978) showed that anthraquinone dyes on nylon
fibers were sensitive to fading from ozone at a concentration of 0.2 ppm at 70
percent relative humidity and 40°C for 16 hr. Moreover, the same degree of
fading occurred in only 4 hr at 90 percent relative humidity. At higher
concentrations, there was a parallel increase in fading. Along with Heuvel et
al. (1978) and Salvin (1969), Haylock and Rush (1976, 1978) noted the importance
of surface area in relation to the degree of fading. In explaining this
relationship, Kamath et al. (1982) found that ozone penetrated into the fiber
itself and caused most of the fading through subsequent diffusion to the
surface.
Field studies by Nipe (1981) and laboratory work by Kamath et al. (1982)
showed a positive association between ozone levels and dye fading of nylon
materials at an ozone concentration of 0.2 ppm and various relative humidities.
In summary, dye fading is a complex function of ozone concentration, relative
humidity, and the presence of other gaseous pollutants. At present, the
available research is insufficient to quantify the amount of damaged material
attributable to ozone alone. Anthraquinone dyes incorporated into cotton and
nylon fibers appear to be the most sensitive to ozone damage.
The degradation of fibers from exposure to ozone is poorly characterized.
In general, most synthetic fibers like modacrylic and polyester are relatively
resistant, whereas cottons nylon, and acrylic fibers have greater but varying
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sensitivities to the gas. Ozone reduces the breaking strength of these fibers,
and the degree of reduction depends on the amount of moisture present. Under
laboratory conditions, Bogaty et al. (1952) found a 20 percent Toss in breaking
strength in cotton textiles under high-moisture conditions after exposure to a
0.06 ppm concentration of ozone for 50 days; they equated these conditions to
a 500- to 600-day exposure under natural conditions. Kerr et al. (1969) found
a net loss of 9 percent in breaking strength of moist cotton fibers exposed to
ozone at a concentration of 1.0 ppm for 60 days. The limited research in this
area indicates that ozone in ambient air may have a minimal effect on textile
fibers, but additional research is needed to verify this conclusion.
The effects of ozone on paint are small in comparison with those.of other
factors. Past studies have shown that, of various paints, only vinyl and
acrylic coil coatings are affected, and that this impact has a negligible
effect on the useful life of the material coated. Preliminary results of
current studies have indicated a statistically significant effect of ozone and
relative humidity on latex house paint, but the final results of those studies
are needed before conclusions can be drawn.
For a number of important reasons, the estimates of economic damage to
materials are far from reliable. Most of the available studies are now out-
dated in terms of the ozone concentrations, technologies, and supply-demand
relationships that prevailed when the studies were conducted. Additionally,
little was (and is) known about the physical damage functions, and cost esti-
mates were simplified to the point of not properly recognizing many of the
scientific complexities of the impact of ozone. Assumptions about exposure to
ozone generally ignored the difference between outdoor and indoor concentrations.
Also, analysts have had difficulty separating ozone damage from other factors
affecting materials maintenance and replacement schedules. For the most part,
the studies of economic cost have not marshalled factual observations on how
materials manufacturers have altered their technologies, materials, and methods
in response to ozone. Rather, the analysts have merely made bold assumptions
in this regard, most of which remain unverified through the present time.
Even more seriously, the studies followed engineering approaches ,that, do
not conform with acceptable methodologies for measuring economic welfare.
Almost without exception, the studies reported one or more types of estimated
or assumed cost increases, borne by materials producers, consumers, or both.
The recognition of cost increase is only a preliminary step, however, towards
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evaluating economic gains and losses. The analysis should then use these cost
data to proceed with supply and demand estimation that will show how materials
prices and production levels are shifted. Because the available studies fail
to do this, there is a serious question as to what they indeed measure.
Increased ozone levels increase sales for some industries even as they
decrease welfare for others. For example, manufacturers of antiozonants for
automobile tires conceivably stand to increase sales as ozone increases, while
purchasers of tires stand to pay higher prices. This is only one illustration
of a fundamental analytical deficiency in the various studies of materials
damage: the absence of a framework for identifying gainers and losers, and the
respective amounts they gain and lose.
Among the various materials studies, research has narrowed the type of
materials most likely to affect the economy from increased ozone exposure.
These include elastomers and textile fibers and dyes. Among these, natural
rubber used for tires is probably the most important economically for the
following reasons: (1) significant ambient air exposure and long use life;
(2) significant unit cost; and (3) large quantities and widespread distribution.
The study by McCarthy et al. (1983) calculated the cost of antiozonants
in tires for protection against ozone along with the economic loss to the
retread industry. While limitations in this study preclude the reliable
estimation of damage costs, the figures indicate the magnitude of potential
damage from exposure to ozone in ambient air.
Research has shown that certain textile fibers and dyes and house paint
are also damaged by ozone, but the absence of reliable damage functions make
accurate economic assessments impossible. Thus, while damage to these materials
is undoubtedly occurring, the actual damage costs cannot be estimated confi-
dently.
It is apparent from the review presented in this chapter that a great
deal of work remains to be done in developing quantitative estimates of mate-
rials damage from photochemical oxidant exposures. This is not meant to
deprecate the years of research reported in this document, for much has been
gained in refining the initial methodologies used for assessing damage. Yocom
et al. (1985) have summarized the current state of knowledge:
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We have learned that some costs may be difficult to quantify either
because they are minimal or because they are overshadowed by other factors,
such as wear or obsolescence. We have learned that damage functions are
complex and are influenced by the presence of other pollutants and by weather.
We have learned that more accurate estimates of materials in place may be
obtained using selective sampling and extrapolation. And we have learned that
a mere cost-accounting of damage does not present a true estimate of economic
cost if it does not account for the welfare effects induced by shifts in the
supply-demand relationship.
1.8 TOXICOLOGICAL EFFECTS OF OZONE AND OTHER PHOTOCHEMICAL OXIDANTS
1.8.1 Introduction
The biological effects of CU have been studied extensively in animals and
a wide array of toxic effects have been ascribed to 0- inhalation. Although
much has been accomplished to improve the existing data base, refine the con-
centration-response relationships and interpret better the mechanisms of 0-
effects, many of the present data were not accumulated with the idea that
quantitative comparisons to man would be drawn. In many cases, only qualita-
tive comparisons can be made. To maximize the extent that animal toxicological
data can be used to estimate the human health risk of exposure to 03, the
qualitative as well as quantitative similarities between the toxicity of 0~ to
animals and man must be considered more carefully in the future. Significant
advances have been made in understanding the toxicity of 03 through appropriate
animal models. This summary highlights the significant results of selected
studies that will provide useful data for better predicting and assessing, in
a scientifically sound manner, the possible human responses to 0,.
Summary figures and tables are presented in the following sections. The
practical purpose of this presentation of the data is to help the reader focus
on what types of effects or responses have been reported, what concentrations
have been tested (1.0 ppm and lower), and as a convenient list of references
with each of the biological parameters measured. Studies were selected for
inclusion in these figures and tables on the basis of specific criteria presen-
ted below:
/
1. Studies have been cited when the reported effects are clearly due to 0^
exposure. Effects due to mixtures of 0, with other pollutants have been
summarized in a separate figure and table. Studies involving exercise,
diet deficiencies, or other possible modifiers of response to Og have not
been included.
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2. Cited studies report the effects of 03 exposure over a broad range of
animal species and strains and for varying lengths of time. Specific
details on animal species, exposure duration, and observed biological
effects can be obtained from the tables in Chapter 9.
3. Each closed symbol on the figures represents one or more studies conducted
at that particular concentration that caused effects. Specific references
can be found in the accompanying tables.
4. Each open circle represents one or more studies that used the given
concentration, but reported no significant effects. No-effect levels are
also indicated by brackets in the accompanying tables.
5. Only pulmonary function effects were divided by short-term (<14 days) and
long-term exposures to follow the discussion in the text.
In order to keep this section brief and concise, it was necessary to be somewhat
selective in determining what and how this information would be presented. A
number of important factors, such as the specific length of exposure, were not
included. Also, the parameter selected to illustrate a specific response was
usually broad and very general. For example, the category "decreases in
maerophage function" includes such diverse endpoints as measurements of lyso-
somal and phagocytic activity, maerophage mobility, or chemotactic response.
These responses may or may not be related to one another. Thus, care must be
taken in how these data are used and interpreted. The only appropriate use is
to gain an overview of the broad array of the effects of ozone and the concen-
trations which did and did not cause these effects.
1.8.2 Regional Dosimetry in the Respiratory Tract
The amount of 0, acting at a given site in the lung is related to the
airway luminal concentration at that level. As a result, 0, does not immediately
interact with cellular components of the respiratory tract. Instead, it first
comes into contact with the mucous or surfactant layer lining the airway. It
should be noted that 0- is quite reactive chemically. Reactions with components
of this layer cause an increase in total absorption of Q~ in the upper airways
and in a reduction of the amount of 0, reaching sensitive tissues. The site
1-104
-------
at which uptake and subsequent interaction occur and the local dose (quantity
of 0, absorbed per unit area per time), along with cellular sensitivity, will
determine the type and extent of the injury. Also, the capacity for responding
to a specific dose may vary between animals and humans because of dissimilari-
ties in detoxification systems, pharmacokinetics, metabolic rates, genetic
makeup, or other factors. Thus, along with the above, a knowledge of the
complex process of gas transport and absorption is crucial to understanding
the effects of 0- and other oxidants in humans.
The animal studies that have been conducted on ozone absorption are
beginning to indicate the quantity and site of 03 uptake in the respiratory
tract. Experiments on the nasopharyngeal removal of 03 in animals suggest
that the fraction of 0^ uptake depends inversely on flow rate, that uptake is
greater for nose than for mouth breathing, and that tracheal and chamber
concentrations are positively correlated. Only one experiment measured 0,
uptake in the lower respiratory tract, finding 80 to 87 percent uptake by the
lower respiratory tract of dogs (Yokoyama and Frank, 1972). At present,
however, there are no reported results for human nasopharyngeal- or lower
respiratory tract absorption. Caution must be used in estimating nasopharyngeal
uptake for normal respiration based upon experiments employing unidirectional
flows.
To further an understanding of 0, absorption, mathematical models have
been developed to simulate the processes involved and to predict 0., uptake by
various regions and sites within the respiratory tract. The model of Aharonson
et al. (1974) has been used to analyze nasopharyngeal uptake data. Applied to
0, data, the model indicates that the average mass transfer coefficient in the
nasopharyngeal region increases with increasing air flow, but the actual
percent uptake decreases.
Three models have been developed to simulate lower respiratory uptake
(McJilton et al., 1972; Hiller et al., 1978b, 1985). These models are very
similar in their treatment of 03 in the airways (taking into account convection,
diffusion, wall losses, and ventilatory patterns) and in their use of morpho-
logical data to define the dimensions of the airways and liquid lining. The
models differ in their treatment of the mechanism of absorption. Both of the
models of Miller and co-workers take into account chemical reactions of 0-
with constituents of the liquid lining, whereas the model of McJilton et al.
does not. The models of Miller et al. differ in their treatment of chemical
1-105
-------
reactions, as well as in the fact that the newer model includes chemical reac-
tions of Og in additional compartments, such as tissue and blood.
Tissue dose is predicted by the models of Miller et al. to be relatively
low in the trachea, to increase to a maximum between the junction of the
conducting airways and the gas-exchange region, and then to decrease distally.
This is not only true for animal simulations (guinea pig and rabbit) but it
is also characteristic of the human simulations (Miller et al., 1978b;
1985).
A comparison of the results of Miller and co-workers with morphological
data (that shows the centriacinar region to be most affected by 0,) indicates
qualitative agreement between predicted tissue doses and observed effects in
the pulmonary region. However, comparisons in the tracheobronchial region
indicate that dose-effect correlations may be improved by considering other
expressions of dose such as total absorption by an airway and by further
partitioning of the mucous layer compartment in mathematical models. Further
research is needed to define toxic mechanisms, as well as to refine our know-
ledge of important chemical, physical, and morphological parameters.
At present, there are few experimental results that are useful in judging
the validity of the modeling efforts. Such results are needed, not only to
understand better the absorption of 0, and its role in toxicity, but also to
support and to lend confidence to the modeling efforts. With experimental
confirmation, models which further our understanding of the role of CL in the
respiratory tract will become practical tools.
The consistency and similarity of the human and animal lower respiratory
tract dose curves obtained thus far lend strong support to the feasibility of
extrapolating to man the results obtained on animals exposed to CU. In the
past, extrapolations have usually been qualitative in nature. With additional
research in areas which are basic to the formulation of dosimetry models,
quantitative dosimetric differences among species can be determined. If in
addition, more information is obtained on species sensitivity to a given dose,
significant advances can be made in quantitative extrapolations and in making
inferences about the likelihood of effects of 0- in man. Since animal studies
are the only available approach for investigating the full array of potential
disease states induced by exposure to 0,, quantitative use of animal data is
in the interest of better establishing 03 levels to which man can safely be
exposed.
1-106
-------
1.8.3 Effects of Ozone on the Respiratory Tract
1.8.3.1 Morphological Effects. The morphological changes which follow exposure
o
to less than 1960 ug/m (1.0 ppm) 0,, are very similar in all species of labora-
tory mammals studied. Of the many specific cell types found in the respiratory
system, two types, ciliated cells and type 1 alveolar epithelial cells, are
the cells most damaged morphologically following 03 inhalation. Ciliated cells
are found in the conducting airways, e.g., trachea, bronchi, and nonrespiratory
bronchioles. Ciliated cells function in the normal clearance of the airways
and the removal of inhaled foreign material. Following 03 exposure of experi-
mental animals, damaged ciliated cells have been reported in all of these
conducting airways (Schwartz et a!., 1976; Castleman et al., 1977). In rats,
damage to ciliated cells appears most severe at the junction of the conducting
airways with the gas exchange area (Stephens et al., 1974a; Schwartz et al.,
1976). Damage to type 1 alveolar epithelial cells is limited to those cells
located near this junction, i.e., the centriacinar or proximal alveolar region
of the pulmonary acinus (Stephens et al., 1974b; Schwartz et al., 1976;
Castleman et al., 1980; Barry et al., 1983; Crapo et al., 1984). Type 1
alveolar cells form most of the blood-air barrier where gas exchange occurs.
Severely damaged ciliated and type I alveolar epithelial cells are shed (sloughed)
from the tissue surface and are replaced by multiplication of other cell types
less damaged by 0» (Evans et al . , 1985), This process has been most extensively
studied in the centriacinar region where nonciliated bronchiolar cells and
type 2 alveolar epithelial cells become more numerous (Evans et al., 1976a,b,c;
Lum et al., 1978). Some of these nonciliated bronchiolar and type 2 cells
differentiate into ciliated and type 1 cells, respectively. Cell multiplication
in bronchioles may be more than that required for replacement of damaged
ciliated cells, and nonciliated bronchiolar cells may become hyperplastic
(Castleman et al., 1977; Ibrahim et al., 1980; Eustis et al., 1981) and sometimes
appear as nodules (Zitnik et al., 1978; Moore and Schwartz, 1981; Fujinaka et
al., 1985). Inflammatory changes characterized by a variety of leukocytes
with alveolar macrophages predominating, intramural edema, and fibrin are also
seen in the centriacinar region (Stephens et al., 1974a; Schwartz et al.,
1976; Castleman et a!.,. 1977; Fujinaka et al., 1985).
The damage to ciliated and centriacinar type 1 alveolar epithelial cells
and the inflammatory changes tend to occur soon after exposure to concentrations
3
of 0~ as low as 392 ug/m (0.2 ppm). Damage to centriacinar type 1 alveolar
1-107
-------
epithelium in rats has been well documented as early as 2 hours after exposure
3
to 03 concentrations of 980 ug/m (0.5 ppm) (Stephens et al., 1974a). In the
same publication the authors report damage to centriacinar type 1 alveolar
3
epithelial cells after 2 hours exposure to 392 ug/m (0.2 ppm) 03, but this
portion of their report is not documented by published micrographs (Stephens
et al., 1974a). Loss of cilia from cells in the rat terminal bronchiole
3
occurs following exposure to 980 ug/m (0,5 ppm) 03 for 2 hours (Stephens et
al., 1974a). Damage to ciliated cells has been seen following exposure of
3
both rats and monkeys to 392 ug/m (0.2 ppm) 0-, 8 hr/day for 7 days (Schwartz
et al., 1976; Castleman et al., 1977). Centriacinar inflammation has been
3
reported as early as 6 hours after exposure to 980 ug/m (0.5 ppm) 0., (Stephens
•a 3
et al., 1974b) and 4 hours after exposure to 1568 ug/m (0.8 ppm) 03 (Castleman
et al,, 1980).
During long-term exposures, the damage to-ciliated cells and to centriacinar
type 1 cells and centriacinar inflammation continue, though at a reduced rate.
Damage to cilia has been reported in monkeys following 90-day exposure to 980
3
ug/m (0.5 ppm) 03, 8 hr/day (Eustis et al.s 1981) and in rats exposed to 980
ug/m3 (0.5 ppm) 03, 24 hr/day for 180 days (Moore and Schwartz, 1981). Damage
to centriacinar type 1 cells was reported following exposure of young rats to
490 ug/m3 (0.25 ppm) 0~, 12 hrs/day for 42 days (Barry et al.s 1983; Crapo et
3
al., 1984). Changes in type 1 cells were not detectable after 392 ug/m (0.2
3
ppm) 03, 8 hr/day for 90 days but were seen in rats exposed to 980 ug/m (0.5
ppm) for the same period (Boorman et al., 1980). Centriacinar inflammatory
3
changes persist during 180-day exposures of rats to 980 ug/m (0.5 ppm) 03, 24
hr/day (Moore and Schwartz, 1981) and one-year exposures of monkeys to 1254
ug/m3 (0.64 ppm) 03, 8 hr/day (Fujinaka et al., 1985).
Remodeling of distal airways and centriacinar regions occurs following
long-term exposures to CU. Rats develop respiratory bronchioles between the
terminal bronchiole to alveolar duct junction seen in control rats (Boorman et
al., 1980; Moore and Schwartz, 1981). In monkeys, distal airway remodeling
results in increased volumes of respiratory bronchioles which have, thicker
walls and a smaller internal diameter (Fujinaka et al., 1985). The walls of
centriacinar alveoli are also thickened (Schwartz et al., 1976; Boorman et
al., 1980; Barry et al., 1983; Crapo et al., 1984; Last et al., 1984a).
Studies of the nature of these thickened interalveolar septa and bronchiolar
walls revealed increases in inflammatory cells, fibroblasts, and amorphous
1-108
-------
extracellular matrix (Last et a!., 1984a; Fujinaka et al., 1985). Three
studies provide morphological evidence of mild fibrosis (i.e., local increase
of collagen) in centriacinar interalveolar septa following exposure to < 1960
3
|jg/m (< 1 ppm) of 03 (Last et al., 1979; Boorman et al,, 1980; Moore and
Schwartz, 1981). Changes in collagen location or amounts, or both, which
occur with the remodeling of the distal airways, were reported in two of those
studies (Boorman et a!., 1980; Moore and Schwartz, 1981).
While morphometry of small pulmonary arteries is not commonly studied in
0-~exposed animals, pulmonary artery walls thickened by muscular hyperplasia
3
and edema were reported in rabbits exposed to 784 n9/m (Q.4 ppm) 03, 6 hr/days
5 days/week for 10 months (P'an et al., 1972). Thickened intima and media in
3
pulmonary arterioles were reported in monkeys exposed to 1254 |jg/m (0.64 ppm)
03, 8 hr/day for 1 year (Fujinaka et al., 1985).
Several of the effects of 03 inhalation persisted after the Qg inhalation
ended and the animals breathed only filtered air several days or weeks. Lungs
from rats exposed to 1568 (jg/m (0.8 ppm) 03 for 72 hours appeared normal 6
days after the end of the exposure (Plopper et al., 1978). However, incomplete
resolution of the nonciliated bronchiolar epithelial hyperplasia was reported
3
in monkeys 7 days after 50 hours exposure to 1568 [jg/m (0.8 ppm) 0- (Castleman
3
et al., 1980) and in mice 10 days after a 20-day exposure to 1568 [jg/m (0.8
ppm) DO, 24 hr/day (Ibrahim et al., 1980). Centriacinar inflammation and
distal airway remodeling were still apparent 62 days after a 180-day exposure
to 980 (jg/m3 (0-5 ppm) 03> 24 hr/day (Moore and Schwartz, 1981).
While not all species of laboratory mammals have been studied following a
single 0- exposure regimen or using the same morphological techniques because
investigators have asked different biological questions, there is a striking
similarity of morphological effects in the respiratory system of all species
studied. The cell types most damaged are the same. One of these cells, the
type 1 alveolar epithelial cell, has a wide distribution in the pulmonary
acinus and yet is damaged only in one specific location in all species studied.
The other, the ciliated cell, appears damaged wherever it is located in the
conducting airways. Damage to these cells is seen within hours after exposure
to concentrations of Q~ much lower than 1 ppm and continues during exposures
of weeks or months. Hyperplasia of other cell types is reported to start
early in the exposure period, to continue throughout a long-term exposure, and
when studied, to persist following postexposure periods of days or weeks.
1-109
-------
Centriacinar inflammation is also seen early and is reported throughout long
exposure periods. Duration of centriacinar inflammation during postexposure
periods has been studied less often and appears dependent upon length of the
exposure period.
Other effects which have been reported in fewer studies or in a more
limited number of species include distal airway remodeling and thickened pul-
monary arteriolar walls. Remodeling of distal airways has only been reported
in rats and monkeys after long-term exposures. In rats, remodeling of distal
airways has been reported to persist for several weeks after the 0, exposure
has ended. Thickened pulmonary arteriolar walls have been reported only
twice, once after long-term exposure of rabbits and once after long-term
exposure of monkeys.
Studies on the morphologic effects of 0~ exposures of experimental animals
are summarized in Figure 1-7 and Table 1-11 (see Section 1.8.1 for criteria
used to summarize the studies).
1.8.3.2 Pulmonary Function. One of the limitations of animal studies is that
many pulmonary function tests comparable to those conducted after acute exposure
of human subjects are difficult to interpret. Methods exist, however, for
obtaining similar measurements of many variables pertinent to understanding
the effects of ozone on the respiratory tract, particularly after longer
exposure periods. A number of newer studies reported here reflect recent
advances in studying the effects of 0., on pulmonary function in small animals.
Changes in lung function following ozone exposure have been studied in
mice, rats, guinea pigs, rabbits, cats, dogs, sheep, and monkeys. Short-term
3
exposure for 2 hr to concentrations of 431 to 980 |jg/m (0.22 to 0.5 ppm)
produces rapid, shallow breathing and increased pulmonary resistance during
exposure (Murphy et al., 1964; Yokoyama, 1969; Watanabe et a!., 1973; Amdur
et al., 1978). The onset of these effects is rapid and the abnormal breathing
pattern usually disappears within 30 min after cessation of exposure. Other
changes in lung function measured following short-term ozone exposures lasting
3 hr to 14 days are usually greatest 1 day following exposure and disappear by
7 to 14 days following exposure. These effects are associated with premature
closure of the small, peripheral airways and include increased residual volume,
closing volume, and closing capacity (Inoue et al., 1979).
1-110
-------
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o
03
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0.1-
0.2-
0.3-
0.4-
0.5-
0.6-
0.7-
0,8-
0.9-
1.0
I
1
<
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Figure 1 -7. Summary of morphological effects in experimental animals
exposed to ozone. See Table 1-11 for reference citations of studies
summarized here.
-------
TABLE 1-11. SUMMARY TABLE: MORPHOLOGICAL EFFECTS OF OZONE
IN EXPERIMENTAL ANIMALS
Effect/response
03 concentration, ppm
References
Damaged ciliated
and type 1 cells
Proliferation of non-
ciliated bronchiolar
and type 2 cells
Centriacinar
inflammation
[0.2], 0.5, 0.8
0.2, 0.5, 0.8
0.2, 0.35
0.25
0.25
0.26, 0.50, 1.0
0.5
0.5
0.5
0.5, 0.8
0.5, 0.8
0.54, 0.88
0.8
0.8
0.85
0.2, 0.35
0.35, 0.50, 0.70,
0.75, 1.0
0.5
0.5
0.5
0.5, 0.8
0.54, 0.88
0.64
0.7
0.8
0.8
0.8
1.0
[0.2], 0.5, 0.8
0.2
0.2, 0.5, 0.8
0.25
0.25
0.35
0.5
0.5
0.5, 0.8
0,5, 0.8
0.5, 0.8
0.54, 0.88
0.54, 0.88
0.64
0.8
1.0
Boorman et al. (1980)
Schwartz et al. (1976)
Castleman et al. (1977)
Barry et al. (1983)
Crapo et al. (1984)
Boatman et al. (1974)
Stephens et al. (1974b)
Moore and Schwartz (1981)
Evans et al. (1985)
Eustis et al. (1981)
Mellick et al. (1975, 1977)
Stephens et al. (1974a)
Castleman et al. (1980)
Plopper et al. (1978)
Stephens et al. (1978)
Castleman et al. (1977)
Evans et al. (1976b)
Evans et al. (1985)
Zitnik et al. (1978)
Moore and Schwartz (1981)
Eustis et al. (1981)
Freeman et al. (1974)
Fujinaka et al. (1985)
Evans et al. (1976a)
Castleman et al. (1980)
Lum et al. (1978)
Ibrahim et al. (1980)
Cavender et al. (1977)
Boorman et al. (1980)
Plopper et al. (1979)
Schwartz et al. (1976)
Barry et al. (1983)
Crapo et al. (1984)
Castleman et al. (1977)
Stephens et al. (1974b)
Moore and Schwartz (1981)
Mellick et al. (1975, 1977)
Brummer et al, (1977)
Last et al. (1979)
Stephens et al. (1974a)
Freeman et al. (1974)
Fujinaka et al. (1985)
Castleman et al. (1980)
Freeman et al. (1973)
1-112
-------
TABLE 1-11 (continued). SUMMARY TABLE: MORPHOLOGICAL EFFECTS OF OZONE
IN EXPERIMENTAL ANIMALS
Effect/response 03 concentration, ppm References
Distal airway [0.2], 0.5, 0.8 Boorman et al. (1980)
remodeling 0.2, 0.5, 0.8 Schwartz et al. (1976)
0.5 Moore and Schwartz (1981)
0.64, 0.96 Last et al. (1984a)
0.64 Fujinaka et al. (1985)
1.0 Freeman et al. (1973)
Thickened pulmonary 0.4 P'an et al. (1972)
arteriolar walls 0.64 Fujinaka et al. (1985)
Studies of airway reactivity following short-term ozone exposure of 1 to
2 hr duration in experimental animals show that 0- increases the reactivity of
the lungs to a number of stimuli. Mild exercise, histamine aerosol inhalation,
and breathing air with reduced oxygen or elevated carbon dioxide concentrations
caused rapid, shallow breathing in conscious dogs immediately following 2-hr
exposures to 1100 to 1666 M9/m3 (0.56 to 0.85 ppm) of Oj (Lee et al., 1979,
1980). Aerosolized ovalbumin caused an increased incidence of anaphylaxis in
o
mice preexposed to 980 or 1568 ug/m (0.5 or 0.8 ppm) of 0- continuously for 3
>, O
to 5 days (Osebold et al., 1980). In addition, increased airway sensitivity
to histamine or cholinomimetic drugs administered by aerosol or injection has
2
been noted in several species after exposure to 980 to 5880 ug/m (0.5 to 3.0
ppm) of 03 (Easton and Murphy, 1967; Lee et al., 1977; Abraham et al.5 1980,
1984a,b; Gordon and Amdur, 1980; Gordon et al., 1981, 1984; Roum and Murlas,
1984). The mechanism responsible for 0~-induced bronchial reactivity is still
uncertain but may involve more than one specific factor. Ozone has been shown
to cause increased sensitivity of vagal sensory endings in the dog airway (Lee
et al., 1977, 1979, 1980). Ozone exposure may also enhance the airway respon-
siveness to bronchoconstrictors by altering sensitivity of the airway smooth
muscle directly or through released cellular mediators (Gordon et al.s 1981,
1984; Abraham et al., 1984a,b). In some species, increased airway hyperreac-
tivity may be explained by increased transepithelial permeability or decreased
thickness of the airway mucosa (Osebold et al., 1980; Abraham et al., 1984b).
Ozone exposure may also decrease airway hyperreactivity by causing mucous
1-113
-------
hypersectetion, thereby limiting the airway penetration of inhaled bronchoeon-
strictors (Abraham et al., 1984a).
The time course of airway hyperreactivity after exposure to 980 to 5880
|jg/m (0.5 to 3.0 ppm) of 0., suggests a possible association with inflammatory
cells and pulmonary inflammation (Holtzman et a!., 1983a,b; Sielczak et a!.,
1983; Fabbri et al., 1984; O'Byrne et al., 1984a,b; Murlas and Roum, 1985).
However, the time course of responsiveness is variable in different species
and the relationships between airway inflammation and reactivity at different
concentrations of 03 are not well understood. Additional studies that demon-
strate increased collateral resistance following 30 min local exposure of 0™
or histamine in sublobar bronchi of dogs (Gertner et al., 1983a,b,c,1984)
suggest that other mechanisms, along with amplification of reflex pathways,
may contribute to changes in airway reactivity depending not only on the
concentration of 0, in the airways but also on the extent of penetration of
ozone into the lung periphery.
The effects of short-term exposures to 0™ on pulmonary function and
airway reactivity in experimental animals are summarized in Figure 1-8 and
Table 1-12 (see Section 1.8.1 for criteria used in developing this summary).
3
Exposures of 4 to 6 weeks to ozone concentrations of 392 to 490 ug/m
(0.2 to 0.25 ppm) increased lung distensibility at high lung volumes in young
rats (Bartlett et al., 1974; Raub et al., 1983). Similar increases in lung
distensibility were found in older rats exposed to 784 to 1568 ug/m (0.4 to
0.8 ppm) for up to 180 days (Moore and Schwartz, 1981; Costa et al., 1983;
o
Martin et al., 1983). Exposure to 0- concentrations of 980 to 1568 ug/m (0.5
to 0.8 ppm) increased pulmonary resistance and caused impaired stability of
the small peripheral airways in both rats and monkeys ( Wegner, 1982; Costa
et al., 1983; Yokoyama et al., 1984; Kotlikoff et al., 1984). The effects in
monkeys were not completely reversed by 3 months following exposure; lung
distensibility had also decreased in the postexposure period, suggesting the
development of lung fibrosis which has also been suggested morphologically and
biochemically.
The effects of long-term exposures to ozone on pulmonary function and
airway reactivity in experimental animals are summarized in Figure 1-9 and
Table 1-13 (see Section 1.8.1 for criteria used in developing this summary).
1-114
-------
E
a
a
c
o
ID
U
C
o
U
e
c
(
I 4
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C
)
1
1 1
1
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x£
5
1
»
I
i
Figure 1 -8. Summary of effects of short-term ozone exposures on
pulmonary function in experimental animals. See Table 1-12 for
reference citations of studies summarized here.
-------
TABLE 1-12. TABLE: EFFECTS ON PULMONARY FUNCTION
OF SHORT-TERM EXPOSURES TO OZONE IN EXPERIMENTAL ANIMALS
Effect/response 03 concentration, ppm
References
Increased breathing
frequency
0.22, 0.41, 0.8
0.34, 0.68, 1.0
0.5
Amdur et al. (1978)
Murphy et al. (1964)
Yokoyama (1969)
Decreased tidal volume 0.34, 0.68, 1.0 Murphy et al. (1964)
Decreased lung
compliance
Increased residual
volume (RV),
closing capacity
(CC), and closing
volume (CV)
Decreased diffusion
capacity
[0.22], 0.41, 0.8
0.26, 0.5, 1.0
1.0
0.24 - 1.0
Amdur et al. (1978)
Watanabe et al. (1973)
Yokoyama (1974)
Inoue et al. (1979)
0.26, 0.5, 1.0
Watanabe et al. (1973)
Increased pulmonary
resistance
Increased airway
reactivity
[0.22]
0.26, 0.5, 1.0
0.5
1.0
[0.1]-0.8
[0.1]-0.8, 1.0
0.5, 1.0
0.7
1.0
Amdur et al. (1978)
Watanabe et al. (1973)
Yokoyama (1969)
Yokoyama (1974)
Gordon and Amdur (1980)
Gordon et al. (1981, 1984)
Abraham et al. (1980, 1984a,b)
Lee et al. (1977)
Holtzman et al. (1983a,b)
1.8.3.3 Biochemical Effects. The lung is metabolically active, and several
key steps in metabolism have been studied after 03 exposure. Since the proce-
dures for such studies are invasive, this research has been conducted only in
animals. Effects, to be summarized below, have been observed on antioxidant
metabolism, oxygen consumption, proteins, lipids, and xenobiotic metabolism.
The lung contains several compounds (e.g., vitamin E, sulfhydryls, gluta-
thione) and enzymes (e.g., glutathione peroxidase, glutathione reductase,
glucose-6-phosphate dehydrogenase, and superoxide dismutase) that function as
antioxidants, thereby defending the lung against oxidant toxicity from the
1-116
-------
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E
Q.
a
a>
o
c
o
o
c
o
N
O
u.u —
0.1_
0.2-
0.3-
0.5-
0.6-
0.7_
0.8_
0.9-
1.O
I I
0 C
I
0 «
i
(
I
1
1 1
1
1
<
1
1 <
. — ,.,. .1 - -III 1 — . . 1
1
1
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1 1
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1
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1
1
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1
I
Figure 1-9. Summary of effects of long-term ozone exposures on
pulmonary function in experimental animals. See Table 1-13 for
reference citations of studies summarized here.
-------
TABLE 1-13. SUMMARY TABLE: EFFECTS ON PULMONARY FUNCTION
OF LONG-TERM EXPOSURES TO OZONE IN EXPERIMENTAL ANIMALS
Effect/response
03 concentration, ppm
References
Increased lung volume
Increased pulmonary
resistance
Decreased lung
compliance
Decreased inspiratory
flow
Decreased forced
expiratory volume
(FEV,) and flow
(PEP)
[0.08], [0.12], 0.25
0.2
[0.2], 0.8
0.4
0.2, 0.8
0.5, 1.0
0.64
0.64
0.5, 0.8
0.64
[0.08], 0.12, 0.25
0.2, 0.8
0.64
Raub et al. (1983)
Bartlett et al. (1974)
Costa et al. (1983)
Martin et al. (1983)
Costa et al. (1983)
Yokoyama et al., 1984
Wegner (1982)
Kotlikoff et al., 1984
Eustis et al. (1981)
Wegner (1982)
Raub et al. (1983)
Costa et al. (1983)
Wegner (1982)
oxygen in air, from oxidants produced during metabolic processes, and from
oxidizing air pollutants such as ozone. Obviously, this protection is only
partial for On since exposure to ozone causes numerous effects on lung struc-
ture, function, and biochemistry. Acute exposure to high ozone levels
3
(2920 ug/m , 2 ppm) typically decreases antioxidant metabolism, whereas repeated
3
exposures to lower levels (between 272 and 1568 (jg/m , 0.2 and 0.8 ppm) in-
creases this metabolism (DeLucia et al., 1975b). In rats maintained on normal
diets, this response has been observed after a week of continuous or intermit-
tent exposure to 392 ug/m3 (0.2 ppm) 03 (Mustafa, 1975; Mustafa and Lee, 1976;
Plopper et al., 1979). Similar responses are seen in monkeys and mice, but at
3
higher concentrations (980 ug/m , 0.5 ppm) (Fukase et al., 1978; Mustafa and
Lee, 1976).
The effects of 0, on oxygen consumption have been studied since oxygen
consumption is a fundamental parameter of cellular metabolism, reflecting
energy production by cells. As with antioxidant metabolism, acute exposure to
1-118
-------
3
high ozone levels (> 3920 |jg/m ; > 2 ppm) decreases metabolism (and thus,
3
oxygen consumption); repeated exposure to lower levels (> 1568 jjg/m , 0.8 ppm)
increases oxygen consumption (Mustafa et al., 1973; Schwartz et a!,, 1976;
Mustafa and Lee, 1976). Effects in rats on normal diets have been observed
3
after a short-term exposure to ozone levels as low as 392 |jg/m (0.2 ppm)
(Schwartz et al., 1976; Mustafa et al., 1973; Mustafa and Lee, 1976). Monkeys
3
are affected at a higher level of ozone (980 pg/m , 0.5 ppm).
Similar patterns of response for both antioxidant metabolism and oxygen
consumption are observed after exposure to ozone. A 7-day exposure to ozone
produces linear concentration-related increases in activities of glutathione
peroxidase, glutathione reductase, glucose-6-phosphate dehydrogenase, and
succinate oxidase (Mustafa and Lee, 1976; Ghow et al., 1974; Schwartz et al.,
1976; Mustafa et al., 1973). Rats on a vitamin E-deficient diet experience an
o
increase in enzyme activities at 196 ng/m (0-1 ppm) ozone as compared to
3
392 jjg/m (0.2 ppm) in animals on normal diets (Chow et al., 1981; Mustafa and
Lee, 1976; Mustafa, 1975). Research on these enzymes has shown that there is
no significant difference in effects from continuous versus intermittent
exposure; this, along with concentration-response data, suggests that the con-
centration of ozone is more important than duration of exposure in causing
these effects (Chow et al., 1974; Schwartz et al., 1976; Mustafa and Lee,
1976).
Duration of exposure still plays a role, however. During exposures up to
1 or 4 weeks, antioxidant metabolism and 0? consumption generally do not
change on the first day of exposure; by about day 2, increases are observed
and by about day 4 a plateau is reached (Mustafa and Lee, 1976; DeLucia et al.,
1975a). Recovery from these effects occurs by 6 days post-exposure (Chow
et al., 1976). This plateauing of effects in the presence of exposure does
not result in long-term tolerance. If rats are re-exposed after recovery is
observed, the increase in enzyme activities is equivalent to that observed in
animals exposed for the first time (Chow et al., 1976).
The influence of age on responsiveness is also similar for antioxidant
metabolism and oxygen consumption (Elsayed et al., 1982a; Tyson et al., 1982;
Lunan et al., 1977). Suckling neonates (5 to 20 days old) generally exhibited
a decrease in enzyme activities; as the animals grew older (up to about 180 days
old), enzyme activities generally increased with age. Species differences may
exist in this response (Mustafa and Lee, 1976; Mustafa et al., 1982; Chow
1-119
-------
et al., 1975; DeLucia et al., 1975a). Studies in which monkeys have been
compared to rats did not include a description of appropriate statistical
considerations applied (if any); thus, no definitive conclusions about respon-
siveness of monkeys versus rats can be made.
The mechanism responsible for the increase in antioxidant metabolism and
oxygen consumption is not known. The response is typically attributed, however,
to concurrent morphological changes, principally the loss of type 1 cells and
an increase in type 2 cells that are richer in the enzymes measured.
Monooxygenases constitute another class of enzymes investigated after
ozone exposure. These enzymes function in the metabolism of both endogenous
(e.g., biogenic amines, hormones) and exogenous (xenobiotic) substances. The
substrates acted upon are either activated or detoxified, depending on the
3
substrate and the enzyme. Acute exposure to 1470 to 1960 ug/m (0.75 to
1 ppm) ozone decreased cytochrome P-450 levels and enzyme activities related
to both cytochrome P-450 and P-448. The health impact of these changes is
uncertain since only a few elements of a complex metabolic system were measured.
The activity of lactate dehydrogenase is increased in lungs of vitamin E-
3
deficient rats receiving a short-term exposure to 196 ug/m (0.1 ppm) ozone
(Chow et al., 1981). Higher levels caused a similar response in rats, but not
in monkeys, on normal diets (Chow et al., 1974, 1977). This enzyme is frequent-
ly used as a marker of cellular damage because it is released upon cytotoxicity.
It is not known, however, whether the increase in this enzyme is a direct
reflection of cytotoxicity or whether it is an indicator of an increased
number of type 2 cells and macrophages in the lungs.
An increase in a few of the measured activities of lysosomal enzymes has
o
been shown in the lungs of rats exposed to > 1372 ug/m (0.7 ppm) ozone (Dillard
et al., 1972; Castleman et al., 1973a; Chow et al. , 1974). This response is
most likely the result of an increase in inflammatory cells in the lungs
rather than an induction of enzymes, since lysosomal enzymes in alveolar
macrophages decrease after ijn vivo or i_n vitro exposure to ozone (Hurst et al.,
1970; Hurst and Coffin, 1971).
As discussed previously, long-term exposure to high 0- concentrations
causes mild lung fibrosis (i.e., local increase of collagen in centriacinar
interalveolar septa). This morphological change has been correlated with
biochemical changes in the activity of prolyl hydroxylase (an enzyme that
catalyzes the production of hydroxyproline) and in hydroxyproline content (a
1-120
-------
component of collagen that is present in excess in fibrosis) (Last et al.,
1979; Bhatnagar et al., 1983). An increase in collagen synthesis has been
2
observed, with 980 |jg/m (0-5 ppm) 03 being the minimally effective concentra-
tion tested (Hussain et al., 1976a,b; Last et al., 1979). During a prolonged
exposure, prolyl hydroxylase activity increases by day 7 and returns to control
levels by 60 days of exposure. When a short-term exposure ceases, prolyl
hydroxylase activity returns to normal by about 10 days post-exposure, but
hydroxyproline levels remain elevated 28 days post-exposure. Thus, the product
of the increased synthesis, collagen, remains relatively stable. One study
(Costa et al., 1983) observed a small decrease in collagen levels of rats at
o
392 and 1568 |jg/m (0.2 and 0.8 ppm) 0, after an intermittent exposure for 62
days.
The effects of 03 on increasing collagen content may be progressive;
i.e., after a 6-week intermittent exposure of rats to 0.64 or 0.96 ppm 0,
ceased, collagen levels 6 week post-exposure were elevated over the levels
immediately after exposure (Last et al. , 1984b). Also, there appears to be
little difference between continuous and intermittent exposure in increasing
collagen levels in rat lungs (Last et al, 1984b). Thus, the intermittent
clean air periods were not sufficient to permit recovery.
Although the ability of 03 to initiate peroxidation of unsaturated fatty
acids ijn vitro is well established, few in, vivo studies of lung lipids have
been conducted. Generally, ozone decreases unsaturated fatty acid content of
the lungs (Roehm et al., 1972) and decreases incorporation of fatty acids into
lecithin (a saturated fatty acid) (Kyei-Aboagye et al., 1973). These altera-
tions, however, apparently do not alter the surface-tension-lowering properties
of lung lipids that are important to breathing (Gardner et al., 1971; Huber
et al., 1971).
One of the earliest demonstrated effects of ozone was that very high
concentrations caused mortality as a result of pulmonary edema. As more
2
sensitive techniques were developed, lower levels (510 ug/m , 0.26 ppm) were
observed to increase the protein content of the lung (Hu et al., 1982). Since
some of the excess protein could be attributed to serum proteins, the interpre-
tation was that edema had occurred. This effect was more pronounced several
hours after exposure ceased. At higher concentrations, a loss of carrier-
mediated transport from the air side of the lung to the blood side was observed
(Williams et al., 1980). These changes imply an effect on the barrier function
1-121
-------
of the lung, which regulates fluxes of various substances with potential
physiological activities across the alveolar walls.
The biochemical effects observed in experimental animals exposed to fl-
are summarized in Figure 1-10 and Table 1-14 (see Section 1.8.1 for criteria
used in developing this summary).
1.8.3.4 Host Defense Mechanisms. Reports over the years have presented
substantial evidence that exposure to ozone impairs the 'antibacterial activity
of the lung, resulting in an impairment of the lung's ability to kill inhaled
microorganisms. Suppression of this biocidal defense of the lung can lead to
microbial proliferation within the lung, resulting in mortality. The mortality
response is concentration-related and is significant at concentrations as low
as 157 to 196 |jg/m3 (0.08 to 0.1 ppm) (Coffin et al., 1967; Ehrlich et al. ,
1977; Miller et al., 1978a; Aranyi et al., 1983). The biological basis for
this response appears to be that ozone or one of its reactive products can
impair or suppress the normal bactericidal functions of the pulmonary defenses,
which results in prolonging the life of the infectious agent, permitting its
multiplication and ultimately, in this animal model, resulting in death. Such
infections can occur because of 0, effects on a complex host defense system
involving alveolar macrophage functioning, lung fluids, and other immune
factors.
The data obtained in various experimental animal studies indicate that
short-term ozone exposure can reduce the effectiveness of several vital defense
systems including (1) the ability of the lung to inactivate bacteria and
viruses (Coffin et al., 1968; Coffin and Gardner, 1972b; Goldstein et al. ,
1974, 1977; Warshauer et al., 1974; Bergers et al; 1983. Schwartz and Christman,
1979; Ehrlich et al., 1979); (2) the mucociliary transport system (Phalen
et al., 1980; Frager et al., 1979; Kenoyer et al., 1981; (3) the immunological
system (Campbell and Hilsenroth, 1976; Fujimaki et al., 1984; Thomas et al.,
1981b; Aranyi et al., 1983; and (4) the pulmonary macrophage (Dowell et al.,
1970; Goldstein et al., 1971a,b, and 1977; Hadley et al., 1977; McAllen et al.,
1981; Witz et al., 1983; Hurst et al., 1970; Hurst and Coffin, 1971; Amoruso
et al., 1981). Studies have also indicated that the activity level of the
test subject and the presence of other airborne chemicals are important vari-
ables that can influence the determination of the lowest effective concen-
tration of the pollutant (Gardner et al., 1977; Aranyi et al., 1983; Ehrlich,
1980, 1983; Grose et al., 1980, 1982; Phalen et al., 1980; Goldstein et al.,
1974; Illing et al., 1980).
1-122
-------
IX)
00
a
a
o
05
O
o
o
05
O
N
O
vc*
.cn
o'V
^ «V*X -'
\<^
v».v/ —
0.1 -
0.2-
0.3-
0.4-
0.5-
0.6-
0.7-
0.8-
0.9-
1.0
•
c
i
i
i
i
<
3
1 C
I
I
.
i C
(
I I
) (
<
3 I
1
I
> (
I
3 C
(
I
i I
<
i
i I
i
C
3
(
)
1 I
\ 4
t
<
1
3 O
i
(
i
1
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i
•
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(
i {
i
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1
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I
Figure 1-10. Summary of biochemical changes in experimental animals
exposed to ozone. See Table 1-14 for reference citations of studies
summarized here.
-------
TABLE 1-14. SUMMARY TABLE: BIOCHEMICAL CHANGES
IN EXPERIMENTAL ANIMALS EXPOSED TO OZONE
Effect/response
03 concentration, ppm
References
Increased Og
consumption
Increased lysosomal
enzyme activities
Increased lung
hydroxyproline
and prolyl
hydroxylase
activity
Altered mucus
glycoprotein
secretions
Increased alveolar
protein and
permeability
changes
Increased LDH
activity
Increased NADPH
cytochrome c
reductase
activity
Increased GSH
metabolism
[0.1], 0.2
[0.1], 0.2, 0.35, 0.5, 0.8
0.2, 0.5, 0.8
0.2, 0.5, 0.8
0.45
0.8
0.8
[0.2], [0.5], 0.8
0.7, 0.8
0.7, 0.8
[0.2], 0.5, 0.8
0.2, 0.8
0.45, 0.8
0.5, 0.64, 0.96
0.5
0.8
[0.2], [0.4], 0.5, 0.6, 0.8
0.5, 0.6, 0.8
0.6, 0.8
[0.1], 0.26,
[0.25], 0.5,
0.6, 1.0
1.0
[0.1]
[0.5], 0.8
0.8
0.51, 1.0
1.0
0.2, 0.35, 0.8
0.2, 0.5, 0.8
0.2, 0.5, 0.8
[0.1]
0.1, 0.2
0.2, 0.35, 0.5.
0.2, 0.5r 0.8
0.2, 0.5, 0.8
0.2, 0.5, 0.8
0.2, 0.5, 1.0
0.32
0.45
0.5
0.8
Mustafa (1975)
Mustafa and Lee (1976)
Mustafa et al. (1973)
Schwartz et al. (1976)
Mustafa et al. (1982)
Chow et al, (1976)
Elsayed et al. (1982a)
Chow et al. (1974)
Oil lard et al. (1972)
Castleman et al, (1973a,b)
Hussain et al. (1976a,b)
Costa et al. (1983)
Bhatnagar et al. (1983)
Last et al. (1979, 1984b)
Last and Greenberg (1980)
Hesterberg and Last (1981)
Last and Kaizu (1980)
Last and Cross (1978)
Last et al. (1977)
Hu et al. (1982)
Alpert et al. (1971a)
Williams et al. (1980)
Reason et al. (1979)
Chow et al. (1981)
Chow et al. (1977)
Chow and Tappel (1973)
Mustafa and Lee (1976)
Schwartz et al. (1976)
DeLucia et al. (1972, 1975a,b)
Chow et al. (1981)
Plopper et al. (1979)
Mustafa and Lee (1976)
Chow et al. (1974)
DeLucia et al. (1972, 1975a,b)
Schwartz et al. (1976)
Fukase et al. (1975)
Moore et al. (1980)
Mustafa et al. (1982)
Chow et al. (1975)
1-124
-------
TABLE 1-14 (continued). SUMMARY TABLE: BIOCHEMICAL CHANGES
IN EXPERIMENTAL ANIMALS EXPOSED TO OZONE
Effect/response
03 concentration, ppm
References
0.5, 1.0
0.7, 0.75, 0.8
Fukase et al. (1978)
Chow and Tappel (1972, 1973)
Increased NPSH
Decreased
unsaturated
fatty acids
0.8
0.8
0.9
0.9
0.1, 0.2
0.2, 0.5, 0.8
0.45
0.8
0.5
. Elsayed et al, (1982a,b;
1983)
Chow et al. (1976)
Tyson et al. (1982)
Lunan et al. (1977)
Plopper et al, (1979)
DeLucia et al. (1975b)
Mustafa et al. (1982)
Chow et al . (1976)
Roehm et al . , 1972
Ciliated cells are damaged by 03 inhalation, as demonstrated by major
morphological changes in these cells including necrosis and ploughing or by
the shortening of the cilia in cells attached to the bronchi. Sufficient
ciliated cell damage should result in decreased transport of viable and non-
viable particles from the lung. Rats exposed to 784, 1568, I960, or 2352
3
|jg/m (0.4, 0.8, 1.0, or 1.2 ppm) for times as short as 4 hr have decreased
short-term clearance of particles from the lung (Phalen et al., 1980; Frager
et al., 1979; Kenoyer et al., 1981). Short-term clearance is mostly due to
mucus transport of particles, and the decreased short-term clearance is an
anticipated functional result predicted from morphological observations. The
mucous glycoprotein production of the trachea is also altered by 0~ exposure.
Mucous glycoprotein biosynthesis, as measured ex vivo in cultured trachea!
explants from exposed rats, was inhibited by short-term continuous exposure to
1568 |jg/m3 (0.8 ppm) of 0, for 3 to 5 days (Last and Cross, 1978; Last and
Kaizu, 1980; Last et al., 1977). Glycoprotein synthesis and secretion recovered
to control values after 5 to 10 days of exposure and increased to greater than
control values after 10 days of exposure. With this increase in production of
mucus, investigators have found that the velocity of the trachea! mucus was
1-125
-------
q
significantly reduced following a 2 hr exposure to 1960 [ig/m (1.0 ppm) (Abraham
et al.» 1980).
A problem remains in assessing the relevance of these animal data to
humans. Green (1984) reviewed the literature and compared the host antibacterial
defense systems of the rodent and man and found that these two species had
defenses that are very similar and thus provide a good basis for a qualitative
extrapolation. Both defenses consist of an aerodynamic filtration system, a
fluid layer lining the respiratory membranes, a transport mechanism for removing
foreign particles, microorganisms, and pulmonary cells, and immune secretions
of lymphocytes and plasma cells. In both rodents and humans, these components
act in concert to maintain the lung free of bacteria.
If the animal models are to be used to reflect the toxicological response
occurring in humans, then the endpoint for comparison of such studies should
be morbidity rather than mortality. A better index of 0, effect in humans
might be the increased prevalence of infectious respiratory illness in the
community. Such a comparison may be proper since both mortality from respira-
tory infections (animals) and morbidity from respiratory infections (humans)
can result from a loss in pulmonary defenses (Gardner, 1984). Whether the
microorganisms used in the various animal studies are comparable to the organ-
isms responsible for the respiratory infections in a community still requires
further investigation.
Ideally, studies of pulmonary liost defenses should be performed in man,
using epidemiological or volunteer methods of study. Unfortunately, such
studies have not been reported yet. Attention must therefore be paid to the
results of host-defense experiments conducted with animals.
In the area of host defense of the lung against infection, present know-
ledge of the physiology, metabolism, and function have come primarily from the
study of various animal systems, but it is generally accepted that the basic
mechanisms of action of these defense cells and systems function similarly in
both animals and man. There are also human data to support this statement,
especially in such areas as immunosuppression, ciliostasis, and alveolar
macrophages. The effects seen in animals represent alterations in basic
biological systems. One can assume that similar alterations in basic defense
mechanisms could occur in humans since they possess equivalent pulmonary
defense systems. It is understood, however, that different exposure levels
may be required to produce similar responses in humans. The concentration of
1-126
-------
0, at which effects become evident can be influenced by a number of factors,
such as preexisting disease, virulence of the infectious agent, dietary factors,
concurrent exposure to other pollutants, exercise, or the presence of other
environmental stresses, or a combination of these. Thus, one could hypothesize
that humans exposed to CL could experience effects on host defense mechanisms.
At the present time, however, one cannot predict the exact concentration at
which effects may occur in man nor the severity of the effects.
The effects of CL on host defense mechanisms in experimental animals are
summarized in Figure 1-11 and Table 1-15 (see Section 1.8.1 for criteria used
in developing this summary),
1.8.3.5 Tolerance. Examination of responses to short-term, repeated exposures
to CL clearly indicates that with some of the parameters measured, animals
have an increased capacity to resist the effects of subsequent exposure. This
tolerance persists for varying times, depending on the degree of development
of the tolerance. Previous exposure to low concentrations of CL will protect
against the effects of subsequent exposure to lethal doses and the development
of lung edema (Stokinger et a!., 1956; Fairehild, 1967; Coffin and Gardner,
1972a; Chow, 1984). The prolongation of mucociliary clearance reported for CL
can also be eliminated by pre-exposure to a lower concentration (Frager et
al., 1979). This effect is demonstrated for a short period of time and is
lost as soon as the mucus secretion rate returns to normal. However, not all
of the toxic effects of 0-,, such as reduced functioning activity of the pulmonary
defense system (Gardner et al., 1972); hyperplasia of the type 2 cells (Evans
et al., 1971, 1976a,b); increased susceptibility to respiratory disease (Gardner
and Graham, 1977); loss of pulmonary enzymatic activity (Chow, 1976, Chow
et al., 1976); and inflammatory response (Gardner et al., 1972) can be totally
prevented by prior treatment with low levels of CL. Dungworth et al. (1975)
and Castleman et al. (1980) have attempted to explain tolerance by careful
examination of the morphological changes that occur with repeated CL exposures.
These investigators suggest that during continuous exposure to CL the injured
cells attempt to initiate early repair of the specific lesion. The repair
phase results in a reduction of the effect first observed but lasts only for a
short time since the recovered cells are as sensitive to re-exposure to CL as
the pre-exposed counterpart (Plopper et al., 1978). This information is an
important observation because it implies that the decrease in susceptibility
to CL persists only as long as the exposure to CL continues. The biochemical
studies of Chow et al. (1976) support this conclusion.
1-127
-------
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CO
E
a
a
*
o
'•P
SS
4-1
0)
o
o
o
0)
u.u —
0.1-
0.2-
0.3-
0.4-
0.5-
0.6-
0.7-
0.8-
0.9-
1.0
<
•
i
<
«
1
)
\ (
> i
4
1
1
. I
(
\
1 1
1
I
1 ,
»
1
I 1
(
1
i 1
(
1 (
(
1 (
(
1
1 1
1
I 1
(
<
1 (
1
<
•
(
1
1
C
(
1
1
(
i 1
1
1 1
)
1
1
(
1
)
1
4
i
i
1
I
I
>
Figure 1-11. Summary of effects of ozone on host defense mechanisms
in experimental animals. See Table 1-15 for reference citations of
studies summarized here.
-------
TABLE 1-15. SUMMARY TABLE: EFFECTS OF OZONE ON HOST DEFENSE
MECHANISMS IN EXPERIMENTAL ANIMALS
Effect/response
03 concentration, ppm
References
Delayed mucociliary
clearance; accelerated
alveolar clearance,
ciliary beating
frequency
Inhibited bactericidal
activity
Altered macrophage
membrane
Decreased macrophage
function
Altered
cells
no. of defense
[0.1]
0.4, 0.8, 1.0
[0.5]
[0.5], 1.0
0.8
1.2
0.4
0.4
0.5
0.62
0.7
0.7
0.99
0.1, 1.0
0.5
0.5
0.5, 1.0
0.25, 0.5
0.5
0.5, 0.67
0.5, 0.67
0.8
1.0
1.0
0.2
0.2, 0.35, 0.5, 0.8
0.2, 0.35
0.2, 0.5, 0.8
0.25
0.5
0.5, 0.88
0.5
0.5, 0.88
0.5, 0.8
0.54, 0.88
0.8
1.0
1.0
Grose et al. (1980)
Kenoyer et al. (1981)
Friberg et al. (1972)
Abraham et al. (1980)
Phalen et al. (1980)
Frager et al. (1979)
Coffin and Gardner (1972b)
Goldstein et al. (1972)
Friberg et al. (1972)
Goldstein et al. (1971b)
Bergers et al. (1983)
Warshauer et al. (1974)
Goldstein et al. (1971a)
Gardner et al. (1971)
Dowel 1 et al. (1970)
Hadley et al. (1977)
Goldstein et al. (1977)
Hurst et al. (1970)
Hurst and Coffin (1971)
Alpert et al. (1971b)
Coffin et al. (1968)
Coffin and Gardner (1972b)
Schwartz and Christman (1979)
Shingu et al. (1980)
McAllen et al. (1981)
Plopper et al. (1979)
Dungworth et al. (1975)
Castleman et al. (1977)
Boorman et al. (1977, 1980)
Barry et al. (1983)
Zitnik et al. (1978)
Stephens et al. (1974a)
Last et al. (1979)
Brummer et al. (1977)
Eustis et al. (1981)
Freeman et al. (1974)
Castleman et al. (1980)
Freeman et al. (1973)
Cavender et al. (1977)
1-129
-------
TABLE 1-15 (continued). SUMMARY TABLE: EFFECTS OF OZONE ON HOST DEFENSE
MECHANISMS IN EXPERIMENTAL ANIMALS
Effect/response
Increased suscepti-
bility to infection
Altered immune
activity
03 concentration, ppm
0.08
0.08, 0.1
0.1
0.1
0.1, 0.3
[0.2], 0.4, 0.7
0.3
0.5
[0.64]
0.7, 0.9
1.0
0.1
0.5, 0.8
0.5, 0.8
0.59
0.8
References
Coffin et al . (1967)
Miller et al . (1978a)
Ehrlich et al. (1977)
Aranyi et al. (1983)
11 ling et al. (1980)
Bergers et al . (1983)
Abraham et al . (1982)
Wolcott et al. (1982)
[Sherwood et al . (1984)]
Coffin and Blommer (1970)
Thomas et al. (1981b)
Aranyi et al . (1983)
Osebold et al . (1979, 1980)
Gershwin et al. (1981)
Campbell and Hilsenroth
(1976)
Fujimaki et al. (1984)
At this time, there are a number of hypotheses proposed to explain the
mechanism of this phenomenon (Mustafa and Tierney, 1978; Schwartz et al. ,
1976; Mustafa et al., 1977; Berliner et al., 1978; Gertner et al. , 1983b;
Bhatnagar et al., 1983). Evidence by Nambu and Yokoyama (1983) indicates that
although the pulmonary antioxidant system (glutathione peroxidase, glutathione
reductase, and glucose-6-phosphate dehydrogenase) may play an active role in
defending the lung against ozone, it does not explain the mechanism of toler-
ance in that the development of tolerance does not coincide with the described
biochemical enhancement of the antioxidant system in the lungs of rats.
From this literature, it would appear that tolerance, as seen in animals,
may not be the result of any one single biological process, but instead may
result from a number of different events, depending on the specific response
measured. Tolerance does not imply complete or absolute protection, because
continuing injury does still occur, which could potentially lead to nonrever-
sible pulmonary changes.
Tolerance may not be long-lasting. During 0, exposure, the increase in
antioxidant metabolism reaches a plateau and recovery occurs a few days after
1-130
-------
exposure ceases. Upon re-exposure, effects observed are similar to those that
occurred during the primary exposure (Chow et al., 1976).
1.8.4 Extrapulmonary Effects of Ozone
It is still believed that 03, on contact with respiratory system tissue,
immediately reacts and thus is not absorbed or transported to extrapulmonary
sites to any significant degree. However, several studies suggest that possibly
products formed by the interaction of 0- and respiratory system fluids or
tissue can produce effects in lymphocytes, erythrocytes, and serum, as well as
in the parathyroid gland, the heart, the liver, and the CMS. Ozone exposure
also produces effects on animal behavior that may be caused by pulmonary
consequences of 0~, or by nonpulmonary (CMS) mechanisms. The mechanism by
which 03 causes extrapulmonary changes is unknown. Mathematical models of 0~
dosimetry predict that very little 0, penetrates to the blood of the alveolar
capillaries. Whether these effects result from 0, or a reaction product of 0,
which penetrates to the blood and is transported is the subject of speculation.
1.8.4.1 Central Nervous System and Behavioral Effects. Ozone significantly
affects the behavior of rats during exposure to concentrations as low as
3
235 (jg/m (0.12 ppm) for 6 hr. With increasing concentrations of 0~, further
decreases in unspecified motor activity and in operant learned behaviors have
been observed (Konigsberg and Bachman, 1970; Tepper et al., 1982; Murphy
et al., 1964; and Weiss et al., 1981). Tolerance to the observed decrease in
motor activity may occur on repeated exposure. At low 0- exposure concentra-
3
tions (490 pg/m , 0.25 ppm), an increase in activity is observed after exposure
3
ends. ^Higher 0- concentrations (980 (jg/m , 0.5 ppm) produce a decrease in
rodent activity that persists for several hours after the end of exposure
(Tepper et al., 1982, 1983).
The mechanism by which behavioral performance is reduced is unknown.
Physically active responses appear to enhance the effects of 0~, although this
may be the result of an enhanced minute volume that increases the effective
concentration delivered to the lung. Several reports indicate that it is
unlikely that animals have reduced physiological capacity to respond, prompt-
ing Weiss et al. (1981) to suggest that 03 impairs the inclination to respond.
Two studies indicate that mice will respond to remove themselves from an
3
atmosphere containing greater than 980 (jg/m (0.5 ppm) (Peterson and Andrews,
1963, Tepper et al. , 1983). These studies suggest that the aversive effects
1-131
-------
of 0- may be due to lung irritation. It is unknown whether lung irritation,
odor, or a direct effect on the CNS causes change in rodent behavior at lower
0- concentrations.
1.8.4.2 Cardiovascular Effects. Studies on the effects of 03 on the cardio-
vascular system are few, and to date there are no reports of attempts to con-
firm these studies. The exposure of rats to 0Q alone or in combination with
3 3
cadmium (1176 ug/m , 0.6 ppm 0~) resulted in measurable increases in systolic
•3 s
pressure and heart rate (Revis et al., 1981). No additive or antagonistic
response was observed with the combined exposure. Pulmonary capillary blood
3
flow and PaOy decreased 30 min following exposure of dogs to 588 (jg/m (0.3
ppm) of 03 (Friedman et al., 1983). The decrease in pulmonary capillary blood
flow persisted for as long as 24 hr following exposure.
1.8.4.3 Hematological and Serum Chemistry Effects. The data base for the
effects of 0- on the hematological system is extensive and indicates that 03
or one of its reactive products can cross the blood-gas barrier, causing
changes in the circulating erythrocytes (RBC) as well as significant differ-
ences in various components of the serum.
Effects of 03 on the circulating RBCs can be readily identified by exa-
mining either morphological and/or biochemical endpoints. These cells are
structually and metabolically well understood and are available through rela-
tively non-invasive methods, which makes them ideal candidates for both human
and animal studies. A wide range of structural effects have been reported in
a variety of species of animals, including an increase in the fragility of
3
RBCs isolated from monkeys exposed to 1470 (jg/m (0.75 ppm) of 03 4 hr/day for
4 days (Clark et al., 1978). A single 4-hr exposure to 392 (jg/m3 (0.2 ppm)
also caused increased fragility as well as sphering of RBCs of rabbits (Brinkman
et al., 1964). An increase in the number of RBCs with Heinz bodies was detected
Q
following a 4-hr exposure to 1666 (jg/m (0.85 ppm). The presence of such
inclusion bodies in RBCs is an indication of oxidant stress (Menzel et al.,
1975a).
These morphological changes are frequently accompanied by a wide range of
3
biochemical effects. RBCs of monkeys exposed to 1470 ug/m (0.75 ppm) of 03
for 4 days also had a decreased level of glutathione (GSH) and decreased
acetylcholinesterase (AChE) activity, an enzyme bound to the RBC membranes.
The RBC GSH activity remained significantly lower 4 days postexposure (Clark
et al., 1978).
1-132
-------
Animals deficient in vitamin E are more sensitive to 0,. The RBCs from
these animals, after being exposed to 0-, had a significant increase in the
activity of GSH peroxidase, pyruvate kinase, and lactic dehydrogenase, but had
3
a decrease in RBC GSH after exposure to 1568 |jg/m (0.8 ppm) for 7 days (Chow
and Kaneko, 1979). Animals with a vitamin E-supplemented diet did not have
any changes .in glucose-6-phosphate dehydrogenase (G-6-PD), superoxide dismutase,
3
or catalase activities. At a lower level (980 ug/m , 0.5 ppm), there were no
changes in GSH level or in the activities of GSH peroxidase or GSH reductase
(Chow et al., 1975). Menzel et al. (1972) also reported a significant increase
in lysis of RBCs from vitamin E-deficient animals after 23 days of exposure to
3
980 ug/m (0.5 ppm). These effects were not observed in vitamin E-supplemented
rats. Mice on a vitamin E-supplemented diet and those on a deficient diet
3
showed an increase in G-6-PD activity after an exposure of 627 ug/m (0.32 ppm)
of Q~ for 6 hr. Decreases observed in AChE activity occurred in both groups
(Moore et al., 1980).
Other blood changes are attributed to 0~. Rabbits exposed for 1 hr to
3
392 ug/m (0.2 ppm) of 0, showed a significant drop in total blood serotonin
3
(Veninga, 1967). Six- and 10-month exposures of rabbits to 784 ug/m (0.4 ppm)
of 0~ produced an increase in serum protein esterase and in serum trypsin
inhibitor. This latter effect may be a result of thickening of the small
pulmonary arteries. The same exposure caused a significant decrease in albumin
levels and an increase in alpha and gamma globulins (P'an and Jegier, 1971,
1976; P'an et al., 1972; Jegier, 1973). Chow et al. (1974) reported that the
serum lysozyme level of rats increased significantly after 3 days of continuous
exposure to 0, but was not affected when the exposure was intermittent (8 hr/day,
3
7 days). The 03 concentration in both studies was 1568 ug/m (0.8 ppm) of 0~.
Short-term exposure to low concentrations of 0~ induced an immediate
•5 -
change in the serum creatine phosphokinase level in mice. In this study, the
03 doses were expressed as the product of concentration and time. The C x T
value for this effect ranged from 0.4 to 4.0 (Veninga et al., 1981).
A few of the hematological effects observed in animals (i.e., decrease in
GSH and AChE activity and the formation of Heinz bodies) following exposure to
0, have also been seen following in vitro exposure of RBCs from humans (Freeman
J —— ——~~~~~~~
and Mudd, 1981; Menzel et al., 1975b; Verweij and Van Steveninck, 1981). A
common effect observed by a number of investigators is that 0« inhibits the
membrane ATPase activity of RBCs (Koontz and Heath, 1979; Kesner et al., 1979;
1-133
-------
Kindya and Chan, 1976; Freeman eta"]., 1979; Verweij and Van Steveninck,
1980). It has been postulated that this inhibition of ATPase could be related
to the spherocytosis and increased fragility of RBCs seen in animal and human
cells.
Although these jr» vitro data are useful in studying mechanisms of action,
it is difficult to extrapolate these data to any effects observed in man. Not
only is the method of exposure not physiological, but the actual concentration
of 0« reaching the RBC cannot be determined with any accuracy.
1.8.4.4 Cytogenetic and Teratogem'c Effects. Uncertainty still exists regard-
ing possible reproductive, teratogenic, and mutational effects of exposure to
ozone. Based on various _1n vitro data, a number of chromosomal effects of
ozone have been described for isolated cultured cell lines, human lymphocytes,
and microorganisms (Fetner, 1962; Hamelin et al., 1977a,b, Hamelin and Chung,
1975a,b, 1978; Scott and Lesher, 1963; Erdman and Hernandez, 1982; Guerrero
et al,, 1979; Dubeau and Chung, 1979, 1982). The interpretation, relevance,
and predictive values of such studies to human health are questionable since
(1) the concentrations used were many-fold greater than what is found in the
ambient air (see Chapter 10); (2) extrapolation of jni vitro exposure concentra-
tions to human exposure dose is not yet possible; and (3) direct exposure of
isolated cells to ozone is highly artifactual since it bypasses all the defenses
of the host that would normally be functioning in protecting the individual
from the inhaled gas. Furthermore, the direct exposure of isolated cells j_n
vitro to ozone may result in chemical reactions between ozone and culture
media that might not occur iji vivo.
Important questions still exist regarding in vivo cytogenetic effects of
ozone in rodents and humans. Zelac et al. (1971a,b) reported chromosomal
abnormalities in peripheral leukocytes of hamsters exposed to 03 (0.2 ppm).
Combined exposures to ozone and radiation (227-233 rads) produced an additive
effect on the number of chromosome breaks in peripheral leukocytes. These
specific findings were not confirmed by Gooch et al. (1976) or by Tice et al.
(1978), but sufficient differences in the various experimental protocols make
a direct comparison difficult. The latter group did report significant increases
in the number of chromatid deletions and achromatic lesions resulting from
exposure to 0.43 ppm ozone.
Because the volume of air inspired during pregnancy is significantly
enhanced, the pregnant animal may be at greater risk to low levels of ozone
1-134
-------
exposure. Early studies on the possible teratogenic effects of ozone have
suggested that exposures as low as 0.2 ppm can reduce infant survival rate and
cause unlimited incisor growth (Brinkman et al., 1964; Veninga, 1967). Kavlock
et al. (1979, 1980) found that pregnant rats exposed to 1.0 and 1.49 ppm ozone
showed a significant increase in embryo resorption rate, slower growth, slower
development of righting reflexes, and delayed grooming and rearing behavior,
but no increase in neonatal mortality was observed.
1.8.4.5 Other Extrapulmonary Effects. A series of studies was conducted to
show that Q~ increases drug-induced sleeping time in a number of species of
animals (Gardner et al., 1974; Graham, 1979; Graham et al,, 1981, 1982a,b,
1983, 1985). At 1960 MQ/m3 (1-0 ppm), effects were observed after 1, 2, and 3
days of exposure. As the concentration of O™ was reduced, increasing numbers
of daily 3-hr exposures were required to produce a significant effect. At the
3
lowest concentration studied (196 \ig/m , 0.1 ppm), the increase was observed
at days 15 and 16 of exposure. It appears that this effect is not specific to
the strain of mouse or to the three species of animals tested, but it is
sex-specific, with females being more susceptible. Recovery was complete
within 24 hr after exposure. Although a number of mechanistic studies have
been conducted, the reason for this effect on pentobarbital-induced sleeping
time is not known. It has been hypothesized that some common aspect related
to liver drug metabolism is quantitatively reduced (Graham et al., 1983).
Several investigators have attempted to elucidate the involvement of the
endocrine system in 0- toxicity. Most of these studies were designed to
investigate the hypothesis that the survival rate of mice and rats exposed to
lethal concentrations of 0- could be increased by use of various thyroid
blocking agents or by thyroidectomy. To follow up these findings, demons and
Garcia (1980a,b) and demons and Wei (1984) investigated the effects of a
3
24-hr exposure to 1960 |jg/m (1.0 ppm) of 0, on the hypothalamo-pituitary-thyroid
system of rats. These three organs regulate the function of each other through
various hormonal feedback mechanisms. Ozone caused decreases in serum concen-
tration of thyroid stimulating hormone (TSH), in circulating thyroid hormones
(T, and T.) and in protein-bound iodine. No alterations were observed in many
other hormone levels measured. Thyroidectomy prevented the effect of 0~ on
TSH and T. and hypophysectomy prevented effects on T«, unless the animals were
supplemented with T- in their drinking water. The thyroid gland itself was
altered (e.g., edema) by 03> The authors hypothesyzed that 03 alters serum
binding of these hormones.
1-135
-------
The extrapulmonary effects of ozone in experimental animals are summarized
in Figure 1-12 and Table 1-16. Criteria used in developing the summary were
presented in Section 1.8.1.
1.8.5 Interaction of Ozone With Other Pollutants
Combined exposure studies in laboratory animals have produced varied
results, depending upon the pollutant combination evaluated and the measured
variables. Additive and/or possibly synergistic effects of 0, exposure in
combination with N02 have been described for increased susceptibility to
bacterial infection (Ehrlich et al., 1977, 1979; Ehrlich, 1980, 1983), morpho-
logical lesions (Freeman et al., 1974), and increased antioxidant metabolism
(Mustafa et al., 1984). Additive or possibly synergistic effects from exposure
to 0~ and HUSO, have also been reported for host defense mechanisms (Gardner
et al., 1977; Last and Cross, 1978; Grose et al., 1982), pulmonary sensitivity
(Osebold et al. 1980), and collagen synthesis (Last et al., 1983), but not for
morphology (Cavender et al., 1977; Moore and Schwartz, 1981). Mixtures of 03
and (NH^Op SO- had synergistic effects on collagen synthesis and morphometry,
including percentage of fibroblasts (Last et al., 1983, 1984a).
Combining 0, with other particulate pollutants produces a variety of
responses, depending on the endpoint measured. Mixtures of 0,, Fe^SO-)^
HpSO-, and (NH^^SO* produced the same effect on clearance rate as exposure to
Q~ alone. However, when measuring changes in host defenses, the combination
of 03 with N02 and ZnS04 or 03 with S02 and (NH4)2S04 produced enhanced effects
that can not be attributed to 03 only.
However, since these issues are complex, they must be addressed experi-
mentally using exposure regimens for combined pollutants that are more represen-
tative of ambient ratios of peak concentrations, frequency, duration, and time
intervals between events.
The interactive effects of 0,, with other pollutants are summarized in
Figure 1-13 and Table 1-17.
1.8.6 Effects of Other Photochemical Oxidants
There have been far too few controlled toxicological studies with the
other oxidants to permit any sound scientific evaluation of their contribution
to the toxic action of photochemical oxidant mixtures. When the effects seen
after exposure to 03 and PAN are examined and compared, it is obvious that the
1-136
-------
X
*
ca
--J
0.1 _
0.2-
E 0.3-
a
a
c 0.4^
.2
ID
l»
S 0.5-
0)
0
e
8 0.6-
0)
c
o
5 0.7-
0,8-
0.9_
1 n
I
i
i
i
i
i
<
i
1
> . 4
i
I <
<
i
i
I
i {
1
t
1
1
b „ — _— ...J
>
1
»
> <
. (
t
<
t
^
t
1
t
i
1
1
t— 4
> 4
>
1 J
(
^
-
\
ft ™ _, ~4
»
^
>
•
j
b m , 1 t -— . ..
Figure 1-12. Summary of extrapulmonary effects of ozone in
experimental animals. See Table 1-16 for reference citations of studies
summarized here.
-------
TABLE 1-16. SUMMARY TABLE: EXTRAPULMONARY EFFECTS OF OZONE
IN EXPERIMENTAL ANIMALS
Effect/response 03 concentration, ppm
CNS effects 0.05, 0.5
0.1 - 1.0
0.12 - 1.0
0.2, 0.3, 0.5, 0.7
0.5
0.5
0.5
0.6
1.0
1.0
References
Konigsberg and Bachman (1970)
Weiss et al. (1981)
Tepper et al. (1982)
Murphy et al. (1964)
Tepper et al. (1983)
Reynolds and Chaffee (1970)
Xintaras et al . (1966)
Peterson and Andrews (1963)
Fletcher and Tappel (1973)
Trams et al. (1972)
Hematological effects
Chromosomal, reproduc-
tive, teratological
effects
Liver effects
Endocrine system
effects
0.06, 0.12, 0.48
0.2
0.2, 1.0
0.25, 0.32, 0.5
0.4
0.4
0.5
0.64
0.75
0.8
0.8
0.85
0.86
1.0
1.0
1.0
0.1
0.2
0.24, 0.3
0.43
0.44
1.0
0.1, 6.25, 0.5, 1.0
0.82
1.0
0.75
0.75
0.75
0.75
1.0
1.0
Calabrese et al. (1983)
Brinkman et al, (1964)
Veninga (1967, 1970)
Veninga et al. (1981)
Moore et al. (1980; 1981a,b)
Jegier (1973)
P'an and Jegier (1972, 1976)
Menzel et al. (1972)
Larkin et al. (1983)
Clark et al. (1978) .
Chow and Kaneko (1979)
Chow et al. (1974)
Menzel et al. (1975a)
Schlipkoter and Bruch (1973)
Dorsey et al. (1983)
Mizoguchi et al. (1973)
Christiansen and Giese (1954)
Brinkman et al. (1964)
Veninga (1967)
Zelac et al. (1971a)
Tice et al. (1978)
Kavlock et al. (1979)
Kavlock et al. (1980)
Graham (1979)
Graham et al. (1981, 1982a,b)
Veninga et al. (1981)
Gardner et al. (1974)
Atwal and Wilson (1974)
Atwal et al. (1975)
Atwal and Pernsingh (1981, 1984)
Pernsingh and Atwal (1983)
demons and Garcia (1980a,b)
demons and Wei (1984)
1-138
-------
E
Q.
a
«.
c
o
0)
o
o
o
0!
O
N
O
'
0.1-
0.2-
0.3-
0,4-
0.5-
0.6-
0.7-
0.8-
0.9-
1.O
c
4
<
i
! I I
3
0
3 • * (
t I
t
'
I
I I
I 0
O
} *
|l
> O
I
Figure 1 -13. Summary of effects in experimental animals exposed to
ozone combined with other pollutants. See Table 1-17 for reference
citations of studies summarized here.
-------
TABLE 1-17. SUMMARY TABLE: INTERACTION OF OZONE WITH OTHER POLLUTANTS
IN EXPERIMENTAL ANIMALS
Effect/response
Pollutant concentrations
References
Increased
pulmonary
lesions
Increased
pulmonary
sensitivity
Increased anti-
oxidant metabolism
and D£ consumption
Altered mucus
secretion
Increased collagen
synthesis
Increased
susceptibility to
respiratory
infections
[0.25 ppm 03
+2.5 ppm N02]
[0.5 ppm 03
+ 1 mg/m3 H2S04]
[0.5 ppm Q3
+ 10 mg/m3 H2S04
0.64, 0.96 ppm 03
+ 5 mg/m3 (NH4)2 S04
0.9 ppm 03
+0.9 ppm N02
1.2 ppm 03
+ 5 mg/m3 (NH4)2S04
0.5 ppm 03
+ 1 mg/m3 H2S04
0.45 ppm 03
+4.8 ppm N02
0.5 ppm 03
+1.1 mg/m3 H2S04
[0.5], [0.8], 1.5 ppm 03
+ 5 mg/m3 (NH4)2S04
0.5 ppm 03
+ 1 mg/m3 H2S04
0.64, 0.96 ppm 03
+ 5 mg/m3 (NH4)2S04
0,05 ppm 03
+ 3760 Mg/m3 (NH4)2S04
0.05 ppm Q3
+ 100-400 Mg/m3 N02
+1.5 mg/m3 ZnS04
0.1 ppm Q3
+0.9 mg/m3 H2S04
(sequential exposure)
0.1 ppm 03
+4.8 mg/m3 H2S04
0.1 ppm 03
+ 940 Mg/m3 N02
0.1 ppm 03
+13.2 mg/m3 S02
+1.0 mg/m3 (NH4)2S04
Freeman et al. (1974)
Moore and Schwartz (1981)
Cavender et al. (1978)
Last et al. (1984a)
Freeman et al. (1974)
Last et al. (1983)
Osebold et al. (1980)
Mustafa et al. (1984)
Last and Cross (1978);
Last and Kaizu (1980)
Last et al. (1983)
Last et al. (1983)
Last et al. (1984a)
Ehrlich et al. (1977, 1979);
Ehrlich (1980)
Ehrlich (1983)
Gardner et al. (1977).
Grose et al. (1982)
Ehrlich (1980)
Aranyi et al. (1983)
1-140
-------
TABLE 1-17 (continued). SUMMARY TABLE: INTERACTION OF OZONE
WITH OTHER POLLUTANTS
Effect/response Pollutant concentrations Referencest
Altered upper [0.1 ppm 03 Grose et al. (1980)
respiratory +1.1 mg/m3 H2S04]
clearance (sequential exposure)
mechanisms 0.4 ppm 03 Goldstein et al. (1974)
+7.0 ppm N02
0.5 ppm 03 Last and Cross (1978)
+ 3 mg/m3 H2S04
[0.8 ppm 03 Phalen et al. (1980)
+3.5 mg/m3
(Fe2(SQ4)3
+ H2S04
+ (NH4)2S04}]
test animals must be exposed to concentrations of PAN much greater than those
needed with 0- to produce a similar effect on lethality, behavior modification,
morphology, or significant alterations in host pulmonary defense system (Campbell
et al., 1967; Dungworth et al., 1969; Thomas et al., 1979, 1981a). The concen-
trations of PAN required to produce these effects are many times greater than
what has been measured in the atmosphere (0.047 ppm).
Similarly, most of the investigations reporting H^O^ toxicity have involved
concentrations much higher than those found in the ambient air, or the investi-
gations were conducted by using various jin vitro techniques for exposure.
Very limited information is available on the health significance of inhalation
exposure to gaseous \\JSy. Because H^Og is highly soluble, it is generally
assumed that it does not penetrate into the alveolar regions of the lung but
is instead deposited on the surface of the upper airways (Last et al., 1982).
Unfortunately, there have not been studies designed to look for possible
effects in this region of the .respiratory tract.
A few in•vitro studies have reported cytotoxic, genotoxic, and biochemical
effects of HpOx when using isolated cells or organs (Stewart et al., 1981;
Bradley et al., 1979; Bradley and Erickson, 1981; Speit et al., 1982; MacRae
and Stich, 1979). Although these studies can provide useful data for studying
possible mechanisms of action, it is not yet possible to extrapolate these
responses to those that might occur in the mammalian system.
1-141
-------
Field and epidemiologies! studies have shown that human health effects
from exposure to ambient mixtures of oxidants and other airborne pollutants
can produce human health effects (Chapter 12). Few such studies have been
conducted with laboratory animals, because testing and measuring of such
mixtures is not only complicated, but extremely costly. In these studies, the
investigators attempted to simulate the photochemical reaction products pro-
duced under natural conditions and to define the cause-effect relationship.
Exposure to complex mixtures of oxidants plus the various components
found in UV-irradiated auto exhaust indicates that certain effects, such as
histopathological changes, increase in susceptibility to infection, a variety
of altered pulmonary functional activities were observed in this oxidant
atmosphere which was not reported in the nonirradiated exhaust (Murphy et al.,
1963; Murphy, 1964; Nakajima eta!., 1972; Hueter etal., 1966). Certain
other biological responses were observed in both treatment groups, including a
decrease in spontaneous activity, a decrease in infant survival rate, fertil-
ity, and certain pulmonary functional abnormalities (Hueter et al., 1966;
Boche and Quilligan, 1960; Lewis et al,, 1967).
Dogs exposed to UV-irradiated auto exhaust containing oxidants either
with or without SO showed significant pulmonary functional abnormalities that
*\
had relatively good correlation with structural changes (Hyde et al., 1978;
Gillespie, 1980; Lewis et al., 1974). There were no significant differences
in the magnitude of the response in these two treatment groups, indicating
that oxidant gases and SO did not interact in any synergistic or additive
y\
manner.
1.9 CONTROLLED HUMAN STUDIES OF THE EFFECTS OF OZONE AND OTHER PHOTOCHEMICAL
OXIDANTS
A number of important controlled studies discussed in this chapter have
reported significant decrements in pulmonary function associated with 0^
exposure (Table 1-18). In most of the studies reported, greatest attention
has been accorded decrements in FEV-, „, as this variable represents a summation
of changes in both volume and resistance. While this is true, it must be
pointed out that for exposure concentrations critical to the standard-setting
process (i.e., <0.3 ppm 03), the observed decrements in FEV^ Q primarily
reflect FVC decrements of similar magnitude, with little or no contribution
from changes in resistance.
1-142
-------
TABLE 1-18. SUMMARY TABLE: CONTROLLED HUMAN EXPOSURE TO OZONE
CO
Ozone® k
concentration Measurement ' Exposure
ug'/i3
HEALTHY
627
1960
980
980
1470
ppm method duration
ADULT SUBJECTS AT REST
0.32 HAST, NBKI 2 hr
1.0
0.5 CHEM, NBKI • 2 hr
0.50 CHEM, NBKI 2 hr
0. 75
Activity*1
level (Vr) Observed effects(s)
R Specific airway resistance increased with
acetylcholine challenge; subjective symptoms
in 3/14 at 0.32 ppm and 8/14 at 1.0 ppm.
R (10) Decrement in forced expiratory volume and
flow.
R (8) Decrement in forced expiratory volume and
flow.
No. and sex
of subjects Reference
v
13 nale Konig et al., 1980
1 female
40 male Folinsbee et al,,
(divided into four 1978
exposure groups)
8 male Horvath et al.,
7 female 1979
EXERCISING HEALTHY ADULTS
235
353
470
588
784
314
470
627
353
470
588
784
392
686
0.12 CHEM, UV 2.5 hr
0.18
0.24
0.30
0.40
0.16 UV, UV 1 hr
0.24
0.32
0.18 CHEM, UV 2.5 hr
0.24
0.30
0.40
0.20 UV, UV 1 hr
0. 35 (mouth-
piece)
IE (65) Decreient in forced expiratory volume and
i 15-nin intervals flow suggested at 0.12 ppm with larger
decrements at >_ 0, 18 ppm; respiratory
frequency and specific airway resistance
increased and tidal volume decreased at
>_ 0.24 ppm; coughing reported at all
concentrations, pain and shortness of
breath at S 0.24 ppm.
CE (57) Small decrements in forced expiratory
volume at 0.16 ppm with larger decrements
at >0,24 ppa; lower-respiratory symptoms
increased at >0. 16 ppm.
IE (65) Individual responses to Os were highly
915-nin intervals reproducible for periods as long as 10
months; large intersubject variability
in response due to intrinsic responsiveness
to Os.
IE (77.5) @ vari- Decrement in forced expiratory volume and
able competitive flow with IE and CE; subjective symptoms
intervals increased with 03 concentration and may
CE (77.5) limit performance; respiratory frequency
increased and tidal volume decreased with
CE.
135 male McDonnell et al.,
(divided into six 1983
exposure groups)
42 male Avol et al,, 1984
8 female
(competitive
bicyclists)
32 male McDonnell et al.,
1985a
10 male Adams and Schelegle,
(distance runners) 1983
-------
TABLE 1-18 (continued). SUWiARY TABLE: CONTROLLED HUMAN EXPOSURE TO OZONE
Ozonta
concentration
Mg/»4
392
823
980
392
490
412
,_, 490
i
*" 588
980
. 725
980
1470
784
784
pp«
0.2
0.42
0.50
0.20
0.25
0.21
0.25
0.3
0.5
0.37
0.50
0.75
0.4
0.4
Measurement * Exposure Activity
tiethod duration level (Vc)
UV, UV 2 hr IE (30 for male,
18 for female
subjects)
@ 15-min intervals
UV, UV 2 hr IE (68)
(4) 14-min periods
UV, UV 1 hr CE (81)
UV, UV 1 hr CE (63)
CHEM, NBKI • 2 hr R (10), IE (31,
50, 67)
@ 15-rain intervals
MAST, NBKI 2 hr R (11) & IE (29) .
@ 15-inin intervals
UV, NBKI 2 hr IE (2xR)
@ 15-rain intervals
CHEM, NBKI & 3 hr IE (4-5xR)
HAST, NBKI
Observed effects(s)
Repeated daily exposure to 0,2 ppm did not
affect response at higher exposure concen-
trations (0.42 or 0.50 ppm); large inter-
subject variability but individual
pulmonary function responses were highly
reproducible. «
Large intersubject variability in response;
significant concentration-response relation-
ships for pulmonary function and respiratory
symptoms.
Decrement in forced expiratory volume and
flow; subjective symptoms may limit per-
formance.
Increased responsiveness to 03 lasts for
24 hr, may persist in some subjects for
48 hr, but is generally lost within 72 hr.
Decrement in forced expiratory volume and
flow; the magnitude of the change was
related to 03 concentration and VV.
Total lung capacity and inspiratory
capacity decreased with IE (50, 67); no
change in airway resistance or residual
volume even at highest IE (67). No-
significant changes in pulmonary function
were observed at 0.1 ppn.
Good correlation between dose (concen-
tration x VV) and decrement in forced
expiratory volume and flow.
Specific airway 'resistance increased with
histamlne challenge; no changes were
observed at concentrations of 0.2 ppm.
Decrement in forced expiratory volume and
SG was greatest on the 2nd of 5 exposure
days; attenuated response by the 4th day
of exposure.
No. and sex
of subjects
8 male
13 female
20 male
6 male
1 female
(distance cyclists)
19 male
7 female
40 male
(divided into four
exposure groups)
20 male
8 female (divided into
six exposure groups)
12 male
7 female
(divided into three
exposure groups)
10 male
4 female
Reference
Gliner et al. , 1983
Kulle et al., 1985
Folinsbee et al. ,
1984
Folinsbee and
Horvath, 1986
Folinsbee et al. ,
1978
Silverman et al, ,
1976
Dimeo et al., 1981
Fa'rrell et al., 1979
-------
TABLE 1-18 (continued). SUMHARY TABLE: CONTROLLED HUMAN EXPOSURE TO OZONE
en
Ozone .
concentration Measurement ' Exposure
ug/nr1 ppm nethod duration
784 0.4 CHEH, UV 3 hr
784 0.4 CHEM, UV 2.5 hr
823 0.42 'UV, UV 2 hr
882 0.45 UV, UV 2 hr
921 0.47 UV, NBKI 2 hr
980 0.5 MAST, NflKI 6 hr
1176 0.6 UV, NBKI 2 hr
(noseclip)
1470 0.75 HAST, NBKI 2 hr
Activity
level (VE)
IE (4-5xR)
for 15 lain
IE (71)
S> 15-min intervals
IE (30)
IE (27)
@ 20-min intervals
IE (3xR)
IE (44) for two
15-m'n periods
IE (2xR)
i 15-min intervals
IE (2xR)
1 15-min intervals
Observed effects(s)
Decrement in forced expiratory volume was
greatest on the 2nd of 5 exposure days;
attenuation of response occurred by the
5th day and persisted for 4 to 7 days.
Enhanced bronchoreactivity with ;
methacholine on the first 3 days;
attenuation of response occurred by
the 4th and 5th day and persisted
for > 7 days.
Atropine pretreatment prevented the
increased R w observed with Oa exposure,
partially blocked the decreased forced
expiratory flow, but did not prevent the
Da-induced decreases in FVC and TLC,
changes in exercise ventilation, or
respiratory symptoms.
Decrement in forced expiratory volume and
flow greatest on the 2nd of 5 exposure
days; attenuation of response occurred by
the 5th day and persisted for < 14 days with
considerable intersubjeet variability.
Increased responsiveness to Oa was found
with a 2nd Oa challenge given 48 hr after
the initial exposure.
Decrement in forced expiratory volume and
flow greatest on the 2nd of 4 exposure
days; attenuation of response occurred by
the 4th day and persisted for 4 days.
Small decrements in forced expiratory
volume and specific airway conductance.
Specific airway resistance increased in 7
nonatopic subjects with histamine and
methacholine and in 9 atopic subjects
with histaaine.
Decrements in spirometric variables
(20&-55SQ; residual volume and closing
capacity increased.
No. and sex
of subjects '
13 male
11 female
(divided intd two
exposure groups)
8 male
24 male
1 male
5 female
8 male
3 female
19 male
1 female
11 male
5 female (divided
by history of atopy)
12 male
Reference 7
Kulle et al., 1982
Beckett et al.
1985 '
Horvath et al., 1981
8ed1 et al., 1985
Linn et al., 1982b
Kerr et al., 1975
Holtzman et al. ,
1979
Hazucha et al . ,
1973
-------
TABLE 1-18. (continued) SUHHARY TABLE: COHTMLLED B1WSH EXPOSURE TO OZONE
Ozone • ,
concentration Measurement '
pgTI3 pp »ethod
Exposure
duration
Activityd
level (VE)
Observed effects(s)
Ko. and sex
of subjects
Reference
EXERCISING HEALTHY CHILDREN
235 0.12 CHEH, UV
2.5 hr
IE (39)
tlS-nrin intervals
Small decrements in forced expiratory
volune, persisting for 24 hr. No subjec-
tive symptoms.
23 male
(8-11 yrs)
McDonnell et al.,
1985b,c
ADULT ASTHMATICS
392 0.2 CHEH, NBKI
490 0.25 CHEK, NBKI
2 hr
2 hr
IE (2xR)
@ 15-ain intervals
R
No significant changes in pulmonary func-
tion. Snail changes in blood biochemistry.
Increase in symptom frequency reported.
No significant changes in pulmonary func-
tion.
20 male
2 feaale
5 nales
12 female
Linn et al., 1978
Silverman, 1979
ADOLESCENT ASTHMATICS
V 235 0.12 UV
i— «
.p.
cr,
SUBJECTS WITH CHRONIC OBSTRUCTIVE
235 0.12 UV, NBKI
353 0.18 UV, NBKI
490 0.25
392 0.2 CHEH, NBKI
588 0.3
784 0.41 UV, UV
1 hr
(mouthpiece)
IUNS DISEASE
1 hr
1 hr
2 hr
3 hr
R
IE (variable)
8 15-min intervals
IE (variable)
@ 15-min intervals
IE (28) for
7.5 min each
half hour
IE (4-5xR)
for 15 Bin
No significant changes in pulmonary function
or symptoms.
No significant changes in forced expiratory
performance or symptoms. Decreased arterial
oxygen saturation during exercise was
observed.
No significant changes in forced expiratory
performance or symptoms. Group mean arterial
oxygen saturation was not altered by Oa
exposure.
No significant changes in pulmonary function
or symptoms. Decreased arterial oxygen
saturation during exposure to 0.2 ppm.
Small decreases in FVC.and FEVa.0.
4 male
6 female
(11-18 yrs)
18 male
7 feiale
15 male
13 feiale
13 male
17 sale
3 female
Koenig et al., 1985
Linn et al., 1982a
Linn et al., 1983
Solic et al., 1982
Kehrl et al., 1983,
1985
Kulle et al., 1984
Ranked by lowest observed effect level.
Measurement method: HAST = Kl-Coulometric (Hast neter); CHEH = gas phase chenriluminescence; UV = ultraviolet photometry.
cCalibration method: NBKI = neutral buffered potassium iodide; UV = ultraviolet photometry.
Minute ventilation reported in L/min or as a multiple of resting ventilation. R = rest; IE = intermittent exercise; CE = continuous exercise.
-------
Results from studies of at-rest exposures to 0, have demonstrated decre-
3
merits in forced expiratory volumes and flows occurring at and above 980 ug/m
(0.5 ppm) of 03 (Folinsbee et a!., 1978; Horvath et al., 1979). Airway resis-
tance is not clearly affected at these 0~ concentrations. At or below 588
3
pg/m (0.3 ppm) of Q3, changes in pulmonary function do not occur during at
rest exposure (Folinsbee et al., 1978), but the occurrence of some 0,,-induced
pulmonary symptoms has been suggested (Kb'nig et al., 1980).
With moderate intermittent exercise at a VE of 30 to 50 L/min, decrements
in forced expiratory volumes and flows have been observed at and above 588
3
(jg/m (0.30 ppm) of 03 (Folinsbee et al., 1978). With heavy intermittent
exercise (V^ = 65 L/min), pulmonary symptoms are present and decrements in
forced expiratory volumes and flows are suggested to occur following 2-hr
3
exposures to 235 pg/m (0.12 ppm) of 0_ (McDonnell et al., 1983). Symptoms
are present and decrements in forced expiratory volumes and flows definitely
3
occur at 314 to 470 ug/m (0.16 to 0.24 ppm) of 0™ following 1 hr of continuous
heavy exercise at a Vp of 57 L/min (Avol et al., 1984) or very heavy exercise
at a V£ of 80 to 90 L/min (Adams and Schelegle, 1983; Folinsbee et al., 1984)
and following 2 hr of intermittent heavy exercise at a VV of 65 to 68 L/min
(McDonnell et al., 1983; Kulle et al., 1985). Airway resistance is only
modestly affected with moderate exercise (Kerr et al., 1975; Parrel! et al.,
3
1979) or even with heavy exercise while exposed at levels as high as 980 |jg/m
(0.50 ppm) 03 (Folinsbee et al., 1978; McDonnell et al., 1983). Increased fR
and decreased VT, while maintaining the same VV, occur with prolonged heavy
exercise when exposed at 392 to 470 ug/m3 (0.20 to 0.24 ppm) of 03 (McDonnell
et al., 1983; Adams and ScheTegle, 1983). While an increase in RV has been
3
reported to result from exposure to 1470 ug/m (0.75 ppm) of 0_ (Hazucha et
al., 1973), changes in RV have not been observed following exposures to 980 jjg/
m (0.50 ppm) of 0-, or less, even with heavy exercise (Folinsbee et al.,
1978). Decreases in TLC and 1C have been observed to result from exposures to
3
980 ug/m (0.50 ppm) of 03 or less, with moderate and heavy exercise (Folinsbee
et al., 1978).
Recovery of the lung from the effects of 03 exposure consists of return
of pulmonary function (FVC, FEV,, and SR ) to preexposure levels. The time
course of this recovery is related to the magnitude of the 03~induced functional
decrement (i.e., recovery from small decrements is rapid). Despite apparent
functional recovery of most subjects within 24 hr, an enhanced responsiveness
1-147
-------
to a second 03 challenge may persist in some subjects for up to 48 hr (Bedi et
al., 1985; Folinsbee and Horvath, 1986).
Group mean decrements in pulmonary function can be predicted with some
degree of accuracy when expressed as a function of effective dose of 0^, the
simple product of Qg concentration, VV, and exposure duration (Silverman et
al . , 1976), The relative contribution of these variables to pulmonary decre-
ments is greater for 0~ concentration than for VV. A greater degree of
predictive accuracy is obtained if the contribution of these variables is
appropriately weighted (Folinsbee et al., 1978). However, several additional
factors make the interpretation of prediction equations more difficult. There
is considerable intersubject variability in the magnitude of individual pulmonary
function responses to 03 (Horvath et al., 1981; Gliner et al , 1983; McDonnell
et al., 1983; Kulle et al., 1985). Individual responses to a given 0- concen-
tration have been shown to be quite reproducible (Gliner et al., 1983; McDonnell
et al,, 1985a)j indicating that some individuals are consistently more respon-
sive to Og than are others. No information is available to account for these
differences. Considering the great variability in individual pulmonary re-
sponses to Q~ exposure, prediction equations that only use some form of effec-
tive dose are not adequate for predicting individual responses to 0^,
In addition to overt changes in pulmonary function, enhanced nonspecific
bronchial reactivity has been observed following exposures to 0- concentrations
~ • o
>588 fjg/m (0.3 ppm) (Holtzman et al., 1979; Konig et al., 1980; Dimeo et al.,
~~
3
1981). Exposure to 392 pg/m (0.2 ppm) of 0,, with intermittent light exercise
does not affect nonspecific bronchial reactivity (Dimeo et al., 1981).
Changes in forced expiratory volumes and flows resulting from 0- exposure
reflect reduced maximal inspiratory position (inspiratory capacity) (Folinsbee
et al., 1978). These changes, as well as altered ventilatory control and the
occurrence of respiratory symptoms, most likely result from sensitization or
stimulation of airway irritant receptors (Folinsbee et al., 1978; Holtzman et
al., 1979; McDonnell et al., 1983). The increased airways resistance observed
following 03 exposure is probably initiated by a similar mechanism. Different
efferent pathways have been proposed (Beckett et al., 1985) to account for the
lack of correlation between individual changes in SR and FVC (McDonnell
sw
et al . , 1983). The increased responsiveness of airways to histamine and
methacholine following 03 exposure most likely results from an Qg-induced
increase in airways permeability or from an alteration of smooth muscle charac-
teristics.
1-148
-------
Decrements in pulmonary function were not observed for adult asthmatics
exposed for 2 hours at rest (Silverman, 1979) or with intermittent light
3
exercise (Linn et a!., 1978) to CL concentrations of 490 |jg/m (0.25 ppm) and
less. Likewise, no significant changes in pulmonary function or symptoms were
found in adolescent asthmatics exposed for 1 hr at rest to 235 ug/m (0.12
ppm) of 03 (Koenig et al., 1985). Although these results indicate that asthma-
tics are not more responsive to 03 than are healthy subjects, experimental-design
considerations in reported studies suggest that this issue is still unresolved.
For patients with COLD performing light to moderate intermittent exercise, no
decrements in pulmonary function are observed for 1- and 2-hr exposures to Q~
Q a
concentrations of 588 |jg/m (0.30 ppm) and less (Linn et al., 1982a, 1983;
Solic et al., 1982; Kehrl et al., 1983, 1985) and only small decreases in
forced expiratory volume are observed for 3-hr exposures of chronic bronchitics
q
to 804 |jg/m (0.41 ppm) (Kulle et al., 1984). Small decreases in Sa02 have
also been observed in some of these studies but not in others; therefore,.
interpretation of these decreases and their clinical significance is uncertain.
Many variables have not been adequately addressed in the available clini-
cal data. Information derived from Oq exposure of smokers and nonsmokers is
sparse and somewhat inconsistent, perhaps partly because of undocumented
variability in smoking histories. Although some degree of attenuation appears
to occur in smokers, all current results should be interpreted with caution.
Further and more precise studies are required to answer the complex problems
associated with personal and ambient pollutant exposures. Possible age differ-
ences in response to CL have not been explored systematically. Young adults
usually provide the subject population, and where subjects of differing age
are combined, the groups studied are often too small in number to make adequate
statistical comparisons. Children (boys, aged 8 to 11 yr) have been the
subjects in only one study (McDonnell et al., 1985b) and nonstatistical compari-
son with adult males exposed under identical conditions has indicated that the
effects of 0™ on lung spirometry were very similar (McDonnell et al., 1985c).
While a few studies have investigated sex differences, they have not conclu-
sively demonstrated that men and women respond differently to 0~, and consid-
eration of differences in pulmonary capacities have not been adequately taken
into account. Environmental conditions such as heat and relative humidity may
enhance subjective symptoms and physiological impairment following 0^ exposure,
but the results so far indicate that the effects are no more than additive.
1-149
-------
In addition, there may be considerable interaction between these variables
that may result in modification of interpretations made based on available
information.
During repeated daily exposures to 03, decrements in pulmonary function
are greatest on the second exposure day (Parrel! et a!., 1979; Horvath et al.,
1981; Kulle et al., 1982; Linn et al., 1982b); thereafter, pulmonary respon-
siveness to CL is attenuated with smaller decrements on each successive day
than on the day before until the fourth or fifth exposure day when small
decrements or no changes are observed. Following a sequence of repeated daily
exposures, this attenuated pulmonary responsiveness persists for 3 (Kulle
et al., 1982; Linn et al.s 1982b) to 7 (Horvath et al., 1981) days. Repeated
daily exposures to a given low effective dose of 03 does not affect the magni-
tude of decrements in pulmonary function resulting from exposure at a higher
effective dose of 03 (Gliner et al., 1983).
There is some evidence suggesting that exercise performance may be limited
by exposure to CU, Decrements in forced expiratory flow occurring with 03
exposure during prolonged heavy exercise (V> = 65 to 81 L/min) along with
increased fR and decreased VT might be expected to produce ventilatory limita-
tions at near maximal exercise. Results from exposure to ozone during high
exercise levels (68 to 75 percent of max VOp) indicate that discomfort associ-
ated with maximal ventilation may be an important factor in limiting perfor-
mance (Adams and Schelegle, 1983; Folinsbee et al., 1984). However, there is
not enough data available to adequately address this issue.
No consistent cytogenetic or functional changes have been demonstrated in
circulating cells from human subjects exposed to 0~ concentrations as high as
3
784 to 1176 ug/m (0.4 to 0.6 ppm). Chromosome or chromatid aberrations would
therefore be unlikely at lower 03 levels. Limited data have indicated that 03
can interfere with biochemical mechanisms in blood erythrocytes and sera but
the physiological significance of these studies is questionable.
No significant enhancement of respiratory effects has been consistently
demonstrated for combined exposures of 03 with S02, N02, and sulfuric acid or
particulate aerosols or with multiple combinations of these pollutants. Most
of the available studies with other photochemical oxidants have been limited
to studies on the effects of peroxyacetyl nitrate (PAN) on healthy young and
middle-aged males during intermittent moderate exercise. No significant
effects were observed at PAN concentrations of 0.25 to 0.30 ppm, which are
1-150
-------
higher than the daily maximum concentrations of PAN reported for relatively
high oxidant areas (0.047 ppm). One study (Drechsler-Parks et al., 1984)
suggested a possible simultaneous effect of PAN and 0~; however, there are not
enough data to evaluate the significance of this effect. Further studies are
also required to evaluate the relationships between 0~ and the more complex
mix of pollutants found in the natural environment.
1.10 FIELD AND EPIDEMIOLOGICAL STUDIES OF THE EFFECTS OF OZONE AND OTHER
PHOTOCHEMICAL OXIDANTS
Field and epidemiological studies offer a unique view of health effects
research because they involve the real world, i.e., the study of human popula-
tions in their natural setting. These studies have attendant limitations,
however, that must be considered in a critical evaluation of their results.
One major problem in singling out the effects of one air pollutant in field
studies of morbidity in populations has been the interference of other environ-
mental variables that are critical. Limitations of epidemiological research
on the health effects of oxidants include: interference by other air pollutants
or interactions between oxidants and other pollutants; meteorological factors
such as temperature and relative humidity; proper exposure assessments, includ-
ing determination of individual activity patterns and adequacy of number and
location of pollutant monitors; difficulty in identifying oxidant species
responsible for observed effects; and characteristics of the populations such
as smoking habits and socioeconomic status.
The most quantitatively useful information of the effects of acute exposure
to photochemical oxidants presented in this chapter comes from the field
studies of symptoms and pulmonary function. These studies offer the advantage
of studying the effects of naturally-occurring, ambient air on a local subject
population using the methods and better experimental control typical of con-
trol led-exposure studies. In addition, the measured responses in ambient air
can be compared to clean, filtered air without pollutants or to filtered air
containing artificially-generated concentrations of 03 that are comparable to
those found in the ambient environment. As shown in Table 1-19, studies by
Linn et al. (1980, 1983) and Avol et al. (1983, 1984, 1985a,b,c) have demon-
strated that respiratory effects in Los Angeles area residents are related to
03 concentration and level of exercise. Such effects include: pulmonary
1-151
-------
TABLE 1-19. SUMMARY TABLE: ACUTE EFFECTS OF OZONE AND OTHER PHOTOCHEMICAL OXIDANTS IH FIELD STUDIES WITH A MOBILE LABORATORY3
en
IX)
Mean ozone
concentration
ug/mj ppi»
282 0.144
300 0,153
306 0,156
323 0.165
341 0.174
Measurement >c
Method
UV,
UV
UV,
UV
UV,
NBKI
UV,
NBKI
UV,
NBKI
Exposure Activity
duration level (Vg) Observed effect(s)
1 hr CE{32) Siall significant decreases in FVC (-2.11), FEV0 75
(-4.0X), FEVt.0 (-4.2%), and PEFR (-4.4X) relative
to control with no recovery during a 1-hr post-
exposure rest; no significant increases in
symptoms.
1 hr CE(53) Hi Id increases in lower respiratory symptom scores
and significant decreases in FEVi (-5.3%) and
FVC; mean changes in ambient air were not statisti-
cally different from those in purified air contain-
ing 0.16 ppm Os.
1 hr CE(38) No significant changes for total symptom score or
forced expiratory performance in normals or
asthmatics; however, FEVi remained low or
decreased further (-31) 3 hr after ambient air
exposure in asthmatics.
1 hr CE(42) Small significant decreases in FEV, (-3.3X) and
FVC with no recovery during a 1-hr postexposure
rest; TLC decreased and M\z increased slightly.
2 hr IE(2 x R) Increased symptom scores and small significant
@ 15-rain decreases in FEVi (-2.4%), FVC, PEFR, and TLC
intervals in both asthmatic and healthy subjects however,
25/34 healthy subjects were allergic and "atypi-
cal ly" reactive to 0$.
No.
of subjects
59 healthy
adolescents
(12-15 yr)
50 healthy
adults (compe-
titive bicy-
clists)
48 healthy
adults
50 asthmatic
adults
60 "healthy"
adults
(7 were
asthmatic)
34 "healthy"
adults
30 asthmatic
adults
Reference
Avol et al., ISBSa.b
Avol et al., 1984, 1985c
Linn et al,, 1983;
Avol et al., 1983
Linn et al., 1983;
Avol et al., 1983
Linn et al., 1980, 1983
Ranked by lowest observed effect level for 03 in ambient air.
Measurement method: UV = ultraviolet photonetry.
cCalibratlon method: UV = ultraviolet photonetry standard; NBKI = neutral buffered potassium iodide.
Tlinute ventilation reported In L/min or as a multiple of resting ventilation. CE = continuous exercise, IE = intermittent exercise.
-------
3
function decrements seen at 03 concentrations of 282 |jg/m (0.144 ppm) in
exercising healthy adolescents; and increased respiratory symptoms and pulmonary
function decrements seen at 0- concentrations of 300 jjg/m (0.153 ppm) in
3
heavily exercising athletes and at 0, concentrations of 341 |jg/m (0.174 ppm)
in lightly exercising normal and asthmatic subjects. The light exercise level
is probably the type most likely to occur in the exposed population of Los
Angeles. The observed effects are typically mild, and generally no substantial
differences were seen in asthmatics versus persons with normal, respiratory
health, although symptoms lasted for a few hours longer in asthmatics. Many
of the normal subjects, however, had a history of allergy and appeared to be
more sensitive to 0~ than "non-allergic" normal subjects. Concerns raised
about the relative contribution to untoward effects in these field studies by
pollutants other than 0, have been diminished by direct comparative findings
in exercising athletes (Avol et al., 1984, 1985c) showing no differences in
response between chamber exposures to oxidant-polluted ambient air with a mean
3
CL concentration of 294 yg/m (0.15 ppm) and purified air containing a con-
3
trolled concentration of generated 03 at 314 (jg/m (0.16 ppm). The relative
importance of exercise level, duration of exposure, and individual variations
in sensitivity in producing the observed effects remains to be more fully
investigated, although the results from field studies relative to those factors
are consistent with results from controlled human exposure studies (Chapter 10).
Studies of the effects of both acute and chronic exposures have been
reported in the epidemiclogical literature on photochemical oxidants. Inves-
tigative approaches comparing communities with high 0, concentrations and
communities with low 0~ concentrations have usually been unsuccessful, often
because actual pollutant levels have not differed enough during the study, or
other important variables have not been adequately controlled. The terms
"oxidant" and "ozone" and their respective association with health effects are
often unclear. Moreover, information about the measurement and calibration
methods used is often lacking. Also, as epidemiological methods improve, the
incorporation of new key variables into the analyses is desirable, such as
the use of individual exposure data (e.g., from the home and workplace).
Analyses employing these variables are lacking for most of the community
studies evaluated.
1-153
-------
Studies of effects associated with acute exposure that are considered to
be qualitatively useful for standard-setting purposes include those on irrita-
tive symptoms, pulmonary function, and aggravation of existing respiratory
disease. Reported effects on the incidence of acute respiratory illness and
on physician, emergency room, and hospital visits are not clearly related with
acute exposure to ambient 0- or oxidants and, therefore, are not useful for
deriving health effects criteria for standard-setting purposes. Likewise, no
convincing association has been demonstrated between daily mortality and daily
oxidant concentrations; rather, the effect correlates most closely with elevated
temperature.
Studies on the irritative effects of 0- have been complicated by the
presence of other photochemical pollutants and their precursors in the ambient
environment and by the lack of adequate control for other pollutants, meteoro-
logical variables, and non-environmental factors in the analysis. Although 0,,
does not cause the eye irritation normally associated with smog, several
studies in the Los Angeles basin have indicated that eye irritation is likely
to occur in ambient air when oxidant levels are about 0.10 ppm. Qualitative
associations between oxidant levels in the ambient air and symptoms such as
eye and throat irritation, chest discomfort, cough, and headache have been
reported at >0.10 ppm in both children and young adults (Hammer et a!., 1974;
Makino and Mizoguchi, 1975). Discomfort caused by irritative symptoms may be
responsible for the impairment of athletic performance reported in high school
students during cross-country track meets in Los Angeles (Wayne et a!., 1967;
Herman, 1972) and is consistent with the evidence from field studies (Section
11.2.1) and from controlled human exposure studies (Section 10.4) indicating
that exercise performance may be limited by exposure to 0.,. Although several
additional studies have shown respiratory irritation apparently related to
exposure to ambient 0~ or oxidants in community populations, none of these
epidemiologies! studies provide satisfactory quantitative data on acute
respiratory illnesses.
Epidemiological studies in children and young adults suggest an association
of decreased peak flow and increased airway resistance with acute ambient air
exposures to daily maximum 1-hr 03 concentrations ranging from 20 to 274 ug/m
(0.01 to 0.14 ppm) over the entire study period ( Lippmann et al., 1983;
Lebowitz et al., 1982, 1983, 1985; Lebowitz, 1984; Bock et al., 1985; Lioy et
al., 1985). None of these studies by themselves can provide satisfactory
1-154
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quantitative data on acute effects of CL because of methodological problems
along with the confounding influence of other pollutants and environmental
conditions in the ambient air. The aggregation of individual studies, however,
provides reasonably good evidence for an association between ambient 0-
exposure and acute pulmonary function effects. This association is strengthened
by the consistency between the findings from the epidemiological studies and
the results from the field studies in exercising adolescents (Avol et a!.,
1985a,b) which have shown small decreases in forced expiratory volume and flow
o
at 282 ug/m (0.144 ppm) of 03 in the ambient air; and with the results from
the controlled human exposure studies in exercising children which have shown
3
small decrements in forced expiratory volume at 235 pg/rn (0.12 ppm) of 0,
(Section 10.2.9.2).
In studies of exacerbation of asthma and chronic lung diseases, the major
problems have been the lack of information on the possible effects of medica-
tions, the absence of records for all days on which symptoms could have oc-
curred, and the possible concurrence of symptomatic attacks resulting from the
presence of other environmental conditions in ambient air. For example,
Whittemore and Korn (1980) and Holguin et al. (1985) found small increases in
the probability of asthma attacks associated with previous attacks, decreased
temperature, and with incremental increases in oxidant and 0, concentrations,
respectively. Lebowitz et al. (1982, 1983, 1985) and Lebowitz (1984) showed
effects in asthmatics, such as decreased peak expiratory flow and increased
respiratory symptoms, that were related to the interaction of 03 and tempera-
ture. All of these studies have questionable effects from other pollutants,
particularly inhalable particles. There have been no consistent findings of
symptom aggravation or changes in lung function in patients with chronic lung
diseases other than asthma.
Only a few prospective studies have been reported on morbidity, mortality,
and chromosomal effects from chronic exposure to 03 or other photochemical
oxidants. The lack of quantitative measures of oxidant exposures seriously
limits the usefulness of many population studies of morbidity and mortality
for standards-setting purposes. .Most of these long-term studies have employed
average annual levels of photochemical oxidants or have involved broad ranges
of pollutants; others have used a simple high-oxidant/low-oxidant dichotomy.
In addition, these population studies are also limited by their inability to
control for the effects of other factors that can potentially contribute to
1-155
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the development and progression of respiratory disease over long periods of
time. Thus, insufficient information is available in the epidemiological
literature on possible exposure-response relationships between ambient 0,, or
other photochemical oxidants and the prevalence of chronic lung disease or the
rates of chronic disease mortality. None of the epidemiological studies
investigating chromosomal changes have found any evidence that ambient 03 or
oxidants affect the peripheral lymphocytes of the exposed population.
1.11 EVALUATION OF HEALTH EFFECTS DATA FOR OZONE AND OTHER PHOTOCHEMICAL
OXIDANTS
1.11.1 Health Effects in the General Human Population
Controlled human studies of at-rest exposures to 03 lasting 2 to 4 hr
have demonstrated decrements in forced expiratory volume and flow occurring at
and above 0.5 ppm of 03 (Chapter 10). Airway resistance was not significantly
changed at these Q3 concentrations. Breathing 03 at rest at concentrations
< 0.5 ppm did not significantly impair pulmonary function although the occur-
rence of some 03-related pulmonary symptoms has been suggested in a number of
studies.
One of the principal modifiers of the magnitude of response to 03 is
minute ventilation (VV), which increases proportionately with increases in
exercise work load. Adjustment by the respiratory system to an increased work
load is characterized by increased frequency and depth of breathing. Consequent
increases in VV not only increase the overall volume of inhaled pollutant, but
the increased tidal volume may lead to a higher concentration of ozone in the
lung regions most sensitive to ozone. These processes are further enhanced at
high work loads (VV > 35 L/min), since the mode of breathing changes at that
VV from nasal to oronasal.
Statistically significant decrements in forced expiratory volume and flow
&
are generally observed in healthy adult subjects (18 to 45 yr old) after 1 to
3 hr of exposure as a function of the level of exercise performed and the
ozone concentration inhaled during the exposure. Group mean data pooled from
numerous controlled human exposure (Chapter 10) and field (Chapter 11) studies
indicate that, on average, pulmonary function decrements occur:
1. At >0.37 ppm QS with light exercise (VV < 23 L/min);
2. At >0.30 ppm Q3 with moderate exercise (V£ = 24-43 L/min);
1-156
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3. At >0.24 ppm 03 with heavy, exercise (VE = 44-63 L/min); and
4. At >0.18 ppm On with very heavy exercise (VV > 64 L/min).
Note, however, that data from specific individual studies indicate that pulmo-
nary function decrements occur with very heavy exercise in healthy adults at
0.15 to 0.16 ppm Q3 (Avol et a!., 1984) and suggest that such effects may
occur in healthy adults at levels as low as 0.12 ppm 0., (McDonnell et a"L,
1983). Also, pulmonary function decrements have been observed in children and
adolescents at concentrations of 0.12 and 0.14 ppm 03 with heavy exercise
(McDonnell et a!., 1985b; Avol et a!., 1985a). At the lower concentrations
(0.12 to 0.15 ppm), the average changes in lung function are generally small
(<5 percent) and are a matter of controversy in regard to their medical signi-
ficance.
In the majority of the studies reported, 15-min intermittent exercise
alternated with 15-min rest was employed for the duration of the exposure.
Figure 1-14 uses the pulmonary function measurement FEV-, to illustrate the
effects of intermittent exercise and 0- concentration during 2-hr exposures.
As noted above, larger decrements in lung function occur at higher exercise
levels and at higher 03 concentrations. The maximum reponse to 0™ exposure
can be observed within 5 to 10 min following the end of each exercise period.
Other measures of spirometric pulmonary function (e.g., FVC and FEF2c_75^) are
consistent with FEV., and, therefore, are not depicted here. It is important
to note, however, that any predictions of average pulmonary function responses
to 03 only apply under the specific set of exposure conditions at which these
data were derived.
Continuous exercise equivalent in duration to the sum of intermittent
exercise periods at comparable ozone concentrations (0.2 to 0.4 ppm) and
.minute ventilation (60 to 80 L/min) seems to elicit greater changes in pulmonary
function (Folinsbee et a!., 1984; Avol et a!., 1984, 1985c) but the differences
between intermittent and continuous exercise are not clearly established.
More experimental data are needed to make any quantitative evaluation of the
differences in effects induced by these two modes of exercise.
Functional recovery, at least from a single exposure with exercise,
appears to progress in two phases: during the initial rapid phase, lasting
between 1 and 3 hr, pulmonary function improves more than 50 percent; this is
followed by a much slower recovery that is usually completed in most subjects
1-157
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110
8 100
a
a
UJ
S
D
90
en
oo
tc.
o
0.
X
UJ
Q
UJ
O
tc.
O
u.
O
UJ
U)
80
70
60
VERY HEAVY
EXERCISE
••-. LIGHT EXERCISE
MODERATE
EXERC|SE
0.2 0.4
OZONE CONCENTRATION, ppm
0.6
0.8
Figure 1-14. Group mean decrements in 1 -sec forced expiratory volume during 2-hr ozone
exposures with different levels of intermittent exercise: light (V"E < 23 L/min); moderate
(VE = 24-43 L/min); heavy (Vg = 44-63 L/min); and very heavy (V"E > 64 L/min).
(Concentration-response curves are taken from Figures 12-2 through 12-5 in Chapter 12,
Volume V.)
-------
within 24 hr. In some individuals, an enhanced responsiveness to a second 0~
challenge may persist for up to 48 hr (Bedi et al., 1985; Folinsbee and Horvath,
1986). In addition, despite apparent functional recovery, other regulatory
systems may still exhibit abnormal responses when stimulated; e.g., airway
hyperreactivity may persist for days.
Group mean changes may be useful for making statistical inferences about
homogeneous populations, but they are not adequate for describing difference
in responsiveness to 0- among individuals. Even in well-controlled experiments
on an apparently homogeneous group of healthy subjects, physiological responses
to the same work and pollutant loads will vary widely among individuals (Horvath
et al., 1981; Gliner et al., 1983; McDonnell et al., 1983; Kulle et al., 1985).
Despite large intersubject variability, individual responsiveness to a given
DO concentration is quite reproducible (Gliner et al., 1983; McDonnell et al.,
1985a). Some individuals, therefore, are consistently more responsive to 0™
than are others. The term "responders" has been used to describe the 5 to 20
percent of the studied population that is most responsive to CU exposure.
There are no clearly established criteria to define this group of subjects.
Likewise, there are no known specific factors responsible for increased or
decreased responsiveness to 0.,. Characterization of individual responses to
03, however, is pertinent since it permits the assessment of a segment of the
general population that is potentially at-risk to 0- exposure (see Section
12.7.3) although statistical treatment of these data is still rudimentary and
their validity is open to question.
A close association has been observed between the occurrence of respiratory
symptoms and changes in pulmonary function in adults acutely exposed in environ-
mental chambers to 03 (Chapter 10) or to ambient air containing 03 as the
predominant pollutant (Chapter 11), This association holds for both the
time-course and magnitude of effects. Studies on children and adolescents
exposed to 0, or ambient air containing 03 under similar conditions have found
no significant increases in symptoms despite significant changes in pulmonary
function (Avol et al., 1985a,b; McDonnell et al., 1985b,c). Epidemiological
studies of exposure to ambient photochemical pollution are of limited use for
quantifying exposure-response relationships for 03 because they have not
adequately controlled for other pollutants, meteorological variables, and
non-environmental factors in the data analysis. Eye irritation, for example,
one of the most common complaints associated with photochemical pollution, is
1-159
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not characteristic of clinical exposures to 03, even at concentrations several
times higher than any likely to be encountered in ambient air. There is
limited qualitative evidence to suggest that at low concentrations of CL,
other respiratory and nonrespiratory symptoms, as well, are more likely to
occur in populations exposed to ambient air pollution than in subjects exposed
in chamber studies (Chapter 11).
Discomfort caused by irritative symptoms may be responsible for the
impairment of athletic performance reported in high school students during
cross-country track meets in Los Angeles (Chapter 11). Only a few control!ed-
exposure studies, however, have been designed to examine the effects of 03 on
exercise performance (Chapter 10). In one study, light intermittent exercise
(Vp = 20-25 L/min) at a high 03 concentration (0.75 ppm) reduced postexposure
maximal exercise capacity by limiting maximal oxygen consumption; submaximal
oxygen consumption changes were not significant. The extent of ventilatory
and respiratory metabolic changes observed during or following the exposure
appears to have been related to the magnitude of pulmonary function impairment.
Whether such changes are consequent to respiratory discomfort (i.e., symptomatic
effects) or are the result of changes in lung mechanics or both is still
unclear and needs to be elucidated.
Environmental conditions such as heat and relative humidity may alter
subjective symptoms and physiological impairment associated with 0~ exposure.
Modification of the effects of 03 by these factors may be attributed to in-
creased ventilation associated with elevated body temperature but there may
also be an independent effect of elevated body temperature on pulmonary function
(e.g., VC).
Numerous additional factors have the potential for altering responsiveness
to ozone. For example, children and older individuals may be more responsive
than young adults. Other factors such as gender differences (at any age),
personal habits such as smoking, nutritional deficiencies, or differences in
imtnunologic status may predispose individuals to susceptibility to ozone. In
addition, social, cultural, or economic factors may be involved. Those actually
known to alter sensitivity, however, are few, largely because they have not
been examined adequately to determine definitively their effects on sensitivity
to 03- The following briefly summarizes what is actually known from the data
regarding the importance of these factors (see Section 12.3.3 for details):
1-160
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1. Age. Although changes in growth and development of the lung with
age have been postulated as one of many factors capable of modifying responsive-
ness to 0~, sufficient numbers of studies have not been performed to provide
any sound conclusions for effects of different age groups on responsiveness to
°3'
2. Sex. Sex differences in responsiveness to ozone have not been
adequately studied. Lung function of women, as assessed by changes in FEV-, QJ
might be affected more than that of men under similar exercise and exposure
conditions, but the possible differences have not been tested systematically.
3. Smoking Status. Differences between smokers and nonsmokers have
been studied often, but the smoking histories of subjects are not documented
well. There is some evidence, however, to suggest that smokers may be less
responsive to 0™ than nonsmokers.
4, Nutritional Status. Antioxidant properties of vitamin E in preventing
ozone-initiated peroxidation J_n vitro are well demonstrated and their protective
effects jji vivo are clearly demonstrated in rats and mice. No evidence indi-
cates, however, that man would benefit from increased vitamin E intake relative
to ambient ozone exposures.
5. Red Blood Cell EnzymeDeficiencies. There have been too few studies
performed to document reliably that individuals with a hereditary deficiency
of glucose-6-phosphate dehydrogenase may be at-risk to significant hematolog-
ical effects from 03 exposure. Even if 0*3 or a reactive product of (k-tissue
interaction were to penetrate the red blood cell after j_n vivo exposure, it is
unlikely that any depletion of glutathione or other reducing compounds would
be of functional significance for the affected individual.
Successive daily brief exposures of healthy human subjects to 0, (<0.7 ppm
for approximately 2 hr) induce a typical temporal pattern of response (Chap-
ter 10, Section 10.3). Maximum functional changes that occur after the first
or second exposure day become progressively attenuated on each of the subsequent
days. By the fourth day of exposure, the average effects are not different
from those observed following control (air) exposure. Individuals need between
3 and 7 days of exposure to develop full attenuation, with more sensitive
subjects requiring more time. The magnitude of a peak response to 03 appears
to be directly related to 03 concentration. It is not known how variations in
the length or frequency of exposure will modify the time course of this altered
1-161
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responsiveness. In addition, concentrations of 0, that have no detectable
effect appear not to invoke changes in response to subsequent exposures at
higher 0, concentrations. Full attenuation, even in ozone-sensitive subjects,
does not persist for more than 3 to 7 days after exposure in most individuals,
while partial attenuation might persist for up to 2 weeks. Although the
severity of symptoms is generally related to the magnitude of the functional
response, partial attenuation of symptoms appears to persist longer, for up to
4 weeks after exposure.
Whether populations exposed to photochemical air pollution develop at
least partial attenuation is unknown. No epidemiological studies have been
designed to test this hypothesis and additional information is required from
controlled laboratory studies before any sound conclusions can be made.
Ozone toxicity, in both humans and laboratory animals, may be mitigated
through altered responses at the cellular and/or subcellular level. At present,
the mechanisms underlying altered responses are unclear and the effectiveness
of such mitigating factors in protecting the long-term health of the individual
against ozone is still uncertain. A growing body of experimental evidence
suggests the involvement of vagal sensory receptors in modulating the acute
responsiveness to ozone. It is highly probable that most of the decrements in
lung volume reported to result from exposures of greatest relevance to standard-
setting (<0.3 ppm On) are caused by the inhibition of maximal inspiration
rather than by changes in airway diameter. None of the experimental evidence,
however, is definitive and additional research is needed to elucidate the
precise mechanism(s) associated with ozone exposure.
1.11.2 Health Effects in Individuals with Preexisting Disease
Currently available evidence indicates that individuals with preexisting
disease respond to 03 exposure to a similar degree as normal, healthy subjects.
Patients with chronic obstructive lung disease and/or asthma have not shown
increased responsiveness to 03 in controlled human exposure studies, but there
is some indication from epidemiological studies that asthmatics may be sympto-
matically and possibly functionally more responsive than healthy individuals
to ambient air exposures. Appropriate inclusion and exclusion criteria for
selection of these subjects, however, especially better clinical diagnoses
validated by pulmonary function, must be considered before their responsiveness
to 0- can be adequately determined. None of these factors has been sufficiently
studied in relation to 0-, exposures to give definitive answers.
1-162
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1.11.3 Extrapolation of Effects Observed In Animals to Human Populations
Animal experiments on a variety of species have demonstrated increased
susceptibility to bacterial respiratory infections following 0™ exposure.
Thus, it could be hypothesized that humans exposed to CU could experience
decrements in their host defenses against infection. At the present time,
however, these effects have not been studied in humans exposed to 03.
Animal studies have also reported a number of extrapulmonary responses to
0,, including cardiovascular, reproductive, and teratological effects, along
with changes in endocrine and metabolic function. The implications of these
findings for human health are difficult to judge at the present time. In
addition, central nervous system effects, alterations in red blood cell
morphology and enzymatic activity, as well as cytogenetic effects on circulating
lymphocytes, have been observed in laboratory .animals following exposure to
0_. While similar effects have been described in circulating cells from human
subjects exposed to high concentrations of 0,,, the results were either incon-
sistent or of questionable physiological significance (Section 12.3.8). It is
not known, therefore, if extrapulmonary responses would be likely to occur in
humans when exposure schedules are used that are representative of exposures
that the population at large might actually experience.
Despite wide variations in study techniques and experimental designs,
acute and subchronic exposures of animals to levels of ozone < 0.5 pptn produce
remarkably similar types of responses in all species examined. A characteristic
ozone lesion occurs at the junction of the conducting airways and the gas-
exchange regions of the lung after acute Q3 exposure. Dosimetry model simula-
tions predict that the maximal tissue dose of 03 occurs in this region of the
lung. Continuation of the inflammatory process during longer 03 exposures is
especially important since it appears to be correlated with increased airway
resistance, increased lung collagen content, and remodeling of the centriacinar
airways, suggesting the development of distal airway narrowing. No convincing
evidence of emphysema in animals chronically exposed to 03 has yet been pub-
lished, but centriacinar inflammation has been shown to occur.
Since substantial animal data exist for 03~induced changes in lung struc-
ture and function, biochemistry, and host defenses, it is conceivable that man
may experience more types of effects from exposure to ozone than have been
established in human clinical studies. It is important to note, however, that
the risks to man from breathing ambient levels of ozone cannot fully be
determined until quantitative extrapolations of animal results can be made.
1-163
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1.11.4 Health Effects of Other Photochemical Oxldants and PollutantMixtures
Controlled human studies have not consistently demonstrated any modifica-
tion of respiratory effects for combined exposures of 03 with S02, NO-, CO, or
HgSO. and other particulate aerosols. Ozone alone is considered to be respon-
sible for the observed effects of those combinations or of multiple mixtures
of these pollutants. Combined exposure studies in laboratory animals have
produced varied results, depending upon the pollutant combination evaluated,
the exposure design, and the measured variables (Section 12.6.3). Thus, no
definitive conclusions can be drawn from animal studies of pollutant interac-
tions. There have been far too few controlled toxicological studies with
other oxidants, such as peroxyacetyl nitrate or hydrogen peroxide, to permit a
sound scientific evaluation of their contribution to the toxic action of
photochemical oxidant mixtures. There is still some concern, however, that
combinations of oxidant pollutants with other pollutants may contribute to the
symptom aggravation and decreased lung function described in epidemiological
studies on individuals with asthma and in children and young adults. For this
reason, the effects of interaction between inhaled oxidant gases and other
environmental pollutants on the lung need to be systematically studied using
exposure regimens that are more closely representative of ambient air ratios
of peak concentrations, frequency, duration, and time intervals between events.
1.11.5 Identificationof Potentially At-Risk Groups
Despite uncertainties that may exist in the data, it is possible to
identify the groups that may be at potential risk from exposure to ozone,
based on known health effects, activity patterns, personal habits, and actual
or potential exposures to ozone.
The first group that appears to be at potential risk from exposure to
ozone is that group of the general population characterized as having pre-
existing respiratory disease. Available data on actual differences in
responsiveness between these and healthy members of the general population
indicate that under the exposure conditions studied to date, individuals with
preexisting disease are as responsive to ozone as healthy individuals. Neverthe-
less, two primary considerations place individuals with preexisting respiratory
disease among groups at potential risk from exposure to ozone. First, it must
be noted that concern with triggering untoward reactions has necessitated the
use of low concentrations and low exercise levels in most studies on subjects
1-164
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with mild, but not severe, preexisting disease. Therefore, few or no data on
responses at higher concentrations, at higher exercise levels, and in subjects
with more severe disease states are available for comparison with responses in
healthy subjects. Thus, definitive data on the modification by preexisting
disease of responses to ozone are not available. Second, however, it must be
emphasized that in individuals with already compromised pulmonary function,
the decrements in function produced by exposure to ozone, while similar to or
even the same as those experienced by normal subjects, represent a further
decline in volumes and flows that are already diminished. It is possible that
such declines may impair further the ability to perform normal activities. In
individuals with preexisting diseases such as asthma or allergies, increases
in symptoms upon exposure to ozone, above and beyond symptoms seen in the
general population, may also impair or further curtail the ability to function
normally.
The second group at potential special risk from exposure to ozone consists
of the general population of normal, healthy individuals. Two specific factors
place members of the general population at potential risk from exposure to
ozone. First unusual responsiveness to ozone has been observed in some individ-
uals ("responders"), not yet characterized medically except by their response
to ozone, who experience greater decrements in lung function from exposure to
ozone than the average response of the groups studied. It is not known if
"responders" are a specific population subgroup or simply represent the upper
5 to 20 percent of the ozone response distribution. As yet no means of deter-
mining in advance those members of the general population who are "responders"
has been devised. Second, data presented in this chapter underscore the
importance of exercise in the potentiation of effects from exposure to ozone.
Thus, the general population potentially at risk from exposure to ozone includes
those individuals whose activities out of doors, whether vocational or
avocational, result in increases in minute ventilation, which is the most
prominent modifier of response to ozone.
Other biological and nonbiological factors have the potential for influenc-
ing responses to ozone. Data remain inconclusive at the present, however,
regarding the importance of age, gender, and other factors in influencing
response to ozone. Thus, at the present time, no other groups are thought to
be biologically predisposed to increased sensitivity to ozone. It must be
emphasized, however, that the final identification of those effects that are
considered "adverse" and the final identification of "at-risk" groups are both
the domain of the Administrator of the U.S. Environmental Protection Agency.
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1.12 REFERENCES
1,12.1 References for Introduction
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1.12.2 References for Properties, Chemistry, and Transport of Ozoneand
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Atkinson, R,; Aschmann, S. M. (1984) Rate constants for the reactions of 03
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Atkinson, R.; Carter, W. P. L. (1984) Kinetics and mechanisms of the reactions
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References for Properties, Chemistry, Transport (cont'd.)
Atkinson, R.; Plum, C. N.; Carter, W. P. L.; Winer, A. M.; Pitts, J. N., Jr.
(1984a) Rate constants for the gas-phase reactions of N03 radicals with .
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Atkinson, R.; Pitts, J. N., Jr.;
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1584-1587.
Aschmann, S. M. (1984b)
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Tropospheric re-
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Atkinson, R.; Aschmann, S. M.; Winer, A. M.; Pitts, J. N. , Jr. (1984c)
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Fehsenfeld, F.
Proceedings of
May; Williams-
Calvert, J. G.; Stockwell, W. R. (1983) Acid generation in the troposphere by
gas-phase chemistry. Environ. Sci. Techno!. 17: 428A-443A.
Calvert, J. G.; Stockwell, W. R. (1984) The mechanism and rates of the gas-phase
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Carter, W. P. L. ; Winer, A. M.; Darnall, K. R. ;
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Pitts, J. N.,
photochemical
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Carter, W. P. L.; Winer, A. M.; Pitts, J. N., Jr. (1981a) Major atmospheric
sink for phenol and the cresols: reaction with the nitrate radical.
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Carter, W. P. L.; Winer, A. M.; Pitts, J. N., Jr. (1981b) Effect of peroxyacetyl
nitrate on the initiation of photochemical smog. Environ. Sci. Techno!.
15: 831-834.
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Zaitsu, K.; Okhura, Y. (1980) New fluorogenic substrates for horseradish pero-
xidase: rapid and sensitive assays for hydrogen peroxide and peroxidase.
Anal. Biochem. 109: 109-113. . •
Zika, R. G.; Saltzman, E. S. (1982) Interaction of ozone and hydrogen peroxide
in water: implications for analysis of H202 -in air. Geophys. Res. Lett.
9: 231-234. • ' -.. v..' , : . i '• •
1.12.4 References for Concentrations of Ozone and Other Photochemical Oxidants
in Ambient Air - . . . , ,•--'., -,-••- ' •.•"-,..
Altshuller, A. P. (1983) Measurements of the products of atmospheric photo-
chemical reactions in laboratory studies and in ambient air—relation-
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1-187
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References for Ambient Air Concentrations (cont'd.)
Berk, J. V.; Young, R. A.; Brown, S. R.; Hollowell, C. D. (1981) Impact of
energy-conserving retrofits on indoor air quality in residential housing.
Presented at: 74th annual meeting of the Air Pollution Control Assoc-
ciation; June; Philadelphia, PA. Pittsburgh, PA: Air Pollution Control
Association; paper no. 81-22.1
Cleveland, W. S.; Graedel, T. E.; Kleiner, B. (1976a) Photochemical air pollu-
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Cleveland, W. S.; Guarino, R.; Kleiner, B.; McRae, J. E.; Warner, J. L. (1976b)
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environment. Pittsburgh, PA: Air Pollution Control Association; pp.
109-120.
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Contant, C. F.; Gehan, B. M.; Stock, T. H.; Holguin, A. H.; Buffler, P. A.
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Davies, T, D.; Ramer, B.; Kaspyzok, G.; Delany, A. C. (1984) Indoor/outdoor
ozone concentrations at a contemporary art gallery. J. Air Pollut. Control
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Evans, G. F. (1985) The National Air Pollution Background Network: final project
report. Research Triangle Park, NC: U.S. Environmental Protection Agency,
Office of Research and Development; EPA report no. EPA-600/4-85-038.
Available from: NTIS, Springfield, VA; PB85-212413.
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References for Ambient Air Concentrations (cont'd.')
Grosjean, D. (1981) Critica-1 evaluation and comparison of measurement methods
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synthesis in the troposphere. JGR J. Geophys. Res. 87: 3045-3051.
Jorgen, R. T.; Meyer, R. A.; Hughes, R. A. (1978) Routine peroxyacetyl nitrate
(PAN) monitoring applied to the Houston Area Oxidant Study. Presented at:
71st annual meeting of the Air Pollution Control Association; June;
Houston, TX. Pittsburgh, PA: Air Pollution Control Association; paper no.
78-50.1.
Kelly, N. A.; Ferman, M. A.; Wolff, G. T. (1986) The chemical and meteorological
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Pollution Control Assoc.
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Ozone/Oxidants Standards, Houston, TX, November 28-30, 1984. Pittsburgh,
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ment in ambient atmospheres. Environ. Sci. Technol. 10: 374-380.
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1-189
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References for Ambient Air Concentrations (cont'd.)
Mayrsohn, H.; Brooks, C. (1965) The analysis of PAN by electron capture gas
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Chemical Society; November; Los Angeles, CA. Los Angeles, CA: California
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Oxidants, 1976 — Analysis of evidence and viewpoints. Part III. The issue
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Oxidants 1976 — Analysis of evidence and viewpoints. Part III. The issue
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rates, and removal of ozone and their relation to establishing clean
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SAROAD, Storage and Retrieval of Aerometric Data [data base]. (1985a) Data
file for 1976. Research Triangle Park, NC: U.S. Environmental Protection
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SAROAD, Storage and Retrieval of Aerometric Data [data base]. (1985b) Data
file for 1979. Research Triangle Park, NC: U.S. Environmental Protection
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SAROAD, Storage and Retrieval of Aerometric Data [data base]. (1985c) Data
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References for Ambient Air Concentrations (cont'd.)
SAROAD, Storage and Retrieval of Aerometric Data [data base]. (1985d) Data
file for 1981. Research Triangle Park, NC: U.S. Environmental Protection
Agency, Office of Air Quality Planning and Standards. Disc; ASCII.
SAROAD, Storage and Retrieval of Aerometric Data [data base]. (1985e) Data
file for 1982. Research Triangle Park, NC: U.S. Environmental Protection
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file for 1983. Research Triangle Park, NC: U.S. Environmental Protection
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Seller, W. ; Fishman, J. (1981) The distribution of carbon monoxide and ozone
in the free troposphere. JGR J. Geophys. Res. 86: 7255-7265.
Singh, H. B. ; Salas, L. J. ; Smith, A. J. ; Sh-igeishi, H. (1981) Measurements of
some potentially hazardous organic chemicals in urban environments.
Atmos. Environ. 15: 601-612.
Singh, H. B. , Salas, L. J. ; Stiles, R.; Shigeishi, H. (1982) Measurements of
hazardous organic chemicals in the ambient atmosphere. Research Triangle
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NTIS, Springfield, VA; PB83-156935.
Smith, W. J. (1981) New York State air monitoring data report for the Northeast
Corridor Regional Modeling Project (NECRMP). Albany, NY:. New York State
Department of Environment and Conservation.
Spicer, C. W.; Gemma, J. L; Joseph, D. W.; Sticksel, P. R.; Ward, G. F. (1976)
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Columbus Laboratories; EPA report no. EPA-600/3-76-018a. Available from:
NTIS, Springfield, VA; PB-253736.
Stock, T. H. ; Holguin, A. H.; Selwyn, B. J.; Hsi, B. P.; Contant, C. F.;
Buffler, P. A.; Kotchmar, D. J. (1983). Exposure estimates for the Houston
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Temple, P. J.; Taylor, 0. C. (1983) World-wide ambient measurements of peroxy-
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17: 1583-1587.
Thompson, C. R. ; Hensel, E. G. ; Kats, G. (1973) Outdoor-indoor levels of six
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1-191
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References for Ambient Air Concentrations (cont'd.)
Tuazon, E. C. ; Winer, A. M. ; Graham, R. A.; Pitts, J. N. , Jr. (1981a) Atmos-
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U.S. Department of Commerce, Bureau of the Census (1982) Statistical Abstract
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U.S. Environmental Protection Agency (1980) Air quality data--1979 annual
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Westberg, H.; Allwine, K.; Robinson, E, (1978) Measurement of light hydrocar-
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1.12.5 References for Effects of Ozone and Other Photochemical Oxidantson
Vegetation
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References for Vegetation Effects (cont'd.)
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•*
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1-194
-------
References for Vegetation Effects (cont'd.)
Davis, D. D. (1977) Response of ponderosa pine primary needles to separate and
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References for Vegetation Effects (cont'd.)
Heagle, A. S.; Johnston, J. W. (1979) Variable responses of soybean to mixtures
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Heagle, A. S.; Cure, W. W.; Rawlings, J. 0. (1985) Response of turnips to
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of dose and injury to plants. Science (Washington, DC) 151: 511-515.
Heck, W. W.; Taylor, 0. C.; Adams, R.; Bingham, G.; Miller, J.; Preston, E.;
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Control Assoc. 32: 353-361.
Heck, W. W.; Adams, R. M.; Cure, W. W.; Heagle, A. S.; Heggestad, H. E.;
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reassessment of crop loss from ozone. Environ. Sci. Techno!.
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Heck, W. W.; Taylor, 0. C.; Adams, R. M.; Miller, J. E.; Weinstein, L (1983b)
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Laboratory, Con/all is, OR.
1-196
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References for Vegetation Effects (cont'd.)
Heck, W. W.; Cure, W. W.; Rawlings, J. 0.; Zaragoza, L. J.; Heagle, A. S.;
Heggestad, H. E.; Kohut, R. J.; Kress, L. W.; Temple, P. J.; (1984)
Assessing impacts of ozone on agricultural crops: II. Crop yield functions
and alternative exposure statistics. J. Air Pollut. Control Assoc. 34:
810-817.
Heggestad, H. E. (1973) Photochemical air pollution injury to potatoes in the
Atlantic coastal states. Am. Potato J. 50: 315-328.
Heggestad, H. E.; Bennett, J. H. (1981) Photochemical oxidants potentiate
yield losses in snap beans attributable to sulfur dioxide. Science
(Washington, DC) 213: 1008-1010.
Henderson, W, R.; Reinert, R. A. (1979) Yield response of four fresh market
tomato cultivars after acute ozone exposure in the seedling stage. J.
Am. Soc. Hortic. Sci. 104: 754-759.
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1-12.7. References for Effectsof Ozone and Other Photochemical Oxidants on
Nonbiological Materials
Beloin, N. J. (1972) Fading of dyed fabrics by air pollution: a field study.
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References for Nonbiologlcal Materials (cont'd.)
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References for Toxlcological Effects (cont'd.)
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Kulle, T. J.; Milman, J. H.; Sauder, L. R.; Kerr, H. D.; Parrel!, B. P.;
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References forControlled Human Studies (cont'd.)
Linn, W. S.; Medway, D. A.; Anzar, U. T.; Valencia, L. M.; Spier, C. E.; Tsao,
F. S-0.; Fischer, D. A.; Hackney, J, D. (1982b) Persistence of adaptation
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chronic obstructive pulmonary disease. Arch. Environ. Health 38: 278-283.
McDonnell, W. F.; Horstmann, D. H.; Hazucha, M. J.; Seal, E., Jr.; Haak, E.
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1.12.10 References forField andEpidemiologicalStudies of the Effects of
Ozone and Other Photochemical Oxidants
Avol, E. L.; Linn, W. S. ; Shamoo, D. A.; Venet, T. G. ; Hackney, J. D. (1983)
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Comparative respiratory effects of ozone and ambient oxidant pollution
exposure during heavy exercise. J. Air Pollut. Control Assoc. 34: 804-809.
5
1-233
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References for Field and Epidemiological Studies (cont'd.)
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Hackney, J. D. (1985b) Short-term health effects of ambient air pollution
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References for EvaluationofHealthEffects Data (cont'd.)
McDonnell, W. F.; Chapman, R. S.; Horstman, D. H.; Leigh, M. W.; Abdul-Salaam,
S. (1985c) A comparison of the responses of children and adults to
acute ozone exposure. In: Lee, S. D,, ed. Evaluation of the scientific
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(APCA international specialty conference transactions: TR-4).
1-237
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