United States 

Environmental Protection February 2006 

Agency EPA EPA 600/R-05/004cF 

Air Quality Criteria for 

Ozone and Related 

Photochemical Oxidants 

Volume III of III
EPA 600/R-05/004cF 

February 2006 

Air Quality Criteria for Ozone and Related 

Photochemical Oxidants 

Volume III 

National Center for Environmental Assessment-RTP Office 

Office of Research and Development 

U.S. Environmental Protection Agency 

Research Triangle Park, NC 
DISCLAIMER 

This document has been reviewed in accordance with U.S. Environmental Protection 

Agency policy and approved for publication. Mention of trade names or commercial products 

does not constitute endorsement or recommendation for use. 

PREFACE 

National Ambient Air Quality Standards (NAAQS) are promulgated by the United States 

Environmental Protection Agency (EPA) to meet requirements set forth in Sections 108 and 109 

of the U.S. Clean Air Act (CAA). Sections 108 and 109 require the EPA Administrator 

(1) to list widespread air pollutants that reasonably may be expected to endanger public health 

or welfare; (2) to issue air quality criteria for them that assess the latest available scientific 

information on nature and effects of ambient exposure to them; (3) to set “primary” NAAQS to 

protect human health with adequate margin of safety and to set “secondary” NAAQS to protect 

against welfare effects (e.g., effects on vegetation, ecosystems, visibility, climate, manmade 

materials, etc); and (5) to periodically review and revise, as appropriate, the criteria and NAAQS 

for a given listed pollutant or class of pollutants. 

In 1971, the U.S. Environmental Protection Agency (EPA) promulgated National Ambient 

Air Quality Standards (NAAQS) to protect the public health and welfare from adverse effects of 

photochemical oxidants. The EPA promulgates the NAAQS on the basis of scientific 

information contained in air quality criteria issued under Section 108 of the Clean Air Act. 

Following the review of criteria as contained in the EPA document, Air Quality Criteria for 

Ozone and Other Photochemical Oxidants published in 1978, the chemical designation of the 

standards was changed from photochemical oxidants to ozone (O3) in 1979 and a 1-hour O3 

NAAQS was set. The 1978 document focused mainly on the air quality criteria for O3 and, to a 

lesser extent, on those for other photochemical oxidants (e.g., hydrogen peroxide and the 

peroxyacyl nitrates), as have subsequent revised versions of the document. 

To meet Clean Air Act requirements noted above for periodic review of criteria and 

NAAQS, the O3 criteria document, Air Quality Criteria for Ozone and Other Photochemical 

Oxidants, was next revised and released in August 1986; and a supplement, Summary of Selected 

III-ii 
New Information on Effects of Ozone on Health and Vegetation, was issued in January 1992. 

These documents were the basis for a March 1993 decision by EPA that revision of the existing 

1-h NAAQS for O3 was not appropriate at that time. That decision, however, did not take into 

account newer scientific data that had become available after completion of the 1986 criteria 

document. Such literature was assessed in the next periodic revision of the O3 air quality criteria 

document (O3 AQCD) which has completed in 1996 and provided scientific bases supporting the 

setting by EPA in 1997 of the current 8-h O3 NAAQS. 

The purpose of this revised air quality criteria document for O3 and related photochemical 

oxidants is to critically evaluate and assess the latest scientific information published since that 

assessed in the above 1996 O3 AQCD, with the main focus being on pertinent new information 

useful in evaluating health and environmental effects data associated with ambient air O3 

exposures. However, other scientific data are also discussed in order to provide a better 

understanding of the nature, sources, distribution, measurement, and concentrations of O3 and 

related photochemical oxidants and their precursors in the environment. The document mainly 

assesses pertinent literature published through 2004, but also includes assessment of a few 

additional important studies published or accepted for publication in 2005. 

A First External Review Draft of this O3 AQCD (dated January 2005) was released for 

public comment and was reviewed by the Clean Air Scientific Advisory Committee (CASAC) 

in May, 2005 to obtain. Public comments and CASAC recommendations were then taken into 

account in making revisions to the document for incorporation into a Second External Review 

Draft (dated August, 2005), which underwent further public comment and CASAC review at a 

December, 2005 public meeting. Public comments and CASAC advice derived from review of 

that Second External Review Draft were considered in making revisions incorporated into this 

final version of the document (dated February, 2006). Evaluations contained in the present 

document will be drawn on to provide inputs to associated O3 Staff Paper analyses prepared by 

EPA’s Office of Air Quality Planning and Standards (OAQPS) to pose options for consideration 

by the EPA Administrator with regard to proposal and, ultimately, promulgation of decisions on 

potential retention or revision, as appropriate, of the current O3 NAAQS. 

Preparation of this document was coordinated by staff of EPA’s National Center for 

Environmental Assessment in Research Triangle Park (NCEA-RTP). NCEA-RTP scientific 

staff, together with experts from other EPA/ORD laboratories and academia, contributed to 

III-iii 
writing of document chapters. Earlier drafts of document materials were reviewed by non-EPA 

experts in peer consultation workshops held by EPA. The document describes the nature, 

sources, distribution, measurement, and concentrations of O3 in outdoor (ambient) and indoor 

environments. It also evaluates the latest data on human exposures to ambient O3 and 

consequent health effects in exposed human populations, to support decision making regarding 

the primary, health-related O3 NAAQS. Lastly, the document also evaluates ambient O3 

environmental effects on vegetation and ecosystems, surface level solar UV radiation flux and 

global climate change, and man-made materials to support decision making on secondary 

O3 NAAQS. 

NCEA acknowledges the valuable contributions provided by authors, contributors, and 

reviewers and the diligence of its staff and contractors in the preparation of this document. 

III-iv 
Air Quality Criteria for Ozone and Related 

Photochemical Oxidants 

VOLUME I

Executive Summary .................................................... E-1

1. INTRODUCTION ................................................ 1-1

2. PHYSICS AND CHEMISTRY OF OZONE IN THE ATMOSPHERE ....... 2-1

3. ENVIRONMENTAL CONCENTRATIONS, PATTERNS, AND 

EXPOSURE ESTIMATES .......................................... 3-1

4. DOSIMETRY, SPECIES HOMOLOGY, SENSITIVITY, AND 

ANIMAL-TO-HUMAN EXTRAPOLATION ........................... 4-1

5. TOXICOLOGICAL EFFECTS OF OZONE AND RELATED

PHOTOCHEMICAL OXIDANTS IN LABORATORY ANIMALS 

AND IN VITRO TEST SYSTEMS ................................... 5-1

6. CONTROLLED HUMAN EXPOSURE STUDIES OF OZONE AND 

RELATED PHOTOCHEMICAL OXIDANTS .......................... 6-1

7. EPIDEMIOLOGICAL STUDIES OF HUMAN HEALTH EFFECTS

ASSOCIATED WITH AMBIENT OZONE EXPOSURE .................. 7-1

8. INTEGRATIVE SYNTHESIS: EXPOSURE AND HEALTH EFFECTS ..... 8-1

9. ENVIRONMENTAL EFFECTS: OZONE EFFECTS ON 

VEGETATION AND ECOSYSTEMS ................................ 9-1

10. TROPOSPHERIC OZONE EFFECTS ON UV-B FLUX AND 

CLIMATE CHANGE PROCESSES ................................. 10-1

11. OZONE EFFECTS ON MAN-MADE MATERIALS .................... 11-1

III-v 
Air Quality Criteria for Ozone and Related 

Photochemical Oxidants 

(cont’d) 

VOLUME II 

CHAPTER 2 ANNEX (ATMOSPHERIC PHYSICS/CHEMISTRY) . . . . . . . . . . AX2-1

CHAPTER 3 ANNEX (AIR QUALITY AND EXPOSURE) . . . . . . . . . . . . . . . . . AX3-1

CHAPTER 4 ANNEX (DOSIMETRY) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . AX4-1

CHAPTER 5 ANNEX (ANIMAL TOXICOLOGY) . . . . . . . . . . . . . . . . . . . . . . . AX5-1

CHAPTER 6 ANNEX (CONTROLLED HUMAN EXPOSURE) . . . . . . . . . . . . . AX6-1

CHAPTER 7 ANNEX (EPIDEMIOLOGY) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . AX7-1

VOLUME III 

CHAPTER 9 ANNEX (ENVIRONMENTAL EFFECTS) . . . . . . . . . . . . . . . . . . . AX9-1

III-vi 
Table of Contents 

Page 

Table of Contents . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . III-vii

List of Tables . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . III-xii

List of Figures . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . III-xiv

Authors, Contributors, and Reviewers . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . III-xvi

U.S. Environmental Protection Agency Project Team for Development of 

Air Quality Criteria for Ozone and Related Photochemical Oxidants ......... III-xix 

U.S. Environmental Protection Agency Science Advisory Board (SAB) 

Staff Office Clean Air Scientific Advisory Committee (CASAC) 

Ozone Review Panel . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . III-xxii 

Abbreviations and Acronyms ............................................... III-xxv

AX9. ENVIRONMENTAL EFFECTS: OZONE EFFECTS ON VEGETATION 

AND ECOSYSTEMS . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . AX9-1

AX9.1 METHODOLOGIES USED IN VEGETATION RESEARCH . . . . . . AX9-1

AX9.1.1 Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . AX9-1

AX9.1.2 Methods Involving Experimental Exposures 

to Ozone . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . AX9-2

AX9.1.2.1 “Indoor”, Controlled Environment, 

and Greenhouse Chambers . . . . . . . . . . AX9-2

AX9.1.2.2 Field Chambers . . . . . . . . . . . . . . . . . . . AX9-4

AX9.1.2.3 Plume Systems . . . . . . . . . . . . . . . . . . . . AX9-7

AX9.1.2.4 Comparative Studies . . . . . . . . . . . . . . . AX9-9

AX9.1.2.5 Ozone Generation Systems . . . . . . . . . AX9-12

AX9.1.2.6 Experimental Exposure Protocols . . . . AX9-12

AX9.1.3 Methods Involving Exposures to Ozone in 

Ambient Air . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . AX9-13

AX9.1.3.1 Air-Exclusion Systems . . . . . . . . . . . . AX9-14

AX9.1.3.2 Natural Gradients . . . . . . . . . . . . . . . . . AX9-14

AX9.1.3.3 Use of Chemical Protectants . . . . . . . . AX9-15

AX9.1.3.4 Biomonitoring . . . . . . . . . . . . . . . . . . . AX9-18

AX9.1.3.5 Calibrated Passive Monitors . . . . . . . . AX9-27

AX9.1.4 Numerical/Statistical Methodologies . . . . . . . . . . . . . . AX9-27

AX9.1.5 Improved Methods for Defining Exposure . . . . . . . . . . AX9-30

AX9.2 SPECIES RESPONSE/MODE-OF-ACTION . . . . . . . . . . . . . . . . . . AX9-30

AX9.2.1 Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . AX9-30

AX9.2.2 Mechanisms of Ozone-Induced Plant Alterations . . . . . AX9-34

AX9.2.2.1 Changes in Metabolic Processes: 

Current Theories . . . . . . . . . . . . . . . . . AX9-35

AX9.2.2.2 Modifications of Plant Physiological 

Processes . . . . . . . . . . . . . . . . . . . . . . . AX9-36

AX9.2.3 Ozone Uptake by Leaves . . . . . . . . . . . . . . . . . . . . . . . . AX9-37

AX9.2.3.1 Possible Reactions Within the Leaf . . . AX9-44

AX9.2.3.2 Toxicants Within the Wall Space . . . . AX9-48

III-vii 
AX9.2.4 

AX9.2.5 

AX9.2.6 

AX9.2.7 

AX9.2.8 

AX9.2.9 

Table of Contents 

(cont’d) 

AX9.2.3.3 Products of Ozone ................ 

AX9.2.3.4 Antioxidants Within the Apoplastic 

Space . . . . . . . . . . . . . . . . . . . . . . . . . . 

Wounding and Pathogen Attack .................. 

AX9.2.4.1 Peroxidases ..................... 

AX9.2.4.2 Jasmonic Acid and Salicylic Acid .... 

AX9.2.4.3 Stress-Induced Alterations in 

Gene Expression . . . . . . . . . . . . . . . . .

Page 

AX9-50 

AX9-57 

AX9-66 

AX9-69 

AX9-71 

AX9-72 

Primary Assimilation by Photosynthesis . . . . . . . . . . . . AX9-75

AX9.2.5.1 Photooxidation: Light Reactions ....

Alteration of Rubisco by Ozone: Dark Reactions ....

Carbohydrate Transformations and Translocation ....

AX9.2.7.1 Lipid Synthesis ..................

Role of Age and Size Influencing Response 

to Ozone . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

Summary ....................................

AX9-75 

AX9-76 

AX9-79 

AX9-82 

AX9-84 

AX9-86 

AX9.3 MODIFICATION OF FUNCTIONAL AND GROWTH RESPONSES . . . . . . . AX9-87 

AX9.3.1 Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . AX9-87 

AX9.3.2 Genetics . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . AX9-91 

AX9.3.2.1 Genetic Basis of Ozone Sensitivity . . . AX9-92 

AX9.3.3 Environmental Biological Factors . . . . . . . . . . . . . . . . . AX9-94 

AX9.3.3.1 Oxidant-Plant-Insect Interactions . . . . AX9-95 

AX9.3.3.2 Oxidant-Plant-Pathogen 

Interactions . . . . . . . . . . . . . . . . . . . . . . AX9-98 

AX9.3.3.3 Oxidant-Plant-Symbiont 

Interactions . . . . . . . . . . . . . . . . . . . . . AX9-103 

AX9.3.3.4 Oxidant-Plant-Plant Interactions: 

Competition . . . . . . . . . . . . . . . . . . . . AX9-104 

AX9.3.4 Physical Factors . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . AX9-106 

AX9.3.4.1 Light . . . . . . . . . . . . . . . . . . . . . . . . . . AX9-107 

AX9.3.4.2 Temperature . . . . . . . . . . . . . . . . . . . . AX9-110 

AX9.3.4.3 Humidity and Surface Wetness . . . . . AX9-114 

AX9.3.4.4 Drought and Salinity . . . . . . . . . . . . . AX9-115 

AX9.3.5 Nutritional Factors . . . . . . . . . . . . . . . . . . . . . . . . . . . . AX9-119 

AX9.3.6 Interactions with Other Pollutants . . . . . . . . . . . . . . . . AX9-121 

AX9.3.6.1 

AX9.3.6.2 

AX9.3.6.3 

AX9.3.6.4 

AX9.3.6.5 

AX9.3.6.6 

Oxidant Mixtures . . . . . . . . . . . . . . . . AX9-121 

Sulfur Dioxide . . . . . . . . . . . . . . . . . . AX9-122 

Nitrogen Oxides, Nitric Acid Vapor, 

and Ammonia . . . . . . . . . . . . . . . . . . . AX9-122 

Hydrogen Fluoride and Other 

Gaseous Pollutants . . . . . . . . . . . . . . . AX9-125 

Acidic Deposition . . . . . . . . . . . . . . . AX9-126 

Heavy Metals . . . . . . . . . . . . . . . . . . . AX9-127 

III-viii 
AX9.3.7 

AX9.3.8 

AX9.3.9 

Table of Contents 

(cont’d) 

AX9.3.6.7 Mixtures of Ozone with Two or 

More Pollutants . . . . . . . . . . . . . . . . . 

Interactions with Agricultural Chemicals .......... 

Factors Associated with Global Climate Change .... 

Page 

AX9-127 

AX9-128 

AX9-128 

AX9.3.8.1 

AX9.3.8.2 

AX9.3.8.3 

Ozone-Carbon DioxideTemperature Interactions . . . . . . . . . . AX9-130 

Ozone-UV-B Interactions . . . . . . . . . AX9-153 

Interactions of Ozone with Multiple 

Climate Change Factors . . . . . . . . . . AX9-155 

Summary - Environmental Factors . . . . . . . . . . . . . . . AX9-155

AX9.4 EFFECTS-BASED AIR QUALITY EXPOSURE INDICES . . . . . AX9-159 

AX9.4.1 Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . AX9-159 

AX9.4.2 Summary of Conclusions from the Previous 

Criteria Document . . . . . . . . . . . . . . . . . . . . . . . . . . . . AX9-160 

AX9.4.3 Evaluation of Various Exposure Indices for 

Describing Ambient Exposure-Response 

Relationships . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . AX9-163 

AX9.4.4 Identifying Exposure Components That Relate 

to Vegetation Effects . . . . . . . . . . . . . . . . . . . . . . . . . . AX9-168 

AX9.4.4.1 

AX9.4.4.2 

AX9.4.4.3 

AX9.4.4.4 

AX9.4.4.5 

AX9.4.4.6 

AX9.4.4.7 

Role of Concentration ............

Role of Duration ................

Patterns of Exposure .............

Frequency of Occurrence of Peak

Concentrations . . . . . . . . . . . . . . . . . .

Canopy Structure ................

Site and Climate Factors ..........

Plant Defense Mechanism - 

Detoxification . . . . . . . . . . . . . . . . . .

AX9.4.5 Ozone Uptake or Effective Dose as an Index ....... 

AX9.4.5.1 Models of Stomatal Conductance . . . 

AX9.4.5.2 Nonlinear Response and 

Developing Flux Indices . . . . . . . . . . 

AX9-169 

AX9-172 

AX9-173 

AX9-176 

AX9-177 

AX9-178 

AX9-178 

AX9-178 

AX9-179 

AX9-183 

AX9.4.5.3 Simulation Models . . . . . . . . . . . . . . . AX9-184 

AX9.4.6 Summary . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . AX9-185 

AX9.5 OZONE EXPOSURE-PLANT RESPONSE 

RELATIONSHIPS . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . AX9-187 

AX9.5.1 Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . AX9-187 

AX9.5.2 Summary of Key Findings/Conclusions from 

Previous Criteria Documents . . . . . . . . . . . . . . . . . . . . AX9-188 

AX9.5.3 Ozone Indices and Ambient Exposure . . . . . . . . . . . . AX9-198 

AX9.5.4 Effects of Ozone on Annual and Biennial Species . . . AX9-216 

III-ix 
Table of Contents 

(cont’d) 

Page 

AX9.5.4.1 Effects on Growth, Biomass, and 

Yield of Individual Species . . . . . . . . AX9-217

AX9.5.4.2 Effects on Plant Quality . . . . . . . . . . . AX9-224

AX9.5.4.3 Effects on Foliar Symptoms . . . . . . . . AX9-226

AX9.5.4.4 Other Effects . . . . . . . . . . . . . . . . . . . AX9-226

AX9.5.4.5 Scaling Experimental Data to 

Field Conditions . . . . . . . . . . . . . . . . . AX9-227

AX9.5.4.6 Summary of Effects on ShortLived Species . . . . . . . . . . . . . . . . . . . AX9-230

AX9.5.5 Effects of Ozone on Long-Lived (Perennial) 

Species . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . AX9-232

AX9.5.5.1 Herbaceous Perennial Species . . . . . . AX9-232

AX9.5.5.2 Deciduous Woody Species . . . . . . . . AX9-237

AX9.5.5.3 European Critical Levels . . . . . . . . . . AX9-243

AX9.5.5.4 Summary of Effects on Deciduous 

Woody Species . . . . . . . . . . . . . . . . . . AX9-244

AX9.5.5.5 Evergreen Woody Species . . . . . . . . . AX9-245

AX9.5.5.6 Summary of Effects on Evergreen 

Woody Species . . . . . . . . . . . . . . . . . . AX9-249

AX9.5.5.7 Scaling Experimental Data to 

Mature Trees . . . . . . . . . . . . . . . . . . . AX9-249

AX9.5.6 Studies With the Chemical EDU . . . . . . . . . . . . . . . . . AX9-253

AX9.5.7 Summary . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . AX9-256

AX9.6 EFFECTS OF OZONE EXPOSURE ON NATURAL 

ECOSYSTEMS . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . AX9-260

AX9.6.1 Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . AX9-260

AX9.6.2 Case Studies . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . AX9-262

AX9.6.2.1 Valley of Mexico . . . . . . . . . . . . . . . . AX9-262

AX9.6.2.2 San Bernardino Mountains . . . . . . . . AX9-270

AX9.6.2.3 Sierra Nevada Mountains . . . . . . . . . . AX9-272

AX9.6.2.4 Appalachian Mountains . . . . . . . . . . . AX9-277

AX9.6.2.5 Plantago Studies in the 

United Kingdom . . . . . . . . . . . . . . . . . AX9-278

AX9.6.2.6 Forest Health in the Carpathian 

Mountains . . . . . . . . . . . . . . . . . . . . . AX9-278

AX9.6.2.7 Field Exposure System (FACE), 

Rhinelander, Wisconsin . . . . . . . . . . . AX9-280

AX9.6.3 Landscape Condition . . . . . . . . . . . . . . . . . . . . . . . . . . AX9-282

AX9.6.4 Biotic Condition . . . . . . . . . . . . . . . . . . . . . . . . . . . . . AX9-284

AX9.6.4.1 Ecosystems and Communities . . . . . . AX9-284

AX9.6.4.2 Species and Populations . . . . . . . . . . . AX9-291

AX9.6.5 Organism Condition . . . . . . . . . . . . . . . . . . . . . . . . . . AX9-294

III-x 
Table of Contents 

(cont’d) 

Page 

AX9.6.6 Ecosystem, Chemical, and Physical Characteristics 

(water, soil) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . AX9-302 

AX9.6.6.1 Nutrient Concentrations, Trace 

Inorganic, and Organic Chemicals . . . AX9-302 

AX9.6.7 Ecological Processes . . . . . . . . . . . . . . . . . . . . . . . . . . AX9-303 

AX9.6.7.1 Energy Flow . . . . . . . . . . . . . . . . . . . . AX9-303 

AX9.6.7.2 Material Flow . . . . . . . . . . . . . . . . . . . AX9-305 

AX9.6.8 Hydrological and Geomorphological . . . . . . . . . . . . . AX9-308 

AX9.6.9 Natural Disturbance Regimes . . . . . . . . . . . . . . . . . . . AX9-308 

AX9.6.10 Scaling to Ecosystem Levels . . . . . . . . . . . . . . . . . . . . AX9-309 

AX9.6.10.1 Scaling from Seedlings to 

Mature Trees . . . . . . . . . . . . . . . . . . . AX9-310 

AX9.6.10.2 Surveys, Growth Correlations, and 

Stand-Level Modeling . . . . . . . . . . . . AX9-312 

AX9.6.11 Summary of Ecological Effects of Ozone Exposure 

on Natural Ecosystems . . . . . . . . . . . . . . . . . . . . . . . . AX9-319 

AX9.7 ECONOMIC EVALUATION OF OZONE EFFECTS 

ON AGRICULTURE, FORESTRY, AND NATURAL 

ECOSYSTEMS . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . AX9-323 

AX9.7.1 Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . AX9-323 

AX9.7.2 The Measurement of Economic Information . . . . . . . AX9-324 

AX9.7.3 Understanding of Air Pollutants Effects on the 

Economic Valuation of Agriculture and Other 

Vegetation in the 1996 Criteria Document . . . . . . . . . AX9-326 

AX9.7.3.1 Agriculture . . . . . . . . . . . . . . . . . . . . . AX9-327 

AX9.7.3.2 Forests (Tree Species) and Natural 

Ecosystems . . . . . . . . . . . . . . . . . . . . . AX9-330 

AX9.7.4 Studies Since 1996 of Ozone Exposure Effects 

on the Economic Value of Agriculture, Forests, 

and Ecosystems . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . AX9-331 

AX9.7.5 Limitations of Scientific Studies and Economic 

Information . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . AX9-333 

AX9.7.6 Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . AX9-336 

REFERENCES . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . AX9-337 

III-xi 
List of Tables 

Number 

AX9-1 

AX9-2 

AX9-3 

AX9-4 

AX9-5 

AX9-6 

AX9-7 

AX9-8 

AX9-9 

AX9-10 

AX9-11 

AX9-12 

AX9-13 

AX9-14 

AX9-15 

Page 

Advantages and Disadvantages of Protective Chemicals Used in 

Assessment of O3 Effects on Plants . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . AX9-19 

Advantages and Disadvantages of Bioindicators Used to Study O3 

Plant Effects . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . AX9-20 

Advantages and Disadvantages of Cultivar Comparisons Used in 

Assessment of O3 Effects on Plants . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . AX9-24 

Advantages and Disadvantages of Various Dendrochronological 

Techniques Used in Assessment of O3 Effects on Plants . . . . . . . . . . . . . AX9-26 

Advantages and Disadvantages of Modeling Techniques Used in 

Assessment of O3 Effects on Plants . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . AX9-29 

The Flow of Ozone into a Leaf and Possible Reactions . . . . . . . . . . . . . .

Some Rates of Reaction of Ozone with Critical Biochemicals . . . . . . . . .

Superoxide Dismutase Isozymes and Isoforms . . . . . . . . . . . . . . . . . . . . .

Gene Families and cDNA Clones Used as Probes for SAR ...........

Proteins Altered by Ozone as Measured by Molecular Biological 

Techniques as mRNA Level or Other Gene Activity Rather than 

Enzyme Activity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

Interactions Involving O3 and Plant Pathogens .....................

Effects of Increased Carbon Dioxide on Ozone-Induced Responses 

AX9-42 

AX9-47 

AX9-64 

AX9-67 

AX9-73 

AX9-99 

of Plants at the Metabolic, Physiological, and Whole-Plant Levels . . . . AX9-132 

Summary of Ozone Exposure Indices Calculated for 3- or 5-Month 

Growing Seasons from 1982 to 1991 . . . . . . . . . . . . . . . . . . . . . . . . . . . . AX9-190 

Ozone Exposure Levels (Using Various Indices) Estimated to Cause 

at Least 10% Crop Loss in 50 and 75% of Experimental Cases . . . . . . . AX9-193 

SUM06 Levels Associated with 10 and 20% Total Biomass Loss for 

50 and 75% of the Seedling Studies . . . . . . . . . . . . . . . . . . . . . . . . . . . . AX9-195 

III-xii 
Number 

AX9-16 

AX9-17 

AX9-18 

AX9-19 

AX9-20 

AX9-21 

AX9-22 

AX9-23 

AX9-24 

AX9-25 

AX9-26 

List of Tables 

(cont’d) 

Page 

Summary of Selected Studies of Ozone Effects on Annual Species . . . . AX9-201 

Summary of Selected Studies of the Effects of Ozone on Perennial 

Herbaceous Plants . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

Summary of Selected Studies of Ozone Effects on Deciduous 

Trees and Shrubs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

Summary of Selected Studies of Ozone Effects on Evergreen Trees 

and Shrubs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

Ozone Exposures at 35 Rural Sites in the Clean Air Status and 

Trends Network in the Central and Eastern United States From

1989 to 1995 ..............................................

Ethylene Diurea Effects on Vegetation Responses to Ozone .........

Essential Ecological Attributes for Natural Ecosystems Affected 

by O3 ....................................................

Case Studies Demonstrating the Ecological Effects of O3 ...........

The Most Comprehensively Studied Effects of O3 on Natural 

AX9-206 

AX9-208 

AX9-213 

AX9-217 

AX9-254 

AX9-263 

AX9-271 

Ecosystem are for the San Bernardino Mountain Forest Ecosystem . . . . AX9-273 

Effects of Ozone, Ozone and N Deposition, and Ozone and Drought 

Stress on Pinus ponderosa and Pinus jeffreyi in the Sierra Nevada 

and the San Bernardino Mountains, California . . . . . . . . . . . . . . . . . . . . AX9-274 

Summary of Responses of Populus tremuloides to Elevated CO2 

(+200 :mol mol-1), O3 (1.5 × ambient), or CO2+O3 Compared with 

Control During 3 Years of Treatments at the Aspen FACE . . . . . . . . . . AX9-281 

III-xiii 
List of Figures 

Number Page

AX9-1 Ozone uptake from the atmosphere . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . AX9-32

AX9-2 Absorption and transformation of O3 within the leaf . . . . . . . . . . . . . . . . . AX9-34

AX9-3 The uptake of O3 into the leaf . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . AX9-38

AX9-4 The microarchitecture of a dicot leaf . . . . . . . . . . . . . . . . . . . . . . . . . . . . . AX9-39

AX9-5 The change in the O3 concentration inside a leaf with time . . . . . . . . . . . AX9-40

AX9-6 Possible transformations of O3 within a leaf . . . . . . . . . . . . . . . . . . . . . . . AX9-43

AX9-7 Possible reactions of O3 within water . . . . . . . . . . . . . . . . . . . . . . . . . . . . AX9-43

AX9-8a,b The Crigee mechanism of O3 attack of a double bond . . . . . . . . . . . . . . . AX9-45

AX9-9 Varied ESR radicals, trapped and not, generated by O3 under 

somewhat physiological conditions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . AX9-49 

AX9-10 Pathogen-Induced Hypersensitivity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . AX9-51 

AX9-11 The interaction of H2O2 and Ca2+ movements with ABA-induced 

stomatal closure . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . AX9-54 

AX9-12 The reaction of ascorbate within the apoplasm of the cell wall and 

its ultimate reduction/oxidations . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . AX9-58 

AX9-13 The pathway leading from phospholipids to jasmonic and 

traumatic acid . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . AX9-71 

AX9-14 The production of Rubisco and its Calvin Cycle pathway reactions . . . . . AX9-77 

AX9-15 Linkage of senescence with hypersensitivity reactions and the first 

event of O3 attack of plants . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . AX9-85 

AX9-16 Diagrammatic representation of several exposure indices, illustrating 

how they weight concentration and accumulate exposure . . . . . . . . . . . . AX9-164 

AX9-17 Trends in May-September 12-h SUM06, peak 1-h ozone concentration

 and number of daily exceedances of 95 ppb for Crestline in 1963-1999 

in relation to trends in mean daily maximum temperature for Crestline 

and daily reactive organic gases (ROG) and oxides of nitrogen (NOx) 

for San Bernardino county . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . AX9-170

III-xiv 
List of Figures 

(cont’d) 

Number Page 

AX9-18 The number of hourly average concentrations between 50 and 89 ppb 

for the period 1980 to 2000 for the Crestline, San Bernardino, CA 

monitoring site . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . AX9-172 

AX9-19 Distribution of biomass loss predictions from Weibull and linear 

exposure-response models that relate biomass to O3 exposure . . . . . . . . AX9-194 

AX9-20 A conceptual diagram of processes and storage pools in sources and 

sinks that are affected by O3 exposure . . . . . . . . . . . . . . . . . . . . . . . . . . . AX9-283 

AX9-21 Common anthropogenic stressors and the essential ecological 

attributes they affect . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . AX9-320 

III-xv 
Authors, Contributors, and Reviewers 

CHAPTER 9 ANNEX (ENVIRONMENTAL EFFECTS) 

Principal Authors 

Dr. Jay Garner—National Center for Environmental Assessment (B243-01), U.S. Environmental 

Protection Agency, Research Triangle Park, NC 27711 (retired 

Dr. Timothy Lewis—National Center for Environmental Assessment (B243-01), 

U.S. Environmental Protection Agency, Research Triangle Park, NC 27711 

Dr. William Hogsett—National Health and Environmental Effects Research Laboratory, 

U.S. Environmental Protection Agency, Corvallis, OR 

Dr. Christian Andersen—National Health and Environmental Effects Research Laboratory, 

U.S. Environmental Protection Agency, Corvallis, OR 

Dr. Allen Lefohn—ASL and Associates, Helena, MT 

Dr. David Karnosky—Forest Resources and Environmental Sciences, Michigan Technological 

University, Houghton, MI 

Dr. Michael Nannini—Center for Aquatic Ecology, Illinois Natural History Survey,

Kinmundy, IL

Dr. Nancy Grulke—Pacific Southwest Research Station Forest Fire Laboratory, USDA Forest

Service, Riverside, CA

Dr. Richard Adams—Department of Agriculture and Resource Economics, Oregon State

University., Corvallis, OR

Dr. Robert Heath—Department of Botany and Plant Sciences, University of California,

Riverside, CA

Dr. Victor Runeckle—Biology Department, University of British Columbia, Vancouver, B.C., 

CN (Retired) 

Dr. Arthur Chappelka—Auburn University, School of Forestry, Auburn, AL 

Dr. William Massman—USDA Forest Service, Ft. Collins, CO 

Dr. Robert Musselman—USDA Forest Service, Fort Collins, CO 

Dr. Peter Woodbury—Cornell University, Ithaca, NY (former USDA Forest Service) 

III-xvi 
Authors, Contributors, and Reviewers 

(cont’d) 

Contributors and Reviewers 

Dr. Fitzgerald Booker—USDA-ARS Plant Science Research Unit, 3908 Inwood Rd., Raleigh, 

NC 27603

Dr. Boris Chevone—Department of Plant Pathology, Virginia Technological University,

Blacksburg, VA 24061

Dr. Alan Davison—School of Biology, Newcastle University, Newcastle on Tyne,

United Kingdom, NE1 7RU

Dr. Bruce L. Dixon—Department of Agricultural Economics, University of Arkansas,

Fayetteville, AR 72701

Dr. David Grantz—Kearney Agricultural Center, University of California at Riverside, Parlier,

CA 93648

Dr. Allen S. Heagle—1216 Scott Pl., Raleigh, NC 27511

Dr. Robert Horst, Jr.—121 Thorwald Dr., Plainsboro, NJ 08536

Dr. John Innes—Forest Sciences Centre, Department of Forest Resources, University of British

Columbia, Vancouver, BC, Canada V6T 1Z4

Dr. Hans-Jürgen Jäger—Heinrich-Buff-Ring 26-32, Institute of Plant Ecology, Justus-Leibig

University, Gessen, Germany D35392

Dr. Robert Kohut— Tower Road, Boyce Thompson Institute, Rm 131,Cornell University,

Ithaca, NY 14853

Dr. Sagar Krupa—1519 Gortner Ave., Department of Plant Pathology, University of Minnesota,

St. Paul, MN 55108

Dr. William Manning—203 Morrill, Department of Microbiology, University of Massachusetts,

Amherst, MA 01003

Dr. Paul Miller—USDA Forest Service, Pacific Southwest Research Station, 4955 Canyon Crest

Drive, Riverside CA 92507

Dr. Howard Neufeld—Rankin Science Bldg., Appalachian State University, Boone, NC 28608

III-xvii 
Authors, Contributors, and Reviewers 

(cont’d) 

Contributors and Reviewers 

(cont’d) 

Dr. Maria-Jose Sanz—Fundacion CEAM, c/Charles Darein, 14-Parque Te Valencia, Spain 

Dr. James Shortle—Department of Ag Econ, Armsby, Pennsylvania State University, 

University Park, PA 16802 

Dr. John Skelly—Department of Plant Pathology, Pennsylvania State University, 

University Park, PA 16803 

III-xviii 
U.S. Environmental Protection Agency Project Team 

for Development of Air Quality Criteria for Ozone 

and Related Photochemical Oxidants 

Executive Direction 

Dr. Lester D. Grant (Director)—National Center for Environmental Assessment-RTP Division,

(B243-01), U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Scientific Staff 

Dr. Lori White(Ozone Team Leader)—National Center for Environmental Assessment

(B243-01), U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. Joseph Pinto—National Center for Environmental Assessment (B243-01), 

U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Ms. Beverly Comfort—National Center for Environmental Assessment (B243-01), 

U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. Brooke Hemming—National Center for Environmental Assessment (B243-01), 

U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. James S. Brown—National Center for Environmental Assessment (B243-01), 

U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. Dennis Kotchmar—National Center for Environmental Assessment (B243-01), 

U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. Jee-Young Kim—National Center for Environmental Assessment (B243-01), 

U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. David Svendsgaard—National Center for Environmental Assessment (B243-01), 

U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. Srikanth Nadadur—National Center for Environmental Assessment (B243-01), 

U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. Timothy Lewis—National Center for Environmental Assessment (B243-01), 

U.S. Environmental Protection Agency, Research Triangle Park, NC 27711 

Dr. Jay Garner—National Center for Environmental Assessment (B243-01), U.S. Environmental

Protection Agency, Research Triangle Park, NC 27711

III-xix 
U.S. Environmental Protection Agency Project Team 

for Development of Air Quality Criteria for Ozone 

and Related Photochemical Oxidants 

(cont’d) 

Scientific Staff 

(cont’d) 

Dr. William Hogsett—National Health and Environmental Effects Research Laboratory, 

U.S. Environmental Protection Agency, Corvallis, OR 

Dr. Christian Andersen—National Health and Environmental Effects Research Laboratory, 

U.S. Environmental Protection Agency, Corvallis, OR 

Mr. Bill Ewald—National Center for Environmental Assessment (B243-01), U.S. Environmental 

Protection Agency, Research Triangle Park, NC 27711 (retired) 

Mr. James Raub—National Center for Environmental Assessment (B243-01), 

U.S. Environmental Protection Agency, Research Triangle Park, NC 27711 (retired) 

Technical Support Staff 

Ms. Nancy Broom—Information Technology Manager, National Center for Environmental 

Assessment (B243-01), U.S. Environmental Protection Agency, Research Triangle Park, NC 

27711 

Mr. Douglas B. Fennell—Technical Information Specialist, National Center for Environmental 

Assessment (B243-01), U.S. Environmental Protection Agency, Research Triangle Park, NC 

27711 

Ms. Emily R. Lee—Management Analyst, National Center for Environmental Assessment 

(B243-01), U.S. Environmental Protection Agency, Research Triangle Park, NC 27711 

Ms. Diane H. Ray—Program Specialist, National Center for Environmental Assessment 

(B243-01), U.S. Environmental Protection Agency, Research Triangle Park, NC 27711 

Ms. Donna Wicker—Administrative Officer, National Center for Environmental Assessment 

(B243-01), U.S. Environmental Protection Agency, Research Triangle Park, NC 27711 (retired) 

Mr. Richard Wilson—Clerk, National Center for Environmental Assessment (B243-01), 

U.S. Environmental Protection Agency, Research Triangle Park, NC 27711 

III-xx 
U.S. Environmental Protection Agency Project Team 

for Development of Air Quality Criteria for Ozone 

and Related Photochemical Oxidants 

(cont’d) 

Document Production Staff 

Ms. Carolyn T. Perry—Manager, Computer Sciences Corporation, 2803 Slater Road, Suite 220,

Morrisville, NC 27560

Mr. John A. Bennett—Technical Information Specialist, Library Associates of Maryland,

11820 Parklawn Drive, Suite 400, Rockville, MD 20852

Mr. William Ellis—Records Management Technician, InfoPro, Inc., 8200 Greensboro Drive,

Suite 1450, McLean, VA 22102

Ms. Sandra L. Hughey—Technical Information Specialist, Library Associates of Maryland,

11820 Parklawn Drive, Suite 400, Rockville, MD 20852

Mr. Matthew Kirk—Graphic Artist, Computer Sciences Corporation, 2803 Slater Road,

Suite 220, Morrisville, NC 27560

Dr. Barbara Liljequist—Technical Editor, Computer Sciences Corporation, 2803 Slater Road,

Suite 220, Morrisville, NC 27560

Ms. Rosemary Procko—Senior Word Processor, TekSystems, 1201 Edwards Mill Road,

Suite 201, Raleigh, NC 27607

Ms. Faye Silliman—Publication/Graphics Specialist, InfoPro, Inc., 8200 Greensboro Drive,

Suite 1450, McLean, VA 22102

Mr. Carlton Witherspoon—Graphic Artist, Computer Sciences Corporation, 2803 Slater Road,

Suite 220, Morrisville, NC 27560

III-xxi 
U.S. Environmental Protection Agency Science Advisory Board (SAB) 

Staff Office Clean Air Scientific Advisory Committee (CASAC) 

Ozone Review Panel 

Chair 

Dr. Rogene Henderson*, Scientist Emeritus, Lovelace Respiratory Research Institute, 2425 

Ridgecrest Drive SE, Albuquerque, NM, 87108, Phone: 505-348-9464, Fax: 505-348-8541, 

(rhenders@lrri.org) (FedEx: Dr. Rogene Henderson, Lovelace Respiratory Research Institute, 

2425 Ridgecrest Drive SE, Albuquerque, NM, 87108, Phone: 505-348-9464) 

Members 

Dr. John Balmes, Professor, Department of Medicine, University of California San Francisco, 

University of California - San Francisco, San Francisco, California, 94143, Phone: 

415-206-8953, Fax: 415-206-8949, (jbalmes@itsa.ucsf.edu) 

Dr. Ellis Cowling*, University Distinguished Professor-at-Large, North Carolina State 

University, Colleges of Natural Resources and Agriculture and Life Sciences, North Carolina 

State University, 1509 Varsity Drive, Raleigh, NC, 27695-7632, Phone: 919-515-7564 , Fax: 

919-515-1700, (ellis_cowling@ncsu.edu) 

Dr. James D. Crapo*, Professor, Department of Medicine, National Jewish Medical and 

Research Center. 1400 Jackson Street, Denver, CO, 80206, Phone: 303-398-1436, Fax: 303-

270-2243, (crapoj@njc.org) 

Dr. William (Jim) Gauderman, Associate Professor, Preventive Medicine, University of 

Southerm California, 1540 Alcazar #220, Los Angeles, CA, 91016, Phone: 323-442-1567, 

Fax: 323-442-2349, (jimg@usc.edu) 

Dr. Henry Gong, Professor of Medicine and Preventive Medicine, Medicine and Preventive 

Medicine, Keck School of Medicine, University of Southern California, Environmental Health 

Service, MSB 51, Rancho Los Amigos NRC, 7601 East Imperial Highway, Downey, CA, 90242, 

Phone: 562-401-7561, Fax: 562-803-6883, (hgong@ladhs.org) 

Dr. Paul J. Hanson, Senior Research and Development Scientist , Environmental Sciences 

Division, Oak Ridge National Laboratory (ORNL), Bethel Valley Road, Building 1062, Oak 

Ridge, TN, 37831-6422, Phone: 865-574-5361, Fax: 865-576-9939, (hansonpz@comcast.net) 

Dr. Jack Harkema, Professor, Department of Pathobiology, College of Veterinary Medicine, 

Michigan State University, 212 Food Safety & Toxicology Center, East Lansing, MI, 48824, 

Phone: 517-353-8627, Fax: 517-353-9902, (harkemaj@msu.edu) 

III-xxii 
U.S. Environmental Protection Agency Science Advisory Board (SAB) 

Staff Office Clean Air Scientific Advisory Committee (CASAC) 

Ozone Review Panel 

(cont’d) 

Members 

(cont’d) 

Dr. Philip Hopke, Bayard D. Clarkson Distinguished Professor, Department of Chemical 

Engineering, Clarkson University, Box 5708, Potsdam, NY, 13699-5708, Phone: 315-268-3861, 

Fax: 315-268-4410, (hopkepk@clarkson.edu) (FedEx: 8 Clarkson Avenue, Potsdam, NY 

136995708) 

Dr. Michael T. Kleinman, Professor, Department of Community & Environmental Medicine, 

100 FRF, University of California - Irvine, Irvine, CA, 92697-1825, Phone: 949-824-4765, Fax: 

949-824-2070, (mtkleinm@uci.edu) 

Dr. Allan Legge, President, Biosphere Solutions, 1601 11th Avenue NW, Calgary, Alberta, 

CANADA, T2N 1H1, Phone: 403-282-4479, Fax: 403-282-4479, (allan.legge@shaw.ca) 

Dr. Morton Lippmann, Professor, Nelson Institute of Environmental Medicine, New York 

University School of Medicine, 57 Old Forge Road, Tuxedo, NY, 10987, Phone: 845-731-3558, 

Fax: 845-351-5472, (lippmann@env.med.nyu.edu) 

Dr. Frederick J. Miller*, Consultant, 911 Queensferry Road, Cary, NC, 27511, Phone: 

919-467-3194, (fjmiller@nc.rr.com) 

Dr. Maria Morandi, Assistant Professor of Environmental Science & Occupational Health, 

Department of Environmental Sciences, School of Public Health, University of Texas - Houston 

Health Science Center, 1200 Herman Pressler Street, Houston, TX, 77030, Phone: 

713-500-9288, Fax: 713-500-9249, (mmorandi@sph.uth.tmc.edu) (FedEx: 1200 Herman 

Pressler, Suite 624) 

Dr. Charles Plopper, Professor, Department of Anatomy, Physiology and Cell Biology, School 

of Veterinary Medicine, University of California - Davis, Davis, California, 95616, Phone: 

530-752-7065, (cgplopper@ucdavis.edu) 

Mr. Richard L. Poirot*, Environmental Analyst, Air Pollution Control Division, Department of 

Environmental Conservation, Vermont Agency of Natural Resources, Bldg. 3 South, 103 South 

Main Street, Waterbury, VT, 05671-0402, Phone: 802-241-3807, Fax: 802-241-2590, 

(rich.poirot@state.vt.us) 

Dr. Armistead (Ted) Russell, Georgia Power Distinguished Professor of Environmental 

Engineering, Environmental Engineering Group, School of Civil and Environmental 

Engineering, Georgia Institute of Technology, 311 Ferst Drive, Room 3310, Atlanta, GA, 

30332-0512, Phone: 404-894-3079, Fax: 404-894-8266, (trussell@ce.gatech.edu) 

III-xxiii 
U.S. Environmental Protection Agency Science Advisory Board (SAB) 

Staff Office Clean Air Scientific Advisory Committee (CASAC) 

Ozone Review Panel 

(cont’d) 

Members 

(cont’d) 

Dr. Elizabeth A. (Lianne) Sheppard, Research Associate Professor, Biostatistics and 

Environmental & Occupational Health Sciences, Public Health and Community Medicine, 

University of Washington, Box 357232, Seattle, WA, 98195-7232, Phone: 206-616-2722, Fax: 

206 616-2724, (sheppard@u.washington.edu) 

Dr. Frank Speizer*, Edward Kass Professor of Medicine, Channing Laboratory, Harvard 

Medical School, 181 Longwood Avenue, Boston, MA, 02115-5804, Phone: 617-525-2275, Fax: 

617-525-2066, (frank.speizer@channing.harvard.edu) 

Dr. James Ultman, Professor, Chemical Engineering, Bioengineering program, Pennsylvania 

State University, 106 Fenske Lab, University Park, PA, 16802, Phone: 814-863-4802, Fax: 

814-865-7846, (jsu@psu.edu) 

Dr. Sverre Vedal, Professor of Medicine, Department of Environmental and Occupational Health 

Sciences, School of Public Health and Community Medicine, University of Washington, 4225 

Roosevelt Way NE, Suite 100, Seattle, WA, 98105-6099, Phone: 206-616-8285, Fax: 

206-685-4696, (svedal@u.washington.edu) 

Dr. James (Jim) Zidek, Professor, Statistics, Science, University of British Columbia, 6856 

Agriculture Rd., Vancouver, BC, Canada, V6T 1Z2, Phone: 604-822-4302, Fax: 604-822-6960, 

(jim@stat.ubc.ca) 

Dr. Barbara Zielinska*, Research Professor , Division of Atmospheric Science, Desert Research 

Institute, 2215 Raggio Parkway, Reno, NV, 89512-1095, Phone: 775-674-7066, Fax: 

775-674-7008, (barbz@dri.edu) 

Science Advisory Board Staff 

Mr. Fred Butterfield, CASAC Designated Federal Officer, 1200 Pennsylvania Avenue, N.W., 

Washington, DC, 20460, Phone: 202-343-9994, Fax: 202-233-0643 (butterfield.fred@epa.gov) 

(Physical/Courier/FedEx Address: Fred A. Butterfield, III, EPA Science Advisory Board Staff 

Office (Mail Code 1400F), Woodies Building, 1025 F Street, N.W., Room 3604, Washington, 

DC 20004, Telephone: 202-343-9994) 

*Members of the statutory Clean Air Scientific Advisory Committee (CASAC) appointed by

 the EPA Administrator 

III-xxiv 
ABBREVIATIONS AND ACRONYMS 

AA ambient air 

ABA abscisic acid 

ABI2 phospho-tyrosine-specific protein phosphatase 

ACC 1-aminocyclopropane-1-carboxylate 

ACS 1-aminocyclopropane-1-carboxylase synthase 

Amax maximum photosynthesis rate 

ANN artificial neural network 

ANOVA analysis of variance 

ANP Acadia National Park 

AOS allene oxide synthase 

AOT40 seasonal sum of the difference between an hourly concentration at the 

threshold value of 40 ppb, minus the threshold value of 40 ppb 

AOT60 seasonal sum of the difference between an hourly concentration at the 

threshold value of 60 ppb, minus the threshold value of 60 ppb 

AOTx family of cumulative, cutoff concentration-based exposure indices 

APX ascorbate peroxidase 

AQCD Air Quality Criteria Document 

Asat photosynthetic assimilation in saturating light 

ASC ascorbate 

ATPase adenosine triphosphatase 

AVG 1-aminoethoxyvinyl-glycine 

AZO azoxystrobin 

BCB blue copper binding protein 

Cab chlorophyll a/b binding protein 

CAT catalase 

CEC controlled environment chambers 

cDNA complementary DNA 

III-xxv 
CV 

CF charcoal-filtered 

CFA charcoal/Purafil-filtered air 

CFI continuous forest inventory 

CHIP Effects of Elevated Carbon Dioxide and Ozone on Potato Tuber Quality 

in the European Multiple Site Experiment 

CO2 carbon dioxide 

CSTR continuous stirred tank reactor 

CU cumulative uptake 

coefficient of variation 

cyt cytochrome 

DG diacylglycerol 

DGDG digalactosyldiacylglycerol 

DHA dehydroascorbate 

DMPO dimethylphrrolise 1-oxide; 5,5-dimethyl-1-pyrroline N-oxide 

DNA deoxyribonucleic acid 

ECM ectomycorrhizal fungi 

EDU ethylenediurea 

EEA essential ecological attribute 

EMEP European Monitoring and Evaluation Program 

EPA U.S. Environmental Protection Agency 

EPO epoxyconazole 

EPR electron paramagnetic resonance; ESR 

ERD1 ethylene response 

ESPACE-wheat European Stress Physiology and Climate Experiment on the Effects of 

Carbon Dioxide and Oxygen on Spring Wheat 

ESR electron spin resonance; EPR 

ET ethylene 

EU European Union 

NPSII,max maximum light-adapted apparent quantum efficiency of Photosystem II 

III-xxvi 
FA fatty acid 

FACE free-air carbon dioxide enrichment (system) 

FFAs free fatty acids 

FHM Forest Health Monitoring (assessment) 

FLAG Federal Land Managers’ Air Quality Related Values Workgroup 

FPM Forest Pest Management 

G plants rooted in ground 

GDP guanosine diphosphate 

GGGT galactolipid:galactolipid galactosyltransferase 

GHG greenhouse gas 

GPx glutathione peroxidase 

GR glutathione reductase 

GRSM Great Smoky Mountains National Park 

GSH glutathione 

GSH-Px glutathione peroxidase 

GSMNP Great Smoky Mountains National Park 

GSSG glutathione disulfide 

GST glutathione synthase 

H+ hydrogen ion 

2HDM, 2ndHDM second-highest daily maximum 1-h concentration 

HF hydrogen fluoride 

HNO3 nitric acid 

H2O2 hydrogen peroxide 

HO3C protonated ozone radical 

HOC hydroxyl radical 

HO2 

C hydroperoxyl; hydroperoxy radical; protonated superoxide 

HPOT 13-hydroperoxide linolenic acid 

HR hypersensitive response 

III-xxvii 
ICP Forests International Cooperative Programme on Assessment of Air Pollutant 

Effects on Forests 

IPCC Intergovernmental Panel on Climate Change 

JA jasmonic acid 

Jmax maximum rate of electron transport for the regeneration of RuBP 

Jsat saturating light 

KROFEX Krauzberg Ozone Fumigation Experiment 

LAI leaf area index 

LOX1 lipoxygenase 

M7 7-hour seasonal mean 

mAOT modified accumulated exposure over the threshold 

MDA malonaldehyde 

MDGD monogalactosyldiacylglycerol 

MGDG monogalactosyldiacylglycerol 

mRNA messenger ribonucleic acid 

MT1 mitochondria 

MV methyl viologen 

NAAQS National Ambient Air Quality Standards 

NAD nicotinamide adenine dinucleotide 

NADH reduced nicotinamide adenine dinucleotide 

NADP+ nicotinamide adenine dinucleotide phosphate 

NADPH, reduced nicotinamide adenine dinucleotide phosphate 

NAD(P)H 

NaE sodium erythorbate 

NCLAN National Crop Loss Assessment Network 

n.d. no data 

NDF neutral detergent fiber 

NF national forest 

NF non-filtered 

III-xxviii 
NH3 ammonia 

(NH4)2SO4 ammonium sulfate 

N2O nitrous oxide 

NO nitric oxide 

NO2 nitrogen dioxide 

NO3 

! nitrate 

NOx nitrogen oxides 

NP national park 

NPP net primary productivity 

n.s. nonsignificant 

O2 

! superoxide 

O2 

C superoxide radical 

1

O2 singlet oxygen 

O3 ozone 

OD outer diameter 

OTC open-top chamber 

p, P probability value 

P plants grown in pots 

PAD pollutant applied dose 

PAL phenylalanine lyase 

PAN peroxyacetyl nitrate 

PAR photosynthetically active radiation 

PC phosphatidylchloline 

PE phosphatidylethanolamine 

PFD photosynthetic flux density 

PG phosphatidylglycerol 

PGSM Plant Growth Stress Model 

PI phosphatidylinositol 

III-xxix 
R

r

POD peroxidase 

ppb parts per billion 

ppm parts per million 

PQH2 plastoquinone 

PR pathogenesis-related (protein) 

PR-1 promotor region 1 

PRYL predicted relative yield (biomass) loss 

PSII Photosystem II 

Pxase peroxidase 

qP photochemical quenching 

2 correlation coefficient 

2 multiple regression correlation coefficient 

rbcL Rubisco large subunit 

rbcS Rubisco small subunit 

RH relative humidity 

RNA ribonucleic acid 

ROG reactive organic gases 

ROS reactive oxygen species 

Rubisco ribulose-1,6-P2-carboxylase/oxygenase1 

RuBP ribulose bisphosphate 

SA salicylic acid 

SAB Science Advisory Board 

SAG21 senescence 

SAR systemic acquired resistance 

SD standard deviation 

SE standard error 

SHEN Shenandoah National Park 

SIGMOID sigmoid weighted summed concentration 

III-xxx 
SO

SLA specific leaf area 

SMD soil moisture deficit 

SNAAQS Secondary National Ambient Air Quality Standards 

SO2 sulfur dioxide 

4

2! sulfate 

SOD superoxide dismutase 

SUM00 sum of all hourly average concentrations 

SUM06 seasonal sum of all hourly average concentrations $ 0.06 ppm 

SUM08 seasonal sum of all hourly average concentrations $ 0.08 ppm 

TMPO tetramethylphrrolise 1-oxide 

TNC total nonstructural carbohydrate 

UDGT UDP galactose-1,2,-diacylglycerol galactosyltransferase 

UNECE United Nations Economic Commission for Europe 

UN ECE ICP- United Nations Economic Commission for Europe International 

Vegetation Cooperative Programme on effects of air pollution and other stresses on 

crops and non-woody plants (UN/ECE-Vegetation; formerly -Crops) 

UNEP United Nations Environment Program 

UDP uridine diphosphate 

USDA U.S. Department of Agriculture 

UV ultraviolet 

UV-B ultraviolet radiation of wavelengths from 280 to 320 nm 

VOC volatile organic compound 

VPD vapor pressure deficit 

W95 cumulative integrated exposure index with a sigmoidal weighting 

function 

W126 cumulative integrated exposure index with a sigmoidal weighting 

function 

ZAPS Zonal Air Pollution System 

III-xxxi 
ANNEX AX9. ENVIRONMENTAL EFFECTS: OZONE

EFFECTS ON VEGETATION AND ECOSYSTEMS

AX9.1 METHODOLOGIES USED IN VEGETATION RESEARCH 

AX9.1.1 Introduction 

The scale of investigations evaluating the direct effects of O3 on plant response ranges 

from subcellular to cellular, organismal, population, community, and ecosystem levels, with 

each level having its own particular experimental methodologies and specialized 

instrumentation, equipment, facilities, and experimental protocols. These investigations generate 

data. Other types of methodologies exist for the handling of data and statistical analysis as well 

as the utilization of data in developing the different exposure metrics or indices used to define 

exposure, quantitative exposure-response relationships, and computer simulation models of these 

exposure-response relationships. The objective of this section is not to provide an updated 

encyclopedia of all the methods that have been used but rather to focus on approaches that have 

(1) led to an improved understanding of the quantitatively measurable growth and 

development responses of plants and plant communities to O3, or 

(2) provided information about the extent and geographic distribution of the responses 

of herbaceous and woody plants, both cultivated and native, to ambient O3 

exposures. 

The first part of the objective is essential for determining dose-response functions used in 

developing impact and risk assessments of the effects of O3; it usually involves treating plants to 

a range of artificial O3 exposures. The second part of the objective is essential for determining 

the geographic distribution of the risk; it usually involves subjecting plants to ambient air O3 

exposures. 

The types of methodologies used by biochemists, molecular biologists, or plant 

physiologists, whose interests lie in determining effects on specific constituents or in 

understanding the mode of action of O3, are not discussed here but are addressed in Section 

AX9.2. Methods used to characterize the O3 content of ambient air and to define exposure and 

exposure-response relations are discussed in Sections AX9.4 and AX9.5, respectively. 

AX9-1
The methodologies for exposure-response studies have involved many different types of 

exposure facilities and protocols and have employed a range of statistical approaches in the 

analysis and interpretation of the data. Most of the studies have been conducted using major 

agricultural crop species. The methodologies have improved over the years as a result of the 

development, availability, or application of new or improved instrumentation, physical systems, 

and numerical approaches to data analysis. Yet equally important to the roles played by these 

advances has been the clearer understanding that has emerged from earlier work identifying the 

type of experimentation needed to achieve realistic assessments of the magnitude and extent of 

the impact of O3 on vegetation of all types. As a result, significantly increased attention is now 

being paid to field observations and biomonitoring, particularly to the responses of forest trees 

and native vegetation. 

Other than in various exploratory studies that have used chamber-based steady-state 

exposure concentrations (so-called “square-wave” exposures), the trend in experimental 

exposure protocols has been to attempt to expose plants under conditions as natural as possible 

to temporal profiles that simulate the real world, either by conducting experiments in the field or 

in elaborately controlled environment facilities that provide simulated field conditions. 

Previous Air Quality Criteria Documents for Ozone and Other Photochemical 

Oxidants (U.S. Environmental Protection Agency, 1986, 1996) described the time course for 

these methodological developments. Although this section provides a brief overview of the 

methodologies used in the past and their limitations, it focuses mainly on those techniques that 

have come into prominence over the last decade. This focus has been aided considerably 

by several compilations of experimental methodologies and facilities, such as the earlier 

comprehensive review for the U.S. Environmental Protection Agency/National Acid 

Precipitation Assessment Program by Hogsett et al. (1987a,b), Manning and Krupa (1992) 

and by more recent reviews by Musselman and Hale (1997) and Karnosky et al. (2001b). 

AX9.1.2 Methods Involving Experimental Exposures to Ozone 

AX9.1.2.1 “Indoor”, Controlled Environment, and Greenhouse Chambers 

The earliest experimental investigations of the effects of O3 on plants utilized simple glass 

or plastic-covered chambers, often located within greenhouses, into which a flow of O3-enriched 

air or oxygen could be passed to provide the exposure. The types, shapes, styles, materials of 

AX9-2
construction, and locations of these chambers were as numerous as the different investigators 

and, in spite of providing little resemblance to real-world conditions, they yielded much of the 

basic information on the visible and physiological effects on plants. The construction and 

performance of more elaborate and better instrumented chambers dating back to the 1960s has 

been well-summarized in Hogsett et al. (1987a), including those installed in greenhouses (with 

or without some control of temperature and light intensity). 

One greenhouse chamber approach that continues to yield useful information on the 

relationships of O3 uptake to both physiological and growth effects employs continuous stirred 

tank reactors (CSTRs) first described by Heck et al. (1978). Although originally developed to 

permit mass-balance studies of O3 flux to plants, their use has more recently widened to include 

short-term physiological and growth studies of O3 × CO2 interactions (e.g., Costa et al., 2001; 

Heagle et al., 1994b; Loats and Rebbeck, 1999; Rao et al., 1995; Reinert and Ho, 1995; and 

Reinert et al., 1997), and of surveys of native plant responses to O3 (Orendovici et al., 2003). 

In many cases, supplementary lighting and temperature control of the surrounding structure have 

been used to control or modify the environmental conditions (e.g., Heagle et al., 1994a). 

Many investigations have utilized commercially available controlled environment 

chambers and walk-in rooms adapted to permit the introduction of a flow of O3 into the 

controlled air-volume. Such chambers continue to find use in genetic screening and in 

physiological and biochemical studies aimed primarily at improving our understanding of modes 

of action. For example, some of the ongoing studies of the O3 responses of Plantago major 

populations have been conducted in controlled environment chambers (Reiling and Davison, 

1994; Whitfield et al., 1996b). 

The environmental conditions provided by indoor chambers of any type will always 

preclude the use of the information obtained with such chambers in predicting O3 effects in the 

natural environment, because the environmental conditions will always be measurably different 

from field conditions. However, highly sophisticated controlled environment chambers such as 

those described by Langebartels et al. (1997), which are subdivided into aerial and root 

compartments with dynamic control of light intensity and photoperiod, air and soil temperature, 

humidity, soil moisture, wind speed, and exposure to O3, may come close to simulating specific 

natural conditions. Such chambers have provided meaningful insights into a wide array of the 

early biochemical responses of plants to O3. They can also minimize confounding factors that 

AX9-3
make indoor chamber studies only rarely able to be extrapolated to field conditions, e.g., that 

shoots and roots develop under different temperature regimes. The applicability of the results 

of many chamber studies may be further limited by their use of container-grown plants. Most 

of the concerns over the applicability of CO2 enrichment studies, as discussed in Section 

AX9.5.7.1, may also be relevant to O3 enrichment studies, as suggested by Whitfield et al. 

(1996a). 

Whitfield et al. (1996a) reported significant interactive effects between O3 and soil volume 

on the growth of Plantago major. They noted that although container size may limit root and, 

hence, plant growth, the reverse may also be true for single plants in large containers, which do 

not experience typical field competition for resources. Other studies found little or no effect of 

rooting volume on plant response to O3. Heagle et al. (1979a,b; 1983) found that four wheat 

cultivars (Triticum aestivum L.) had similar proportional suppression of seed yield to seasonlong O3 exposure whether plants were grown in the ground or in 3.8-L pots. Similarly, 

proportional O3 injury and yield response of field corn (Zea mays L.) (Heagle et al., 1979a) and 

soybean (Glycine max (L.) Merr.) (Heagle et al., 1983) was similar whether the plants were 

grown in 15-L pots or in the ground. In a two-year experiment with soybean, the relative effects 

of CO2 and O3 on above-ground biomass and seed yield were similar whether the plants were 

grown in pots (15 and 21 L) or grown in the ground (Booker et al., 2005). Collectively, the 

results suggest that while planting density and rooting environment affect plant morphology and 

growth, the relative responses of seed yield to elevated O3 may be similar whether plants are 

grown in pots or in the ground. 

AX9.1.2.2 Field Chambers 

Although some types of closed field chambers have largely fallen out of favor in recent 

years, closed “Solardome” field chambers (Lucas et al., 1987; Rafarel and Ashenden, 1991) have 

been successfully used in studies of O3 × acid mist interactions (Ashenden et al., 1995, 1996). 

Concern over the need to establish realistic plant-litter-soil relationships as a prerequisite to 

studies of the effects of O3 and CO2 enrichment on ponderosa pine (Pinus ponderosa) seedlings 

led Tingey et al. (1996) to develop closed, partially environmentally controlled, sun-lit chambers 

(“terracosms”) incorporating 1-m-deep lysimeters containing forest soil in which the appropriate 

horizon structure was retained. 

AX9-4
In general, field chamber studies are dominated by the use of various versions of the opentop chamber (OTC) design, first described by Heagle et al. (1973) and Mandl et al. (1973). Most 

chambers are ~3 m in diameter with 2.5-m-high walls. Hogsett et al. (1987a) described in detail 

many of the various modifications to the original OTC designs that appeared subsequently, e.g., 

the use of larger chambers to permit exposing small trees (Kats et al., 1985a) and grapevines 

(Mandl et al., 1989), the addition of a conical baffle at the top to improve ventilation (Kats et al., 

1976), a frustrum at the top to reduce ambient air incursions, and a plastic rain-cap to exclude 

precipitation (Hogsett et al., 1985b). All of these modifications included the discharge of air via 

ports in annular ducting or interiorly perforated double-layered walls at the base of the chambers 

to provide turbulent mixing and the upward mass flow of air. 

Wiltshire et al. (1992) described a large OTC suitable for small trees with roll-up sides that 

permitted the trees to be readily subjected from time to time to episodic, normal, “chamberless” 

environmental conditions. In the 6-m-high OTCs described by Seufert and Arndt (1985) used 

with Norway spruce (Picea abies) trees, a second zone of annular enrichment was also provided 

between 4 and 5 m. The use of OTCs was adopted for the large European Stress Physiology and 

Climate Experiment on the effects of CO2 and O3 on spring wheat (ESPACE-wheat), conducted 

over 1994 to 1996 at field sites in eight countries (Jäger et al., 1999). However, typical 

European chambers have the introduction of O3-enriched air at or above canopy height. The 

relatively low costs of fabrication, operation, and maintenance has favored OTC use in field 

studies (Fangmeier et al., 1992; Musselman and Hale, 1997). The air supplied to the chambers 

can be readily filtered through activated charcoal to reduce the O3 concentration, or it can be 

enriched with O3 to provide a range of exposures. 

All field chambers create internal environments that differ from ambient air, giving rise to 

so-called “chamber effects” with the modification of microclimatic variables, including reduced 

and uneven light intensity, uneven rainfall, constant wind speed, reduced dew formation, and 

increased air temperatures (Fuhrer, 1994; Manning and Krupa, 1992). Because of the constant 

wind speed and delivery systems, OTCs can provide a more definable exposure than free-air 

systems can due to the lack of “hot-spots”, where exposures are essentially undefined, in free-air 

systems. Nonetheless, there are several characteristics of the OTC design and operation that can 

lead to unrealistic exposures. First, the plants are subjected to constant turbulence, which, 

through increased uptake resulting from the consequently low boundary layer resistance to 

AX9-5
diffusion, may lead to overestimation of cause-effect relationships (Krupa et al., 1995; Legge 

et al., 1995). However, in at least one case where canopy resistances were quantified in OTCs 

and in the field, it was determined that gaseous pollutant exposure to crops in OTCs was similar 

to that which would have occurred at the same concentration in the field (Unsworth et al., 1984a, 

1984b). 

A second concern is that the introduction of the O3-enriched air into the lower part of 

chambers as described by Heagle et al. (1973) and Mandl et al. (1973) results in a O3 

concentration gradient that decreases with increasing height, the converse of the situation 

observed in ambient air in which the O3 concentration decreases markedly from above a plant 

canopy to ground level (Grünhage and Jäger, 1994; Pleijel et al., 1995b, 1996). Concern that 

studies conducted in such OTCs may somewhat overestimate the effects of O3 led to the 

European design, which provides a decreasing downward gradient. It seems unlikely that the 

“chamber effects” produced by the two designs will be the same. These issues are discussed 

more fully in Section AX9.1.2.4. 

It should also be noted that, although OTCs were originally developed for exposing row 

crops in the field, many recent studies employing OTCs have used potted plants in order to 

include or control edaphic or nutritional factors or water relations within the experimental 

design. Therefore, some caution should be used when extrapolating results of pot studies to the 

field as noted above (Section AX9.2.2.1). 

The difficulties faced in the experimental exposure of forest trees to air pollutants in 

chambers (e.g., Seufert and Arndt [1985]) led to the development of branch chambers such as 

those described by Ennis et al. (1990), Houpis et al. (1991), and Teskey et al. (1991). These 

chambers are essentially large cuvettes without temperature control and, as noted by Musselman 

and Hale (1997), share many of the characteristics of CSTRs, i.e., transparent walls, internal 

fans, and inlet and outlet monitors to permit the determination of O3 uptake, CO2 exchange, and 

transpiration. Although they make it possible to expose whole branches to different O3 regimes, 

the relevance of the data they yield in regard to the whole tree may be questionable. As noted by 

Saxe et al. (1998), the inevitable change in environmental conditions resulting from the isolation 

of the branch may cause different responses from those that would be obtained if the whole tree 

had been subjected to the same environmental conditions. 

AX9-6
AX9.1.2.3 Plume Systems 

Plume systems are chamberless exposure facilities in which the atmosphere surrounding 

plants in the field is modified by the injection of pollutant gas into the air above or around them 

from multiple orifices spaced to permit diffusion and turbulence so as to establish relatively 

homogeneous conditions as the individual plumes disperse and mix with the ambient air. As 

pointed out by Manning and Krupa (1992), they can only be used to increase the O3 levels in the 

ambient air. The volume of air to be modified is unconfined, and three approaches have been 

used to achieve desired pollutant concentrations in the air passing over the plants, producing 

various systems that 

(1) achieve a concentration gradient, in most instances dependent upon the direction of 

the prevailing wind; 

(2) achieve spatially uniform concentrations over a plot, dependent upon wind 

direction; and 

(3) seek to achieve spatially uniform concentrations over a plot, independent of wind 

speed and direction. 

Gradient systems created by dispensing a pollutant gas into the air at canopy level from 

perforated horizontal pipes arranged at right angles to the prevailing wind were described 

for SO2 studies in the early 1980s. A modified gradient system for O3 was used by Bytnerowicz 

et al. (1988) to study effects on desert species, but there appear to have been no recent 

applications of the method. A gradient O3-exclusion system is discussed in Section AX9.2.3.1. 

Systems designed to achieve spatially uniform pollutant levels by ensuring that the release 

of a pollutant is always on the upwind side of the study site were also originally described 

for SO2 studies (e.g., Greenwood et al. [1982]). However, the adaptation of these concepts as 

introduced by McLeod et al. (1985) in constructing a large circular field site for exposing crops 

to SO2 led to the subsequent development of both the large-scale O3 and SO2 fumigation system 

for forest trees in the United Kingdom in 1985 (the Liphook Forest Fumigation Project) (1992), 

the smaller system for O3 fumigation constructed at Kuopio, Finland in 1990 (Wulff et al., 

1992), and the free-air carbon-dioxide enrichment (FACE) systems of gas dispersal over crops 

(Hendrey and Kimball, 1994) and forest trees (Hendrey et al., 1999). Although originally 

designed to provide chamberless field facilities for studying the CO2 effects of climate change, 

large forest tree FACE systems have recently been adapted to include the dispensing of O3 

AX9-7
(Karnosky et al., 1999). Volk et al. (2003) recently described a system for exposing grasslands 

that uses 7-m diameter plots. FACE systems discharge the pollutant gas (and/or CO2) through 

orifices spaced along an annular ring (or torus) or at different heights on a ring of vertical pipes. 

Computer-controlled feedback from the monitoring of gas concentration regulates the feed rate 

of enriched air to the dispersion pipes. Feedback of wind speed and direction information 

ensures that the discharges only occur upwind of the treatment plots, and that discharge is 

restricted or closed down during periods of low wind speed or calm conditions. The diameter of 

the arrays and their heights (25 to 30 m) in some FACE systems requires large throughputs of 

enriched air per plot, particularly in forest tree systems. The cost of the throughputs tends to 

limit the number of enrichment treatments, although Hendrey et al. (1999) argued that the cost 

on an enriched volume basis is comparable to that of chamber systems. 

An alternative to the FACE system to free-air fumigation uses a horizontal grid system 

through which pollutant-enriched air is discharged over the canopies of plants in field plots. The 

original design, termed the Zonal Air Pollution System (ZAPS), was developed for studying the 

effects of SO2 on native grasslands (Lee et al., 1975), and it was later modified by Runeckles 

et al. (1990) by randomly dividing each of three treatment plots into four subplots, each with 

different numbers of discharge orifices to provide various levels of O3 enrichment. With the 

ZAPS system, changes in wind direction and speed result in varying degrees of carryover from 

subplot to subplot, effectively resulting in 12 stochastically different seasonal exposures. The 

system was used for studies of growth effects on field crops and 2- to 4-year old Douglas fir 

(Pseudotsuga menziesii) saplings (Runeckles and Wright, 1996). A larger ZAPS design was 

used by Wilbourn et al. (1995) on a grass (Lolium perenne)-clover (Trifolium repens) mixture 

and by Ollerenshaw et al. (1999) on oilseed rape (Brassica napus), whereby four replicate field 

plots were exposed to intermittent constant additions of O3 to ambient air. A ZAPS design with 

eight spatially separated treatment plots was also developed to obtain crop response data used in 

assessing crop losses in the Fraser Valley, British Columbia, Canada (Runeckles and Bowen, 

2000). 

Another recent adaptation of the FACE design was constructed to fumigate soybean 

with CO2 and O3 in combination (Morgan et al., 2004; Rogers et al., 2004). This modified 

FACE design was based on those of Miglietta et al. (2001) and does not force air through the 

canopy; instead, it relies on wind to disperse air across the fumigation plot. 

AX9-8
The FACE-type facility developed for the Kranzberg Ozone Fumigation Experiment 

(KROFEX) in Germany begun in 2000 (Werner and Fabian, 2002; Nunn et al., 2002) to study 

the effects of O3 on mature stands of beech (Fagus sylvatica) and spruce (Picea abies) trees is 

more truly a zonal system that functions independently of wind direction. The enrichment of a 

large volume of the ambient air immediately above the canopy takes place via orifices in vertical 

tubes suspended from a horizontal grid supported above the canopy. 

Recognizing the difficulties of modifying the aerial environments of large trees, Tjoelker 

et al. (1994) devised a free-air system for exposing branches of sugar maple (Acer saccharum) 

trees to O3. Near the ends of up to 10 branches, enriched air was discharged through small holes 

in 38-cm-diameter loops of 0.635-cm-OD (outer diameter) teflon tubes positioned 20 to 30 cm 

below the terminal foliage cluster. 

Although plume systems make virtually none of the modifications to the physical 

environment that are inevitable with chambers, their successful use depends on selecting the 

appropriate numbers, sizes, and orientations of the discharge orifices to avoid hot-spots resulting 

from the direct impingement of jets of pollutant-enriched air on plant foliage (Werner and 

Fabian, 2002). However, because mixing is unassisted and completely dependent on wind 

turbulence and diffusion, local gradients are inevitable even in large-scale FACE systems. Both 

FACE and ZAPS systems have provisions for shutting down under low wind speed or calm 

conditions and for an experimental area that is usually defined within a generous border in order 

to strive for homogeneity of the exposure concentrations within the treatment area. They are 

also both dependent upon continuous computer-controlled feedback of the O3 concentrations in 

the mixed treated air and of the meteorological conditions. 

AX9.1.2.4 Comparative Studies 

All experimental approaches to the exposure of plants to O3 have shortcomings. The use of 

laboratory, greenhouse, or field chambers raises concerns for the roles of chamber effects on 

micrometeorology, as well as the constant turbulence over and within the plant canopy during 

chamber operation, in modifying O3 uptake and subsequent plant response. In contrast, plume 

systems suffer from relatively poor control of exposure levels and an inability to reduce O3 

levels below ambient in areas where O3 is phytotoxic. 

AX9-9
Although chamber effects vary, one concern is the rise in temperature associated with 

enclosing plants in a chamber. Still, it is not clear whether these effects are directly related to 

temperature or are the result of temperature interactions with other environmental variables. 

For example, Olszyk et al. (1992) undertook a 3-year study of the impact of O3 on Valencia 

orange trees (Citrus sinesis (L.) Osbeck) in large OTCs to determine if “insidious differences in 

microclimatic conditions could alter plant growth responses and susceptibility to pollutant 

stress.” Nonfiltered chambers were found to have somewhat lower average O3 concentrations 

than the ambient air, and fewer hourly exceedances of 100 ppb. In cool seasons, stomatal 

conductance was also lower, implying lower O3 uptake. However, the cumulative fruit yields 

were doubled in the chamber trees even though photosynthetically active radiation was 

consistently reduced by about 19% while leaf temperatures averaged more than 2 °C higher. 

These data may be somewhat extreme, but they emphasize the need to be cautious when 

interpreting OTC yield response data, particularly since, as in this study, no O3 enrichment was 

involved as a complicating factor. 

While it is clear that chambers can alter some aspects of plant growth, the question to be 

answered is whether or not these differences affect plant response to O3. As noted in the 

1996 O3 AQCD (U.S. Environmental Protection Agency, 1996), evidence from the comparative 

studies of OTCs and from closed chamber and O3-exclusion exposure systems on the growth of 

alfalfa (Medicago sativa) by Olszyk et al. (1986a) suggested that, since significant differences 

were found for fewer than 10% of the growth parameters measured, the responses were, in 

general, essentially the same regardless of exposure system used and chamber effects did not 

significantly affect response. In 1988, Heagle et al. (1988) concluded: “Although chamber 

effects on yield are common, there are no results showing that this will result in a changed yield 

response to O3.” A more recent study of chamber effects examined the responses of tolerant and 

sensitive white clover clones (Trifolium repens) to ambient O3 in greenhouse, open-top, and 

ambient plots (Heagle et al., 1996). For individual harvests, greenhouse O3 exposure reduced the 

forage weight of the sensitive clone 7 to 23% more than in OTCs. However, the response in 

OTCs was the same as in ambient plots. Several studies have shown very similar yield response 

to O3 for plants grown in pots or in the ground, suggesting that even such a significant change in 

environment does not alter the proportional response to O3, at least as long as the plants are well 

watered (Heagle, 1979; Heagle et al., 1983). 

AX9-10
Recent evidence obtained using free-air exposure systems and OTCs supports results 

observed previously in OTC studies (Table AX9-16, Figure AX9-1). Specifically, a series of 

studies undertaken using free-air O3 enrichment in Rhinelander, WI (Isebrands et al., 2000, 

2001) showed that O3-symptom expression was generally similar in OTCs, FACE, and 

ambient-O3 gradient sites, supporting the previously observed variation among trembling aspen 

clones (Populus tremuloides L.) using OTCs (Karnosky et al., 1999). The FACE study 

evaluated the effects of 3 years of exposure to combinations of elevated CO2 and O3 on growth 

responses in mixture of five trembling aspen clones (Isebrands et al., 2000, 2001). Height, 

diameter, and stem volume (diameter2

 × height) were decreased by elevated O3. On average for 

all clones, stem volume was decreased by 20% over the 3 years in the elevated O3 treatment as 

compared with the 1×-ambient treatment. This FACE facility study is important, because it 

confirms responses reported previously with the same clones grown in pots or soil in OTCs 

without the alterations of microclimate induced by chambers. Currently, this is the only U.S. 

study using this technology to have examined the effects of O3 under these conditions. This 

study is also significant, because the elevated O3-exposure pattern used was intended to 

reproduce the 6-year average pattern from Washtenaw County, Michigan (Karnosky et al., 

1999). 

Chambered systems such as OTCs provide a charcoal-filtered (CF), clean-air control for O3 

experiments, while FACE and some other plume systems do not. Depending on experimental 

intent, a replicated, clean-air control treatment is an essential component in many experimental 

designs. This control cannot be provided by FACE systems where ambient O3 levels are 

phytotoxic. This is especially relevant in CO2 × O3 experiments in which phytotoxic effects of 

ambient O3 can be suppressed due to CO2-induced reductions in stomatal conductance and O3 

uptake (Booker et al., 1997, 2004, 2005; Heagle et al., 1998b; Fiscus et al., 1997, 2002, 2005). 

Plume systems avoid chamber effects, but because they rely solely upon diffusion and 

natural turbulence to modify the ambient O3 concentration, they may fail to achieve homogeneity 

of the air to which the plants are exposed and may give rise to hot spots in which the enriched air 

jets are inadequately diluted and impinge directly on foliage. A further deterrent to their 

widespread use is the large-scale generation of O3 needed, which has, in most cases, limited the 

numbers of treatments that can be included in an experimental design. 

AX9-11
In spite of the various advantages and disadvantages of the two systems, there is still little 

experimental evidence that allows a direct comparison of OTCs to the free-air plume systems or 

a determination of the degree to which chamber effects alter plant response to O3. The evidence 

that is available suggests that chamber effects do not fundamentally alter the response of 

plants to O3; therefore, chambers remain a useful tool for testing species sensitivity and 

developing O3-response relationships. However, chamber effects have the potential to alter O3 

uptake (Nussbaum and Fuhrer, 2000), so it is important to fully characterize temperature, light, 

turbulence, and other chamber characteristics during exposure to allow extrapolation of the 

results. 

AX9.1.2.5 Ozone Generation Systems 

Two approaches have been used to generate the O3 needed for enrichment from air or 

oxygen: (1) high-voltage static discharge and (2) high-intensity UV-irradiation. Using gaseous 

oxygen as feedstock, both generate O3-enriched oxygen, free from other impurities. However, 

the use of high-voltage discharge equipment with air as feedstock requires that the output be 

scrubbed with water to remove appreciable amounts of the higher oxides of nitrogen (especially 

nitric acid vapor) that form concurrently with O3 (Brown and Roberts, 1988; Taylor et al., 1993). 

AX9.1.2.6 Experimental Exposure Protocols 

A few recent chamber studies of physiological or biochemical effects have continued to 

use square-wave exposure profiles typified by a rapid rise to, and falling off from, a steady target 

concentration. However, during the last 20 years, most approaches into studying O3 effects on 

plant growth and development have employed either simulations of the diurnal ambient O3 

profile or enhancement/reduction of the ambient O3 concentrations. 

Hogsett et al. (1985b), Lefohn et al. (1986), and others have described the use in controlled 

chambers of daily exposure profiles based on observed ambient O3 profiles. Such profiles were 

used in the elaborately controlled chamber studies of Langebartels et al. (1997), while several 

recent chamber studies have used simpler computer-controlled half- or full-cosine wave profiles 

to simulate the typical daily rise and fall in ambient O3 levels (Mazarura, 1997; McKee et al., 

1997a,b). 

AX9-12
The early studies with OTCs involved adding constant levels of O3 to ambient air O3 

concentrations, but all recent studies have used enrichment delivery systems that maintain 

proportionality to, and track, ambient O3 concentrations to produce levels that more closely 

resemble field observations. Both FACE and ZAPS studies have used proportional enrichment 

to provide a range of treatments, although Wilbourn et al. (1995) and Ollerenshaw et al. (1999) 

adjusted their systems manually to obtain a relatively constant target concentration during 

exposure episodes. 

AX9.1.3 Methods Involving Exposures to Ozone in Ambient Air 

The experimental methods discussed above are largely aimed at developing quantitative 

growth-response functions to permit the estimation of the effects of different ambient O3 

scenarios. Because such methodologies usually involve exposures to higher than ambient O3 

levels, the applicability of the functions obtained may, to some extent, be relevant only to 

locations that are naturally subjected to high ambient O3 levels. Furthermore, as pointed out in 

Section AX9.4, the response functions that they generate rarely incorporate other environmental, 

genetic, and physiological factors, many of which can severely modify the magnitude of the 

response to O3. The consequences of ignoring such modifications have been well stated by 

De Santis (1999). The European level for protecting crops (based on the AOT40 index; see 

Section AX9.4) was derived from OTC studies of O3-induced yield loss in wheat observed in 

experiments conducted mostly in non-Mediterranean locations. However, the impact of 

ambient O3 on wheat yields in the Po Valley of northern Italy is much less than the devastatingly 

high loss (>60%) suggested by the seasonal exceedances of the level. On a similar note, 

Manning (2003) has recently urged the absolute necessity of seeking “ground truth” as 

verification of the nature and magnitude of impacts on vegetation as suggested by response 

functions using ambient O3 monitoring data. 

Such concerns clearly show that attention needs to be focused on incorporating 

consideration of environmental and other factors into the response functions upon which 

standards are based. This will require the development of improved simulation response models. 

These concerns have also led to increasing attention being paid to seeking and developing 

alternative approaches to assessing of impact, and the geographic extent of such 

impact—approaches that are based on in situ exposures to ambient or sub-ambient O3 levels. 

AX9-13
Although one approach, the use of air-exclusion systems, requires experimental facilities, the 

other approaches are generally based on simple field observations or measurements and, hence, 

can be undertaken on a wide geographic scale. 

AX9.1.3.1 Air-Exclusion Systems 

The term, air-exclusion system, usually refers to a chamberless field system specifically 

designed to protect plants from exposure to polluted air by blowing filtered air through their 

canopies. Hogsett et al. (1987a,b) described several dedicated systems developed in the 1960s 

and 1970s, but there appear to have been no recent O3-exclusion studies using systems 

specifically designed for this purpose since those described by Olszyk et al. (1986a,b). Their 

system, a modification of the earlier system of Jones et al. (1977), consisted of perforated 31.8-

cm OD inflatable polyethylene tubes laid between crop rows and supplied with CF air. 

By increasing the size of the orifices progressively in sections along the 9-m length of the tubes, 

an exclusion gradient was created with a progressive decrease in O3 levels in the air surrounding 

the crop from one end of the system to the other. The system was used for studies on alfalfa 

comparing plant response in OTCs, closed field chambers, the air-exclusion system, and ambient 

air plots (as discussed above in Section AX9.1.2.4). 

An air-exclusion component has also been part of the overall design of many OTC 

experiments which added CF air or mixtures of CF and ambient air to chambers as part of the 

overall design. 

AX9.1.3.2 Natural Gradients 

Naturally occurring O3 gradients hold potential for the examination of plant responses 

along the gradient. However, few such gradients can be found which meet the rigorous 

statistical requirements for comparable site characteristics such as soil type, temperature, 

rainfall, radiation, and aspect (Manning and Krupa, 1992); although with small plants, soil 

variability can be avoided by the use of potted plants. The use of soil monoliths transported to 

various locations along natural O3 gradients is another possible approach to overcome 

differences in soils; however, again this approach is limited to small plants. 

Studies in the 1970s used the natural gradients occurring in southern California to assess 

yield losses of alfalfa and tomato (Lycopersicon esculentum L.) (Oshima et al., 1976, 1977). 

AX9-14
A transect study of the impact of O3 on the growth of white clover and barley (Hordeum vulgare 

L.) in the United Kingdom was confounded by differences in the concurrent gradients of SO2 

and NO2 pollution (Ashmore et al., 1988). Studies of forest tree species in national parks in the 

eastern United States (Winner et al., 1989) revealed increasing gradients of O3 and visible foliar 

injury with increased elevation. 

AX9.1.3.3 Use of Chemical Protectants 

The use of protective chemicals is a relatively inexpensive, promising alternative to 

experimental field exposures in chambers or free-air systems for determining plant response 

to O3. These chemicals have recently been used in studies of different plant species, both in the 

United States (Bergweiler and Manning, 1999; Kuehler and Flagler, 1999) and in Europe 

(Bortier et al., 2001a; Pleijel et al., 1999; Wu and Tiedemann, 2002), to determine if ambient O3 

concentrations affect plant growth and productivity or are just exacerbating foliar injury. 

Several chemical compounds (e.g., antioxidants, antisenescence agents, fungicides, pesticides) 

have been known for many years to provide plants some protection from photochemical oxidants 

such as O3 (Manning and Krupa, 1992). Most of these chemicals were originally used as a onetime application to reduce visible injury caused by acute O3 exposures. The most widely used 

and popular of these has been ethylenediurea (EDU). Carnahan et al. (1978) reported that EDU 

protected pinto bean (Phaseolus vulgaris) from acute O3 injury. After this initial investigation, 

EDU was shown to suppress visible O3 injury on several species of plants under both controlled 

and field conditions (Brennan et al., 1987; Clarke et al., 1983). However, due to lack of a 

commercial market for this product, its commercial manufacture was largely discontinued. 

Other chemicals, including benomyl (Manning et al., 1974), carboxin (Rich et al., 1974), 

ascorbic acid (Dass and Weaver, 1968), and others (Manning and Krupa, 1992), also exhibited 

some beneficial effects in reducing visible O3 injury. 

Several recent studies have used EDU in assessing the response of several plant species 

to O3 to help validate the proposed critical level (AOT40 = 3000 ppbAh; see Section AX9.5) for 

crop protection in Europe (Ball et al., 1998; Ribas and Penuelas, 2000; Tonneijck and Van Dijk, 

2002a,b). EDU appeared to provide protection from visible foliar injury, but the results 

regarding yield and biomass reductions were mixed. In a 3-year study over 12 sites throughout 

Europe, Ball et al. (1998) used the ratio of EDU-treated versus non-treated white clover biomass 

AX9-15
and did not find a significant relationship between biomass reductions and AOT40 level. 

However, an artificial neural network (ANN) model including vapor pressure deficit (VPD), 

temperature, longitude, year, and altitude explained much more of the variance (r2

 = 0.79). The 

authors suggested that the greater sensitivity at certain sites in Germany may have been due to 

occurrence of other pollutants. This meta-analysis indicates that EDU effects may be influenced 

substantially by environmental factors. 

In another study, Tonneijck and Van Dijk (2002b) assessed the relationship of visible 

injury of subterranean clover (Trifolium subterraneum) to ambient O3 at four sites over three 

growing seasons in the Netherlands, using EDU-treated and nontreated plants. Visible injury 

varied by site and year, but was reduced to near zero by EDU treatment. However, no 

relationship indicative of a protective effect of EDU with this plant species was observed for 

biomass. Tonneijck and Van Dijk (2002a) also reported similar results with pinto bean. Both 

EDU-treated and nontreated plants were exposed to ambient O3 at three locations in Spain over 

one growing season (Ribas and Penuelas, 2000). Reductions in yield and biomass were 

correlated with O3 concentration and EDU provided some protective effect, although results 

varied by location and with meteorological conditions. 

Chemicals have also been used to assess the effects of O3 on tree species. Bortier et al. 

(2001b) injected seedlings of an O3-sensitive poplar (Populus nigra) clone with EDU and 

measured growth over a 1-year period at a field site near Brussels, Belgium. Over the growing 

season, stem diameter increment was significantly higher (16%), biomass was increased (9%), 

and foliar O3 symptoms were slightly less for the EDU-treated seedlings. Ozone levels were 

reported to be low (AOT40 = 6170 ppbAh, May to September) during the exposure period. 

In another study, Manning et al. (2003) applied EDU (foliar spray) and sodium erythorbate 

(NaE) at various concentrations, biweekly for three growing seasons to loblolly pine (Pinus 

taeda) at a field site in east Texas. After 3 years, the trees were harvested and biomass 

measured. Neither EDU nor NaE prevented foliar O3 injury, but EDU applications at 450 ppm 

resulted in increases both in stem diameter and height and in total above-ground biomass. These 

measures of growth also tended to slightly increase with applications of NaE, but the effects 

were statistically nonsignificant. 

The mechanisms by which protective chemicals, especially EDU, protect plants are poorly 

understood. However, Wu and von Tiedemann (2002) reported that applications of two recently 

AX9-16
developed fungicides (azoxystrobin and epoxiconazole) provided protection to spring barley to 

relatively high O3 exposures (150 to 250 ppb, 5 days, 7 h/day) and resulted in increases in leaf 

soluble protein content as well as the activity of several antioxidative enzymes (e.g., superoxide 

dismutase, catalase, ascorbate-peroxidase, and glutathione reductase). In addition, the increase 

in these enzymes reduced superoxide levels in the leaves. 

Despite advances in the use of protective chemicals, a number of hurdles remain in using 

them for assessing O3 effects. The phytotoxicity of EDU is well known, and the point has been 

made repeatedly that for a particular species or cultivar, tests under a range of environmental 

conditions and O3 exposures must be made to establish the efficacy of EDU for quantifying O3 

effects (Heggestad, 1988; Kostka-Rick and Manning, 1992a). Unfortunately, although many 

studies with EDU have been conducted in recent decades, very few have used multiple EDU 

application levels along with multiple O3 exposures to characterize the EDU system for a given 

plant species. 

Recent studies have also shown that EDU does not always have greater effects at higher O3 

exposures. For pinto bean grown in pots in studies in Spain and in the Netherlands, EDU 

increased pod yield (Ribas and Penuelas, 2000; Tonneijck and Van Dijk, 1997). However, this 

effect was not greater at sites with higher O3 exposure despite consistent experimental protocols 

at all sites, including growing the same cultivar in pots with adequate water (Ribas and Penuelas, 

2000; Tonneijck and Van Dijk, 1997). Such results suggest that it may be difficult to quantify 

ambient O3 effects using EDU, because the amount of plant growth or yield expected at a low 

(background) O3 concentration cannot be inferred from EDU-treated plants grown at locations 

with higher O3 exposures. 

Several studies suggest that EDU has effects other than its antioxidant protection and 

phytotoxicity and show that environmental conditions affect the degree of protection afforded by 

protective chemicals. In one study, even low concentrations of EDU (8 to 32 mg L!1

 soil), 

decreased soybean yield under low O3 exposure (7-h mean of 19 ppb) in CF OTCs (Miller et al., 

1994). This study also demonstrated that phytotoxicity (both foliar symptoms and growth 

effects) can differ even in the same series of experiments, apparently due to changes in 

environmental conditions, and that EDU can suppress yield at application rates that do not 

always cause foliar symptoms (Miller et al., 1994). Finally, this study found that EDU altered 

biomass partitioning by increasing vegetative growth and decreasing reproductive growth. 

AX9-17
A study of pinto bean grown in OTCs in Germany found that EDU treatment in CF OTCs 

significantly increased yield, while EDU had no significant effect on yield in other O3 treatments 

(Brunschon-Harti et al., 1995). In this study, O3 significantly reduced the mass of pods, shoots, 

and roots. EDU increased root, leaf, and shoot mass across O3 treatments. However, the only 

statistically significant interaction occurred with O3 × root mass. This study indicates that EDU 

can stimulate above-ground growth and/or delay senescence regardless of O3 treatment. 

The EDU approach for assessing the impact of ambient O3 exposures is potentially useful, 

because it provides a separate line of evidence from other methods. Before using these 

chemicals in a field setting, preliminary investigations under controlled conditions (e.g., 

chambers) should be done to evaluate the methods and timing of application, as well as proper 

application rates, so as to avoid any potential toxic effects (Manning, 2000; Manning and Krupa, 

1992). Unfortunately, such characterization has so far been limited, although substantial 

progress has been made for radish (Raphanus sativus L.) (Kostka-Rick et al., 1993; Kostka-Rick 

and Manning, 1992a, 1993). Thus, it is difficult to use data from existing EDU studies to 

develop exposure-response relationships or to quantify the effects of ambient O3 exposure. 

Despite these limitations, the EDU studies reviewed in previous O3 AQCDs (U.S. Environmental 

Protection Agency, 1986, 1996) and the more recent studies summarized in Table AX9-1 

(Section AX9.5) provide another line of evidence that ambient O3 exposures occurring in many 

regions of the United States may be reducing the growth of crops and trees. 

AX9.1.3.4 Biomonitoring 

Bioindicators 

The use of biological indicators to detect the presence of O3 injury to plants is a 

longstanding and effective methodology (Chappelka and Samuelson, 1998; Manning and Krupa, 

1992). A bioindicator can be defined as a vascular or nonvascular plant exhibiting a typical and 

verifiable response when exposed to a plant stress such as an air pollutant (Manning et al., 2003). 

To be considered a good indicator species, plants must 

(1) exhibit a distinct, verified response, 

(2) have few or no confounding disease or pest problems, and 

(3) exhibit genetic stability. 

AX9-18 
Table AX9-1. Advantages and Disadvantages of Protective Chemicals Used in 

Assessment of O3 Effects on Plants 

Advantages

 No chambers required. Plants exposed to ambient conditions of O3, light, temperature, etc.

 Can conduct studies “in situ.” Equipment needs are minimal. No “chamber effects”

 A high degree of replication possible both within and among locations

Disadvantages

 Exposure-response studies require inclusion of other methodologies (OTCs, etc.)

 Need measurements of ambient O3 and other meteorological variables (temp, rainfall, etc)

 Many are toxic; have to conduct preliminary toxicology studies to determine proper rate, 

 timing etc.

 Species response can vary; need to screen for proper species to use

 Mode of action not fully understood; may alter growth and biomass partitioning

Sources: Manning and Krupa (1992); Heggestad (1988); Kostka-Rick and Manning (1992b);

Miller and Pursley (1994).

Such sensitive plants can be used to detect the presence of a specific air pollutant such 

as O3 in the ambient air at a specific location or region and, as a result of the magnitude of their 

response, provide unique information regarding specific ambient air quality. Bioindicators can 

be either introduced sentinels, such as the widely used tobacco (Nicotiana tabacum) variety 

Bel W3, or detectors, which are sensitive native plant species (e.g., milkweed [Asclepias 

syriaca]). The approach is especially useful in areas where O3 monitors are not operated 

(Manning et al., 2003). For example, in remote wilderness areas where instrument monitoring is 

generally not available, the use of bioindicator surveys in conjunction with the use of passive 

samplers (Krupa et al., 2001) is a particularly useful methodology (Manning et al., 2003). 

However, the method requires expertise or training in recognizing those signs and symptoms 

uniquely attributable to exposure to O3 as well as in their quantitative assessment. 

Since the 1996 O3 AQCD (U.S. Environmental Protection Agency, 1996), many new 

sensitive species have been identified from controlled exposure studies and verified in the field 

(Flagler, 1998; Innes et al., 2001). In addition, several new uses of this methodology have been 

demonstrated, including a national O3 bioindicator network, studies in wilderness areas, and 

AX9-19
mature tree studies. Although it has been difficult to find robust relationships between the foliar 

injury symptoms caused by O3 and effects on plant productivity or ecosystem function, visible 

injury correlations with growth responses have been reported (Table AX9-2) (1998) (2003) 

(2003). One workshop on the utility of bioindicators of air pollutants led to a useful series of 

peer-reviewed publications in Environmental Pollution (Skelly, 2003). 

Table AX9-2. Advantages and Disadvantages of Bioindicators Used to Study 

O3 Plant Effects 

Advantages

 No chambers required. Plants exposed to ambient conditions of O3, light, temperature, etc.

 Relatively inexpensive. Equipment needs are minimal. No “chamber effects”

 A high degree of replication possible (sentinels) both within and among locations

Disadvantages

 Results are generally correlative in nature with no true control

 Individuals need to be trained and experienced in O3 symptom recognition

 Need adequate numbers of plants (detectors) to ensure valid results

 Need preliminary tests to insure a constant symptomotology of material used

 Need to use more than one indicator species (detector) per area if possible

 Need to quantify site characteristics (soils, light) that may influence symptom expression

 Need measurements of ambient O3 (active or passive) and other meteorological variables

 (temp, rainfall, etc)

 Need to ensure that cultural (sentinels) practices (soil, irrigation, fertilization, etc.)

 are similar among sites

National network 

The U.S. Forest Service in cooperation with other federal and state agencies developed a 

network of O3 bioindicators to detect the presence of O3 in forested systems throughout the 

United States (Smith et al., 2003). This ongoing program was initiated in 1994; and 33 states 

currently participate. In a coordinated effort, a systematic grid system is used as the basis of plot 

AX9-20 
selection, and field crews are trained to evaluate O3 symptoms on sensitive plant species within 

the plots (Coulston et al., 2003; Smith et al., 2003). 

The network has provided evidence of O3 concentrations high enough to induce visible 

symptoms on sensitive vegetation. From repeated observations and measurements made over a 

number of years, specific patterns of areas experiencing visible O3 injury symptoms can be 

identified. Coulston et al. (2003) used information gathered over a 6-year period (1994 to 1999) 

from the network to identify several species that were sensitive to O3 over a regional scale 

including sweetgum (Liquidambar styraciflua), loblolly pine, and black cherry (Prunus 

serotina). 

Wilderness areas 

The use of bioindicator species as detectors has proven to be an effective technique for 

deriving a relative estimate of O3 injury in wilderness areas in both the United States and Europe 

(Chappelka et al., 1997, 2003; Manning et al., 2002). However, to be truly effective, these 

regional and national bioindicator studies need the inclusion of air quality data and related 

growth studies to determine effects on productivity and ecosystem function (Bytnerowicz et al., 

2002; Manning et al., 2003; Smith et al., 2003). In addition, O3 often co-occurs with other air 

borne pollutants, so it is important to consider that, in some areas, other pollutants may be 

playing a role as well. 

Chappelka et al. (1997, 2003) conducted surveys of foliar injury on several native plant 

species throughout the Great Smoky Mountains National Park (GRSM), including black cherry 

(Prunus serotina), tall milkweed (Asclepias exaltata), cutleaf coneflower (Rudbeckia laciniata), 

and crownbeard (Verbesina occidentalis). Visible foliar symptoms were prevalent throughout 

the Park, indicating that injury-producing O3 levels were widespread in GRSM. 

Manning et al. (2002) recently summarized a multiyear (1993 to 2000) bioindicator project 

in the Carpathian Mountain range in eastern Europe. They evaluated numerous trees, shrubs, 

forbs, and vines for possible symptoms of O3 injury. Observations were made at plots located in 

the vicinity of either active or passive O3 monitors (Bytnerowicz et al., 2002). Approximately 

30 species of native plant detectors were identified as possible bioindicators, the majority of 

which (21) were shrubs (Manning et al., 2002). Based on these observations, it was concluded 

that O3 concentrations were sufficiently high to impact ecosystems in the region. Similar 

AX9-21
investigations regarding the sensitivity of native species have been conducted in Switzerland 

(Novak et al., 2003) and Spain (Orendovici et al., 2003). 

Mature tree detectors 

Many studies have reported visible injury of mature coniferous trees caused by O3, 

primarily in the western United States (Arbaugh et al., 1998) and, to a lesser extent, to mature 

deciduous trees in eastern North America. In an effort to determine the extent and magnitude of 

visible injury in mature tree canopies, Hildebrand et al. (1996) and Chappelka et al. (1999b) 

conducted independent studies in the GRSM and the Shenandoah National Park (SHEN). 

The species examined were sassafras (Sassafras albidum), black cherry, and yellow-poplar 

(Liriodendron tulipifera) in GRSM and white ash (Fraxinus americana), black cherry, and 

yellow-poplar in SHEN. Protocols were similar at both parks, and trees were located near O3 

monitors at three different areas in each park. Results from both studies indicated that symptoms 

of O3 injury were present in the trees and correlated with O3 exposure both spatially and 

temporally. Ozone injury tended to be most severe at the highest elevation, except with 

yellow-poplar. 

Hildebrand et al. (1996) observed significant O3 exposure-plant response relationships with 

black cherry. The best relationships were found between foliar injury and the cumulative 

exposure statistics SUM06 and W126 (see Section AX9.5), indicating that higher O3 

concentrations were important in eliciting a response in black cherry. No O3 exposure-plant 

response relationships were found with any species tested in GRSM (Chappelka et al., 1999b); 

but, when the data were combined for both parks, a significant correlation (r = 0.72) with black 

cherry was found for both SUM06 and W126, and injury was the greatest (r = 0.87) at the higher 

elevations (Chappelka et al., 1999a). 

Based on a study in which visible symptoms of O3 injury were characterized for large, 

mature yellow-poplar and black cherry trees in GRSM (Chappelka et al., 1999a), Somers et al. 

(1998) compared radial growth differences among trees classified as sensitive or nonsensitive 

based on the severity of visible foliar injury observed over a 3-year period (1991 to 1993). 

Significantly more radial growth was observed over both a 5- and a 10-year period for the 

nonsensitive compared to the sensitive trees. No significant relationship was found for black 

cherry tree growth. 

AX9-22
Vollenweider et al. (2003a), using data collected from continuous forest inventory (CFI) 

plots across Massachusetts, compared growth rates among either symptomatic or asymptomatic 

mature black cherry trees. Of the 120 trees sampled in 1996, 47% exhibited visible foliar injury. 

Using CFI data, growth rates were compared over a 31-year period. The growth rates for 

symptomatic trees were reduced by 28% compared with asymptomatic trees. 

Because these studies (Somers et al., 1998; Vollenweider et al., 2003a) were not controlled 

studies and used a small sample of trees, they cannot validly be used to characterize cause and 

effects related to the visible symptoms and radial growth they describe. However, the results 

indicate the possibility that O3 is correlated with growth losses in some sensitive genotypes, 

illustrating the potential usefulness of this visible O3 injury methodology in assessing effects on 

the growth rates of mature deciduous trees. 

Cultivar comparisons 

The idea of using cultivars or isogenic lines of crop species that differed in O3 sensitivity as 

sentinels to determine the ambient effects of O3 in the field was presented in the 1996 O3 AQCD 

(U.S. Environmental Protection Agency, 1996). The rationale was that comparing the ratio of 

injury scores or some measure of growth between two different cultivars varying in O3 

sensitivity should be indicative of the relative amount of ambient stress to plants at a given 

location. A sensitive:resistant ratio close to unity would indicate relatively low O3 

concentrations; and a low ratio, higher O3 levels. Results from locations differing in O3 

concentrations could be evaluated to develop exposure-response models. The original protocol 

was derived using two isogenic lines of white clover (Trifolium repens) differing in O3 

sensitivity (Heagle et al., 1994b, 1995). 

This white clover model system has been used in several multi-location studies in the 

United States (Heagle and Stefanski, 2000) and Europe (Ball et al., 2000; Bermejo et al., 2002; 

Mills et al., 2000). Heagle and Stefanski (2000) compared results from eight sites over a 2-year 

period with various exposure indices (SUM00, SUM06, W126, and others) to determine a 

best-fit regression. They found that most of the indices preformed similarly. The highest r2 

values (0.87 to 0.93) were obtained using only the later harvests and a 6 h day-1 index (1000 to 

1600 h). Similar multiple-comparison studies conducted in Europe using the AOT40 index (Ball 

et al., 2000; Mills et al., 2000) yielded poorer r2

 values. Factors such as air temperature, NOx 

AX9-23
(high levels at some sites), and lower O3 concentrations in Europe were suggested to account in 

part for the differences between U.S. and European study results. Bermejo et al. (2002), in a 

study in Spain, improved the model by comparing the biomass ratio of these white clover 

isolines to measures of O3 uptake (flux) rather than an exposure index (AOT40). Together, these 

studies indicate that systems such as the white clover model can help reveal O3 exposureresponse relationships and provide valuable information regarding ambient O3 conditions in a 

given location. Table AX9-3 lists the advantages and disadvantages of the use of cultivar 

comparisons in assessing O3 effects of plants. 

Table AX9-3. Advantages and Disadvantages of Cultivar Comparisons Used in 

Assessment of O3 Effects on Plants 

Advantages

 No chambers required. Plants exposed to ambient conditions of O3, light, temperature, etc.

 Relatively inexpensive. Equipment needs are minimal. No “chamber effects”

 A high degree of replication possible both within and among locations

 Can conduct studies “in situ”

Disadvantages

 Need preliminary tests to insure sensitivity and growth patterns of genotypes used are consistent

 Need measurements of ambient O3 and other meteorological variables (temp, rainfall, etc)

 Have to ensure cultural practices (soil, irrigation, fertilization, etc.) are similar among sites

 Need to closely monitor plants for presence of other factors that may cause a misinterpretation

 of results 

Dendrochronological techniques 

It has been difficult to determine whether O3 significantly affects tree growth and 

productivity in the field, because O3 concentrations are omnipresent and tree response to this 

pollutant is altered by many factors. The use of dendrochronological techniques to answer 

questions regarding ambient O3 effects on forest growth and ecosystem function has recently 

emerged as a very useful biomonitoring methodology (Cook, 1990; McLaughlin et al., 2002). 

AX9-24
The technique is useful when either instrument or passive O3 monitoring methods are used to 

determine ambient O3 conditions. 

Initial experiments were primarily correlative in nature and attempted to relate symptoms 

of visible injury with growth losses as revealed by tree ring analysis (Arbaugh et al., 1998; 

Benoit et al., 1983; Peterson et al., 1995; Somers et al., 1998; Swank and Vose, 1990). These 

studies evaluated radial growth patterns determined by cores removed from trees in the presence 

or absence of overt O3 injury symptoms. 

The method has also been adapted to better understand forest ecosystem function 

(McLaughlin and Downing, 1995; Bartholomay et al., 1997; McLaughlin et al., 2003). The 

response of mature loblolly pine growing in eastern Tennessee to ambient O3 and moisture stress 

was evaluated by McLaughlin and Downing (1995, 1996). They made radial growth 

measurements from 12 to 37 times per year using dendrometer bands and determined 

relationships between O3, moisture stress, and radial growth. Exposures to O3 concentrations 

$0.04 ppm with high temperatures and low soil moisture resulted in short-term depression in 

radial growth. Reductions in growth were estimated to vary from 0 to 15% per year and 

averaged approximately 5% per year. 

Bartholomay et al. (1997) examined white pine (Pinus strobus) radial growth in eight 

stands throughout Acadia National Park, Maine over a 10-year period from 1983 to 1992. They 

related growth rates to several factors, including O3 concentration. Ozone levels were negatively 

correlated with radial growth in seven of the eight stands. Site characteristics were important in 

the relationship: stands growing on shallow, poorly drained soils were most sensitive to O3 in 

the late portion of the growing season, possibly due to premature senescence of foliage. 

However, litterfall measurements were not reported. Trees growing on better sites were more 

sensitive to O3 during the entire growing season, indicating the possibility of high O3 uptake 

rates throughout the growing season. Although these field studies (Bartholomay et al., 1997, 

1996; McLaughlin and Downing, 1995) did not compare the direct effects of O3 on the two pine 

species, they indicate that potential interactions exist among O3 and other climatic and edaphic 

factors, such as temperature and soil moisture. 

Using both automated and manual dendrometer bands, McLaughlin et al. (2002) examined 

the growth response of yellow-poplar trees recently released from competition. In addition to 

measuring growth, sap flow measurements were conducted and soil moisture was measured in 

AX9-25
the vicinity of the trees. They were not able to detect O3 effects in this 1-year study. Advantages 

and disadvantages of dendrochronology techniques for evaluating whole-tree physiological 

responses for individual trees and forest stands are listed in Table AX9-4. 

Table AX9-4. Advantages and Disadvantages of Various Dendrochronological 

Techniques Used in Assessment of O3 Effects on Plants 

Advantages

 Provide information regarding growth effects under ambient conditions

 Good historical information regarding O3 effects

 Can provide data on daily and seasonal growth and O3 patterns and correlate with 

 physiological function

 Provide information on forest function related to ambient O3 concentrations

 Can link data with process-level growth models 

Disadvantages

 Results are generally correlative in nature with no true control

 Need background O3 and meteorological data (historical records) 

 Need to account for other factors such as competition, in analyzing data

 Individuals need to be trained in counting growth rings

 Replication can be difficult (expensive and technological limitations)

 Complicated statistical analyses are sometimes required

 Can be expensive, especially if using automated growth (dendrometer) bands 

The use and evolution of various dendrochronological methods in the field of air pollution 

effects research is reviewed in detail by McLaughlin et al. (2002). Automated dendrometer 

bands provide a powerful tool for measuring radial growth responses of trees on an hourly or 

daily basis. Diurnal patterns of growth can be related to water use and O3 concentrations using 

time-series analyses. The major drawbacks of the method are that it is expensive and time 

consuming. 

AX9-26
AX9.1.3.5 Calibrated Passive Monitors 

Many studies have used passive monitors in the mapping of ambient O3 concentrations, 

especially in remote areas (Cox and Malcolm, 1999; Grosjean et al., 1995; Krupa et al., 2001). 

Because they are cumulative recording devices, they do not record short-term variations in O3 

concentration but only the total exposure over a given interval, usually between 1 to 4 weeks. 

Thus, they produce a measurement that resembles the instrumentally derived exposure index 

SUM00. However, it is common to divide the cumulative exposure by the number of hours of 

exposure to get an hourly average. In addition, Krupa et al. (2001, 2003) were able to estimate 

the underlying frequency distribution of hourly O3 concentrations from passive samplers using 

models based on a collocated O3 monitor, showing the potential for passive samplers to provide 

estimates beyond total O3 sum. 

Runeckles and Bowen (2000) used the ZAPS system described in Section AX9.2.2.3 to 

subject both crops and passive monitors (Williams, 1994) to a range of exposures. Passive 

monitors were also exposed at 16 agricultural field sites along a transect through the Fraser 

Valley, British Columbia, Canada. Most field sites were downwind of the Greater Vancouver 

metropolitan area. All passive monitors were replaced at weekly intervals and the data from 

those in the ZAPS plots were “calibrated” to crop responses by means of Weibull exposureresponse functions. Since the meteorological conditions throughout the valley were reasonably 

consistent from site to site, the use of these functions with data from the network passive 

monitors as inputs permitted the estimation of crop losses at the network sites. The overall 

method was, thus, a hybrid of several methodologies. 

Although based on a single study, the use of passive monitors has potential for assessing 

crop losses at sites removed from locations with known ambient O3 concentrations. Provided 

that the network and calibration sites have similar meteorological conditions, the method yields 

crop loss estimates that are responses to local ambient O3 levels as influenced by local 

meteorological conditions. 

AX9.1.4 Numerical/Statistical Methodologies 

Proper experimental design strategies including replication, randomization, and 

experimental protocols are paramount in O3-effects research. These have been discussed in 

detail in previous O3 AQCDs (U.S. Environmental Protection Agency, 1996, 1986), as have the 

AX9-27
different statistical analytical procedures used to determine the probable significance of results. 

However, new investigative approaches have demanded the adoption of new analytical methods. 

For example, the use of dendrochronological techniques has led to the use of time-series analysis 

(McLaughlin et al., 2003) and linear aggregate models (Cook [1990], as reviewed by 

McLaughlin et al. [2002]). 

In spite of the rigors of the analyses, many differences occur in the published literature for 

almost any plant response to O3 stress. Differences inevitably result from different researchers 

studying different locations, using different experimental methodologies and genetically 

different plant material even when using a common species. The techniques of meta-analysis 

can be used to consolidate and extract a summary of significant responses from a selection of 

such data. 

Despite the differences in responses in the 53 primary studies used, a recent meta-analysis 

by Morgan et al. (2003) of the effects of O3 on photosynthesis, growth, and yield of soybean 

showed “overwhelming evidence for a significant decrease in photosynthesis, dry matter 

production and yield. . . .across all the reported studies on effects of chronic O3 treatment.” The 

meta-analysis defined O3 stress as exposure to ~70 ppb O3 for at least 7 days and found average 

shoot biomass and seed yield decreases of 34% and 24%, respectively. Furthermore, although 

other stress factors such as drought and UV-B did not affect the O3 responses, elevated CO2 was 

found to significantly decrease O3-induced losses. 

The meta-analysis method clearly has the potential to consolidate and refine the 

quantitative exposure-response models for many species. The majority of the reported growth 

and physiological responses related to O3 stress are for individual plants, primarily in various 

types of exposure chambers. It is difficult to extrapolate these responses to stand/community, 

ecosystem, or region-wide assessments, particularly in view of the importance of the significant 

interactions that may occur between plant responses O3 and other environmental stresses. Along 

with the shift in effects research to a more ecological approach, these concerns necessitate a 

move from simple regression analysis to more complex mathematical approaches to handle a 

wider array of independent input variables than O3 exposure alone. Other independent input 

variables that must be accounted for include air and soil temperatures, soil moisture, relative 

humidity, wind speed, and, particularly in the case of natural systems, biotic factors such as pests 

and pathogens, plant density/spacing, and measures of plant competition. 

AX9-28
Artificial neural network methodology was used by Balls et al. (1995) for “unraveling the 

complex interactions between microclimate, ozone dose, and ozone injury in clover” and in the 

study with the protectant chemical EDU, discussed in Section AX9.2.3.3 (Ball et al., 1998). The 

multi-factor model for predicting the effects of ambient O3 on white clover developed by Mills 

et al. (2000) utilized both ANN and multiple linear regression methods. 

Models incorporating ANNs are of the “regression” type (Luxmoore, 1988) in contrast to 

“mechanistic” or “phenomenological” models which have wider applicability. Process-level 

models of either type have been developed at the organelle, individual plant (Constable and 

Taylor, 1997; Weinstein et al., 1998), canopy (Amthor et al., 1994), and stand level (Ollinger 

et al., 1997; Weinstein et al., 2001) and provide estimates of the rate of change of response 

variables as affected by O3 over time. However, as pointed out in the 1996 O3 AQCD (U.S. 

Environmental Protection Agency, 1996), mechanistic process models lack the precision of 

regression models as well as their ability to estimate the likelihoods of responses. In their 

extensive reviews, Kickert and Krupa (1991) and Kickert et al. (1999) summarized the 

advantages and shortcomings of many different models and made the important point that most 

of the models that have been described provide consequence assessments that quantify the 

magnitudes of effects, but not risk assessments that quantify the likelihoods of such effects. 

Descriptions of several specific models are provided in other sections of this criteria document, 

and advantages and disadvantages of modeling techniques used in assessing O3 effects on plants 

are summarized in Table AX9-5. 

Table AX9-5. Advantages and Disadvantages of Modeling Techniques Used in 

Assessment of O3 Effects on Plants 

Advantages 

C Provide an understanding of cause-effect relationships over time 

Disadvantages 

C Have to make assumptions based on a paucity of data 

C Most models are very complex and difficult to understand 

C Need to be evaluated for predictive validity 

AX9-29
AX9.1.5 Improved Methods for Defining Exposure 

Ambient air quality is defined in terms of the measured O3 concentrations in the air at plant 

height above ground level. Compilations of such concentration data have long been used as 

surrogates of the exposures to which plants are subjected. However, as long ago as 1965, field 

research provided evidence that plant response was a function, not of ambient O3 concentration 

per se, but of the estimated flux of O3 to the plant canopy (Mukammal, 1965). Subsequently, 

Runeckles (1974) introduced the term “effective dose” to define that part of the ambient 

exposure that was taken up by a plant. Fowler and Cape (1982) later referred it as “pollutant 

applied dose” (PAD), defined as the product of concentration, time and stomatal (or canopy) 

conductance, with units g m!2

. Such estimates of O3 uptake or flux provide a more biologically 

relevant description of exposure than the simple product of concentration and time alone, and 

they formed the basis of Reich’s 1983 “unifying theory” of plant response to O3 (Reich, 1983). 

However, it was not until the early 1990s that the inherent advantages of using O3 flux 

rather than O3 concentration as a basis for determining response effects began to be widely 

accepted, as demonstrated by the subsequent increase in publications involving flux 

measurements and modeling (e.g., Fuhrer et al. [1997]; Grünhage and Jäger [2003)]; Grünhage 

et al. [1993; 1997]; Massman et al. [2000)]; Musselman and Massman [1999]; Pleijel [1998]). 

A key requirement for flux determination is the measurement of stomatal or canopy 

conductances, using established porometer/cuvette techniques or eddy correlation methods. 

The usefulness and relevance of flux as a measure of exposure are discussed in detail in Section 

AX9.4. 

Efforts to develop regional-scale models of O3 deposition and stomatal uptake are currently 

under way with a view to providing improved assessments of the risks to vegetation across 

Europe (Emberson et al., 2000; Simpson et al., 2001, 2003). 

AX9.2 SPECIES RESPONSE/MODE-OF-ACTION 

AX9.2.1 Introduction 

The evaluation of O3 risk to vegetation requires fundamental understanding of both the 

functioning of the vegetation and how external environmental influences can alter that function. 

For biological organisms subjected to atmospheric O3, those alterations can be complex and 

AX9-30
multiple. In addition, biological organisms have plasticity to external interactions due to their 

complex internal, self-correcting systems, making the task of identifying their “correct” 

functioning difficult. This section emphasizes reactions of O3 with the cell and tissue, rather 

than the whole plant, to describe the fundamental mechanisms known to govern the response of 

the plant to O3 exposure. 

The many regulatory systems contained in leaves change both as a function of leaf 

development and in response to various environmental stresses. Leaves function as the major 

regulators of anatomical and morphological development of the shoot and control the 

translocation of carbohydrates to the whole plant (Dickson and Isebrands, 1991). This section 

discusses the movement of O3 into plant leaves and their biochemical and physiological 

responses to O3. 

The 1996 criteria document (U.S. Environmental Protection Agency, 1996) assessed the 

information available at that time concerning the biochemical and physiological responses to the 

movement of O3 into plant leaves. This information continues to be valid. Ozone uptake in a 

plant canopy is a complex process involving adsorption to surfaces (leaves, stems, and soil) and 

absorption into leaves (Figure AX9-1). However, the initial biochemical changes that result 

within leaf cells after the entry of O3 and how these changes interact to produce plant responses 

remain unclear. The response of vascular plants to O3 may be viewed as the culmination of a 

sequence of physical, biochemical, and physiological events. Only the O3 that diffuses into a 

plant through the stomata (which exert some control on O3 uptake) to the active sites within a 

leaf impairs plant processes or performance. An effect will occur only if sufficient amounts 

of O3 reach sensitive cellular sites that are subject to the various physiological and biochemical 

controls within the leaf cells. Ozone injury will not occur if (1) the rate and amount of O3 uptake 

is small enough for the plant to detoxify or metabolize O3 or its metabolites or (2) the plant is 

able to repair or compensate for the O3 impacts (Tingey and Taylor, 1982; U.S. Environmental 

Protection Agency, 1996). Therefore, a precondition for O3 to affect plant function is that it 

must enter the stomata and be absorbed into the water lining the mesophyll cell walls. The 

response of each plant is determined by the amount of O3 entering the leaves, which varies from 

leaf to leaf. 

Some potentially significant processes have been investigated since the 1996 criteria 

document, especially detoxification and compensatory processes. The role of detoxification in 

AX9-31
Figure AX9-1. Ozone uptake from the atmosphere. Ozone moves from the atmosphere 

above the canopy boundary layer into the canopy primarily by turbulent 

air flow. Canopy conductance, controlled by the complexity of the canopy 

architecture, is a measure of the ease with which gases move into the 

canopy. Within the canopy, O3 is adsorbed onto surfaces as well as being 

absorbed into the foliage. Foliage absorption is controlled by two 

conductances, leaf boundary layer and stomatal, which together determine 

leaf conductance. The solid black arrows denote O3 flow; dotted arrows 

indicate processes affecting uptake or response to O3. Boxes at the left 

with double borders are those processes described in the figure. 

providing a level of resistance to O3 has been investigated; however, it is still not clear as to what 

extent detoxification can protect against O3 injury. Data are needed especially on the potential 

rates of antioxidant production and on the subcellular localization of the antioxidants. Potential 

rates of antioxidant production are needed to assess whether they are sufficient to detoxify the O3 

as it enters the cell. The subcellular location(s) is needed to assess whether the antioxidants are 

in cell wall or plasmalemma locations that permit contact with the O3 before it has a chance to 

damage subcellular systems. Although these processes divert resources away from other sinks, 

AX9-32
detoxification and compensation processes may counteract the reduction in canopy carbon 

fixation caused by O3. The quantitative importance of these processes requires investigation. 

As a result of the research since the 1996 criteria document (U.S. Environmental Protection 

Agency, 1996), the way in which O3 exposure reduces photosynthesis, especially its effects on 

the central carboxylating enzyme, Rubisco (ribulose-1,5-bisphosphate/carboxylase), is 

better understood. The rate of leaf senescence has been shown to increase as a function of 

increasing O3 exposure. The mechanism of the increased senescence is not known, and, hence, 

it deserves further study. 

Finally, the role that changes in allocation of resources play in plant response to O3 is now 

better understood. Most studies have shown that O3 decreases allocation of photosynthate to 

roots. In some cases, allocation to leaf production has increased. Whether these changes are 

driven entirely by changes in carbohydrate availability or are controlled by other factors (e.g., 

hormones) is not known. Physiological effects within the leaves inhibit photosynthesis; alter the 

assimilation of photosynthate and shift its allocation patterns; and can lead to reduced biomass 

production, growth, and yield (U.S. Environmental Protection Agency, 1986, 1996). 

The major problem facing researchers trying to predict long-term O3 effects on plants is 

determining how plants integrate the responses to O3 exposures into the overall naturally 

occurring responses to environmental stressors. Little is now known about how plant responses 

to O3 exposures change with increasing age and size, but this information is crucial to predicting 

the long-term consequence of O3 exposure in forested ecosystems. 

This section focuses on reactions of O3 within cells and cellular tissue, in order to explain 

known mechanisms that govern plant responses. The processes that occur at cell and tissue 

levels within the leaf will be divided into several steps beginning with O3 uptake and its initial 

chemical transformations into a series of currently unknown, but suspected toxic, chemicals 

(Figure AX9-2). The discussion will then focus sequentially upon various cell regions, their 

general physiology, and the changes that may occur within a plant after O3 exposure. This is 

important because the varying responses of the different plant species in a community ultimately 

lead to an ecosystem response. Finally, a general summary is presented that discusses the known 

or suspected changes that occur within the whole plant. 

AX9-33
Figure AX9-2. Absorption and transformation of O3 within the leaf. The varied 

processes are broken down in to smaller mechanistic steps that lead from 

uptake of atmospheric O3 into the alterations which may occur within the 

individual plant. Each plant responds to the O3 level and therefore 

interacts with the total ecological setting to generate an ecosystem 

response due to the O3. 

AX9.2.2 Mechanisms of Ozone-Induced Plant Alterations 

Plants can survive O3 stress through exclusion or tolerance mechanisms (Levitt, 1972) 

(Tingey and Taylor, 1982). Ozone may be excluded from tissues or cells via stomatal closure, 

by extracellular oxidants, or by membrane impermeability to O3 or its products. Past 

investigations of O3 injury have indicated that physiological and metabolic changes occur (Harris 

and Bailey-Serres, 1994; Heath, 1988; Heath and Taylor, 1997; Reddy et al., 1993). Many of 

these changes are likely initiated via gene expression. During the last decade, our understanding 

of the cellular processes within plants has increased. Although the fundamental hypotheses 

concerning O3-induced changes in physiology have not changed, a more complete development 

of the theories is now nearing possibility. 

AX9-34
AX9.2.2.1 Changes in Metabolic Processes: Current Theories 

The current hypotheses regarding the biochemical response to O3 exposure revolve about 

injury and its prevention. These are well discussed by Pell et al. (1997) and are listed below in 

no order of importance. Although they are listed separately, some may be interlinked and related 

to each other. 

(1) Membrane Dysfunction. The membrane is altered by O3, principally via protein 

changes not involving the lipid portions of the membrane (except at extremely high 

levels of O3). These alterations involve increased permeability with perhaps lessened 

selectivity, declines in active transport, and changes in the trigger mechanisms of 

signal transduction pathways such that the signals are no longer suitable for the state 

of the cell. The cellular pools and transport systems of Ca2+/K+

/H+ are the primary 

suspects. 

(2) Antioxidant Protectants. Varied antioxidants (both as metabolites and enzyme 

systems) can eliminate the oxidant or its products, if present at time of fumigation 

and in sufficient abundance. However, oxidant entry that occurs rapidly can 

overwhelm the antioxidant response. 

(3) Stress Ethylene Interactions. Visible injury is caused by the interaction of O3 with 

stress-induced ethylene, either by direct chemical transformation to a toxic product or 

by alteration of the biochemical relations at the ethylene binding site. 

(4) Impairment of Photosynthesis. A product of O3 (and less probably, O3 itself) enters 

the cell, causing a decline in the mRNA for Rubisco (especially the message RNA 

species of rbcS and rbcL) such that Rubisco levels slowly decline within the 

chloroplast, leading to a lowered rate of CO2 fixation and productivity. This process 

is very similar to early senescence and may be linked to general senescence. 

Alternatively, a false signal is generated at the cell membrane which lowers the 

transcription of DNA to mRNA. Ozone alters the normal ionic and water relations of 

guard cells and subsidiary cells, causing the stomata to close and limit CO2 fixation. 

In any case, the response of the stomata to the current environment does not promote 

efficient photosynthesis. 

(5) Translocation Disruption. One of the biochemical systems most sensitive to O3 

exposure is the translocation of sugars, such that even a mild exposure inhibits the 

translocation of carbohydrate (Grantz and Farrar, 1999, 2000). 

(6) General Impairment/Disruption of Varied Pathways of Metabolism. This is the 

oldest and most vague concept of how O3 alters metabolism. It is based upon early 

work in which the enzymes and metabolites that could be assayed were. Thus, these 

results were based upon what could be done, rather than on a coherent hypothesis. 

The best examples are listed in Dugger and Ting (1970). 

AX9-35
The latter two theories can be restated as a loss of productivity with three possible 

somewhat-independent causes: (a) a reduced production of the basic building blocks of growth 

and, hence, a slowing of growth in at least one organ; (b) a reduced ability to reproduce, leading 

to a decreased production of viable seeds or of fruits and nuts; and (c) a decreased ability to 

mount a defense against pathogens or insects, leading to weaker plants, which are more liable to 

be overcome by other stresses. It is important to separate out effects that may be detrimental or 

disfiguring, such as the production of visible injury, but which have not been shown to lead 

directly to a loss of productivity due to possible compensation by the remaining tissue. 

AX9.2.2.2 Modifications of Plant Physiological Processes 

The discussion that follows will focus on physiological processes rather than on speciesspecific responses; in most cases, the mechanisms of response are similar regardless of the 

degree of sensitivity of the species. Therefore, Arabidopsis, whose physiology and genome 

continue to be studied and described by a large number of scientists is an appropriate model 

plant for studying O3 injury. Though the responses of mature trees and understory plants are 

critical to understanding plant interactions at an ecosystem level, the time required for trees to 

reach maturity makes using them to study biological mechanisms an inefficient choice. 

The high levels of O3 used for some investigations do not automatically invalidate the 

results obtained in those studies. Typically when a new hypothesis is being investigated, 

extreme levels of the toxicant are used to determine its effects clearly. The older studies that 

used concentrations as high 1 ppm, an extreme level, helped to define current studies. Later 

experiments have used concentrations nearer ambient levels. Many of the current studies on 

physiology use exposures between 0.15 and 0.25 ppm, which though higher than ambient levels 

in some areas of the country, bypass confounding changes but allow for rapid experiments. 

Three forms of air pollutant-induced injury patterns are currently known to exist: (a) acute 

stress, generated by high atmospheric concentrations of pollutants for short periods of time; 

(b) chronic stress, generated by lower concentrations of pollutants for long periods of time; and 

(c) accelerated senescence, generated by very low concentrations of pollutants for very long

periods of time. At higher levels, distinct visible injury generally occurs due to cellular and 

tissue death of regions of leaf mesophyll cells. This leads to a decline in the total area of 

metabolically active tissue, with consequent loss of membrane integrity, loss of metabolites into 

AX9-36
the extracellular tissue space, and formation of oxidative products. When no visible injury is 

observed, lowered rates of photosynthesis or productivity are often used to document injury. 

Under these conditions, metabolism is altered and the pool sizes of many metabolites are 

changed. More importantly, the altered biochemical states within the tissue lead to the inability 

of the plant to respond properly to existing environmental conditions and to other stressors 

(Heath, 1988, 1994b; Koziol and Whatley, 1984; Manning and Keane, 1988; Schulte-Hosted 

et al., 1988). 

AX9.2.3 Ozone Uptake by Leaves 

Plants respond to O3 similarly to other stressors on the levels of exclusion, tolerance, and 

repair (Levitt, 1972). The response mechanism depends upon the O3 concentration, 

environmental conditions, and the developmental and metabolic state of the plant (Guzy and 

Heath, 1993). These responses are detrimental to plant productivity, because they cost the plant 

metabolic resources. In some cases, the stomata close under the O3 exposure, excluding the 

pollutant from the leaf interior and preventing injury. However, if this happens too often, CO2 

fixation is also inhibited and plant productivity suffers. 

Atmospheric O3 does not cause injury, but rather it is the O3 that enters the plant that 

causes an effect (Guzy and Heath, 1993; Tingey and Taylor, 1982). Three well-defined, 

sequential processes control the movement of O3 from the atmosphere into the sites of action 

within the leaf and must occur to trigger O3 stress (Heath, 1980). The processes are (1) entry of 

O3 into the leaf, (2) reactions of O3 and its possible reaction product(s) in the water phase at cell 

surfaces, and (3) movement of an O3 reaction product(s) into the cell with enzymatic or chemical 

transformation of those products in the cell. 

Process 1. Entry of O3 into the leaf. Often incorrectly, the external concentration of O3 is 

used to give an indication of “dose” (Heath, 1994a). Ozone-induced changes on a plant’s cuticle 

are minimal, and O3 does not penetrate the cuticle (Kerstiens and Lendzian, 1989) to cause an 

effect. As O3 has no easily measured isotope, virtually no measurements have been done on an 

actual dose of O3, i.e., the amount of O3 which reacts with individual biochemicals in the leaf. 

Yet the measurement of dose will be the amount of O3 expected to penetrate into the tissue 

through the stomata. Dose is expressed as a rate of delivery to a surface area (mol/m2

 s!1). 

AX9-37
Whether dose or total accumulation (mole/m2

, rate integrated over exposure time) is most critical 

for the development of injury remains a major question. 

Ozone uptake includes gaseous diffusion through the leaf boundary layer and stomata into 

the substomatal cavity (Figure AX9-3). Although the movement of pollutants through a 

boundary layer into the stomata region is known to be important, and even rate limiting in many 

cases of low wind velocity, its description has been defined from aeronautical concepts and 

usually relates to smooth surfaces that are not typical of leaf-surface morphology; however, it is 

nearly the only treatment available (Gates, 1968). Once through the boundary layer, the gas 

must enter the leaf through the stomata. The entry of gases into a leaf is dependent upon the 

physical and chemical processes of gas phase and surfaces and is a well-defined path that 

approximately follows a linear flux law of: 

(AX9-1) 

where the flux, j, into the internal space of a leaf is related to the conductance, g, through the 

boundary layer and stomata and the gradient of concentration of gas from the outside, Co, 

inward, Ci

. This formulation has been used for years for both water and CO2 (Figure AX9-4), 

and for regions of varied CO2 concentration that correspond to Co (CO2 of the atmospheric air, 

below the leaf proper) and Ci

 (CO2 near the leaf’s spongy mesophyll cells) (Ball, 1987; Farquhar 

and Sharkey, 1982). 

Figure AX9-3. The uptake of O3 into the leaf. Each of the individual concentration 

layers of O3 represents a different process of movement and of 

plant/microenvironmental interaction. This figure leads into Table 9-6, 

in which the amounts of O3 along the pathway are calculated. 

AX9-38
Figure AX9-4. The microarchitecture of a dicot leaf. While details among species vary, 

the general overview remains the same. Light that drives photosynthesis 

generally falls upon the upper (adaxial) leaf surface. Carbon dioxide and 

O3 enters through the stomata on the lower (abaxial) leaf surface, while 

water vapor exits through the stomata (transpiration). 

In the past, the internal concentration of O3 has been assumed to be zero (Laisk et al., 

1989), due to early studies that found that virtually no O3 could pass through a leaf. That was 

expected because O3 is extremely reactive with cellular biochemicals. If the assumption that the 

internal concentration zero is correct, then the effective delivery rate for O3 is given as g × Co, 

with stomatal conductance being the major regulatory control (Amiro et al., 1984; Taylor et al., 

1982). However, a recent study by Moldau and Bichele (2002) indicated that the internal O3 

concentration may not be zero as previous assumed. Moldau and Bichele (2002) permitted 

leaves of Phaseolus vulgaris L., which have stomata on both upper and lower leaf surfaces, to 

take up O3 at a high rate for 3 to 5 min. Exposure of the lower leaf surface resulted in up to 5% 

of the O3 that was taken up to be diffused through the leaf, emerging from the stomata on the 

upper surface. This suggested the presence of above-zero concentrations of O3 in the 

intercellular leaf air spaces. The descriptive calculations and plots of Moldau and Bichele 

(2002) indicate that the rise in internal O3 level (from both sources of external O3 concentration) 

within the first few minutes of exposure is due to its reaction with an antioxidant, most probably 

absorbate, within the apoplastic space of the leaf (Figure AX9-5). The rate of rise is probably 

AX9-39
Figure AX9-5. The change in the O3 concentration inside a leaf with time. Data are from 

O3 exposures at two different concentrations. 

Source: Derived from data in Moldau and Bichele (2002). 

due to more complete penetration of O3 with a concurrent depletion of the external antioxidant. 

The rise peaks at about 2 min for 0.82 ppm and 3 min for 0.34 ppm and then falls to a lower 

level. This may be due to a replenishment of the antioxidant. The authors saw no injury to the 

plasmalemma (as measured by penetration of a dye) and no change in the stomatal conductance 

for the lower concentration of O3 (Moldau and Bichele, 2002). The higher level (0.88 ppm) 

caused the plasmalemma of the mesophyll cells to pass a dye, and a slight decline in stomatal 

conductance resulted at about 2.5 minutes. These data suggest that the antioxidant hypothesis is 

correct. 

Gaseous pollutants flow from the substomatal cavity within the leaf through the cell wall 

into the cell. It is suspected that the internal concentration of the pollutant is not uniform within 

the cavity (Taylor, and Hanson, 1992). From within the wall, an equilibrium between the gas 

and aqueous phase must occur at the interface where the gaseous species dissolve into the water 

according to Henry’s Law (Heath, 1980, 1987; Wellburn, 1990). It is important to understand 

AX9-40
exactly how much O3 could move into the tissue of the leaves. Calculations in Table AX9-6 

give an indication of the amount of O3 which may end up near the surface of cells within the 

leaf. The calculation is done for a standard temperature (25 °C), an ambient concentration of O3 

(0.10 ppm), and for nonspecific leaves. For example, 0.3 ppm would be the same general 

numbers but multiplied by 3. Similarly for more closed stomata, the value of 1.0 cm/s 

(equivalent to about 400 μmole!2

-leaf area s!1) for a conductance would be reduced and the 

smaller values would lead to a smaller amount of O3 moving into the tissue. Nonetheless, these 

values give some indication of what sort of chemical concentration can be expected. Under 

these conditions, a delivery rate of O3 into the substomatal cavity near the spongy mesophyll 

tissue of about 0.42 nmol/(L@h) appears to be reasonable. 

Process 2. Ozone diffuses into the leaf air spaces and reacts either with varied biochemical 

compounds that are exposed to the air (path 1) or is solubilized into the water lining the cell wall 

of the air spaces (path 2). As shown in Figure AX9-6, each reaction has the possibility of 

transforming O3 into another chemical species (a toxicant) which, in turn, may react with other 

chemical species and lead to a cascade of reactions. 

Within the stomata, gases react with the water at the cell’s surface and generate new 

species with the components within the cell wall region. The possible varied pathways are 

depicted in Figure AX9-7. Although these chemical reactions are poorly understood, some of 

the fundamentals are known (Heath, 1987, 1988; Wellburn, 1990). Ozone reacts with organic 

molecules at the double bonds to form carbonyl groups and, under certain circumstances, 

generates peroxides, such as hydrogen peroxides (H2O2), superoxide (O2 

!

) and its protonated 

form (HO2 

C

), hydroxyl radicals (HOC), and peroxy radicals (HO2 

C

). Other chemicals present in 

the water phase can lead to many other oxygenated moieties (Figure AX9-6). Each of the steps 

is generally pH dependent (Jans and Hoigne, 2000; Walcek et al., 1997). 

Sulfhydryls are particularly easy targets, with the formation of disulfide bridges or sulfones 

(Mudd and Kozlowski, 1975). In water, the reactions become more confusing, but some 

! products have been described by Heath and Castillo (1987), such as H2O2, HOC

, and O2 

(Figure AX9-7). Effective detoxification reactions can occur here via antioxidant metabolites 

and enzymes such as ascorbate, glutathione (GSH), and superoxide dismutase (SOD) if they are 

present at high enough concentrations (Castillo et al., 1987; Matters and Scandalios, 1987). 

AX9-41
Table AX9-6. The Flow of Ozone into a Leaf and Possible Reactions 

The level of O3 in the atmosphere is chosen to be close to a standard and yet make calculations to other amounts easy. 

The same concept will be used for all standard parameters for these calculations. 

DESCRIPTION VALUES 

The atmospheric level of O3 is given as: 

For an air temperature of: 

The perfect gas law (pV = nRT) is used to convert the O3 

level into standard mks. Further, the volume for a mole of 

gas (Vo = 22400 m3

) will be used, from the perfect gas law 

with R = 8.3144 

Thus, the concentration of O3 within the atmosphere is: 

The stomatal conductance of the gas must be chosen to be 

standard but adjustable. The number should be as large as 

typically measured but allow for easy conversion, if 

necessary. For a stomatal conductance of: 

The amount of O3 that will penetrate inside the leaf (for a 

typical concentration of nearly zero inside the leaf), is: 

In terms of amount of water within the leaf, we can assume 

that about 85% of the weight is water and the density of 

water is 1 g/mL. A typical leaf has a wet weight/area: 

Thus, the square surface area of the leaf will translate into 

water space (for concentration of chemicals), as: 

The maximum amount of toxic compound that will be 

generated, assuming all the O3 is converted, is given below. 

Here the units of the leaf area weight are converted into the 

mks system and the water space units are converted into L, 

such that the concentrations calculated will be in mol/(L hr). 

The final units assume that the O3 is present (and no back 

reactions occur) for one hour (short but typical units 

of exposure). 

Thus, the maximum amount of toxic chemicals generated 

per hour in a leaf would be: 

O3 a = 0.1 ppm 

Ta = 25 °C 

CO3 = 4.873 ×10!12 moles/m3 

gswv = l cm/s 

O3 L = 2.984 ×10!14 mol/(m2@s) 

FWL = 30 mg/cm2 

ArL = 0.255 L/m2 

O

O3Lc = 4.21 × 10!10 mol/(L@h) 

3Lc = 0.42 nmol/(L@h) 

Possible errors in these calculations (aside from the input numbers) are (1) the O3 within the leaf does not react 

uniformly within the leaf space; (2) the O3 within the leaf does not totally convert to any one species; (3) varied 

products of O3 react, leading to innocuous chemicals; and (4) O3 reactions can be catalytic and generate more reactions 

by radical reaction cycling. 

AX9-42 
Figure AX9-6. Possible transformations of O3 within a leaf. 

Figure AX9-7. Possible reactions of O3 within water. (a) Ozone reacts at the double 

bonds to form carbonyl groups. (b) Under certain circumstances, 

peroxides are generated. 

AX9-43
If the levels are low, it is believed that stimulation of their production is a response to O3, albeit a 

slow one (Harris and Bailey-Serres, 1994). Certainly it is possible that chemical modification of 

wall-specific biochemicals (Castillo et al., 1987) such as glucan synthase (Ordin et al., 1969) and 

diamine oxidase (Peters et al., 1988) occurs. 

Process 3. Movement of reaction product(s) into and enzymatic or chemical 

transformations within the cell. It is believed that the initial site of O3 injury is near or within the 

plasma membrane. Certainly, membrane functions, such as membrane fluidity (Pauls and 

Thompson, 1980), permeability (Elkiey and Ormrod, 1979), K+

-exchange via ATPase reactions 

(Dominy and Heath, 1985), and Ca2+ exclusion (Castillo and Heath, 1990), are changed. The 

similarity of wounding responses (Langebartels et al., 1991) and O3-induced membrane 

disruption suggests the induction of normal wound-regulated genes (Mehlhorn et al., 1991; 

Sandermann, 1998). This implies that O3 can react with cell-wall components that are connected 

to the cytoplasm through the cell wall and membrane by membrane-specific proteins that are not 

directly involved with transport. 

Ozone is soluble in water and once having entered the aqueous phase, it can be rapidly 

altered to form oxidative products that can diffuse more readily into and through the cell and 

react with many biochemicals. Again, the presence of an internal antioxidant would be critical 

to reduce the concentration of most oxidants. A toxic product of O3 may migrate through the 

cytoplast to react with photosynthetic processes, or a spurious signal generated at the membrane 

may affect some control process or signal transduction pathway (Schraudner et al., 1998; 

Overmyer et al., 2000, 2003; DeCaria et al., 2000; Rao et al., 2002; Booker et al., 2004; Leitao 

et al., 2003; Rao and Davis, 2001; Sandermann, 2000; Vahala et al., 2003b). 

AX9.2.3.1 Possible Reactions Within the Leaf 

Ozone can react with many compounds within the substomatal cavity of the leaf1

 to 

produce a variety of oxidizing and toxic chemicals. Some of the possible reactions that will 

generate H2O2, HOC

, and SO2 

!

, as well as charged O3 intermediates, are indicated in Figure 

AX9-8. Many of these complex reactions have been studied within water solutions through 

1

The volume of the substomatal cavity (that are within the leaf immediately below the stomata) must be 

regarded as the region in which most O3 reactions occur. That volume, at a relative humidity of near 100%, 

possesses many diverse surfaces with varied bonding, which could alter the fate of O3. 

AX9-44 
Figure AX9-8a,b. The Crigee mechanism of O3 attack of a double bond. (a) The typical 

Crigee mechanism is shown in which several reactions paths from the 

initial product is shown. (b) Typical reaction of ascorbic acid with O3. 

Source: Adapted from Mudd (1996). 

AX9-45 
research of O3-induced water purification and are very dependent upon solutes present with the 

solutions, including H+

 (see Von Gunten [2003]). An important point is that in alkaline 

media, O3 forms H2O2, but in acid media, O3 is relatively stable in the absence of free metal ions. 

The rates of reaction of O3 with several important compounds, including those with a 

double bond, the so-called Crigee Mechanism shown in Figure AX9-8, can be calculated from 

the reaction coefficient as given by Atkinson (1990) (Table AX9-7). The double bond of the 

ascorbate molecule is particularly sensitive to O3 attack. Because of the ring formation of the 

ascorbate molecule, an unstable ozonide product is formed, which then accelerates the breakage 

of the double bond, leading to the formation of two products. These products are relatively 

unstable and can lead to further reactions not shown in Figure AX9-8. The rates of reactions can 

be calculated (Heath, 1987). At a local concentration of 25 μM O3, it would take 5000 s 

(83 min) for all of the O3 to react if there was no further flow of O3. Clearly, O3 does not react 

rapidly with the compounds in Table AX9-7 and, although some of the products would be 

formed through the Crigee Mechanism (see Figure AX9-8a), they would be low in 

concentration2

. While other radicals, such as the hydroxyl radical (see Figure AX9-8b) can 

attack double bonds, the products differ. Of particular note for later discussion, is the reaction 

of O3 with ascorbate (see Mudd (1996) (see Figure AX9-8b), which will cleave the double bond 

in the ring. Unfortunately, little work has been done to characterize possible products within the 

leaf (but see next section). 

In a paper discussing the stability and reactivity of O3 in the pulmonary air/tissue 

boundary, Pryor (1992) calculated that O3 has a half-life of about 7 × 10!8

 s in a bilayer. 

However, the transit time through the lung lining fluid layer is about 2 × 10!6

 s, based upon a 

reasonable estimate for the diffusion of O3. This means that O3 would suffer nearly 29 halflives3

 in passage through the layer, reducing it to about 3 × 10!9 of the original concentration — 

zero for all practical aspects. In the same publication, Pryor points out that any sulfhydryl or 

ascorbate would interact strongly with O3, further reducing its net concentration. The reactivity 

of cysteine is 109

, while the reactivity of tryptophan, methionine, polyunsaturated fatty acids, 

2

For example, hydroxylmethyl hydroperoxide would be expected to be formed by the reaction of O3 with 

ethylene and its effects have been tested on peroxidases (Polle and Junkermann, 1994). Unfortunately, the 

concentration of required for inhibition is much higher than would be expected to be formed within the leaf. 

3

Here a half-life is the time that it takes the reactive species to travel a distance in which it loses 50% of its 

initial concentration. Therefore for a 29 half-life, the concentration has been reduced by 2!29 or about a 10!9

 decline. 

AX9-46
Table AX9-7. Some Rates of Reaction of Ozone With Critical Biochemicals 

[a] Double bond reactions. The second column is taken from Atkinson (1990) and transformed into 

Column 3. Those rate coefficients are used to calculate the rate of reaction at a concentration of 

10 ppm for the organic and 0.1 ppm for O3 in the air stream within the leaf (localized concentration 

of about 25 mM, see Table AX9-6). 

Compound × 10!18 cm3

/molecules s!1 Rate coefficient (L/mole s!1) Rate of reaction (M/s) 

Ethane 1.7 1.02 × 103 4.3 × 10!11 

Propene 11.3 6.80 × 103 2.8 × 10!10 

1-butene 11 5.91 × 103 2.5 × 10!10 

trans-2 Butene 200 1.20 × 105 5.0 × 10!9 

"-pinene 85 5.12 × 104 2.1 × 10!9 

[b] Possible Oxidative Species. Another possibility is given by the reactions below from 

Walcek et al. (1997). 

Reactions Rate constants 

(1) O3 + OH!

 + H2O ÿ H2O2 + O2 + OH! k1 = 3.67 × 10 mole!1 L s!1 

(2) O3+ O2 

! ÿ HO.

 + 2 O2 + OH! k2 = 1.26 × 109 mole!1 L s!1 

(3) O3+ HO2 

! ÿ HO.

 + O2 

!

 + O2 k3 = 2.09 × 106 mole!1 L s!1 

[c] Possible Concentrations of Other Oxidative Species. Table from Heath (1987). 

Based upon 100 ppm O3 in gas stream. 

Concentration (M) 

Species pH 7 pH 9 Molecules within wall 

C Superoxide Radical (O2 ) 8.75 × 10!15 1 × 10!12 5.5 × 10!6 6.3 × 10!4 

Ozone Radical 4.16 × 10!15 5 × 10!14 

C Protonated O3 radical (HO3 ) 1.48 × 10!16 1 × 10!18 

Number of molecules within apoplastic space of (10!12 L) at 0.1 ppm O3. 

and tyrosine is about 2 × 106 and that of phenylalanine is only 103. These numbers are similar to 

what has been found for O3 reactivity with amino acids and proteins in aqueous solutions. 

In glycophorin (Banerjee and Mudd, 1992) and cytochrome C (Mudd et al., 1997b; Mudd et al., 

1996) in aqueous solutions, only the methionine was oxidized by O3, producing sulfoxide. 

In other proteins lacking methionine, tryptophans were oxidized only if they were in an exposed 

position on the surface of the proteins (Mudd et al., 1997b). Treatment of red blood cell ghosts 

AX9-47
with O3, oxidized peripheral proteins of the plasma membrane before it oxidized lipids (Mudd 

et al., 1997a). 

AX9.2.3.2 Toxicants Within the Wall Space 

While Mehlhorn et al. (1990) are often thought to have shown that free radicals were 

formed in plant leaves under O3 exposure, careful reading of that paper clearly shows that there 

was no real evidence of free radicals induced by O3. Living tissues have many free radical 

signals, making it difficult to observe changes in free radicals. Further, the work of Grimes et al. 

(1983) has also been cited as showing the presence of free radicals in living tissues due to O3 

exposure; however, no radical signals were found unless certain organic acids (e.g., caffeic acid) 

were added to the tissue with the O3 exposure. They used the radical trap TMPO 

(tetramethylphrrolise 1-oxide) which reacts with many types of free radicals to form a stable 

radical that can be used to “trap” or increase the amount of radical present (see Figure AX9-9a). 

Ozone would directly react with this trap only if it were bubbled into the solution, not passed 

over the top of the solution. In the presence of sorbitol or caffeic acid, the trap would indicate 

the presence of OH radical, which would mean that O3 ÷ HOC

. Superoxide dismutase, catalase, 

or EDU had no effect upon this signal, suggesting O2 

!

 and H2O2 were not involved in the above 

sequence. Both O3 and O3 plus caffeic acid had no effect upon the protoplasts’ intactness or 

viability. Thus, 10!5

 M HOC and/or 0.30 to 0.40 ppm O3 did not react with the cell membrane. 

They found no signal in normal cells after subjecting the leaf to O3 and concluded that the 

radicals were produced via a concerted mechanism with the acid. This does not fit with the 

mechanism postulated by Mehlhorn et al. (1991), which involved a reaction of wound-induced 

ethylene and O3 at the wall level to generate some free radicals. 

The hypothesis that the production of wound-induced ethylene by O3 exposure and its 

reaction with O3 would result in the production of radicals was tested by Mehlhorn et al. (1990), 

using electron paramagnetic resonance spectroscopy. After 4 h of 300 ppb, an EPR signal of a 

compound was detected which resembled a butonyl radical (Figure AX9-9b [Mehlhorn et al., 

1990]). Using 70 ppb, the signal was reduced by about one third that of an ethyl radical4

, 

4

The reaction would be: O3 + H2C = CH2 ÷ varied C-1 compounds, due to double bond cleavage, at a 

rate constant of 1.7 × 10!18 cm3

/molecule sec = 1.02 × 103 M!1 s!1 (Atkinson, 1990). This should be compared 

with a reaction of the hydroxyl radical with ethylene, which has a rate constant of 8.52 × 10!12 cm3

 molecule!1 s!1, 

or 106 × faster. 

AX9-48
Figure AX9-9. Varied ESR radicals, trapped and not, generated by O3 under somewhat 

physiological conditions. (a) The generation of a DMPO-trapped radical 

with caffeic acid in water solution (Grimes et al., 1983). (b) The 

generation of a DMPO-trapped radical within bean (Phaseolus vulgaris cv. 

Pinto) exposed to 70 nL/L O3 for 4 h. The lower trace is the ESR signal 

produced with 300 nL/L O3 (Mehlhorn et al., 1990). (c) The EPR signal 

produced within a bluegrass leaf exposed to 1000 μg m-3 of O3 for 1 h 

(Runeckles and Vaartnou, 1997). Although no trapping agent was used in 

this experiment, the signal is complex, because of various free radicals 

normally present within the illuminated leaf. 

leading to injury. However, the spraying of the plant with 1-aminoethoxyvinyl-Gly (AVG), 

which reduces the production of ethylene and visible injury’, had no effect upon the EPR signal, 

suggesting that the radical is not a direct sequela to visible injury. 

Runeckles and Vaartnou (1997) (Figure AX9-9c) discovered a signal by subtracting other 

EPR signals of the leaf, which seemed to be due to an O3 reaction with plant material, using 

! 0.48 ppm O3. This difference signal looked very much like O2 . At a lower concentration, they 

observed that this signal still occurred but accumulated more slowly. Both bluegrass and 

ryegrass leaves seemed to saturate after about 5 h of exposure at 22 to 28 units of signal, while 

radish leaves reached a maximum of 7 units at 3 h and then declined. The problem, which is 

AX9-49
typical of any of these methods, was that the detached leaf had to be rolled and placed into the 

EPR detection cavity. Reichenauer et al. (1998) also detected an undefined free radical signal 

that seemed to be related to a Mn(II) spectrum. The Nandu and Perlo cultivars of wheat were 

more sensitive to O3 than Extradur (according to growth rate and closure of stomata under an O3 

exposure of 80 ppb for 8 h/day, 7 days/week over 100 days), and these more sensitive cultivars 

had a greater, but insignificant (P = 15%), EPR signal. Thus, data showing any production of a 

free radical must be approached with some skepticism. 

With an O3 delivery rate of about 25 μM/h (Table AX9-6), only 250 μM would be found 

after a full day, if all of the O3 were stable. While the use of free radical traps is the best method 

available to observe any build-up of radicals, the traps are not as specific to individual radicals. 

Currently, studies should be looking for hydroxyl radicals, superoxide, hyroperoxides, ethylene 

radicals, and ascorbate radicals. 

AX9.2.3.3 Products of Ozone 

Ozone should reach a certain concentration in the substomatal cavity, which is dependent 

upon its entry speed and its reactivity with the wall constituents. Once near the apoplastic 

space, O3 moves in two different pathways (Figure AX9-6). It can react with constituents that 

are within the wall as a gas in reaction 1A (path 1); or it can solubilize into a water space and 

travel to another region within the water space and react through reaction 2A (path 2). 

Hydrogen Peroxide 

Hydrogen peroxide, until recently was thought to be purely a toxic compound for cells. 

However, it is now clear that it functions as a signaling molecule in plants and mediates 

responses to abiotic and biotic stressors (Figure AX9-10). Generation of H2O2 is increased in 

response to various stressors, implicating it as a key factor mediating the phenomena of 

acclimation and cross tolerance, in which exposure to one stressor can induce tolerance of 

subsequent exposure to the same or different stresses (Neill et al., 2002). The signaling response 

to attack by invading pathogens using H2O2 has been described (Mehdy, 1994; Simon-Plas et al., 

1997). The reactions leading to hypersensitive cell death are caused by a pathogen recognition 

step (Figure AX9-10a), probably due to the plant cell wall releasing oligosaccharides in response 

to the pathogen enzymatically breaking down the cell wall to penetrate it. A feed-forward 

AX9-50
Figure AX9-10. Pathogen-Induced Hypersensitivity. (a) The reactions leading to 

hypersensitive cell death and the formation of a global response of 

salicylic acid. (b) The cascade of the elicitor-induced reactions within 

the cell. 

step in which H2O2 increases the level of benzoic acid leads to the activation of the hydroxylase 

step in the production of salicylic acid and to a feedback step in which the salicylic acid 

increases the production of H2O2 (Leon et al., 1995). 

AX9-51
An elicitor, e.g., a bacterial or fungal pathogen, induces a cascade of reactions within a cell 

(Figure AX9-10b). Some of the lipid reactions are thought to be due to the opening of the Ca2+ 

channels and the alkalination of the cell wall region. The oxidative burst due to H2O2 production 

is believed to lead to the transformation of a small population of lipids into jasmonic acid, which 

is a secondary messenger. 

Hydrogen peroxide also has an oxidative role in lignification (Schopfer, 1994). In the 

interaction of lignification and the beginning processes of hypersensitivity, pectinase produced 

by the pathogen disrupts pectin and dissolves the cell wall. Fragments of the dissolved cell wall 

trigger an increase in the transcription of peroxidases within the remaining cell wall, leading to 

signification, which is a cross-linking of the cell wall that does not use pectin. This prevents 

further pathogen disruption of the wall and reduces its further entry into the plant cell. 

It is believed that the first species generated through a one-electron reduction of molecular 

! oxygen is SO2 . That generation is carried out using a cytochrome b6 by the NAD(P)H oxidase 

! located on the cell membrane (Auh and Murphy, 1995). In the acid region of the cell wall, SO2 

is converted by a protonation and dismutation to H2O2. The induced oxidative burst is believed 

to play a role in stimulating the Cl!

 and K+ efflux, generating an alkalinization of the 

extracellular space (Cazale et al., 1998). In the wall region, H2O2 is not especially toxic, as no 

necrosis was reported in tobacco when 500 mM peroxide was infiltrated into the leaf tissue. 

However, the production of salicylic acid and benzoic 2-hydroxylase can be induced with only 

30 and 0.3 mM H2O2, respectively, indicating some metabolic signaling (Leon et al., 1995). 

On the other hand, 1M H2O2 infiltrated into soybean will generate lipid peroxidation after 1 h 

with a peroxidation rate of 15 nmol/g-FW@h (Degousee et al., 1994). Cells react to the system5 

and generate peroxide scavenging compounds within 1.5 to 2 hours, which appear to “mop up” 

the excess H2O2 (Baker et al., 1995). 

After O3 exposure in birch, H2O2 has been found in the wall (Pellinen et al., 1999). By 

using CeCl2 as a cellular stain for H2O2 (as a cerium perhydroxide precipitate), Liu et al. (1995) 

observed a gradual development of stain after 8 h of O3 exposure (at 150 ppb). After 2 h 

exposure, H2O2 stain was visible on the surfaces of both sets of mesophyll cells. Accumulation 

of H2O2 stain continued for 16 h after exposure, suggesting a triggered-reaction rather than O3 

5

Soybean suspension cells were inoculated with Pseudomonas syringae pv syringae, which generate an 

active oxygen response. Light emission by luminol, reacting with H2O2, was the assay for the peroxide. 

AX9-52 
decomposition itself. H2O2 stain was present in the mitochondria, peroxisomes, and cytoplasm 

but not in the chloroplast. If methyl viologen (MV) was given to the leaves and then the leaves 

were exposed to light, H2O2 stain could be observed within the chloroplast. This indicated that 

the stain worked within the chloroplast if H2O2 were generated by the Mehler reaction 

(MV + O2 

!

). Thus, apparently, for birch, O3 exposure does not generate excess H2O2 within the 

chloroplast. Furthermore, these sets of experiments indicate that O3 per se does not generate the 

H2O2, but rather triggers stress-related H2O2 formation similar to what occurs in a pathogen 

attack (the Reactive Oxidative Species or ROS reaction). 

The presence of higher than normal levels of H2O2 within the apoplastic space is a potential 

trigger for the normal, well-studied pathogen defense pathway. Figure AX9-10b depicts such a 

pathway and suggests that all the events and activation of pathways/genes caused by pathogen 

defense could be observed when plants are fumigated with O3. The events shown in Figure 

AX9-10b will be alluded to in later sections. 

H2O2 has been linked to the hormone ABA-induced closure of the stomata by activating the 

calcium influx in guard cells (Pel et al., 2000). The addition of H2O2 at a level of only 5 mM to a 

guard cell preparation will cause a dramatic increase (ca. 9×) in electrical current at the 

hyperpolarizing potential of –200 mV. Amounts as low 50 mM H2O2 will cause a less, but still 

sizable, increase. Membrane stability is unaffected by the H2O2 and the activation of the channel 

requires only about 2 to 3 minutes. Pel et al. (2000) also found that ABA induced the production 

of H2O2 through ROS accumulation (also see Zhang et al. [2001]). 

Certain levels of ABA within the leaf lead to stomatal closure. The inactivation of a 

phospho-tyrosine-specific protein phosphatase (ABI2) is an inhibitor of stomatal opening 

induced by ABA but that enzyme is inhibited by H2O2 (Meinhard et al., 2002). This means that 

H2O2 shifts the sensitivity of the stomatal opening to ABA (Figure AX9-11), making the 

stomatal complex more sensitive to ABA. Thus, for a given level of ABA present in the guard 

cell complex due to environmental factors (e.g., low humidity, high air temperature, or low soil 

water potential), the generation of H2O2 would (by inhibiting ABI2) induce a closure of the 

stomata by increasing the sensitivity of the guard cells to ABA. In the past, it has been difficult 

to understand why O3 would often decrease conductance in some cases, but not always (Heath, 

1994b). This interaction between H2O2 and ABA could help understand this complexity. 

AX9-53
Figure AX9-11. The interaction of H2O2 and Ca2+ movements with ABA-induced stomatal 

closure. It is well known that certain levels of ABA within the leaf lead to 

closure of the stomata within the leaf. That level, however, can be shifted 

to make closure more or less sensitive to a given level of ABA. Recently it 

has been shown that H2O2 (externally or produced by the plant) within 

the cell wall region can shift that sensitivity. Here ABA stimulates the 

production of H2O2, which in turn increases the rate of Ca2+ moving from 

the wall region into the cytoplasm. That shift in internal Ca2+ level 

increases the closure of the stomata. Hydrogen peroxide also blocks the 

activation of a polypeptide (ABI2) that inhibits stomatal closure 

seemingly induced by ABA. 

Sources: From Assmann (2003); Assmann and Wang (2001); and Zhang et al. (2001). 

Ethylene Reactivity 

Ethylene (ET) is produced when plants are subjected to biotic stressors, e.g., attacks by 

insects, fungi, and bacteria or abiotic stressors such as wounding or environmental stressors such 

as heat, cold, or oxidative stress and O3. If an O3 stress has induced a wounding response with 

ET release, then ET within the substomatal cavity could react with O3, generating some 

relatively noxious chemicals (see Figure AX9-6). The relationship between O3 injury and 

wounding is supported by the observation by Mehlhorn et al. (1991) that an inhibitor of ET 

formation, AVG (an inhibitor of ACC synthase, a committed step to ET production), would 

block ET formation and inhibit visible injury. Other studies with polyamine (which is closely 

linked to ET production), including those of Ormrod and Beckerson (1986) who fed polyamines 

to the transpirational stream and prevented visible injury, suggested a close involvement of both 

AX9-54
pathways to the production of visible injury. Both the lack of ET production and an increased 

level of polyamines slowed or prevented visible injury. 

This concept was taken another step by Langerbartles (1991). The linkage to the pathogen 

wound responses and visible injury is well established (Sandermann, 1996). Sandermann (1998) 

used a system of Bell B and W3 tobacco, plants with differential O3 sensitivities, in which the O3 

exposure level was chosen such that the sensitive cultivar was injured, while the tolerant one was 

not. This led to a marvelous control that could be used to their advantage. They followed a time 

sequence to show that the rise of varied systems followed the same order as seen for a pathogen 

attack (Heath, 1994a). 

More recent studies, however, indicate that O3 responses resemble components of the 

hypersensitive response (HR) observed in incompatible plant-pathogen interactions 

(Sandermann, 1998). The similarity to the HR response may be related to the occurrence of 

ROS in the apoplast. The O3-derived ROS apparently trigger an oxidative burst in the affected 

cells by an as yet unknown mechanism. An oxidative burst is similar to one of the earliest 

responses of plants to microbial pathogens and is an integral component in HR-related cell death 

(Overmyer et al., 2000). 

In plants exposed to O3, ET synthesis is a result of the specific ET induction of the genes 

encoding 1-aminocyclopropane-1-carboxylase synthase (ACS), one of the fastest and most 

obvious responses to O3, which has been mechanistically linked to the regulation of O3 lesion 

formation. Biosynthesis of ET inhibited, with ACS inhibitors significantly reduced, the 

induction of lesion formation in plant leaves exposed to O3 (Mehlhorn et al., 1991; Mehlhorn and 

Wellburn, 1987; Vahala et al., 2003b). Ethylene biosynthesis correlates best with O3 exposure 

(Overmyer et al., 2000; Vahala et al., 2003b). These data support the concept that elimination of 

ET formation will prevent visible injury. 

Ethylene-Interaction with Injury and Conductance 

Increased ET production by plants exposed to O3 stress was identified as a consistent 

marker for O3 exposure decades ago (Tingey et al., 1976a). They exposed more than 20 plant 

species and cultivars to O3 to determine whether the production of O3-induced stress-ET could be 

used to determine differences in plant sensitivity to O3. Their studies suggested that increased 

AX9-55
production of stress-ET correlated well with the degree of foliar injury that developed within 

hours or days after O3 exposure. The amount of ET released was exponentially related to the O3 

exposure. Furthermore, the amount of O3-induced ET declined with repeated exposure, 

indicating an acclimatization to O3. This acclimatization effect associated with repeated 

wounding has not yet been well described. The release of wound-induced ET is not linear with 

time, but declines after the initial response (Stan et al., 1981), as is also seen after O3 exposure 

(Stan and Schicker, 1982). The stress-induced ET production correlates better with O3 exposure 

level than with exposure duration. In other words, peaks of high O3 (rather than accumulated 

dose) generate a higher rate of ET release, at least for a single O3 exposure under an acute dose. 

The production of ET after an O3 exposure is thought to be a typical wounding response 

(Tingey et al., 1975). Prevention of ET release may prevent the formation of visible injury 

(Mehlhorn and Wellburn, 1987). However, the question arises as to whether this effect was 

limited to the prevention of visible injury or if the chemicals used to prevent ET release closed 

the stomata. Using Glycine max L., Taylor et al. (1988b) showed clearly that AVG did not 

necessarily close stomata nor inhibit carbon assimilation per se. 

The correlation of ET release with O3-induced visible injury was likewise shown in pea 

cultivars (Dijak and Ormrod, 1982). With O3 exposure (generally 6 h at 0.3 ppm), the stomata 

closed by ~50% within 3 h after a dose of 3 × 10!5

 mol cm!2 (with an average rate of 2 × 10!9 

moles cm!2 s!1, as calculated from their data). Both sensitive and insensitive cultivars had a 

visible-linked-injury ET release, but sensitive cultivars scored higher both in visible injury and 

in ET release after a given exposure. 

Gunderson and Taylor (1988, 1991) used exogenous ET to alter the gas exchange of 

Glycine max and found an exponential, but not simultaneous, decline of both stomatal 

conductance and carbon assimilation with ET. Interestingly, the exogenous ET caused a slight 

rise in difference of CO2 within and without the leaf, indicating a lowering of internal CO2, 

which was not observed in the experiments of Farage et al. (1991) for O3 exposure. Ethylene 

inhibits both stomatal conductance and carbon assimilation to some extent (Taylor et al., 1988b). 

Thus, one could postulate that O3 generates a wounding response with a production of ET, which 

would, in turn, generate the change in stomatal conductance and photosynthesis. Clearly, these 

multiple events may have confounded some earlier studies. 

AX9-56
AX9.2.3.4 Antioxidants Within the Apoplastic Space 

The first line of defense against O3 is a closure of the stomata to exclude its uptake. This is 

counterproductive for efficient photosynthesis, but some amount of closure limits the rate of O3 

deposition into the leaf tissue to allow for a secondary line of defense to detoxify the O3. The 

secondary line of defense involves a range of antioxidants, which are highly reactive to the types 

of chemicals that can be generated by O3. Several antioxidant proteins are stimulated by O3 in 

Plumbagini folia, including glutathione peroxidase (GSH-Px), SOD, and catalase. The 

timescales for changes in their levels vary: some rise rapidly, while others rise more slowly. 

The pattern of changes in these particular proteins varies greatly among different species and 

conditions. 

Ascorbate Within the Cell Wall 

Most of the recent reports indicate that ascorbate within the cell wall is the real first line of 

all defense. Ascorbate within the wall declines when the tissue is exposed to O3 (Luwe et al., 

1993; Moldau, 1998). This decline appears to be closely linked to the amount of O3 penetrating 

the leaf tissue. 

It has long been suspected that intracellular antioxidants play a role in preventing O3­

induced injury to plant cells. Variation in the types of biochemical compounds present in the 

apoplastic space can give rise to a multiplicity of reactions with O3, but the predominant 

biochemical species is ascorbate. Ascorbate is water soluble, present in the solution where O3 

can dissolve, and is highly reactive. Unfortunately, a variety of antioxidants are found 

throughout the cell and any measurements of one particular type within the total leaf tissue can 

give misleading results. For example, ascorbate is present within the cell wall, cytoplasm, and 

chloroplasts (Burkey, 1999; Moldau, 1998); and ascorbate can move between the cytoplasm and 

the cell wall with relative ease (Figure AX9-12) (Bichele et al., 2000). The total of all ascorbate 

pools is measured when the tissue is ground and assayed. If the cell wall ascorbate concentration 

drops by 50% due to O3 exposure but all other tissue concentrations remain the same, the 

measurement of the total loss is dependent upon the amount of ascorbate within the cell wall. 

Turcsányi et al. (2000) showed that, compared to the concentration of apoplastic ascorbate, the 

rest of the cells contained about 38 times as much. So a 50% loss of apoplastic ascorbate would 

be converted into only 2 to 3% loss of the total ascorbate. 

AX9-57
Figure AX9-12. The reaction of ascorbate within the apoplasm of the cell wall and its 

ultimate reduction/oxidations. 

(a) Movements of reducing power (from Dietz [(1997]). 

(b) The use of glutathione to maintain the level of ascorbate within 

the cell wall region (from Horemans et al. [2000]). 

AX9-58
The ascorbate deficient Arabidopsis thaliana mutant has proven to be a powerful tool in 

furthering the understanding of ascorbate biosynthesis in plants (Smirnoff et al., 2001). Three 

classes of mutants were formed when A. thaliana seed was mutagenized with ethyl 

methanesulfonate: (a) those deficient in SOD, (b) those that failed to accumulate more 

antioxidant proteins upon increased O3 exposure, and (c) those that were deficient (but not 

depleted) in ascorbate. The low-ascorbate mutant type had 50 to 60% less ascorbate than the 

wild type and displayed more foliar injury. This mutant is involved with the coding of the 

GDP-D-Mannose pyrophosphorylase enzyme6

 in the Smirnoff-Wheeler pathway for ascorbate 

biosynthesis. Smirnoff et al. (2001) also suggested that other pathways can produce ascorbate 

without relying upon the pyrophosphorylase step, but most probably at a slower through-put rate, 

because any fully ascorbate-deficient mutant would be lethal, perhaps because of ascorbate use 

as a cofactor rather than its antioxidant properties. In addition to its lowered antioxidant 

capacity, the ascorbate-deficient Arabidopsis mutant vtc1 (Conklin et al., 1996; Conklin and 

Barth, 2004) may show suppressed growth due to lower mannose levels that are necessary for 

cell wall formation. Ozone thus may suppress growth in these mutants through interference with 

cell wall biosynthesis as well as through lower antioxidant protection. 

The ascorbate peroxidase (APX, which uses ascorbate to detoxify peroxides) family 

consists of at least five different isoforms, with isozymes in the apoplastic and cytosolic space. 

Furthermore, most forms of ascorbate can move through the plasma membrane (Bichele et al., 

2000), making the levels of all forms of ascorbate interdependent and able to at least partially 

influence each other. Dehydroascorbate (DHA) can be broken down into other smaller 

fragments easily in vivo and represents a continuous loss of ascorbate from varied parts of the 

cell if ascorbate is allowed to remain in the oxidized form in some regions. In fact, the turnover 

rate in leaves is estimated to be from 2 to 13% per hour, depending upon species and 

developmental age (Smirnoff et al., 2001). There are apparently three pathways for ascorbate 

turnover (Figure AX9-12a). The typical reaction is a reduction of DHA into ascorbate from 

which an oxidative step generates DHA. Pathway I requires a reductive step using NADH 

external to the plasma membrane generated from internal malate using a malate/oxaloacetate 

6 EC 2.7.713, Mannose-1 phosphate guanylyltransferase; mannose + GTP ÿ GDP-mannose + ppi; this 

product leads into cell wall polysaccharide synthesis and protein glycosylation through GDP-galactose and 

GDP-fucose and, ultimately, through galactose into ascorbate synthesis. 

AX9-59 
transporter. Pathway II uses a direct transporter of ascorbate/DHA. Pathway III moves the 

required electron(s) through a cytochrome b system, maintaining two separate pools of 

ascorbate/DHA (within the cytoplasm and within the wall). Each of the pathways (Dietz, 1997) 

represented by Roman Numerals in the Figure AX9-12, require only one NAD(P)H molecule to 

reduce the DHA molecule back to ascorbate. However, the transport properties and redox 

potential of the cell differ for each pathway. The efficiency of the reduction of DHA is 

dependent upon the redox coupling and the region in which the chemical species is located. 

Turcsányi et al. (2000) exposed broad bean (Vicia faba) grown under two regimes in 

duplicate controlled chambers: charcoal/Purafil filtered air (CFA) or (CFA) plus 0.075 ppm O3 

for 7 h/day for 28 days (chronic exposure) or exposed to 0.150 ppm for 8 h (acute exposure). 

Responses of the two sets of plants were similar except for stomatal conductance, which was 

50% lower in the chronically exposed plants. Plants grown under acute exposures developed 

visible injury, while plants grown under chronic conditions developed no visible injury. Within 

an hour of the start of the acute exposure, the stomatal conductance was reduced by nearly 40% 

and assimilation was reduced by nearly 18% in the clean air plants; the reduction in conductance 

was only 21% and assimilation 16% in the plants subjected to chronic O3 exposures. The 

assimilation was affected similarly in both cases, while the conductance showed less of a 

percentage drop in the chronic O3-exposed plants and began at a lower O3 level. The similarity 

of the assimilation indicated that the stomata were not limiting assimilation in either case before 

acute exposure. More to the point, the decline in ascorbate in the apoplastic space due to the O3 

exposure was “...more often than not, on the borderlines of statistical significance.” However, a 

30% decline in ascorbate after 4 h of acute O3 exposure in both cases was observed. This lack of 

significance may be due to a relatively large standard error of the data, which in turn may be due 

to the difficulty of extracting and measuring ascorbate from the apoplastic space in quantitative 

terms. 

The chemical reaction7

 of ascorbate and O3 is given by the molecular rate constant of 

4.8 × 107

 M!1 s!1. This is some 50,000× the rate constant for ET. Of course, it depends upon the 

relative concentration of ascorbate and ET, but it is likely that ascorbate is in higher 

7

These chemical rate constants are those constants within a bulk solution. In the apoplasm, the possibility 

exists for the chemicals to be preferentially oriented near a surface; so the constants may not be the same as for bulk 

solutions. 

AX9-60 
concentration than ET. One would then expect that the rate reaction of ascorbate with O3 would 

greatly dominate any possible reaction of O3 with ET. For a concentration of ascorbate in the 

range of 1 mM and for an O3 concentration of about 0.1 ppm or 4.2 × 10!9

 M, the detoxification 

rate would be 4.8 × 107

 × 10!3 × 4.2 × 10!9 M s!1 = 2.0 × 10!4 M s!1. Turcsányi et al. (2000) 

calculated an O3 flux of about 1.6 × 10!9

 moles m!2 sec!1. With a wall thickness of 0.12 × 10!6 m 

and all the O3 flux going into the wall region, this would give about 1.3 × 10!2

 mol m!3 s!1 or 

1.3 × 10!5

 M s!1 flux, which is less than 10% the detoxification rate. 

Glutathione 

Many of the initial studies of O3 exposure used high concentrations and measured only the 

total sulfhydryl contents of the tissues. For example, in some of the earlier work, exposures of 

tobacco to 1 ppm O3 for 30 min induced a 15% loss of the total sulfhydryls (0.74 μmole/g-FW) 

(Tomlinson and Rich, 1968). These results are similar to other studies at high O3 levels (Dugger 

and Ting, 1970). It is now suspected that the severe injury in their studies resulted in a massive 

collapse of the cells and release of most of their internal constituents. Much of the oxidation 

thus observed may have been the result of chemical oxidations of the O3 that subsequently 

entered the damaged tissue. Even under milder conditions, changes in sulfhydryl components 

have still been noted and any sulfhydryl on the surface of the cell would be at risk due to its high 

reactivity with O3 (Mudd et al., 1997b; Mudd et al., 1969). For example, the level of sulfhydryl 

compounds within the protein of isolated chloroplasts declined about 66% when the chloroplasts 

were subjected to O3 (about 1 μmole O3) exposure (Mudd et al., 1971). 

At this stage, it is important to note that there are inherent problems with metabolic studies 

of full tissues. The first is that most organs have several different types of tissues. For example, 

leaves have, at the minimum, epidermal and vascular tissues and two types of mesophyll cells. 

Each type of cell may be metabolizing quite differently and producing very different levels of 

metabolites and enzymes. Furthermore, most pathways are well regulated and after any small 

disruption, the pathway tends to return to near its former stability. Changes in the level of 

enzymes are likewise difficult to measure. Many enzymes function below their maximum 

activities. Their speeds of reactions are often increased through regulation, rather than through 

the production of more enzyme. 

AX9-61
Glutathione is a three-amino acid peptide, which has antioxidant properties due to its 

free reducing sulfhydryl group (G-SH). Glutathione is generally kept in its reduced form 

by +glutathione reductase (GR) with the reaction: 

(AX9-2) 

GR has six isoforms8 within the chloroplast and six isoforms outside. The optimum activity 

occurs at pH 7.8, suggesting it is located within the stroma of the chloroplast or in the cytoplasm 

rather than in the cell wall, which is at pH 4-5 (Madamanchi et al., 1992). Clearly, an increased 

expression of GR (generated through transgenic implants) is important within the chloroplast to 

prevent some oxidations9

 (Aono et al., 1995). 

Catalase 

Catalase, even though it breaks down H2O2, does not appear to protect plants from O3 

exposures. Two principal reasons may cause this lack of reactivity: (1) catalase has a high Km 

for H2O2 and a low rate coefficient, and (2) catalase seems not to occur within the cell wall 

regions but rather in the cytoplasm and peroxisomes (Buchanan et al., 2000). While a few 

reports suggest that catalase is increased by exposure to O3 (Azevedo et al., 1998), Booker et al. 

(1997) found no effect of catalase activity in soybean until late in the growing season, and others 

have found decreased catalase activity in wheat in response to O3 (McKee et al., 1997). 

Unfortunately H2O2 induced by some forms of wounding in mesophyll cells can lead to 

induction of an increase in GSH and the transient production of catalase (Vanacker et al., 2000). 

In general, it seems that catalase is not really involved primarily in the defense of the cell due to 

O3 attack but rather may be a secondary response. The reaction of catalase (Scandalios, 1993) is 

as follows: 

8

An isoform is the same enzyme, with the same structure and perhaps within the same organelle, but its 

promoter region has different DNA codes. Thus, each protein segment is induced by different signals, and so its 

enzyme can be formed in response to different environments. This is in contrast to isozymes, which classically are 

similarly reacting, but structurally different, enzymes in different compartments. 

9

Typically this protection is observed in the paraquat sensitivity of plants. In this assay, added paraquat, 

the herbicide which intercepts electrons from the reducing end of photosystem I in the chloroplast, caused 

oxidations, chlorophyll loss, and death due to the buildup of superoxide and peroxides. 

AX9-62 
(AX9-3) 

Superoxide Dismutase 

The varied compounds that O3 can produce upon entering an aqueous solution are very 

similar to those involved in the HR when plants are infected by an avirulent pathogen 

(Figure AX9-10). The sequence of the plant response to the pathogen is (1) recognition of the 

gene products of the pathogen by the plant (elicitor), (2) generation of an immediate 

phytoresponse to attempt to localize the attack and its products, and (3) generation of a systemic 

acquired resistance (SAR) to subsequent attack by the pathogen. Inducible defense responses are 

phytoalexin synthesis and production of pathogenesis-related proteins (PR). One aspect of this 

total response is the production of O2 

!

 and H2O2 by the cell (Lamb and Dixon, 1997). The 

elicitor can generate a transient alkalinization of the apoplast, up to pH 7.2, caused by a lowering 

of the H+

-pump rate and an increase in the H+-influx/K+-efflux exchange. Other effects include a 

weak accumulation of transcripts for PAL (phenylalanine lyase); a larger and rapid induction of 

glutathione S-transferase, GSH-Px; oxidative cross linking of cell wall proteins which is blocked 

by ascorbate acid; generation of localized apoptosis; and rapid influx of Ca2+, which activates 

apoptosis among other pathways (Lamb and Dixon, 1997). These effects seem to be very similar 

to those induced by O3 exposure (Sandermann, 1996, 1998). 

The putative antioxidant enzyme SOD (Equation AX9-4 and Table AX9-8) catalyzes the 

oxidoreductase reaction, which eliminates SO2 

!

 by dismutation (Bowler et al., 1992): 

(AX9-4) 

The number, as well as the activity, of isozymes of each type of SOD in Table AX9-3 can 

vary with plant species. However, the isozymes that have been tabulated are Cu-Zn SODs, in 

cytosol and chloroplast; Fe-SOD, active in chloroplast stroma; and Mn-SOD, active in 

mitochondrial matrix (Karpinski et al., 1993). In the experiment demonstrating the activation of 

varied SODs, there were three Cu-Zn SODs (csd1, csd2, csd3), three Fe-SODs (fsd1, fsd2, fsd3), 

and one Mn-SOD (msd1) (Kliebenstein et al., 1998). Ozone sensitivity was determined by 

AX9-63
Table AX9-8. Superoxide Dismutase Isozymes and Isoforms

 Reaction: 2H+ + 2 O2 

! 

6 H2O2 + O2

 Isozymes M.W. Isoforms Cytolocation

 Cu-Zn 20 kDa csd1 

csd2 Plastid 

csd3 Peroxisomal

 Fe 23 kDa fsd1 Mitochondrial 

sd2 

fsd3 Plastid

 Mn 23 kDa msd1 Mitochondrial 

exposure of plants to 8 h of 0.33 ppm of O3. csd1 induced by O3 and UV-B was one of the 

earliest SOD increases and most pronounced responses for mRNA and protein. Also, some 

increase in csd3 (thought to be peroxisomal) was induced when the plants were exposed to a 

high-intensity light pulse; msd1 was unresponsive to the environmental stressors used here, 

including O3; and csd2 (thought to be chloroplast) showed little increase. The fsd1 isozyme 

(present in the apoplasm) showed a slight decrease. On the other hand, an early report on snap 

beans in which the experimenters used EDU, N-[2-(2-oxo-1-imidazolidinyl)ethyl]-N-phenylurea 

(Carnahan et al., 1978; Kostka-Rick and Manning, 1993), to prevent visible injury by O3, 4-h O3 

exposure at 0.45 ppm was correlated with an increase in general enzyme activity of SOD, i.e., 

the level rose nearly 2.5× in 2 weeks at a level of 50 mg EDU per pot (Lee and Bennett, 1982). 

It is believed that EDU may induce SOD, which then protects the plant. While gross assays of 

enzyme activity have not proven to be very useful in understanding the mechanism of O3 action, 

in a well-crafted, long-term study involving ponderosa pine clones. Benes et al. (1995) stated 

that “changes in antioxidant enzyme activity were not a consistent response to the O3 fumigation, 

but when observed, they occurred most often in the O3-sensitive clone and in symptomatic, 

fumigated branches...total (intra- and extracellular) activities of the antioxidant enzymes did not 

appear to be good indicators of O3 tolerance....” 

AX9-64
Ozone exposure (70 ppb for 7 h/day for 14 to 42 days of exposure)10 caused an increase in 

POD and a decline in SOD with no change in APX. No GSH was detected, but the 

concentration of (ASC + DHA) was at 20 to 25 nmol/g-FW of extracellular fluid, compared with 

2.4 to 3.0 mmol/g-FW of cell fluid. Glutathione within the cell was only 100 to 170 nmol/g-FW 

of cell fluid. While these results are what one might expect for POD, the decline in SOD and 

lack of change in APX are not what would be expected if protection was provided by SOD and 

ascorbate. Yet as noted, because the rate of SOD reaction is many times higher than the rate of 

O3 entry, there may be no pressure to increase the SOD level. 

Some protection against visible injury (induced by 59 ppb daily mean O3 for 14 h/day for 

7 days) was observed in genetically modified tobacco plants with excess chloroplast SOD11 (2 to 

4 times higher), but less protection was observed in plants that had an excess of mitochondrial 

SOD (8× higher) (Van Camp et al., 1994). In all lines, the conductance of the leaves dropped 

about 50%, compared with the unmodified plants. There was a correlation with age of leaf (less 

injury in younger leaves) that corresponded to that found in spruce trees in which the amount of 

SOD declined in relation to the longer that needles were held on the tree (Polle et al., 1989). 

A slightly different study, however, found no O3 protection with varied SOD within the needles 

(Polle and Rennenberg, 1991). Interestingly, in maize, the synthesis of SOD (any form) was not 

stimulated by O3 exposure (at 0.50 or 0.75 ppm for 8 h, variable times thereafter) but was by 

exposure to 90% O2 (Matters and Scandalios, 1987). It may be that this high level of O3 does not 

affect the SOD, or perhaps it stimulates and degrades the enzyme simultaneously. 

The conclusions to be drawn from these results are not obvious. There seems to be SOD 

(a Cu-Zn form) present in the apoplastic space of some plants, but it does not necessarily rise 

with O3 exposure. Thus, either its concentration is sufficient to provide protection or it is not 

needed. Over expression of any SOD in other organelles may play a role, especially in the 

chloroplast (Cu-Zn or Fe forms); however, it may be playing a secondary role due to other 

effects of O3 that generate conditions in which light can overload the chloroplast and generate 

10At a level of 70 ppb, the concentration of O3 in air was about 3.06 × 10!6

 mol/m3, which with the m2 conductance of 0.042 mol/m2

 s, gives a flux rate of O3 of 1.27 × 10!8 mole/ s. Converting the SOD rate of 

23 units/g-FW into a SOD rate within the apoplastic space of 6.9 × 10!3

 mol/m2 s, or about 500,000 times the 

entry rate of O3. 

11The SOD enzymes were from Nicotiana plumbaginifolia with appropriate transit sequence for targeting 

the correct organelle and expressed under control of cauliflower mosaic virus 35S promoter. 

AX9-65 
! detrimental circumstances, including the production of SO2 . In addition, SOD is 

developmentally expressed in varied concentrations, so that long-term exposure to O3 may alter 

each leaf’s developmental age and, in turn, alter what level of SOD is observed. In any case, 

SOD does not seem to be the primary antioxidant system to protect against O3. 

Changes to the Plasmalemma 

Reports of “peroxidation” generally occur within unicellular organisms subjected to very 

high levels of O3 (in Chlorella [Frederick and Heath, 1970] and in Euglena [Chevrier et al., 

1990]). Heath (1987) determined that by the time biochemical events were altered and MDA 

was produced in Chlorella, little permeability remained in the cells and most metabolic 

pathways were greatly disrupted by the subsequent loss of substrates. In fact, MDA production 

was concurrent with a high O3 uptake by the cell, indicating a complete opening of the cell and 

associated with the concurrent inability to plate the cells on a glucose medium (indicative of cell 

death). Heath (1987) reached the conclusion that no one had proven that lipid oxidation was in 

any way a part of the initial reactions of O3 with the cell, a conclusion confirmed by Mudd et al. 

(1997a). An excellent review regarding the initial action induced by O3 within a plant 

(Kangasjärvi et al., 1994) should be consulted. There is little data to show that lipids are 

attacked by O3 in any living system that was not previously severely injured by O3. Most of the 

data suggesting lipid attack by O3 has been demonstrated in plants subjected to O3 concentrations 

of 0.5 to 1.0 ppm for several hours, during which gross wilting of the plant tissues usually 

occurs, suggesting extreme water loss. It is not surprising that lipid and protein injury are 

observed under these conditions. While those reports were useful in the 1960s and 1970s, they 

are not especially insightful now when ambient levels of O3 are rarely above 0.2 ppm. 

AX9.2.4 Wounding and Pathogen Attack 

The decline of an enzyme is more difficult to measure than the rise of a new enzyme; an 

increase from 0 to 2% may be within the precision of any assay, but a decrease from 100 to 98% 

is often masked by simple variation of the assay. Thus, measuring enzymes, which are in great 

abundance in prefumigated tissue, is a risky operation. On the other hand, if O3 induces a 

general physiological change that has characteristics similar to other well-studied, stress-induced 

AX9-66
changes, then O3 studies could “piggyback” onto those studies to gain insight into the full scope 

of metabolic alterations. It is now becoming clear that wounding and pathogen attack of plants 

are similar to O3-induced changes in plants, and a reasonable hypothesis is that O3 must induce 

one or more of the first steps seen in the wounding/pathogen-attack response. 

Systemic acquired resistance (SAR) has been heavily investigated, and DNA probes have 

existed for some time for a series of expressed genes (see Table AX9-9). Several enzyme 

classes are associated with O3 injury, including glucanases and peroxidases and others, such as 

the PR proteins and chitinases. Thus, strong evidence exists from enzyme function and genetic 

material that O3 induces an activation of a SAR-like response. 

Table AX9-9. Gene Families and cDNA Clones Used as Probes for SAR 

(Ward et al., 1991) 

Probe Relevant Properties of Encoded Protein Reference 

PR-1 Acidic, extracellular; function unknown most 

abundant PR protein in tobacco; >90% identical to 

PR-1b and PR-1c 

Payne et al. (1989) 

PR-2 Acidic, extracellular b-1,3-glucanase, >90% identical 

to PR-N and PR-O 

Ward et al. (1991) 

PR-3 Acidic, extracellular chitinase; also known as PR-O; 

>90% identical to PR-P 

Payne et al. (1990) 

PR-4 Acidic, extracellular; unknown function; homologous 

to C-terminal domain of Win1 and Win2 of potato 

Friedrich et al. (1991) 

PR-5 Acidic, extracellular; homologous to thaumatin and 

bifunctional amylase/proteinase inhibitor of maize; 

also known as PR-R or PR-S 

Payne et al. (1989) 

PR-1 basic Basic isoform of acidic PR-1 Payne et al. (1989) 

PR-O' Acidic, extracellular b-1,3-glucanase; approximately 

55% identical to PR-2 group 

Payne et al. (1990) 

Basic, glucanase Vacuolar; approximately 55% identical to PR-2 group 

and PR-O' 

Shinshi et al. (1988) 

Basic chitinase Vacuolar; approximately 65% identical to PR-3 group Shinshi et al. (1987) 

Acidic peroxidase Extracellular; lignin-forming Lagrimini and Rothstein (1987) 

AX9-67
Mehdy (1994) described a model of how an elicitor produced by the pathogen attack 

activates a G-protein, which opens the inward-flowing Ca2+ channel. The flow of Ca2+ into the 

cytoplasm raises the internal level (at the μM level) and activates a protein kinase that increases 

! the activity of the plasma membrane NAD(P)H oxidase and generates O2 . Superoxide 

dismutase converts O2 

!

 into H2O2. Both O2 

!

 and H2O2 are responsible for the active oxygen 

species response, which is believed to be a defense mechanism to kill the pathogen. In this 

normal defensive reaction, a subsequent system induces either localized lipid peroxidation per se 

or a membrane lipase to produce jasmonic acid or inositol triphosphate, which act as secondary 

messages to activate the defense gene products. 

Booker et al. (2004) found that G-proteins might be involved in the perception of O3 in the 

extracellular region using A. thaliana G-protein null mutants. The activation of a passive inward 

flow of Ca2+, e.g., by an O3-induced response, would serve the same function as activation of the 

G-protein. Once the level of cytoplasmic Ca2+ rises, all else follows. It is suspected that 

exposure of plants to O3 does just that, as Castillo and Heath (1990) demonstrated — the in vivo 

fumigation of bean plants both inhibits the outward-directed ATP-requiring Ca2+ pump and 

increases the passive permeability of Ca2+. It was thought that the calcium transporter system 

has a sensitive sulfhydryl group which, if oxidized, would alter normal Ca2+ movements. 

In addition, Dominy and Heath (1985) observed that the K+

-activated ATPase (believed to be 

involved in K+ transport) was inactivated by in vivo exposure to O3 and that inactivation was 

traced to a sensitivity sulfhydryl. Mudd et al. (1996) argued that several amino acids are very 

sensitive to O3, including any with an exposed sulfhydryl. Thus, the O3-induced change in Ca2+ 

permeability may be the trigger to most, if not all, the wounding responses. However, the 

difficult problem of proving that the cytoplasmic Ca2+ change is the first event in O3 injury 

remains. 

Some wound- and pathogen-induced genes that are activated or repressed in Arabidopsis 

thalania are found with DNA arrays (Cheong et al., 2002). While these responses may not be 

uniform for all plants, they suggest the possibility of wide-ranging gene changes that may occur 

with a simple wound and that those changes are wide-ranging and diverse. As an example, these 

responses are related to hormonal responses that are related to jasmonic acid, ET, and auxin 

pathways; signal transduction responses; and transcription factors for a variety of pathways. 

The involvement of ET in wounding and pathogen attacks is discussed in Section 9.2.3.3. 

AX9-68
AX9.2.4.1 Peroxidases 

Increases in cytosolic and apoplastic peroxidase activity in response to O3 are often 

observed, but the reasons and outcomes of these changes have yet to be fully explained. 

Increased activity is frequently correlated with O3 injury. Dass and Weaver (1972) observed that 

increases in peroxidase after O3 injury was similar to that observed for plant infection by a virus. 

Tingey et al. (1975) observed a 35% decrease in peroxidase activity immediately following O3 

exposure; however, within 24 to 48 h, activity had increased significantly and was above control 

level and remained there throughout the remainder of the study. Dijak and Ormrod (1982) also 

observed increases in peroxidase activity when two O3-sensitive and two O3-resistant varieties of 

garden peas (Pisum sativum) were exposed to O3. Peroxidase activity was not related to cultivar 

sensitivity nor to visible injury. Unfortunately, there are many peroxidases (Birecka et al., 

1976); therefore, any general increase is not specific. In ET-treated leaves, peroxidase reaction 

products were found between the plasma membrane and the cell wall, suggesting that ET itself 

could induce peroxidase activity (Abeles et al., 1989a,b; Birecka et al., 1976). 

At the same time, others examined peroxidase reactions in general and found two types of 

peroxidases (designated as acidic or anionic and basic or cationic, EC 1.11.1.7, but also listed as 

EC 1.14.18.1). Many types of peroxidases are located in diverse organelles, and each seems to 

be activated by different conditions (e.g., pH for anionic and cation types and substrates such as 

guaiacol, syringaldazine, and ascorbate). Peroxidases belong to at least two groups, which 

catalyze two separate reactions: (1) the reaction of H2O2 with ascorbate to form DHA, discussed 

earlier (Thom and Maretzki, 1985), which is regenerated by plasma membrane electron transport 

using a dehydrogenase (Gross and Janse, 1977) now believed to be a malate/oxaloacetate shuttle 

through the membrane coupled to a NAD(P)H-cytochrome-b-reductoxidase; and (2) the reaction 

with coniferyl alcohol (from phenylalanine through phenylalanine ammonia lyase) to form lignin 

within the wall. The anionic peroxidase thought to be involved with lignification is within the 

cell wall (Buchanan et al., 2000; Taiz and Zeiger, 2002). Some basic peroxidases are maintained 

within the cell, while some are external to the cell. After wounding (Gasper et al., 1985; 

Lagrimini and Rothstein, 1987), some basic peroxidases can be activated by processes leading to 

the synthesis of stress ET (Yang and Hoffman, 1984) and/or by excess Ca2+ (Gasper et al., 1985). 

Elicitor treatment of plants change a series of peroxidases, some of which are similar to those 

seen in O3-induced changes (see Table AX9-9). 

AX9-69
The formation of lignin is due to the phenylpropanoid metabolism (Buchanan et al., 2000). 

Tyrosine and phenylalanine are converted to cinnamic and p-coumaric acid, which are, in turn, 

converted to p-coumaryl, coniferyl, and sinapyl alcohols, and then into lignins. Hence, the 

peroxidase activity is often measured by one of these substrates (Espelie et al., 1986; Gasper 

et al., 1985). However, it is questionable whether apoplastic peroxidase activity is limiting for 

lignification; laccases have a prominent role as well. Also, availability of monolignols is critical 

for core lignin formation, and it is unclear whether levels of these metabolites change in 

response to O3. Studies by Booker and others (Booker et al., 1991, 1996; Booker, 2000; Booker 

and Miller, 1998) indicated that O3 did not increase core lignin concentrations in foliage of 

loblolly pine, soybean, or cotton; although levels of phenolic polymers and cell wall-bound 

phenolics were elevated in soybean. Increased phenolic polymers appear to be lignin in acidinsoluble lignin assays and may well be responsible, along with polyphenol oxidase, for the 

stippling injury observed in O3-treated plants. Cell wall function implies the transport of 

peroxidase molecules out of the cell and, most likely, the regulation of their activities within the 

wall space. These extracellular peroxidases may be observed by vacuum infiltration of buffer 

into leaf air spaces and subsequent centrifugation of the tissues to remove the buffer with the 

apoplastic enzymes that wash out (Castillo and Greppin, 1986). However, exposure to O3 

induces important changes in the plant. For example, extracellular peroxidase activity in Sedum 

album leaves increased nearly 3-fold over that in the control plants after 2-h exposure to 0.40 

ppm O3 (Castillo et al., 1984). This O3-induced increase in extracellular peroxidase appears to 

be under the control of Ca2+ (Castillo et al., 1984; Heath and Castillo, 1987). Initially, no effect 

on the anionic activity as measured with syringaldazine (specific electron donor for lignifying 

peroxidases) was observed, yet 21 h later, the anionic peroxidase activity was increased, whereas 

the cationic (ascorbate measured) peroxidase activity was decreased, in O3-treated plants. This 

suggests an immediate response (ascorbate peroxidase activation) and a secondary response that 

activates the lignifying peroxidase via gene activation. 

The rapid response of cationic peroxidase after O3 exposure may not result from de novo 

protein synthesis but from the secretion and direct activation by Ca2+ ions of enzyme molecules 

already present in the tissue. Cationic peroxidases might attack the peroxides and, in this 

manner, act as a detoxifying agent with ascorbate as the substrate in the apoplasm. The effect of 

Ca2+ upon peroxidase activity is stronger at low H2O2 concentrations (Penel, 1986). Thus, one 

AX9-70
can imagine that, when the H2O2 concentration is low, this peroxidase activation would have a 

greater in vivo importance. Furthermore, the secretion of cationic peroxidases into the free 

spaces as a result of O3 treatment is accompanied by a simultaneous release of at least one of its 

natural substrates (ascorbic acid); this cationic peroxidase exhibits a much higher affinity toward 

ascorbate (up to 6-fold) than the anionic isozyme (Castillo and Greppin, 1986). 

AX9.2.4.2 Jasmonic Acid and Salicylic Acid 

Jasmonic acid (JA) and salicylic acid (SA) are considered to be regulators of the plant 

defense response (Figure AX9-13) (Buchanan et al., 2000). They tend to respond more slowly 

than ET, causing widespread effects in the plant tissues. Both seem to be heavily involved in 

responses of the plant to O3, once again linking the pathogen/wounding defense to O3-induced 

injury; however, their roles are far from clear. 

Figure AX9-13. The pathway leading from phospholipids to jasmonic and traumatic acid. 

The role of lipoxygenase and the production of a hydroperoxyl moiety 

from the unsaturated fatty acid are clearly demonstrated. More 

importantly, several of the enzymes within this pathway have been 

shown to be activated by oxidative conditions, including O3 exposure. 

The production of both of these acid species could lead to a general 

global response of a whole plant to the O3 exposure of a single leaf. 

Sources: Howe et al. (2000); Croft et al. (1993); Buchanan et al. (2000). 

AX9-71 
One of the lipoxidase isoforms is activated by pathogen infection (POTLX-3) within 6 h 

and accumulates for a week (Kolomiets et al., 2000). This enzyme is the first stage of the JA 

pathway which leads to 13-hydroperoxide linolcenic acid (HPOT) which is converted either to 

allene oxide through AOS or to C6 aldehydes through hydroperoxide lyase. These aldehydes act 

as signaling agents via systemic (Sivasankar et al., 2000) or volatile odiferous compounds 

(oxylipins) that have been implicated as antimicrobial toxins (Froehlich et al., 2001). 

Interestingly, these compounds seem to target the chloroplast envelop where they interact with 

its metabolism. As HPOT and AOs are both implicated in plant defense and are activated by O3, 

these interactions may be related to how chloroplast enzymes and their mRNAs are involved 

in O3-induced injury. 

AX9.2.4.3 Stress-Induced Alterations in Gene Expression 

Early studies addressed the qualitative and quantitative effects of O3 on protein metabolism 

(Harris and Bailey-Serres, 1994). Subsequent reports suggested that the physiologic and 

metabolic consequences of exposure to O3 were, in part, mediated by increased gene expression. 

A summary of the gene-linked changes in proteins induced by SAR was provided earlier in 

Table AX9-9. Of particular note are the productions of PR proteins, chitinase, glucanase, and 

acidic peroxidases that appear to be common markers used in many O3 studies. A summary of 

varied proteins as measured by changes in the mRNA in A. thalania induced by O3 exposure is 

shown in Table AX9-10. While studies on Arabidopsis thaliana required high concentrations of 

O3 to produce a response, the levels reported in most of the studies did not induce visible injury. 

The types of messages induced included glutathione S-transferase, PAL, ACC synthase, SOD, 

and some PRs. Slower increases in messages are seen for other PR and SAR-senescence 

proteins. Declines in messages were observed for varied chloroplast enzymes, including those 

for Rubisco and chlorophyll binding proteins. A few new proteins were found — a casein kinase 

and three plasma membrane proteins. It is interesting to note that few messages for “new” 

proteins were generated by O3 exposure. 

The working hypothesis is that O3, which is not eliminated by antioxidants in the cell wall, 

alters the properties of the plasma membrane. Specific polypeptides, indicative of these 

antioxidants, are induced. If specific receptor molecules or channels on the membrane are 

affected, the ionic balance within the cytoplasm is changed, leading to altered transcription or 

AX9-72
Table AX9-10. Proteins Altered by Ozone as Measured by Molecular Biological Techniques as mR

Other Gene Activity Rather than Enzyme Activity 

AX9-73

Identified Proteins Fast Increase Slow Increase Examined, Unknow

Exposure Physiological Events Response Response Decline But No Change Protein

150/300 ppb for Leaf curling; reduced GST, PAL Pxase, SOD CAT, LOX1 6 h daily growth 300 ppb for 6 h daily 10 bands of 10 RNA AtOZI1 

kinase I

200/500/1000 ppb Wilting (8 h); premature 3 plasmfor 2 h senescence membra

proteins35-kDa 

350 ppb for 1-6 h Ethylene production; ACS-6 ACS-1, -2, -4, -5 

downward curvature; water logging 

300 ppb for 6 h Necrosis in NahG and Cvi-0 Chl SOD, Chl GPX Cab mRNA, 

(accumulating SA), not in cytAPX, GST1 cyto SOD, Col-0 chl GR 

150 ppm for 6 Downward rolling of leaf; BCB, ERD1, Cab, rbcS Atgsr2, MT1, h/8 and 14 days early senescence SAG21 SAG 12, SAG 13, 

SAG 19, SAG 20 

160 ppb for 3-72 h Early senescence GST1,VSP2 MT1 CCH (a) 250 ppb for 8 h; (a) little chlorosis or lesions; GST Apx, PAT1 Fe-SODl GR, (b) 250 ppb for 2 h; (c) growth retardation CuZn-SOD cab, rbs (c) 175 ppb for 8 h/4 days 
Table AX9-10 (cont’d). Proteins Altered by Ozone as Measured by Molecular Biological Techniques a

Other Gene Activity Rather than Enzyme Activity 

Identified Proteins Fast Increase Slow Increase Examined, Unknow

Exposure Physiological Events Response Response Decline But No Change Protein

200 ppb for 24 h 

PR-1, PR-2a, PR-5, AtEDS1, AtGST1, AtGST2 

PR-3b, PR-4 

LOX2, AtOZI1, PAL, Lhcb, PAT1, HSP 

rcd1, on 

250 ppb for 6 h Lesion initiated on margin and spread inward chromo

single Mtrait 

AX9-74

Abbreviations used in Tables 9-9 and 9-10. 

GST = Glutathionine synthase PAL = Phenylalanine ligase PR-1 = Promoter region 1 Pxase = Peroxidase CAT = Catalase LOX1 = Lipoxygenase ACS-6 = ACC synthatase SOD = Superoxide dismutase CuZn-SOD = cyto SOD Fe-SODl = Chl SOD 

cyt APX = Ascorbate peroxidase 

Chl GPX = Glutathione peroxidase 

Cab mRNA = Chlorophyll a/b binding protein 

chl GR = Gluthatione reductase 

BCB = Blue copper binding protein 

ERD1 = Ethylene response 

SAG21 = Senescence 

rbcS = Rubisco small subunit 

MT1 = Mitochondria 
translation of the genes controlling those and other types of polypeptides. Once this membrane 

disruption occurs, the cell must mobilize repair systems to overcome the injury. Thus, carbon 

and energy sources once destined for productivity, must be used in repair processes. Some of 

these repairs are thought to result from the induction of specific genes. Photosynthesis is 

inhibited by direct inhibition of some of the enzymes, through byproducts of O3 attack or 

by altered ionic balance. At the very least, the decrease in photosynthesis is a result of 

an O3-induced decrease in rbcS mRNA. 

AX9.2.5 Primary Assimilation by Photosynthesis 

AX9.2.5.1 Photooxidation: Light Reactions 

Photooxidation refers to the oxidation of chlorophyll within the light reaction due to an 

imbalance between light absorption and the CO2 use to produce carbohydrates. It was 

discovered in the 1920s and studied under the concept of chlorophyll bleaching and photoautooxidation (Asada, 1999; Rabinowitch, 1945). What generally occurs is that electron transfer 

from H2O to NADPH declines, and a light reaction overload occurs. The slowdown of electron 

transfer may also be due to inhibition of the dark reactions, through the poor use of small 

molecular weight carbohydrates or a lowered amount of the fundamental substrate CO2. 

To counteract these detrimental reactions, a series of “antioxidant” reactions exist, which 

eliminate the buildup of oxidative intermediates. 

A lowered CO2 level, which can be caused by stomatal closure (Heath, 1996), blocks the 

use of reduced plastoquinone (PQH2) in Photosystem II through NADP reduction in 

Photosystem I (Hankamer et al., 1997). The buildup of PQH2 reduces the amount of QA, 

resulting in a buildup of P680

+

|Pheo! species (the primary photoact). The inability to reduce this 

radical leads to injury to the D1 protein (32 kDa) and its fragmentation into 23-, 16-, and 10-kDa 

fragments (Hankamer et al., 1997). Ozone exposure of bean plants leads directly to the loss of 

this D1 protein (Pino et al., 1995). The loss of D1 stimulates the production of new D1 (and its 

mRNA). Also, the production of the oxidized form of P680 (P680

+

) is harmful to the plant, because 

electron flow from water to P680

+

 is limited, generating a P680

T

 (the triplet form of P680), which is 

highly oxidizing and can lead to dangerous reactions. One form of protection is the use 

of $-carotene to convert the triplet form back to its normal state; however, that reaction can lead 

to the loss of $-carotene. Without the protection of $-carotene, oxygen reacts with oxidized 

AX9-75
products to produce singlet state of O2. This, in turn, can react with chlorophyll, leading to ring 

breakage that, in essence, leads to chlorosis. These types of reactions do not seem to occur 

often, but chlorosis is one form of visible injury, and loss of $-carotene has been reported. 

Using a FACE system to expose soybean to elevated O3, Morgan et al. (2004)found that, 

in leaves at the top of the canopy, there was no effect on the maximum light-adapted apparent 

quantum efficiency of PSII (NPSII,max), electron transport at growth [CO2], and saturating 

light (Jsat) nor on the probability of an absorbed photon reaching an open PSII reaction 

center (FvN/FmN; the quantum yield). There was a small, but significant, decrease in 

photochemcial quenching (qP) at the top of the canopy. As leaves aged, the decrease in qP 

was significant as leaves began to senescence, likely due to losses of chlorophyll. The results 

of these studies and others suggest that alterations to the dark reactions are much more common 

than to light reactions (Farage et al., 1991; Farage and Long, 1999). 

AX9.2.6 Alteration of Rubisco by Ozone: Dark Reactions 

A large body of literature published since 1996 shows that O3 exposure affects Rubisco 

concentrations (Pell et al., 1997). Treatment of a variety of plants with O3 at near-ambient levels 

results in a loss of Rubisco and of the mRNA coding for both subunits of Rubisco (rbcS, small 

and rbcL, large). In addition, oxidation of Rubisco by O3-generated ROS may be an important 

factor in suppressing photosynthesis (Pell et al., 1997). Increased carbonyl concentrations of 

Rubisco are correlated with O3 injury (Kanoun et al., 2002; Leitao et al., 2003). Because 

Rubisco plays such an important role in the production of carbohydrates (Figure AX9-14), any 

loss may have severe consequences for the plant’s productivity. 

Chronic O3 exposure, both with and without elevated CO2, significantly lowered 

assimilation and leaf conductance of soybean in aging mature leaves (Fiscus et al., 1997; Reid 

and Fiscus, 1998), which was associated with significant decreases in Rubisco content in aging 

leaves. Noormets et al. (2001b) also found that O3 had the greatest effect on older leaves of 

aspen clones using a FACE exposure facility in which areas of ambient CO2 (daytime 360 ppm) 

and ambient with added CO2 (560 ppm), with added O3 (97.8 ppb), and with added CO2 and O3, 

were used. They found an O3- induced decline in assimilation and in conductance, and 

subsequently confirmed that the internal CO2 (calculated for within the leaf) is not affected by O3 

exposure. Higher levels of CO2 increased the assimilation and lowered the conductance, 

AX9-76
Figure AX9-14. The production of Rubisco and its Calvin Cycle pathway reactions. Two 

peptides are used to build Rubisco: rbcS, the small subunit produced by 

DNA within the nucleus; and rbcL, the large subunit produced by DNA 

within the chloroplast itself. Clearly both polypeptides must be closely 

regulated to produce the enzyme in a coherent manner. Furthermore, at 

least five isoforms of DNA can produce rbcS, each of which is regulated 

by a different promotor region. 

maintaining the internal to external CO2 ratio identical to that found with the ambient CO2 level, 

corresponding to the theory of Farquhar et al. (1980). More to the point was that the stomatal 

limitation12 was not altered by O3 exposure, with or without excess CO2. It is critical to point out 

that mesophyll conductance is directly linked to internal CO2 level13. So if Ci

/Co (Co = CO2 

outside) is constant and gs declines, then gm must likewise decline. If, as it is argued, Rubisco 

levels are constant or at least increasing, then a regeneration of RuBP must be the cause of the 

12The limitation was defined as the ratio of stomatal resistance to the total resistance, which included the 

operating point of the assimilation (A) verses internal CO2 concentration (Ci

) curve and the resistance of the 

boundary layer. The operating point of the curve was defined as the internal CO2 level, which is calculated by the 

conductance and assimilation. The resistance of this operating point was calculated as the cotangent of the slope to 

the operating point. Unfortunately, the slope is not a dimensionless parameter but is rather moles of air per area of 

leaf –

s of time and, thus, it is unclear whether the slope changes with added CO2 and O3. 

13Respiration is generally small at saturating A and often is ignored. By transforming the term {A = gs (Co 

– Ci)} into {A = gs Co – gm Co = (gs – gm ) Co} where gm = gs (Ci/Co) or the mesophyll conductance in earlier 

literature. 

AX9-77 
decline in gm. Farquhar et al. (1980) were more concerned with high levels of CO2 and had little 

to say about O3 exposure. 

In a similar study, Morgan et al. (2004) examined elevated O3 using a FACE system to 

increase O3 by 20% over the entire growing season. They examined O3 effects from emergence 

through the entire life cycle to senescence. Leaf photosynthetic performance was measured 

using a LI-COR 6400 with integrated chlorophyll fluorescence capabilities to examine both dark 

and light reactions. This study found no effect of elevated O3 on newly expanded leaves over the 

growing season. There were little O3-induced changes in the light reactions; however, as leaves 

aged, there were significant changes in the dark reactions. For example, there were significant 

losses in the maximal photosynthetic assimilation in saturating light (Asat) and the maximum rate 

of carboxylation (an in vivo measure of Rubisco efficiency), and maximum rate of electron 

transport for the regeneration of RuBP (Jmax). The findings showed the greatest impact of O3 on 

the oldest leaves and demonstrated the significant impact on seed production. 

The level of carbohydrate within the cell has an effect upon the amount of mRNA for 

Rubisco (rbcS). Experiments by Krapp et al. (1993) indicated that a decline in carbohydrate 

levels is probably due to the increased production of control metabolites, such as fructose 

2,6-bisphosphate, which can shut down important sugar production pathways. This report also 

leads to a measure of half-time for the decline in rbcS of about 2 days14 when 50 mM glucose is 

added to a cell suspension of Chenopodium. Also, the carbohydrate level was increased by cold 

girdling the petioles of intact tobacco and potato plants. The levels of carbohydrate nearly 

doubled in 5 days and the level of rbcS declined rapidly (reaching 25% after 12 h). A decline in 

Rubisco followed, but more slowly (with an estimated half-time of about 108 h after a lag of at 

least 12 hours). This, of course, is expected; the level of the enzyme would decline slowly with 

a lag after a loss of the message. 

A better estimation of the half-life of rbcS can be found in the Jiang et al. (1993) study of 

the destabilization of the message by an antisense message. The wild-type rbcS in tobacco had a 

half-life of about 5 h compared to that in the mutant with the antisense. It was argued that the 

antisense message increased the degradation of the normal rbcS. The estimated half-life of rbcS 

under O3 fumigation is about 1 h (Pell et al., 1994). Although comparisons of these diverse 

systems cannot be easily made, the normal half-life of rbcS may be closer to 5 to 10 h; and O3 

14The amount of Rubisco drops from an initial 0.12 to a final amount of 0.04 μmole/g-FW s in 6 days. 

AX9-78 
fumigation does not simply stop the transcription of DNA, but rather it alters the rate of 

degradation, either independently of, or simultaneously with, transcription. 

Williams et al. (1994) developed a correlation between the levels of ABA after water stress 

in Arabidosis thaliana leaves and the loss of rbcS. Although their data were not quantitative, the 

level of ABA had a half-time rise of about 1 to 2 h and the level of rbcS had a half-life decline of 

about 2 to 4 h. Their work suggests that drought stress may alter the CO2 metabolism by 

changing enzyme relationships much more than by merely closing the stomata. If an ABA rise 

is lowering rbcS, rbcS may not be a good marker of O3 fumigation except under highly 

controlled conditions. 

AX9.2.7 Carbohydrate Transformations and Translocation 

The question of whether translocation of the sugars out of the leaf is inhibited by O3 

exposure arises, because productivity is often dramatically inhibited by O3 fumigation. Though 

nearly 35 years have passed since Dugger and Ting (1970) investigated the question of sugar 

transport within the leaf, the question has since been little studied. Translocation (Cooley and 

Manning, 1987) appears to be inhibited, because root functions are impaired by O3 exposure. 

Many observed events suggest that while carbon assimilation within the leaf declines, 

translocation of carbon is inhibited even more so, because plant growth points are inhibited and 

root/shoot ratios are altered (Dugger and Ting, 1970; Gerant et al., 1996; Tjoelker et al., 1995). 

Many of the experiments with O3 fumigation indicate that O3 exposure decreases the net 

growth or dry mass of the plant, but the mechanism is poorly understood. Generally the 

decrease in assimilation is much less than the decrease in growth, but not always. Under many 

conditions, the stomata will close partially, decreasing assimilation by a smaller factor. Only a 

long exposure, or high levels of exposure for a short time, generate enough decline in Rubisco to 

make the growth of the plant problematical. No convincing argument has linked the decrease in 

growth with a small decline in assimilation, either by a conductance- or Rubisco-limitation. 

Measures of assimilation with crops are frequently done on upper canopy leaves, which are the 

last leaves to exhibit O3 injury, while leaves deeper in the canopy exhibit injury and early 

senescence. Crop root growth must be sensitive to these and other O3 effects, because root 

biomass is often suppressed early by elevated O3. 

AX9-79
Volin et al. (1998) found O3 exposures statistically decreased leaf area ratio, specific leaf 

area, leaf weight ratio, and root weight ratio in Populus tremuloides and two C3 grasses 

(Agropyron smithii and Koeleria cristata) but not in Quercus rubra and in the C4 grasses 

Bouteloua curtipendula and Schizachyrium scoparium. There was no statistically significant 

change in any species in leaf conductance (4% level decline in K. cristata) nor in assimilation 

(although there was a decline in assimilation at the 6% level for P. tremuloides and a decline at 

the 1% level in B. curtipendula). They also reported a correlation between growth decline and 

decreased stomatal conductance among all species. 

Tradeoffs are made by plants. Birch grown in highly fertilized conditions exhibited a 

greater leaf turnover when exposed to O3, in that leaves not only formed faster but abscised 

faster, presumably due to early senescence; whereas birch grown under poorer fertilized 

conditions retained their leaves longer and had a greater respiration rate within those leaves 

(Maurer and Matyssek, 1997). Again, one must be careful in comparing short-term versus longterm exposures. Grulke et al. (2001) observed that maximum concentrations of carbohydrates in 

1-year-old needles that had not abscissed due to early senescence declined when subjected to 

year-long exposures along an increasing pollution gradient. Furthermore, the monosaccharide 

concentrations (along with starch) in fine roots were largely decreased, suggesting that needle 

sugars were limiting, leading to root-sugar limitations. However, determination of the total 

productivity and detailed balance of carbohydrate was impossible, because these were older, 

larger trees and the data were taken over a full growing season. For a shorter-term exposure of 

9 days, Smeulders et al. (1995) observed that O3 appeared to increase the retention of labeled 

photosynthates within the needles of Douglas Fir, and, at higher exposures (400 versus 200 or 

0 μg/m3

), the total starch within the needles decreased, suggesting that less carbohydrate was 

produced within the cell or perhaps that it was in compounds not measured. 

Studies with Pima cotton (Gossypium barbadense), aspen (Populus spp.) and bean 

seedlings (Phaseolus vulgaris) indicate that acute O3 exposures inhibit export of the current 

assimilate that provides carbohydrates to the roots from source leaves of cotton as well as recent 

assimilate from the older leaves of aspen and bean (Grantz and Yang, 2000). Grantz and Yang 

(2000) attempted to distinguish between potential mechanisms of O3 phytotoxicity operating at 

the level of the whole plant. Four hypotheses were tested by fumigating cotton: (1) O3 exposure 

reduces leaf pools of soluble sugars; (2) pruning leaf area and reducing source strength to match 

AX9-80
that of O3-treated plants reproduces O3 effects; (3) pruning lower leaf area more closely 

reproduces O3 effects than pruning the upper leaf area; and (4) manipulating plant age and, 

thereby, plant size to match O3-treated plants reproduces O3 effects. All were shown to be 

incorrect. Under each of the above conditions, Grantz and Yang (2000) reduced the amount of 

foliage to match that caused by O3 injury. While the treatments reduced the biomass and leaf 

area, they did not alter biomass allocation nor root function. They concluded that a simple loss 

of foliage does not induce the changes in translocation to the roots to the same extent as does O3 

injury. 

This finding by Grantz and Yang (2000) is important in that it suggests that O3 can trigger 

a plant-wide response that may be linked to alterations in signal transduction and the generation 

of whole plant signals. Stitt (1996) suggested that “...allocation is regulated by long-distance 

signals that act to influence growth of selected sinks and to modify the delivery of resources to 

these sinks in parallel.” Cooley and Manning (1987), citing McLaughlin and McConathy 

(1983), suggested three possible ways that O3 fumigation might alter translocation: 

(1) malfunction of the phloem loading process, (2) increased translocation to leaf injury repair, 

and (3) an altered balance between the leaf and sinks caused by reduced carbon fixation and a 

greater demand for assimilate in the leaf. 

Ethylene has been shown to reverse this sugar inhibition of development and to be 

antagonistic to the ABA effect (Finkelstein and Gibson, 2002). However, these effects depend 

greatly upon the developmental stage of the plant. Thus, the balance of the effectors (sugars, 

ABA, and ET) may interact to generate the variation observed in the O3-induced productivity 

decline. For example, O3 fumigation can induce a shift in the carbon transfer between roots and 

shoot, and this shift can be amplified by mild drought (Gerant et al., 1996). Furthermore a 

regulation of source-sink relations with the defense responses induced by elicitors was observed 

by wounding the leaves of Chenopodium rubrum (Roitsch, 1999). Ethylene appears to be able to 

repress the expression of extracellular invertase, which is critical for control and downloading of 

sucrose derived from the translocational stream (Roitsch, 1999). In addition, the development of 

Arabidopsis at high concentrations of glucose or sucrose is arrested by increasing the ABA level 

(Coruzzi and Zhou, 2001). 

AX9-81
Clearly more work is needed on the interactions between assimilation, translocation, and 

source/sink relations with O3 exposure. In these interactions, one must be aware of the 

developmental age of the plants and their phytohormonal status. 

AX9.2.7.1 Lipid Synthesis 

Heath (1984) summarized several early reports of O3-exposure induced lipid alterations. 

Most concerned the production of MDA (malonyldialdehyde) as a measure of lipid oxidation as 

well as the loss of unsaturated fatty acids. However, a series of experiments by Sakaki and 

coworkers (Sakaki et al., 1983, 1985) concentrated on one type of fumigation system and one 

metabolic pathway. This literature provides the best, most complete story with regard to lipid 

metabolism and O3 fumigation and suggests that O3 injures cellular membrane systems via lipid 

destruction. 

Sakaki and coworkers (Sakaki et al., 1983, 1985) used spinach, which is a sensitive plant 

but which has not been much evaluated with respect to O3 fumigation. While the O3 level was 

high (0.5 ppm), enough work has been done to be able to discern what is happening. The first 

paper showed that chlorophyll bleaching does not begin until the plants have been exposed to O3 

for more than 10 h, whereas some MDA production begins with as little as 6 h exposure (Sakaki 

et al., 1983). Consistent production of MDA, indicative of gross disruptions, occurred only after 

8 h exposure (Sakaki et al., 1985), within the timescale when chlorophyll and carotenoid levels 

begin to decline. Concurrently, the total fatty acid (FA) level decreased from ~481 to 358 

nmol/cm2 as the MDA level increased from 0.6 to 2.4 nmol/cm2

, indicating FA peroxidation 

(Sakaki et al., 1985). 

Sakaki et al. (1983) also studied the development of changes by cutting disks from exposed 

leaves and floating them on water solutions for varied time periods (up to 24 h). This permitted 

feeding experiments to be done easily, whereas the cutting gives rise to an additional wound 

response and eliminates metabolite movement to and from other portions of the plant. The 

floating experiments indicated that, after exposure, scavengers of singlet oxygen (1

O2), such 

as D2O, and of hydroxyl radicals, such as benzoate and formate, have no effect on development 

of the MDA response after 8 h of in vivo fumigation, while scavengers of (O2 

!

), such as tiron 

and ascorbate, lowered the amount of MDA formed. By measuring metabolites immediately 

after cessation of fumigation, they were able to show that ascorbate loss began with the onset of 

AX9-82
fumigation, as did SOD loss. A lag time associated with the production of DHA suggested that 

the reaction of ascorbate with fumigation did not immediately produce the oxidation product. 

The first 4 h of exposure yielded 30 nmole/cm2

 of ascorbate loss with 5 nmole/cm2 of DHA 

production, whereas the second 4 h of exposure yielded 20 nmole/cm2

 of ascorbate loss with 20 

nmole/cm2

 of DHA production. 

Nouchi and Toyama (1988) exposed Japanese morning glory (Ipomea nil) and kidney 

bean (Phaseolus vulgaris) to 0.15 ppm O3 for 8 h. Under these conditions, little visible injury 

was found with up to 4 h of exposure, while injury increased by ~50% after 8 h of exposure. 

Morning glory produced more MDA than kidney bean, which produced the same as the 

zero-time control. Morning glory also demonstrated a slight (5%) drop in MDGD 

(monogalactosyldiacylcerol), with increases in PC (phosphatidylcholine), PG 

(phosohatidylglycerol), PI (phosphatidylinositol), and PE (phosphatidylethanolamine) after 4 h. 

Twenty-four hours later, the drop in MGDG (mongalactosyldiacyglyerol) was much larger and 

was thought to be related to an inhibition of UDP-galactose galactotransferase due to a rise in 

free fatty acids (FFAs) in the chloroplast. Note that the two distinct timescales involved in O3 

fumigation, immediately postfumigation and a day or so later, allows for comparison after the 

plant metabolism responds to the fumigation event. 

The pathway for the formation of MGDG and DGDG (digalactosyldiacylglycerol) is 

located on the chloroplast envelope. Diacylglycerol (DG) arrives from either the endomembrane 

system or the stroma and the enzyme UDP galactose-1,2-diacylglycerol galactosyltransferase 

(UDGT) forms MGDG with galactose from UDP-galactose. Sakaki et al. (1990) suggested that 

the O3-induced inhibition of UDGT was due to a release of FFAs from within the chloroplast. 

These FFAs are inhibitory to UDGT, but not to GGGT, which is stimulated by high 

concentrations of Mg2+ (Sakaki et al., 1990). The Sakaki et al. (1990) data indicate that the in 

vivo measured activities of both enzymes isolated after fumigation are not affected by O3 

fumigation. Both enzymes have sensitivity sulfhydryls, and both are located on the envelope. 

Ozone, if it reaches those sulfhydryls, should inhibit these enzymes; yet inhibition was not seen. 

It has been thought for years that tocopherols functioned as antioxidants in biological 

systems (Tappel, 1972). Hausladen et al. (1990) examined the role of antioxidants in red spruce 

(Picea rubens) by following seasonal changes. They fit the level of tocopherol within the 

needles (/g-FW) to the time of the year and found little change (fit as level = A + Bt + Ct2

). 

AX9-83
From this empirical fit, they found that the constant A was lower with higher levels of O3. The 

seasonal variation coefficients, B and C, were also lower, suggesting year-long low tocopherol 

levels. Variation with the season is not particularly surprising, given that phytochrome action 

may be linked to tocopherol biosynthesis (Lichtenthaler, 1977). Hausladen et al. (1990) reported 

a significant (p < 0.05) trend in the difference between the high and low level of treatment; 

although there was no discussion of why it occurred or what it meant in relation to metabolism. 

Their major conclusion was that the antioxidant changes due to O3 exposure may decrease cold 

hardiness. 

Sterols, believed to act as membrane stabilizers, have been investigated by several groups 

with mixed results. Tomlinson and Rich (1971), who exposed common bean at 0.25 ppm for 

3 h, and Grunwald and Endress (1985), who exposed soybean at 0.07 ppm for 6 h for 48 days, 

reported an increase in free sterols and a decline in esterified sterols. However, Trevathen et al. 

(1979) exposed tobacco at 0.3 ppm for 6 h and reported opposite results. None of these 

investigators believed that O3 had attacked the sterols directly, instead, they believed that these 

changes involved metabolism and membrane stability. If O3 induced a metabolic shift that 

disturbed the polar lipid to sterol balance, membrane reactions to other stressors, such as cold 

tolerance, would certainly also be affected, perhaps detrimentally. 

AX9.2.8 Role of Age and Size Influencing Response to Ozone 

Clearly many changes occur with O3 exposure can be observed within hours, or perhaps 

days, of the exposure. This document has argued that many of those events are connected with 

wounding and elicitor-induced changes in gene expression, but those are relatively fast acting 

changes (a timescale of tens of hours). Two other effects due to O3 take longer to occur and tend 

to become most obvious under long periods of low-O3 concentrations. These have been linked 

to senescence or some other physiological response very closely linked to senescence. These 

two responses, separated by a time sequence, are shown diagrammatically in Figure AX9-15. 

The understanding of how O3 affects long-term growth and resistance to other biotic and 

abiotic insults in long-lived trees is unclear. Often, the conditions to which a tree is subjected to 

in one year will affect the response of that tree in the next year. This has been called “memory 

effect”, although the term “carry-over” is preferred. In other words, a condition in an earlier 

year sets the stage for a reaction in the next year; thereby giving a “cause-effect” scenario. 

AX9-84
Figure AX9-15. Linkage of senescence with hypersensitivity reactions and the first event 

of O3 attack of plants. 

In perennial plant species, growth affected by a reduction in storage carbohydrates may 

result in the limitation of growth the following year (carry-over effects) (Andersen et al., 1997). 

Carry-over effects have been documented in the growth of tree seedlings (Hogsett et al., 1989; 

Sasek et al., 1991; Temple et al., 1993; U.S. Environmental Protection Agency, 1996) and in 

roots (Andersen et al., 1991; U.S. Environmental Protection Agency, 1996). Accumulation of 

the carry-over effects over time will affect survival and reproduction. Data on the cumulative 

effects of multiple years of O3 exposures have been, for the most part, the result of 2- to 3-year 

seedling studies. The difficulty of experimentally exposing large trees to O3 has lead to the tacit 

assumption that seedling response to O3 is a good predictor of large-tree response to O3 (U.S. 

Environmental Protection Agency, 1996). 

The carry-over effects of O3 exposures as observed in tree seedlings cited above by Hogsett 

et al. (1989) have been termed “memory effects” by Langebartels et al. (1997) and proposed by 

Schmieden and Wild (1995) to explain the sensitivity of spruce seedlings to frost in the winter 

after having been exposed to O3 during the previous summer. Norway spruce (Picea abies L.) 

exposed to 80 ppb O3 for a whole growing season, demonstrated visible injury symptoms the 

following year when the new needle flush appeared (Langebartels et al., 1997). Additional 

studies using Norway spruce and Scots pine (Pinus sylvestris L.) seedlings have shown similarly 

delayed responses following O3 exposures. Carry-over symptoms were noted to develop at 

AX9-85
different times of the year, depending on the species of seedling exposed: in early spring for 

Norway spruce, and in early autumn for Scots pine (Lange et al., 1989). Visible effects of O3 

exposures on spruce and pine may develop after a substantial delay during the “sensitive” 

periods of the year when chlorophyll and needle loss normally occur. Norway spruce and Scots 

pine differ in their sensitive periods because of the different needle classes normally remaining 

on the tree (Langebartels et al., 1997). 

Nutrient status of the tree during the overwintering phase of its life (Schmieden and Wild, 

1995) and chronic exposure to ambient O3 (less severe with fewer peaks of very high levels) 

induce (1) mineral nutrient deficiency; (2) alterations of normal metabolism, including allocation 

of carbohydrates and probably nitrogen; and (3) disturbance of normal transpiration and diurnal 

cycling, leading to water stress. This condition, termed “Montane yellowing”, appears to be 

related to nutrient deficiencies rather than senescence (although early loss of leaves and needles 

occurs). While generalized low nutrient concentrations may not occur within the foliage, 

localized deficiencies may. However, they are hard to observe or prove without a great deal of 

work involving all portions of a tree and without a general hypothesis of what is occurring. 

AX9.2.9 Summary 

As the understanding of wounding responses of plants and more genome details and varied 

plant mutants become available, the cellular and physiological responses of plants to O3 

exposures are slowly becoming clearer. However, more studies are needed on a larger variety of 

species. Nevertheless, several key findings and conclusions can be highlighted: 

(1) The entrance of O3 into the leaf through the stomata remains the critical step in O3 

sensitivity. Not only does O3 modify the opening of the stomata, usually closing it 

partially, but O3 also appears to alter the response of stomata to other stressful 

situations, including a lowering of water potential and ABA responses. The 

concentration of O3 within the leaf is not the same as the external concentration due 

to reactions within the leaf, but it is not “zero”. 

(2) The initial reactions of O3 within the leaf are still unclear, but the involvement of 

H2O2 is clearly indicated. The detection of possible products by EPR spectroscopy 

has progressed, but has not reached the point where any products can be identified. 

Nonetheless, reaction of O3 (or its product) with ascorbate and possibly other 

antioxidants present in the apoplastic space of the mesophyll cells is clear and 

serves to lower the amount of O3 or product available to alter the plasma membrane 

of the cells. 

AX9-86
(3) The initial sites of membrane reactions seem to involve transport properties and, 

possibly, the external signal transducer molecules. The alteration and mechanism of 

the alteration of the varied carriers of K+ and Ca2+ is far from clear, but it would 

seem that one of the primary triggers of O3-induced cell responses is a change in 

internal Ca2+ levels. 

(4) The primary set of metabolic reactions that O3 triggers now clearly includes those 

typical of “wounding” responses generated by cutting of the leaf or by 

pathogen/insect attack. Again, this seems to be due to a rise in cytoplasmic Ca2+ 

levels. Ethylene release and alteration of peroxidases and PAL activities, as well as 

activation of many wound-derived genes, seem to be linked to some of the primary 

reactions. 

(5) The alteration of normal metabolism due to wounding has effects outside of the 

cytoplasm. What effects are due to the “spreading of the problem” to other cellular 

organelles is less clear. One of the secondary reactions is linked to an activation of 

a senescence response. The loss of Rubisco and its messenger RNA is linked to an 

early senescence or a speeding up of normal development leading to senescence. 

The loss of photosynthetic capacity is linked to the lowered productivity of plants, 

and problems with efficient translocation are indicated, although photosynthesis and 

translocation still occur at a reasonable rate. The loss of productivity is not yet 

clearly explained. 

It is important to note that the dramatic strides in understanding the genetic makeup of 

plants, gene control, and signal transduction/control over the last few years will likely accelerate 

in the future. That understanding will translate into better models of the hypotheses listed above 

and into more detailed schemes of how O3 alters much of basic plant metabolism. Thus, while 

understanding of how O3 interacts with the plant at a cellular level has dramatically improved, 

the translation of those mechanisms into how O3 is involved with altered cell metabolism, with 

whole plant productivity, and with other physiological facts remain to be more fully elucidated. 

AX9.3 MODIFICATION OF FUNCTIONAL AND GROWTH RESPONSES 

AX9.3.1 Introduction 

The responses of plants to any air pollutant may be significantly influenced by a wide 

range of biological, chemical, and physical factors. A plant’s genetic makeup is an important 

inherent biological determinant of its response, but response can also be modified by other 

biological agents such as disease-causing organisms, insects and other pests, and by other higher 

AX9-87
plant species with which it may be competing for resources. Chemical factors that may 

influence response to an air pollutant range from mineral nutrients obtained from the soil to other 

air pollutants and agricultural chemicals. Physical factors that may influence response include 

light, temperature and the availability of moisture, which are components of climate and climate 

change. 

Some environmental factors can be controlled, to some degree, by man, while others 

cannot. The biological factors (e.g., pests, diseases, symbioses, competition) are partly 

controllable in agriculture but much less so (if at all) in natural ecosystems. It is possible to 

control agricultural soil fertility and the use of agricultural chemicals, as well as to exercise some 

control over the supply of water and airborne chemical factors. In contrast, the physical factors 

(e.g., light and temperature) are uncontrolled in the field even though they may be controllable in 

specialized situations such as greenhouses or shade houses. Although light and temperature are 

components of climate, they are initially reviewed as individual physical factors, even though 

temperature effects are revisited to some extent in the discussion of interactions with climate 

change. 

The impacts of these various factors on plant response to O3 and other oxidants were 

extensively reviewed in the 1996 O3 AQCD (U.S. Environmental Protection Agency, 1996). 

It was noted in that document that, since any combination of these factors may come into play at 

some time in a plant’s life history, “response will be dictated by the plant’s present and past 

environmental milieu, which also includes the temporal pattern of exposure and the plant’s stage 

of development.” That document also stressed that both the impact of environmental factors on 

response to oxidants and the corollary effects of oxidants on responses to environmental factors 

have to be considered in determining the impact of oxidants on vegetation in the field. The 

variability observed in plant responses to defined exposures to O3, particularly under field 

conditions, is a consequence of the influences of genetics and the range of environmental 

variables. 

In view of the large number of factors to be considered, this section focuses mainly on 

situations in which there is clear evidence that environmental factors truly interact with oxidant 

effects, i.e., they magnify or diminish the impact of O3 and are not merely additive to it. 

Conversely, it will cover situations where O3 acts synergistically or antagonistically, but not 

additively, with effects induced by other factors. It will also emphasize those interactions as a 

AX9-88
result of which overall plant growth, development, and yield are adversely affected, rather than 

the details of interactions at the mechanistic level, unless the latter are deemed to be essential to 

an understanding of larger-scale effects. 

Few studies reported since the 1996 document have systematically investigated 

quantitative responses to O3 exposure concurrent with other variables. Although the 1996 

document cited almost 300 references pertaining to environmental interactions, and the present 

review cites more than 350 new references, the bulk of the recently published work has 

continued to be specific and frequently narrowly focused. Hence, the new findings are scattered 

and far from uniformly distributed among the various subtopics. In some instances, little or no 

new research has been published that adds to our understanding since the 1996 document. In 

such cases, the present review is, therefore, restricted to summarizing the understanding that was 

current in 1996. 

A few reviews have appeared since the early 1990s dealing with various environmental 

interactions, and these are cited in relevant sections below. More general recent reviews are 

those of Wellburn (1994); multi-authored volumes edited by Alscher and Wellburn, 1994; Yunus 

and Iqbal, 1996; De Kok and Stulen, 1998; and Bell and Treshow, 2002); and reports by the 

United Nations Environment Programme (UNEP, 1999) and the Intergovernmental Panel on 

Climate Change (IPCC, 2001). Several biotic and abiotic interactions involving forest trees are 

discussed in the review by Johnson et al. (1996b). 

Although many reports have provided quantitative information on interactive effects, 

in most cases the information simply describes a specific situation involving only two or three 

levels of a variable. While this may be adequate to provide statistical information about the 

existence of interactions with environmental factors, it does not permit the development of 

response surfaces or models to show the form that any influence of such factors might take 

on O3 exposure-response relationships or how O3 might quantitatively influence responses to the 

factors in question. This, together with the fragmented information available on the effects of 

most factors, has contributed to the relative lack of development of simulation models of 

oxidant-environmental factor interactions. Yet, as noted by Taylor et al. (1994), the large 

number of variables constrains the assessment of pollution effects by experimentation alone. 

The only alternative is to use mathematical models to attempt to predict the outcome of 

different O3 and environmental factor scenarios, building up their complexity in stages. 

AX9-89
The few models thus far used to investigate O3 stress have been adapted from existing process 

models of crop or tree growth that include limited numbers of physical or chemical variables 

such as temperature, soil water stress, or nutrient deficiency. Taylor et al. (1994) provided a 

listing of several simulation models developed for trees at the individual, stand, and regional 

levels; these and many other models have been critically reviewed by Kickert and Krupa (1991) 

and Kickert et al. (1999). However, regardless of whether such models are descriptive/empirical 

or process/mechanistic, their outputs will always be associated with varying degrees of 

uncertainty and require validation against observable responses wherever possible. Kickert et al. 

(1999) also noted that very few of the models that have been described provide risk assessments 

that address likelihood, in contrast to consequence assessments that address the magnitudes of 

effects. Thus, even though capable simulation models of plant response to O3 involving complex 

mixes of many biological, physical, and chemical factors may be out of reach at the present time, 

the use of newer mathematical approaches such as artificial neural networks (ANNs) has 

enabled insightful analyses to be performed in several field studies involving numerous 

micrometeorological and other environmental variables (e.g., Balls et al., 1996; Mills et al., 

2000). 

Because the ensuing subsections deal with studies of O3 interactions involving an 

extremely wide array of biological, physical, and chemical factors in the plant’s environment, it 

is inevitable that many different exposure facilities and regimes have been used in these studies. 

To provide specific information regarding the O3 exposure concentrations, profiles, hours and 

days of exposure (as well as the types of systems and facilities used for the exposures) would 

add a wealth of detail that would do little to assist our understanding of the roles of 

environmental factors in modifying the impact of O3 on vegetation or to facilitate our ability to 

estimate the magnitudes of any such modifications. Thus, only experiments in which the 

exposure levels and regimes were within the bounds of ambient experience in North America are 

discussed in the ensuing subsections, regardless of the type of exposure profile used. The cutoffs 

used have been ~200 ppb for peak hourly concentrations or for short-term exposures, ~100 ppb 

for daytime means involving prolonged exposures for several hours, or a doubling of ambient 

levels in cases in which enriched exposure levels were a function of ambient levels. Actual 

details of the exposure regimes and conditions can, of course, be obtained from the original 

references but are only stated here when any distinction is required between the effects of 

AX9-90
different exposure levels. Hence, it should be understood that ensuing statements such as “. . .it 

was found that O3 caused. . .” should always be read as “. . .it was found that exposures to O3 

[within the range of those that have been measured in ambient air] caused.” 

AX9.3.2 Genetics 

The response of individual plants to O3 is affected by several factors, including the 

environment in which it is growing, competition with neighboring plants, ontogeny, and 

genetics. This section examines the role of genetics in plant response to O3. In addition, major 

knowledge gaps in the understanding of genetic aspects of O3 response are pointed out. 

It is well known that species vary greatly in their responsiveness to O3 (U.S. Environmental 

Protection Agency, 1996). This again has been recently demonstrated for grassland species 

(Bungener et al., 1999b; Bungener et al., 1999a; Franzaring et al., 2002; Nussbaum et al., 2000a; 

Pleijel and Danielsson, 1997; Warwick and Taylor, 1995), wild herbaceous plants (Bergmann 

et al., 1999; Danielsson et al., 1999), agricultural crops (Benton et al., 2000; Elagöz and 

Manning, 2002; Fumagalli et al., 2001; Heagle and Stefanski, 2000; Köllner and Krause, 2003; 

Nali et al., 2002; Nussbaum et al., 2000b; Ollerenshaw et al., 1999; Postiglione et al., 2000; 

Renaud et al., 1997), horticultural shrubs and trees (Findley et al., 1997; Hormaza et al., 1996), 

and forest trees (Bortier et al., 2000a; Guidi et al., 2001; Landolt et al., 2000; Matsumura, 2001; 

Momen et al., 2002; Nali et al., 2002; Oksanen and Rousi, 2001; Pääkkönen et al., 1997; Pell 

et al., 1999; Saitanis and Karandinos, 2001; Volin et al., 1998; Zhang et al., 2001). These 

studies have shown a wide range of responses to O3, from growth stimulation by a few species 

such as Festuca ovina L. (Pleijel and Danielsson, 1997) and Silene dioica and Chrysanthemum 

leucanthemum (Bungener et al., 1999b; Bungener et al., 1999a) to significant growth reduction, 

depending on environmental conditions and exposure dose. 

While determining the explanation for differences in species sensitivity to O3 remains one 

of the challenges facing plant biologists (Pell et al., 1999), a number of hypotheses have been 

suggested. Reich (1987) proposed that variation in O3 sensitivity could be explained by variation 

in total uptake of the gas. Others have suggested that (1) fast-growing species are more sensitive 

than slower-growing ones (Bortier et al., 2000b), (2) overall O3 sensitivity may be closely linked 

to root responses to O3 (Warwick and Taylor, 1995), or (3) the relative ability of species to 

detoxify O3-generated reactive oxygen free radicals may determine O3 sensitivity (Alscher et al., 

AX9-91
1997; Pell et al., 1999). Volin et al. (1998) suggested that the relative rate of stomatal 

conductance and the photosynthesis rate at a given conductance both contribute strongly to 

determining a species sensitivity to O3. Likely, there is more than one mechanism determining 

sensitivity, even in a single species. 

Within a given species, individual genotypes or populations can also vary significantly 

in O3 sensitivity (U.S. Environmental Protection Agency, 1996). For example, the intraspecific 

variation in O3 sensitivity was a factor of two for Phleum pratense (Danielsson et al., 1999) and 

Trifolium repens L. (Postiglione et al., 2000). A similar range of intraspecific variations in O3 

responses was demonstrated for clonal differences in Betula pendula by Pääkkönen et al. (1997) 

and Prunus serotina (Lee et al., 2002). These examples of wide ranges within species responses 

suffice to show that caution should be taken when ranking species categorically as having an 

absolute degree of O3 sensitivity (Davison and Barnes, 1998). 

AX9.3.2.1 Genetic Basis of Ozone Sensitivity 

Plant response to O3 is determined by genes that are directly related to oxidant stress and to 

an unknown number of genes that are not specifically related to oxidants. The latter includes 

genes that control leaf and cell wall thickness, stomatal conductance, and the internal 

architecture of the air spaces. Although there is currently a great emphasis on individual 

antioxidants that can be manipulated by molecular methods, the challenge is to determine the 

relative contributions of all of the components to plant response and to understand the interplay 

between them. Recent studies using molecular biological tools are beginning to increase the 

understanding of O3 toxicity and differences in O3 sensitivity. 

While much of the research in developing the understanding of O3 responses has been 

correlative in nature, recent studies with transgenic plants have begun to positively verify the 

role of various genes and gene products in O3 tolerance. The finding that the overexpression of 

MnSOD (manganese superoxide dismutase) in transgenic tobacco plant chloroplasts increased 

O3 tolerance (Van Camp et al., 1994) provided the first definitive proof of antioxidants key role 

in O3 tolerance. Subsequently, Broadbent et al. (1995) showed that the simultaneous 

overexpression of pea glutathione reductase in both chloroplasts and mitochondria enhanced O3 

tolerance in transgenic tobacco. Similarly, increased O3 tolerance to O3-induced foliar necrosis 

was shown for transgenic tobacco plants overexpressing the cytosolic Cu/Zn-SOD gene (Pitcher 

AX9-92
and Zilinskas, 1996). Transgenic tobacco plants expressing antisense RNA for cytosolic 

ascorbate peroxidase, which reduces ascorbate peroxidase production, showed increased 

susceptibility to O3 injury, suggesting a key role in O3 tolerance for the antioxidant ascorbate 

peroxidase (Örvar and Ellis, 1997). 

The consensus among molecular studies of O3 sensitivity is pointing to O3 as triggering 

salicylic acid, ethylene, and jasmonic acid production and that the signaling of these molecules 

determines, in some cases, the O3 susceptibility of plants (DeCaria et al., 2000; Moeder et al., 

2002; Nunn et al., 2002; Overmyer et al., 2000; Rao and Davis, 1999; Tamaoki et al., 2003; 

Vahala et al., 2003a,b). Increased levels of jasmonic acid production in O3-tolerant compared to 

O3-sensitive plants has been shown for Arabidopsis (Overmyer et al., 2000) and Populus (Koch 

et al., 1998, 2000). Blockage of ethylene production by using antisense methods with 

1-aminocyclopropane-1-carboxylate (ACC) synthase and ACC oxidase suggest strongly that 

ethylene synthesis and perception are required for H2O2 production and cell death following O3 

exposure of Lycopersicon esculentum (Moeder et al., 2002). Ethylene signaling may have 

multiple roles in O3 tolerance determination as was demonstrated recently by Vahala et al. 

(2003a,b) who found that, in Populus tremula × P. tremuloides hybrid clones differing in O3 

sensitivity, ethylene accelerated leaf senescence in sensitive plants under low O3, but under acute 

O3, ethylene seemed to be required for protection from cell death. 

While changing the expression of single antioxidant genes has proven very useful in 

identifying possible mechanisms of O3 sensitivity and tolerance (Kuzniak, 2002), it should be 

noted that increased O3 tolerance has not been shown in some studies of transgenic plants with 

enhanced antioxidant production (Strohm et al., 1999; Strohm et al., 2002; Torsethaugen et al., 

1997). Clearly, ethylene production plays a role in O3 sensitivity, but the roles of various 

antioxidants in O3 tolerance regulation are yet to be fully elucidated (Wellburn and Wellburn, 

1996). It is unlikely that single genes are responsible for O3 tolerance responses, except in rare 

exceptions (Engle and Gabelman, 1966). Regulation of stomatal opening and leaf structure 

(Elagöz and Manning, 2002; Fujita et al., 1992) are likely to play key roles in O3 tolerance in 

plants. Newly developing opportunities to examine simultaneous regulation of larger numbers 

of genes are also likely to yield more clarification of the genes controlling O3 tolerance (Desikan 

et al., 2001; Matsuyama et al., 2002). 

AX9-93
Attempts to demonstrate conclusive changes in antioxidant and protective pigments for O3 

sensitive and tolerant mature trees growing in the field have largely been unsuccessful (Tausz 

et al., 1999a,b). However, evidence for antioxidant expression differences contributing to 

differences in O3 sensitivity of 4-year-old Populus tremuloides trees has been found (Wustman 

et al., 2001). 

AX9.3.3 Environmental Biological Factors 

The biological factors within the plant’s environment that may directly or indirectly 

influence its response to O3 in a positive or negative manner encompass insects and other animal 

pests, diseases, weeds, and other competing plant species. Although such interactions are 

ecological in nature, those involving individual pests, plant pathogens, or weeds, or agricultural 

crop or forest tree species are considered in this section. More complex ecological interspecies 

interactions are dealt with in Section AX9.5. 

The different types of biological factors are dealt with separately, as in the 1996 O3 AQCD 

(U.S. Environmental Protection Agency, 1996). Still, it is important to recognize certain general 

features of relationships of plants with the biological components in their environments: 

C Successful infestation or infection involves complex interactions among the target or host 

species, the causal organism and environmental factors. 

C Infestations and infections may co-occur. 

C The successful development and spread of a pest, pathogen, or weed require favorable

environmental factors.

C Significant losses to crops and forest trees result from pests and pathogens. 

C Significant losses to crops and seedling trees result from weed competition. 

Ozone and other photochemical oxidants may influence the severity of a disease or 

infestation by a pest or weed, either by direct effects on the causal species, or indirectly by 

affecting the host, or both. In addition, the interaction between O3, a plant, and a pest, pathogen, 

or weed may influence the response of the target host species to O3. A perceptive overview of 

the possible interactions of O3-exposure with insect pests and fungal diseases has been provided 

by Jones et al. (1994), based on a model system involving two insects and two pathogens 

affecting cottonwood (Populus deltoides). Their study also included effects on the 

decomposition of leaf litter. 

AX9-94
In contrast to detrimental biological interactions, there are mutually beneficial relationships 

or symbioses involving higher plants and bacteria or fungi. These include (1) the nitrogen-fixing 

species Rhizobium and Frankia that nodulate the roots of legumes and alder and (2) the 

mycorrhizae that infect the roots of many crop and tree species, all of which may be affected by 

exposure of the host plants to O3. 

In addition to the interactions involving animal pests, O3 may also have indirect effects on 

higher herbivorous animals, e.g., livestock, due to O3-induced changes in feed quality. 

AX9.3.3.1 Oxidant-Plant-Insect Interactions 

The 1996 O3 AQCD (U.S. Environmental Protection Agency, 1996) stressed the variability 

in the reported effects of O3 on host plant-insect interactions. Since relatively few plant-insect 

systems have been studied, few consistent patterns of response have emerged, as noted in other 

reviews such as those of Colls and Unsworth (1992), Heliövaara and Väisänen (1993), Whittaker 

(1994), Docherty et al. (1997), and, most recently, Flückiger et al. (2002). 

None of the studies reported in the past decade have clarified the situation in terms of 

clearly consistent effects. A 1997 review by Docherty et al. (1997), for example, examined 

17 reports of studies of aphid species on a range of hosts and classed the O3 effects on aphid 

performance as 35% positive, 41% negative, and 24% showing no significant effect. 

A tabulation of 19 studies by Flückiger et al. (2002) gave the corresponding figures: 42%, 21%, 

and 37%. 

Other recent studies with the aphids Schizolachnus pineti and Cinara pinea on Scots pine 

(Pinus sylvestris) and Cinara pilicornis on Norway spruce (Picea abies) have also yielded 

variable results, but suggested that O3 enhances aphid density on pine and aphid performance on 

spruce (Holopainen et al., 1997; Kainulainen et al., 2000a). In an earlier study with 

Schizolachnus pineti on Scots pine, Kainulainen et al. (1994) had observed no significant effects 

of O3-treatment on aphid performance. However, more recent observations of long-term effects 

on aphid populations on aspen (Populus tremuloides) exposed to O3 in a FACE system revealed 

that O3 significantly increased aphid populations while decreasing the populations of predatory 

insects (Percy et al., 2002). 

AX9-95
The observations of Brown et al. (1993) and Jackson (1995) led Whittaker (1994) and 

Brown (1995) to suggest that aphid response was dependent on ambient temperature as well as 

the dynamics of O3 exposure and that growth tended to be stimulated with maximum 

temperatures below ~20 °C but was reduced at higher temperatures. The present situation with 

plant-aphid responses, therefore, remains confused and, although numerous suggestions have 

been offered to explain specific findings, they are difficult to assemble into a coherent picture. 

Variability has also been found with the interactions involving chewing insects. For 

example, Lindroth et al. (1993) reported a small negative O3 effect (8% reduction) on the growth 

of gypsy moth larvae (Limantra dispar) on hybrid poplar (Populus tristis × P. balsamifera) but 

no effect when growing on sugar maple (Acer saccharum). Ozone exposure reduced the growth 

rate of the larvae of the bug Lygus rugulipennis on Scots pine, but enhanced the growth of larvae 

of the sawfly Gilpinia pallida (Manninen et al., 2000). Costa et al. (2001) observed no 

significant O3 effects on the growth and fecundity of the Colorado potato beetle (Leptinotarsa 

decemlincata) on potato (Solanum tuberosum L.) in greenhouse and field experiments. 

Fortin et al. (1997), in a 2-year study of the forest tent caterpillar (Malacosoma disstria) on 

sugar maple, observed that O3 exposure increased the growth rate of female larvae in only one 

year; fourth- and fifth-instar larvae also showed a feeding preference for treated foliage in that 

year. However, studies based on open-air exposures of aspen indicated O3-enhanced growth of 

M. disstria in terms of pupal weight (Percy et al., 2002) and larval performance (Kopper and 

Lindroth, 2003b). Jackson et al. (2000) observed inconsistency in studies on the larva of the 

tobacco hornworm (Manduca sexta) on tobacco (Nicotiana tabacum). In one year, feeding 

on O3-treated foliage resulted in significantly greater larval weight, whereas the increase was not 

statistically significant in a second year although survival was increased. Also, oviposition by 

hornworm moths was increased if ambient O3 levels were increased by 70% and returned to 

normal in ambient O3 levels (Jackson et al., 1999). 

Studies of the two-spotted spider mite (Tetranychus urticae) on white clover (Trifolium 

repens) and peanut (Arachis hypogeae) by Heagle et al. (1994a) and Hummel et al. (1998) 

showed that O3-exposure stimulated mite populations on an O3-sensitive clover clone and on 

peanut. The lack of significant effects on mites on the O3-resistant clover clone suggests that the 

responses were host-mediated. 

AX9-96
There, therefore, appears to be a clearer indication of the likelihood that increased chewing 

insect and mite performance will result from O3-induced changes in the host plant. However, 

negative effects continue to be reported, indicating that the response is probably also being 

determined, in part, by other environmental, genetic, or temporal variables. 

Reported O3-induced enhancement of attack by bark beetles (Dendroctonus brevicomis) on 

Ponderosa pine (Pinus ponderosa) has been suggested by Dahlsten et al. (1997) to be due to 

greater brood development on injured trees, possibly related to decreased numbers of predators 

and parasitoids. This view gains some support from the observation that O3 exposure adversely 

affected the searching behavior of the parasitoid Asobara tabida for larvae of Drosophila 

subobscura which led to fewer parasitized fly larvae (Gate et al., 1995). Such observations 

reveal another level of complexity in the O3-plant-insect interrelationship: O3 may reduce the 

effectiveness of the natural control of insect pests. The phenomenon is probably related to 

effects on olfactory cues, as it was shown by Arndt (1995) that O3 can affect fly behavior by 

modifying the pheromones involved in fly aggregation. 

These reports focus on the direct or indirect effects on the insect or mite feeding on foliage 

previously or currently exposed to O3. They provide little, if any, information on the host plant 

effects other than qualitative references to the injury caused by the O3 exposure. Enhanced pest 

development will ultimately lead to increased adverse effects on the hosts in the long term, but 

the only report of an O3-plant-insect interaction directly affecting the host plant in the short term 

still appears to be that of Rosen and Runeckles (1976). They found that infestation by the 

greenhouse whitefly (Trialeurodes vaporariorum) sensitized bean plants (Phaseolus vulgaris) to 

injury by otherwise noninjurious low levels of O3, leading to premature senescence of the leaves. 

The overall picture regarding possible O3 effects on plant-insect relations, therefore, 

continues to be far from clear. Only a few of the very large number of such interactions that 

may affect crops, forest trees, and other natural vegetation have been studied. The trend 

suggested in the 1996 O3 AQCD (U.S. Environmental Protection Agency, 1996) that O3 may 

enhance insect attack has received some support from a few recent studies. However, the 

variability noted in most of the studies makes it clear that we are still far from being able to 

predict the nature of any particular O3-plant-insect interaction or its magnitude or severity. 

AX9-97
AX9.3.3.2 Oxidant-Plant-Pathogen Interactions 

Plant diseases are caused by pathogenic organisms, e.g., fungi, bacteria, mycoplasmas, 

viruses, and nematodes. Ozone impacts on disease are briefly discussed in earlier reviews by 

Ayres (1991) and Colls and Unsworth (1992) and, more recently, by Flückiger et al. (2002). 

Biotic interactions with forest trees have been reviewed by Chappelka and Samuelson (1998); 

Sandermann (1996) and Schraudner et al. (1996) have summarized molecular similarities and 

interrelationships between necrotic O3 injury to leaves and pathogen attack. A few recent 

publications have added to our fragmented knowledge of O3-plant-disease interactions and the 

mechanisms involved, but there appear to have been no reports to date of studies involving 

mycoplasmal diseases. 

The 1996 O3 AQCD (U.S. Environmental Protection Agency, 1996) noted the concept put 

forward by Dowding, (1988) “that pathogens and pests which can benefit from damaged host 

cells and from disordered transport mechanisms are enhanced by pollution insult to their hosts, 

whereas those pathogens and other symbionts which require a healthy mature host for successful 

invasion are depressed by pollutant stress to their host.” The pathogens of the first type are 

mostly facultative necrotrophic fungal parasites, whereas the second type are largely obligate 

biotrophic fungi, bacteria, and viruses. Based on this distinction, the majority of the cases cited 

in the 1996 document supported Dowding’s (1988) view, as have several more recent studies 

summarized in Table AX9-11. However, there are also some contradictions. 

Most investigations have focused on the incidence and development of disease on plants 

previously or concurrently exposed to O3, rather than on the corollary effect of disease on the 

response to O3. In all of the studies of facultative pathogens and the nematode studies, exposure 

to O3 tended to result in increased disease severity through increased spore germination or 

increased fungal growth and development; although in the case of grey mold (Botrytis cinerea) 

on kidney bean (Phaseolus vulgaris), this was only observed on an O3-sensitive cultivar 

inoculated with conidia (Tonneijck, 1994). After mycelial inoculation, O3 exposure reduced 

disease development in the O3-sensitive cultivar, but no satisfactory explanation was offered to 

account for the difference in response. With poplar leaf spot, Marssonina tremulae, on hybrid 

poplar (Populus trichocarpa × balsamifera), low level exposures to O3 increased disease (in 

agreement with theory) but higher levels (200 ppb, 8 h per day for 15 days) reduced conidial 

germination (Beare et al., 1999b). 

AX9-98
Table AX9-11. Interactions Involving O3 and Plant Pathogens 

Host Plant Pathogen Effect of O3 on Disease Effect of Disease on O3 Resp

Obligate Biotrophs Bottle gourd (Lagenaria siceraria) Powdery mildew 

(Sphaerotheca fulginea) 

Increased in 50ppb O3; 

decreased in 100+ppb 

Decreased; partial protection 

Cucumber (Cucumis sativa) Powdery mildew 

(Sphaerotheca fulginea; 

Increased in 50ppb O3; 

decreased in 100+ppb 

Synergistic increase in 50ppb O

antagonistic decrease in 100+pppartial protection 

Pea (Pisum sativum) Powdery mildew 

(Erysiphe polygoni) 

Decreased infection Decreased; partial protection 

Aspen (Populus tremuloides) Leaf rust (Melampsora 

medusae f. sp. tremuloidae) 

Increased severity Not reported 

Hybrid poplar (Populus trichocarpa × balsamifera) 

Leaf rust (Melampsora larici-populina or M. allii-populina) 

Increased infection and severity 

Increased sensitivity (synergist

Broad bean (Vicia faba) Bean rust (Uromyces 

viciae-fabae) 

Not reported Decreased; partial protection 

AX9-99

Facultative Necrotrophs Kidney bean (Phaseolus vulgaris) Grey mold (Botrytis Increased from conidia on O3 - Not reported 

cinerea) sensitive cultivar; decreased 

from mycelium 

Grey mold (Botrytis Increased infection Not reported cinerea) White mold (Sclerotinia Increased infection Not reported sclerotiorum) 

Scots pine (Pinus sylvestris) Annosus root and butt rot Increased development* Not reported 

(Heterobasidion annosum) 
Table AX9-11 (cont’d). Interactions Involving O3 and Plant Pathogens 

Host Plant Pathogen Effect of O3 on Disease Effect of Disease on O3 Resp

Facultative Necrotrophs (cont’d) Loblolly pine (Pinus taeda) Pitch canker (Fusarium Increased development Increased sensitivity 

subglutinans) 

Hybrid poplar (Populus deltoides × Canker (Septoria musiva Increased incidence Not reported 

nigra) [=Mycosphaerella 

populinum]) 

Hybrid poplar (Populus trichocarpa Leaf spot (Marssonina Increased spore germination Not reported × balsamifera) tremulae) and lesion growth after 

100ppb O3 (30 days); 

decreased germination after 200ppb (15 days) 

Wheat (Triticum aestivum) Blotch (Septoria Increased infection Not reported 

nodorum) 

Tan spot (Pyrenophora Increased infection of Not reported tritici-repentis) disease-susceptible genotypes 

Nematodes Tomato (Lycopersicon esculentum) Root-knot nematode Increased development Increased foliar injury; reduced 

(Meloidogyne incognita) growth (synergistic) 

AX9-100

* Increase completely countered by mycorrhizae (Hebeloma crustuliniforme). 
The situation with obligate biotrophic pathogens is less consistent. The effects on powdery 

mildew (Sphaerotheca fulginea) on both bottle gourd (Lagenaria siceraria) and cucumber 

(Cucumis sativa) resembled the situation with the necrotrophic poplar leaf spot disease, since 

low O3 exposures increased disease severity (in disagreement with theory), although higher 

levels decreased it. The decreased infection in the pea-powdery mildew (Erysiphe polygoni) 

situation agrees with theory, but the situations with leaf rust (Melampsora sp.) on hybrid poplar 

or aspen do not. However, these reports are in contrast to earlier reports included in the 1996 O3 

AQCD (U.S. Environmental Protection Agency, 1996) of observations with other species of 

Erysiphe (Tiedemann et al., 1991) and Melampsora (Coleman et al., 1987). In contrast to the 

recent report of a synergism with Melampsora on poplar, infections caused by the other 

biotrophs (Sphaerotheca, Erysiphe, Uromyces spp.) reduced the severity of injury caused by O3 

(in agreement with numerous earlier reports), but only at high O3 exposures in the case of 

Sphaerotheca on cucumber. At low exposure levels, the disease and O3 exposure acted 

synergistically. The only other recent observations of such disease-related synergisms are the 

tomato-nematode reports of Khan and Khan (Khan and Khan, 1997, 1998b). 

It is, therefore, clear that the type and magnitude of exposure to O3 plays an important role 

in determining both the responses of both the disease organism and the host. 

No recent studies involving interactions between O3 and bacterial diseases appear to have 

been reported since 1996. With regard to viruses, a laboratory study by Yalpani et al. (1994) 

added to several reports of O3 decreasing the severity of tobacco mosaic virus infection of 

tobacco; and Jimenez et al. (2001) reported that previous O3 exposure resulted in increased 

adverse effects on tomato yield attributed to several viral diseases. 

Similarities between the sensitivities of different cultivars or clones to O3 and to specific 

diseases have been noted. For example, Sah et al. (1993) found that the severity of injury caused 

by tar spot and O3 exposures of 12 wheat (Triticum aestivum L.) cultivars were closely correlated 

(R2

 = 0.986). Such similarities appear to have a mechanistic basis, as several studies have noted 

similarities in the molecular and biochemical changes that occur in plants infected with 

pathogens and in O3 exposed plants. Schraudner et al. (1992), Ernst et al. (1992), EckeyKaltenbach et al. (1994a,b), Yalpani et al. (1994), and Bahl et al. (1995) have presented evidence 

that O3 exposures result in responses such as increased levels of salicylic acid, the signaling 

agent for increased induced resistance to pathogens. This, in turn, leads to the activation of the 

AX9-101
genes that encode defense proteins, including the so-called pathogenesis-related proteins. The 

induction of such proteins might account for the decreased infection with Sphaerotheca and 

Melampsora at higher O3 exposures but does not account for increased infections seen at lower 

exposure levels. The issue is discussed more fully by Sandermann (1996) and Schraudner et al. 

(1996). More recently Sandermann (2000) has extended the theory relating O3 exposure and 

disease by suggesting that, because of O3 “memory effects” in affected host plants that may 

persist over weeks or months, analysis for various induced biomarkers of gene activation may 

provide a useful tool for improving our ability to predict the outcome of O3-plant-pathogen 

interactions. 

There have been no reports of O3 studies with mixed infections by pathogens, but the 

complete suppression of Heterobasidion butt and root rot of Scots pine by the mycorrhizal 

symbiont Hebeloma crustuliniforme indicates the possibility of interactions involving more than 

one fungus (see Section AX9.3.4.3.3 below). 

In summary, our understanding of oxidant-plant-disease interactions is far from complete. 

However, a combined tabulation of the evidence presented in the 1996 O3 AQCD (U.S. 

Environmental Protection Agency, 1996) and that noted in Table AX9-11 leads to the following 

summary of O3 effects on plant diseases and corollary effects of infection on plant response 

to O3, as indicated by the number of studies showing increases or decreases in disease or 

susceptibility. 

For obligate biotrophic fungi, bacteria, nematodes: 

O3 increased disease: 9 Increased susceptibility to O3: 3 

O3 decreased disease: 15 Decreased susceptibility to O3: 9 

For facultative necrotrophic fungi: 

O3 increased disease: 25 Increased susceptibility to O3: 2 

O3 decreased disease: 3 Decreased susceptibility to O3: 4 

Thus, although O3 may reduce the severity, but not the incidence, of some of the diseases caused 

by the obligate pathogens, the evidence overall indicates that with most diseases, their severity is 

more likely to be increased by O3 than not. However, the actual consequences will be specific to 

the disease and level of exposure, and, most importantly, will be determined by environmental 

suitability and epidemiological requirements for disease to develop. Conversely, some evidence 

AX9-102
suggests that infection by obligate pathogens may confer some degree of “protection” against O3, 

a dubious benefit from the plant’s point of view. 

AX9.3.3.3 Oxidant-Plant-Symbiont Interactions 

No further studies have appeared regarding O3 effects on the important bacterial symbiont 

of legumes, Rhizobium, since those summarized in the 1996 O3 AQCD (U.S. Environmental 

Protection Agency, 1996). Hence, our present understanding is that, although relatively high 

levels of exposure (>200 ppb) can result in severe (>40%) reductions in nodulation (and 

therefore nitrogen-fixation) on soybean roots, lower O3 exposures may cause lesser reductions in 

nitrogen fixation. However, the data are inadequate to attempt to define any quantitative 

exposure-response relationships. 

There have been a few recent reports on O3-plant-mycorrhizae interrelationships. These 

have mostly involved seedlings of coniferous tree species. A transient O3-induced stimulation of 

mycorrhiza on Scots pine roots reported by Kasurinen et al. (1999) was not observed in a later 

study by Kainulainen et al. (2000b). Studies of the mycorrhiza Paxillus involutus on birch 

(Betula pendula) seedlings showed that, although O3 reduced mycorrhizal growth rate, it led to 

greater extension growth which in turn resulted in greater mycorrhizal infection of neighboring 

Aleppo pine (Pinus halepensis) seedlings (Kytöviita et al., 1999). However, O3 reduced nitrogen 

acquisition by P. halepensis from its mycorrhizal symbiont (Kytöviita et al., 2001). The 

complex interrelationships that may occur in the rhizosphere were revealed by the observation 

by Bonello et al. (1993) that the mycorrhiza Hebeloma crustuliniforme could overcome 

the O3-stimulated severity of root rot on Scots pine caused by the fungus Heterobasidion 

annosum (noted in Section AX9.3.4.3.2). 

In summary, the available evidence is far too fragmented and contradictory to permit 

drawing any general conclusions about mycorrhizal impacts. The negative effects of O3 on 

mycorrhizae and their functioning that have been reported have not necessarily been found to 

lead to deleterious effects on the growth of host plants. Thus, little has changed from 1991 when 

Dighton and Jansen (1991) asked: “Atmospheric Pollutants and Ectomycorrhizae: More 

Questions than Answers?”. Because of their important roles in ecosystems, mycorrhizae are 

further discussed in Section AX9.6. 

AX9-103
AX9.3.3.4 Oxidant-Plant-Plant Interactions: Competition 

Plant competition involves the ability of individual plants to acquire the environmental 

resources needed for growth and development: light, water, nutrients, and space. Intraspecific 

competition involves individuals of the same species, typically in monocultural crop situations, 

while interspecific competition refers to the interference exerted by individuals of different 

species on each other when they are in a mixed culture. 

In cropping situations, optimal cultural practices for row spacing and plant density/row 

tend to balance the negative effects of intraspecific competition with the goal of maximum yield. 

Although interspecific competition is agriculturally undesirable when it involves weak 

infestations, the use of mixed plantings may be agriculturally deliberate, e.g., grass-legume 

mixtures used for pasture or forage. In natural plant communities, monocultures are rare, and 

complex interspecific competition is the norm. 

Although weak competition is the largest global cause of crop losses, little is known about 

the impact of O3 on crop-weed interactions. The topic does not appear to have been investigated 

in recent years. We can only speculate as to the possible consequences of O3 exposure on weed 

competition based on our limited understanding of the effects on a few, mostly two-component 

mixtures of cultivated species. 

The tendency for O3-exposure to shift the biomass of grass-legume mixtures in favor of 

grass species, reported in the 1996 O3 AQCD (U.S. Environmental Protection Agency, 1996) has 

been confirmed by recent studies. In a ryegrass (Lolium perenne) + clover (Trifolium repens) 

mixture grown in an open-air fumigation system, clover growth was impaired by extended 

exposures to above-ambient O3, leaving patches for weed invasion (Wilbourn et al., 1995). An 

open-top chamber study (OTC) by Nussbaum et al. (1995b) using the same species confirmed 

the greater effect on clover but observed that the magnitude of the effect depended highly on the 

pattern of O3-exposures over extended growing periods. Low-level exposures shifted species 

composition in favor of Lolium, but exposures to higher peak O3 levels depressed total mixture 

yield. With an alfalfa (Medicago sativa) + timothy (Phleum pratense) mixture, Johnson et al. 

(1996a) noted that O3 decreased alfalfa root growth and increased timothy shoot growth and 

height. Nussbaum et al. (2000a) reported that, with increased exposure to O3, well-watered red 

clover (Trifolium pratense) plants suffered from increased competition from the grass Trisetum 

flavescens, but the O3 exposure 

AX9-104
also negatively affected grass growth, depressing overall total yield. However, a greater adverse 

effect on Trisetum resulted from O3-induced increased competition when grown with brown 

knapweed (Centaurea jacea), a weed species. 

Andersen et al. (2001) demonstrated the potential for competition and O3 exposure to work 

together to affect the growth of tree seedlings. Ozone had no direct adverse effect on pine 

growth in a 3-year study of ponderosa pine (Andersen et al., 2001) seedlings grown in 

mesocosms with three densities of blue wild-rye grass (Elymus glaucus), but O3 exposure 

increased the competitive pressure of the grass which caused a major reduction in pine growth. 

Three studies have been reported on more complex plant associations. Ashmore and 

Ainsworth (1995) studied mixed plantings of two grasses, Agrostis capilaris and Festuca rubra, 

with two forbs15, Trifolium repens (a legume) and Veronica chamaedrys, exposed to O3 in OTCs. 

The proportion of forbs, Trifolium in particular, declined, especially when cut at biweekly 

intervals. In a related study, Ashmore et al. (1995) used artificial mixtures of grasses and forbs 

and transplanted swards of native calcareous grassland species and found that, regardless of 

whether total biomass was adversely affected by exposures to O3, higher exposures progressively 

shifted species composition, usually at the expense of the forb species. The observed shifts in 

competitive balance in favor of grasses is consistent with observations that many grass species 

are less sensitive to O3 than forbs. However, as previously shown by Evans and Ashmore 

(1992), knowledge of the relative sensitivities to O3 of the component species grown in isolation 

or in monoculture does not always predict the impact of O3 on the components in a mixed 

culture. 

Barbo et al. (1998) exposed an early successional plant community to O3 in OTCs for two 

growing seasons. Ozone decreased community structure features such as height of canopy, 

vertical canopy density (layers of foliage), and species diversity and evenness. Surprisingly, 

blackberry (Rubus cuneifolius), a species considered to be O3-sensitive, replaced sumac (Rhus 

copallina) canopy dominance. Barbo et al. (2002) also demonstrated the role of competition in 

determining the impact of O3 on loblolly pine (Pinus taeda). They reported that the increased 

growth of natural competitors in OTCs using charcoal-filtered air to reduce the ambient O3 

concentrations resulted in decreased pine growth. They noted that this is contrary to the 

15 Forb: any non-grassy herbaceous species on which animals feed. 

AX9-105 
frequently reported increased growth observed in reduced O3 levels in the absence of 

interspecific competition. 

McDonald et al. (2002) classified four clones of aspen (Populus tremuloides) as either 

competitively advantaged or disadvantaged, based on their height relative to the height of 

neighboring trees, and exposed them to 1.5× ambient O3 in a FACE facility over a 4-year period. 

Competitively disadvantaged trees were proportionately more adversely affected by O3 than 

competitively advantaged or neutral trees (McDonald et al., 2002). However, one clone of the 

disadvantaged trees demonstrated enhanced growth. 

In summary, our present knowledge of how O3 may affect the competitive interspecific 

plant-plant relationships typifying the agricultural and natural worlds is very limited. However, 

as noted in the 1996 O3 AQCD (U.S. Environmental Protection Agency, 1996), “the 

development and use of field exposure systems have permitted many recent studies of crop 

species to be conducted at normal planting densities and hence have incorporated intraspecific 

competition as an environmental factor.” Such facilities were used in most of the studies of 

interspecific competition discussed above. But we are still far from being able to use small 

model competing systems to extrapolate to the realities of natural ecosystem complexity. 

AX9.3.4 Physical Factors 

The physical features of a plant’s aerial and edaphic environments exercise numerous 

controls over its growth and development. Thus, many of their effects may be modified by 

exposure to atmospheric oxidants and, alternatively, plants may modify responses to such 

exposures. As in the 1996 O3 AQCD (U.S. Environmental Protection Agency, 1996), this 

section focuses on the defining features of plant microclimate: light, temperature, relative 

humidity (HR, or saturation vapor pressure deficit), and the presence and availability of water, 

especially in the soil. Monteith and Elston (1993) suggested that light energy and mass of water 

should be viewed as climatic resources and that the other two elements (temperature and 

saturation vapor pressure deficit) be viewed as rate modifiers that determine how fast the 

resources are used. The modifications of plant response by physical environmental factors has 

recently been reviewed by Mills (2002). 

Another physical feature of the microclimate, wind and air turbulence, which affects the 

thicknesses of the boundary layers over leaves and canopies and, hence, affects gas exchange 

AX9-106
rates (including the fluxes of O3 and other oxidants into the leaves) is discussed elsewhere 

(Section AX9.4). 

Physical features of the environment are also important components of larger-scale 

regional and global climates. However, the following discussions are confined to issues related 

to individual factors at the plant level; meso-scale effects are reviewed in Section AX9.3.4.8, 

which addresses the issues of climate change interactions. 

AX9.3.4.1 Light 

Plants are the primary producers of biomass on the planet through their ability to capture 

light energy (by the process of photosynthesis) and convert it to the many forms of chemical 

energy that sustain their own growth and that of secondary consumers and decomposers. Light 

intensity is critical because the availability of light energy (a resource, sensu Monteith and Elston 

(1993) governs the rate at which photosynthesis can occur, while light duration (i.e., 

photoperiod) profoundly effects development in many species. Although light quality (i.e., the 

distribution of incident wavelengths) may also affect some physiological plant processes, there is 

no evidence to indicate that such effects are of relevance to concerns over oxidant pollution, 

except at the short wavelengths of UV-B. This topic is discussed in the context of climate 

change in Section AX9.3.8.2, and as a stress factor per se affected by atmospheric O3 in 

Chapter 10. However, as noted above and in the 1996 O3 AQCD (U.S. Environmental 

Protection Agency, 1996), none of these features is controllable in natural field situations. 

A brief discussion of light intensity-O3 interactions is included in the review by Chappelka and 

Samuelson (1998). 

The conclusion in the 1996 O3 AQCD (U.S. Environmental Protection Agency, 1996) that 

low light intensities and short photoperiods tended to increase susceptibility to foliar O3-injury 

may still be valid, but this may or may not translate into adverse effects on growth. For 

example, Tjoelker et al. (1993) found that, when seedlings of sugar maple, a shade-tolerant 

species, were grown in 7% full sunlight, O3 reduced shoot and root growth, but had no 

significant effect in 45% sunlight (a 6-fold increase). In contrast, the reverse was observed with 

a shade-intolerant hybrid poplar, with the greater impact of O3 occurring in the higher light 

intensity treatment. 

AX9-107
The greater sensitivity of maple in low light has also been confirmed in other studies. 

Tjoelker et al. (1995) noted a greater O3-induced inhibition of photosynthetic CO2 assimilation in 

shaded leaves than in leaves in full sunlight. However, in the absence of differences in stomatal 

conductance, the effect was considered to be independent of O3 flux; it appeared to be a 

consequence of reduced chlorophyll contents and quantum efficiencies induced by O3. In 

contrast, Bäck et al. (1999), who also observed a greater inhibition of net photosynthesis by O3 in 

shaded leaves, reported decreased stomatal conductance. Although reduced conductance might 

suggest reduced O3 flux and, therefore, decreased adverse effects, the authors concluded that the 

effects of reduced conductance were offset by long-term changes in leaf structure, leading to less 

densely packed mesophyll cells and greater internal air space within the leaves. Morphological 

differences between lower and upper crown leaves of black cherry (Prunus serotina) have been 

suggested as the basis for the greater O3-susceptibility of the lower crown leaves (Fredericksen 

et al., 1995). Bäck et al. (1999) also observed accelerated foliar senescence induced by O3 on 

shaded leaves, a response also noted by Topa et al. (2001). Sensitivity to O3 was found to be 

increased in shade- but not sun-leaves of shade-tolerant red oak (Quercus rubra) (Samuelson and 

Edwards, 1993). Similarly, Mortensen (1999) observed that seedlings of mountain birch (Betula 

pubescens) grown in 50% shade suffered greater foliar injury from O3 than those grown in full 

sunlight. 

Not all shade-intolerant species exhibit greater reductions in photosynthesis and growth 

due to O3 when grown in full sunlight. Higher than ambient levels of O3 failed to inhibit 

photosynthesis in leaves of shade-intolerant yellow poplar (Liriodendron tulipifera) grown in 

nearly full sunlight (Tjoelker and Luxmoore, 1991). Greater foliar injury in the lower, shaded 

leaves of shade-intolerant black cherry trees and saplings, was attributed to higher stomatal 

conductance and greater O3 uptake relative to net photosynthetic rate (Fredericksen et al., 

1996a). However, in a 3-year study of Norway spruce seedlings in OTCs, Wallin et al. (1992) 

observed that photosynthetic efficiency was more adversely affected by O3 in high than in 

low light. 

The suggestion of greater sensitivity to O3 of shade-tolerant species in low-light conditions 

and the greater sensitivity of shade-intolerant species in high light is somewhat of an 

oversimplification when dealing with mature trees, for which light intensity varies considerably 

within the canopy because of shading. Chappelka and Samuelson (1998) noted that the 

AX9-108
interaction between sensitivity to O3 and the light environment in forest trees is further 

complicated by developmental stage, with seedlings, saplings, and mature trees frequently giving 

different results. Topa et al. (2001) also cautioned that O3 effects on leaf-level photosynthesis 

may be poor predictors of the growth responses of sugar maple in different light environments. 

In high-light intensities, many species exhibit some degree of photoinhibition of the 

photosynthetic process through the overloading of the mechanisms that protect the 

photosynthetic reaction centers in the chloroplasts. Guidi et al. (2000) reported complex 

interactions between high-light intensities (inducing photoinhibition) and O3 exposures in kidney 

bean with high intensities tending to enhance the detrimental effect of O3 on photosynthesis. 

One of the studies in the extensive European Stress Physiology and Climate Experiment-wheat 

(ESPACE-wheat) program (Bender et al., 1999), conducted in 1994 and 1995, included an 

investigation of the effects of climatic variables on yield response to O3 using two simulation 

models, AFRCWHEAT2-O3 and LINTULCC (Ewart et al., 1999; Van Oijen and Ewart, 1999). 

Among the observed trends, it was noted that relative yield loss of wheat due to elevated O3 

tended to increase with light intensity. In contrast, Balls et al. (1996) used ANNS to investigate 

microclimatic influences on injury caused by O3 to clover (Trifolium subterraneum) and found 

that, especially at mid-range cumulative O3 exposures (350 to 500 ppb-h), injury tended to 

decrease with increasing light intensity. Similar observations by Davison et al. (2003) of foliar 

injury to wild populations of cutleaf cone flower (Rudbeckia laciniata) exhibiting a range of 

PAR levels within their canopies led the authors to conclude that the variation in injury 

symptoms observed was “unlikely to be due to differences in ozone flux and more likely to be 

due to variation in light.” Antonielli et al. (1997) found evidence indicating that the high 

sensitivity of the bioindicator tobacco cultivar Nicotiana tabacum cv. Bel-W3 is partly 

determined by its high photosynthetic electron transport rates at high-light intensities, which 

exceed the capabilities of the plant to dissipate energy and oxyradicals. 

The 1996 O3 AQCD referred to the important role of light in controlling stomatal opening 

and suggested that light duration (i.e., photoperiod) might dictate the actual uptake of O3 to some 

degree. However, it should also be noted that Sild et al. (1999) found that clover plants could 

suffer foliar injury even if they were exposed to O3 during the dark period of the day-night cycle, 

when stomatal conductance is at its lowest. 

AX9-109
A possible indirect effect of light intensity was noted by Reiling and Davison (1992c) in 

their study of the O3-tolerance of common plantain (Plantago major L.) plants grown from seeds 

collected from populations at 28 different sites in Britain. Ozone-tolerance, defined in terms of 

plant growth, was found to be a function of both previous O3-exposure history and hours of 

bright sunshine during the year before the seeds were collected. However, the authors cautioned 

that, since tropospheric O3-formation is itself dependent upon irradiation, the observation does 

not necessarily imply a direct effect of light intensity on the plants’ response to O3. 

The only recent studies concerning interactions with light quality appear to be those 

involving O3 and UV-B as a component of climate change. These are dealt with in Section 

AX9.3.4.8.2. The effects of photoperiod on response to O3 or the converse do not appear to have 

received any recent attention. 

Although the intensity, quality, and duration of light are not controllable in the natural 

world, the interactions of O3 with light intensity, in particular, clearly have relevance to the 

growth of shade-tolerant and shade-intolerant species in mixed forest stands. It appears that the 

nature of light intensity-O3 interactions may depend upon the type of light environment to which 

the species are best adapted, with increased light intensity increasing the sensitivity of lighttolerant and decreasing the sensitivity of shade-tolerant species to O3. Although there is 

certainly some evidence to the contrary, this hypothesis is a reasonable summation of current 

understanding with regard to O3-light intensity interactions. 

AX9.3.4.2 Temperature 

“Temperature determines the start and finish and rate and duration of organ growth and 

development” (Lawlor, 1998). Such processes depend on fundamental physiological activities 

that are mostly enzyme-mediated and whose kinetics are directly affected by temperature. Since 

the processes of enzyme deactivation and protein denaturation also increase as temperatures rise, 

each enzymatic process has a unique optimum temperature range for maximal function. 

However, the optima for different processes within the plant vary appreciably and, hence, the 

optimum temperature range for overall plant growth is one within which all of the individual 

reactions and vital processes are collectively functioning optimally, not necessarily maximally. 

Furthermore, individual features of plant development (e.g., shoot and root growth, flowering, 

pollen tube growth, fruit set, seed development) have different specific optima, so that 

AX9-110
differential responses to temperature occur, leading to temperature-induced developmental 

changes. For example, despite increased assimilation, increased temperatures may result in 

decreased grain yields of crops such as wheat, because the growing season is effectively 

shortened by a more rapid onset of senescence (Van Oijen and Ewart, 1999). 

Rowland-Bamford (2000) noted that a plant’s response to temperature changes will depend 

upon whether it is growing at its near optimum temperature for growth or its near maximum 

temperature and whether any increase in mean temperature results in temperatures rising above 

the threshold for beneficial responses. Impairment by O3 of any process may be thought of as 

being analogous to a downward shift below and away from the temperature optimum or an 

upward shift above and away from the optimum. Since a temperature rise toward the optimum 

would result in a rate increase, the combined effects of O3 and such an increase might neutralize 

each other, while the effects of O3 and a decrease in temperature would likely be additively 

negative. Above the optimum temperature, the situations would be reversed with the effects of 

increased temperatures and O3 being additively negative, and decreasing temperatures 

counteracting any negative effect of O3. Thus, it is difficult to generalize about the interactions 

of temperature and O3 on overall plant responses such as growth in which the different 

temperature-rate relationships of different growth components are merged, because they depend 

upon the relationship of any temperature changes to the optimum for a species. 

Studies of the effects of temperature on the impact of O3 have increased recently because 

of an increased need to understand the consequences of global warming as a component of 

climate change. Direct interactions of temperature with O3 are reviewed here, but the issues are 

addressed again in Section AX9.3.8.1 in relation to changes in atmospheric CO2 levels. 

The 1996 O3 AQCD (U.S. Environmental Protection Agency, 1996) stressed the 

interdependence of the temperature within the tissues of the leaf (where the various temperaturesensitive processes occur) on three distinct components: the ambient air temperature, the heating 

effect of incident infrared radiation during the photoperiod, and the evaporative cooling effect 

caused by transpirational loss of water. It also cautioned that, especially in experiments using 

controlled environment chambers, the effects of temperature could well be confounded with 

those of humidity/vapor pressure deficit (VPD). Temperature and VPD are strongly interrelated, 

and VPD plays an important role in regulating stomatal transpiration. Because of the role that 

evaporative cooling plays in determining internal leaf temperatures, any factor that causes 

AX9-111
stomatal closure and reduced conductance inevitably leads to increased leaf temperatures. Such 

interactions add to the difficulties in distinguishing the effects of temperature from those of other 

factors, as actual leaf temperatures are rarely measured and reported. 

Despite these caveats, there is some evidence that temperature per se influences plant 

response to O3. For example, in rapid-cycling Brassica (Brassica rapa) and radish (Raphanus 

sativus), marked O3-inhibitions of growth were observed at low root temperatures (13 °C) but 

not at 18 °C (Kleier et al., 1998, 2001). With regard to air temperature, this was included in the 

range of micrometeorological variables studied in several recent extensive field studies and was 

found to have a significant effect on response to O3 in most cases. Balls et al. (1996) used ANNs 

in an analysis of the growth of clover (Trifolium subterraneum) and concluded that light and 

VPD had greater influences than temperature on the visible injury response to O3. However, in 

three studies with different cultivars of white clover (Trifolium repens), temperature was found 

to be important to the growth response. Ball et al. (1998) exposed T. repens cv. Menna to 

ambient O3 in OTCs at 12 European sites at a range of latitudes and altitudes from 1994 to 1996. 

The impact of O3 on growth was determined as the ratio of growth with and without treatment 

with the O3-protectant, EDU (see Section AX9.2). Artificial neural network analysis showed 

that O3 exposure (measured as the AOT40 index, see Section AX9.3.6), VPD, and temperature 

were consistently the three most important variables governing response to O3 over a range of 

different ANN models. However, the authors did not describe the form of the O3-response 

relationship with temperature. Similar observations were reported by Ball et al. (2000) and Mills 

et al. (2000) for O3-sensitive and -tolerant clones of T. repens cv. Regal, grown at 14 to 18 

European locations from 1995 to 1998. In both studies, the impact of O3 was measured as the 

sensitive/tolerant growth (biomass) ratio. Although Ball et al. (2000) found temperature to be 

less important than O3 exposure and VPD, Mills et al. (2000) found temperature to be the most 

important input variable after O3 exposure (AOT40). In both cases, the adverse effect of O3 

increased with increasing temperature. 

A study of black cherry seedlings and mature trees in Pennsylvania, using 

micrometeorological variables aimed to predict O3 uptake, found temperature to be unimportant 

(Fredericksen et al., 1996b), but in the study of populations of common plantain referred to in 

Section AX9.3.4.1, Reiling and Davison (1992c) noted a weak, positive correlation between 

mean temperature at the collection site and O3 tolerance (based on growth rate) of the different 

AX9-112
populations. In contrast, Danielsson et al. (1999) collected genotypes of Phleum arvense from a 

wide range of Nordic locations and found a positive effect of temperature on the growth of 

genotypes from locations with higher summer temperatures, but sensitivity to O3 did not vary 

systematically with geographic location. 

Van Oijen and Ewart (1999) studied the effects of climatic variables on the response to O3 

in the ESPACE-wheat program, based on two distinctive simulation models (AFRCWHEAT2­

O3 and LINTULCC [Ewart et al., 1999]) and noted that although the relative yield loss of wheat 

due to elevated O3 tended to increase with temperature, the effect was of minor significance. 

In contrast to the variable results obtained in studies of the effects of temperature on 

response to O3, the corollary effect of O3 exposure on subsequent sensitivity to low temperature 

stress, noted in the 1996 criteria document, is well recognized. In reviewing low temperature-O3 

interactions, Colls and Unsworth (1992) noted that winter conditions produce three kinds of 

stress: desiccation, chilling or freezing temperatures, and photooxidation of pigments. Of these, 

they suggested that while the first two were important, the last may play a particularly significant 

role because the “combination of high irradiance and low temperatures permits a buildup of free 

radicals in leaf tissue, and these free radicals then attack chlorophyll.” Chappelka and FreerSmith (1995) suggested that the injury and losses to trees caused by this delayed impact of O3 

may be equally or more important than the direct impacts of O3 on foliage of visible injury and 

necrosis, or the disruption of key physiological processes such as photosynthesis. In this 

context, the 1996 O3 AQCD (U.S. Environmental Protection Agency, 1996) referred to the 

conceptual framework of Eamus and Murray (1991), which is still valid: brief periods of mild 

temperatures in the severest winters result in dehardening; O3 decreases frost hardiness per se, 

but also increases the predisposition to dehardening; dehardening places O3-exposed trees at 

greater risk from subsequent low temperatures. However, no quantified models of these effects 

have yet appeared. 

The 1996 O3 AQCD also noted that O3 adversely affects cold hardiness of herbaceous 

species. More recently, Foot et al. (1996, 1997) observed winter injury and decreased growth in 

low-growing perennial heather Calluna vulgaris exposed to O3 (70 ppb, 8 h/day, 5 days/week for 

6 months) during the winter (6.8 °C mean), but found no significant effects from the same 

exposures during the summer (12.3 °C mean). Although Potter et al. (1996) observed a similar 

AX9-113
situation with the moss Polytrichum commune, the reverse was found with the moss Sphagnum 

recurvum. 

In summary, unequivocal evidence exists that O3 causes sensitization to the adverse effects 

of low temperatures, but there is no clear pattern in the evidence regarding the effects of 

temperature on O3 response. The many contradictory responses to temperature and O3 probably 

reflect our lack of detailed knowledge of the temperature optima for the different growth 

components of the studied species. The topic of temperature-oxidant interactions is revisited 

later in Section AX9.3.4.8 in the context of global warming as a feature of climate change. 

AX9.3.4.3 Humidity and Surface Wetness 

The moisture content of the ambient air (or its VPD) is a rate modifier (sensu Monteith and 

Elston (1993)) and an environmental regulator of stomatal conductance. Both of the previous O3 

AQCDs (U.S. Environmental Protection Agency, 1986, 1996) concluded that the weight of 

evidence indicated that high RH (=low VPD) tended to increase the adverse effects of O3, 

principally because the stomatal closure induced in most situations by O3 is inhibited by high 

RH, leading to increased O3 flux into the leaves. 

Recent reports have confirmed this role of RH. The studies by Balls et al. (1995, 1996) 

and Ball et al. (1998) showed that VPD was an important determinant of O3-induced injury 

and reduced growth in two species of clover, Trifolium repens cv. Menna and T. subterraneum. 

However, Mills et al. (2000) found it to be unimportant in the case of T. repens cv. Regal. Such 

difference between cultivars is not unexpected , because considerable differences also occur 

among species and genera. For example, Bungener et al. (1999a) studied 26 Swiss grassland 

species and found clear evidence that O3 injury increased with decreased VPD (i.e., increased 

RH) in only eight species. However, the 1995 data from the European cooperative study of O3 

injury, which involved 28 sites in 15 countries and six crop species, led to the development of 

two 5-day critical-level scenarios involving O3-exposure (calculated as the AOT40 index) and 

mean VPD (0930-1630h): 200 ppb-h at >1.5 kPa, and 500 ppb-h at < 0.6 kPa (Benton et al., 

2000). 

With forest tree species, Fredericksen et al. (1996b) found significant correlations between 

stomatal conductance of black cherry leaves and RH (+ve) and VPD (!ve), and studies on freestanding Norway spruce and larch (Larix decidua) showed that although ambient VPD was 

AX9-114
highly positively correlated with ambient O3 concentration, increased VPD caused stomatal 

closure, reducing O3 uptake and impact (Wieser and Havranek, 1993, 1995). 

Surface wetness may affect O3 response through its direct effects on deposition to the 

surface and through changes in RH. Effects on the deposition of O3 have been reviewed by Cape 

(1996). A surface film of water on leaves was found to increase O3 deposition in four studies 

involving field-grown grape (Vitis vinifera) (Grantz et al., 1995), red maple (Acer rubrum) 

(Fuentes and Gillespie, 1992), deciduous forest dominated by largetooth aspen (Populus 

grandidentata) and red maple (Fuentes et al., 1992), and clover-grass mixed pasture (Trifolium 

pratense, Phleum pratense, and Festuca pratensis) (Pleijel et al., 1995a). In each case, the 

increased deposition could be attributed partly to an increased stomatal conductance through the 

abaxial (lower) surface and partly to uptake into the aqueous film on the adaxial (upper) surface. 

In contrast, decreased deposition was noted by Grantz et al. (1997) with field-grown cotton 

(Gossypium hirsutum). Since cotton is amphistomatous, with functional stomata on both leaf 

surfaces, it was suggested that, in this case, the water layer effectively sealed the adaxial surface 

stomata, more than offsetting any increase in conductivity of the stomata in the abaxial surface. 

However, none of the studies investigated the consequences of the differences in deposition. 

Although it could be inferred that, with part of any increased deposition being the result of 

increased O3 flux into the leaves, there would be the likelihood of increased O3 adverse effects, 

as suggested by earlier studies (Elkiey and Ormrod, 1981) that, by misting bluegrass (Poa 

pratensis) during exposure to O3, injury was significantly increased. 

To conclude, the effects of high RH (low VPD) and surface wetness have much in 

common, as they both tend to enhance the uptake of O3, largely through effects on stomata 

leading to increased impact. 

AX9.3.4.4 Drought and Salinity 

The 1996 O3 AQCD (U.S. Environmental Protection Agency, 1996) concluded that the 

available evidence clearly indicated that exposure to drought conditions could reduce the adverse 

effects of O3 on the growth of herbaceous and woody plants, but it also noted that no quantitative 

models of the O3-soil moisture deficit (SMD) interaction had yet appeared in print. 

Nevertheless, the “protective” effect was inconsistent, and only appeared when SMD was 

accompanied by high evaporative demand. Since that time, further studies have confirmed the 

AX9-115
interaction, and simulation models have begun to appear. Mills (2002) has recently provided a 

brief review of the topic. 

With regard to herbaceous species, Vozzo et al. (1995) observed less O3-induced injury and 

suppression of net photosynthesis and growth in water-deficient soybean (Glycine max) than in 

well-watered plants. In several studies with wheat (Triticum aestivum), on the other hand, 

although adverse effects of both O3 and SMD were noted, they were consistently additive 

(Bender et al., 1999; Fangmeier et al., 1994a,b; Ommen et al., 1999). 

In attempting to model the stomatal conductance of wheat in relation to O3 and soil 

moisture, Grüters et al. (1995) found that although O3-induced stomatal closure was enhanced by 

SMD, reducing O3 uptake, the R2 of the overall model was only 0.40, indicating that other 

significant factors or relationships were involved. 

With regard to native vegetation, Bungener et al. (1999a) used mixed plantings of 24 Swiss 

grasses, herbs, and legumes and observed that, although O3-drought interactions were speciesspecific, they tended to reflect stomatal functioning. They found that SMD reduced O3 injury in 

two clovers (Trifolium repens and T. pratensis) and two grasses (Trisetum flavescens and 

Bromus erectus), but noted no interactions in the other 20 species. With relative growth rate as 

the measure of response to O3, interactions with SMD were noted in only three species: 

Trifolium repens and two weedy herbs, Knautia arvense and Plantago lanceolata (Bungener 

et al., 1999b). Although this variability in response among species was noted in the review by 

Davison and Barnes (1998), they also pointed out that in severely droughted regions of Europe, 

notably in Greece and Spain, O3-induced injury and growth reductions were common on many 

(usually irrigated) crops, but there were virtually no records of injury symptoms in wild species. 

Thus, the situation with herbaceous species is essentially unchanged from 1988 when 

Heagle et al. (1988) summarized the extensive NCLAN experiments that incorporated water 

stress as a variable: “SMD can reduce the response of crops to O3 under some conditions but not 

under other conditions. Probably the occurrence of O3 by SMD interactions was dependent on 

the degree of SMD-induced plant moisture stress.” 

With regard to trees, O3 interactions with soil water availability have been discussed in 

several recent reviews: Chappelka and Freer-Smith (1995), who focused on O3-induced 

predisposition to drought stress (Johnson et al., 1996b; Chappelka and Samuelson, 1998; and 

Skärby et al., 1998). 

AX9-116
Several recent studies with conifers have yielded mixed results. No interactions with 

drought were observed by Broadmeadow and Jackson (2000) on Scots pine, by Karlsson et al. 

(2002) on Norway spruce, or by Pelloux et al. (2001) on Aleppo pine. More recently, Le Thiec 

and Manninen (2003) reported that drought reduced O3-induced growth suppression of Aleppo 

pine seedlings. Panek and Goldstein (2001) inferred less impact of O3 on droughted Ponderosa 

pine, and Van Den Driessche and Langebartels (1994) reported that drought reduced injury and 

O3-induced ethylene release by Norway spruce. But Karlsson et al. (1997), in a comparative 

study of fast- and slow-growing clones of P. abies, only observed a drought-induced reduction of 

O3-inhibited root growth in the fast-growing clone. In contrast, Grulke et al. (2002b) noted a 

synergistic interaction between O3 and drought stress on gross photosynthesis of Pinus 

ponderosa, and Wallin et al. (2002) reported a synergistic growth response of Norway spruce in 

the third year of a 4-year study. A similar response was noted by Dixon et al. (1998) with the 

Istebna strain of Norway spruce. 

With broad-leaved trees, studies of Durmast oak (Quercus petraea) (Broadmeadow et al., 

1999; Broadmeadow and Jackson, 2000) and European ash (Fraxinus excelsior) (Broadmeadow 

and Jackson, 2000; Reiner et al., 1996) showed that drought provided partial protection 

against O3-induced growth reduction. Although European beech (Fagus sylvatica) is reportedly 

an O3- and drought-sensitive species, neither Pearson and Mansfield (1994) nor Broadmeadow 

et al. (1999) observed any interactions between these stresses, while Dixon et al. (1998) 

observed partial protection. Pääkönen et al. (1998) observed only additive effects in a sensitive 

clone of birch (Betula pendula). However, the experiments of Schaub et al. (2003) and the 

survey by Vollenweider et al. (2003a) on black cherry clearly indicate antagonism between 

drought and O3 stresses on this species. 

With regard to the converse effect, in a critical review of the evidence for predisposition to 

drought stress being caused by O3, Maier-Maercker (1998) supported the hypothesis and 

suggested that the effects were caused by the direct effects of O3 on the walls of the stomatal 

guard and subsidiary cells in the leaf epidermis, leading to stomatal dysfunction. 

The Plant Growth Stress Model (PGSW) developed by Chen et al. (1994) is a physiologybased process model which includes drought among several environmental variables. 

Simulations for Ponderosa pine incorporated antagonistic effects between O3 and drought 

stresses, i.e., partial protection, although Karlsson et al. (2000) have since emphasized that 

AX9-117
drought-induced “memory effects” should be considered when developing simulation models 

incorporating stomatal conductance. 

Retzlaff et al. (2000) used the single-tree model, TREEGRO, to simulate the combined 

effects of O3 and drought on white fir (Abies concolor). Although simulated reductions in 

precipitation $25% reduced growth, they also reduced O3 uptake (and impact). But lesser 

reductions in precipitation combined synergistically with O3 stress to reduce growth, leading the 

authors to conclude that moderate drought may not ameliorate the response of white fir to O3. 

On a much larger scale with a modified forest ecosystem model (PnEt-II) incorporating 

O3-response relationships for hardwood species, Ollinger et al. (1997) showed how predicted 

changes in net primary production and mean wood production in the northeastern U.S. hardwood 

forests due to O3 would be reduced (but not countered or reversed) by drought stress, particularly 

in the southern part of the region. This geographic distribution of the effect was substantiated by 

the work of Lefohn et al. (1997) on the risk to forest trees in the southern Appalachian 

Mountains, based on localized estimates of O3 levels and SMD. The TREEGRO and ZELIG 

models were combined by Laurence et al. (2001) to predict the impacts of O3 and moisture (as 

precipitation) on the growth of loblolly pine and yellow poplar. Based on O3 and precipitation 

data from three sites in the eastern United States, the six model regressions developed for the 

two species included both positive and negative coefficients for O3 exposure and precipitation as 

determinants of growth. 

As noted in the 1996 O3 AQCD (U.S. Environmental Protection Agency, 1996), the effects 

of soil salinity are similar to those of SMD. In a study of rice (Oryza sativa) cultivars of 

differing sensitivity to salinity, Welfare et al. (1996) noted that although both O3 and salinity 

reduced many features of growth additively, antagonistic interactions were only seen for leaf 

length and potassium accumulation. Similarly, a recent study on chickpea (Cicer arietinum) 

found no interactions with regard to most components of biomass accumulation (the effects 

of O3 and salinity were additive), but with root growth, salinity suppressed the adverse effects 

of O3. 

In summary, the recently described interactions of O3 and drought/salinity stresses are 

consistent with the view that, in many species, drought/salinity reduces the impact of O3, but O3 

increases sensitivity to drought stress, i.e., the type of response is determined by the sequence of 

stresses. However, synergisms have also been observed and any antagonisms are speciesAX9-118
specific and unpredictable in the absence of experimental evidence. In no case has an 

antagonism been found to provide complete protection. 

AX9.3.5 Nutritional Factors 

The 1996 O3 AQCD (U.S. Environmental Protection Agency, 1996) noted that the large 

number of macro- and micronutrients and the wide range of species had limited the number of 

experimental investigation to all but a few cases of nutrient-O3 interactions and most of these 

concerned nitrogen (N) and crops or forest tree species. The document also provided a 

comprehensive tabulation of the results of the relevant studies up to 1992. 

The suboptimal supply of mineral nutrients to plants leads to various types of growth 

reductions. The consequences of suboptimal nutrition might, therefore, be expected to have 

some similarities to those of O3 exposure. One might expect nutritional levels below the 

optimum either to amplify any effects of O3 or at least lead to additive responses. The difficulty 

with this suggestion is that the available information has mostly been obtained from 

experimentation conducted using two or more arbitrarily selected levels of fertility with little or 

no regard to optima. Hence, it is not surprising that there have been contradictory reports, even 

among studies with the same species or cultivars conducted by different workers at different 

locations using different soils or soil mixes. 

There appear to have been no recent studies on O3 interactions with specific mineral 

nutrients other than N. Hence, the previous conclusions are still valid, viz. that increasing levels 

of the major elements potassium (K) and sulfur (S) usually reduce the impact of O3, or, 

deficiency increases susceptibility, whereas increased phosphorus (P) usually increases injury, 

or, deficiency decreases susceptibility. 

However, with N, a relationship to the optimum is usually demonstrable. Several earlier 

studies of O3 × N interactions reported that the adverse effects of O3 on growth were greatest at 

the optimum and decreased with increasing N-deficiency, a finding supported by the work on 

aspen of Pell et al. (1995), who also confirmed that excess N decreased O3 impact on growth. 

Similarly, the adverse effects of O3 on growth rate in wheat diminished with decreased N supply 

(Cardoso-Vilhena and Barnes, 2001). However, the effects of N are far from consistent. For 

example, Greitner et al. (1994) reported that O3 and N-deficiency acted additively in aspen in 

reducing leaf surface area and rate of photosynthesis, Bielenberg et al. (2001) reported that the 

AX9-119
rate of O3-induced senescence was increased by N-deficiency in hybrid poplar (Populus 

trichocarpa × P. maximovizii), and Pääkkönen and Holopainen (1995) observed the least adverse 

effects of O3 on European white birch (Betula pendula) at optimum N-fertility levels. With 

cotton, increased N-levels more than overcame the adverse effect of O3 on growth and boll yield 

(Heagle et al., 1999b). In view of these contradictions, one may conclude that other, unrecorded 

factors may have contributed to the various findings. Thus, much remains unclear about O3 × Nfertility interactions. 

There have been two recent studies on the effects of overall soil fertility. Whitfield et al. 

(1998) observed that low general fertility increased O3 sensitivity in selections of common 

plantain. At the biochemical level, well-fertilized European white birch saplings were found to 

be less adversely affected by O3 than nutrient-stressed plants (Landolt et al., 1997). 

TREEGRO model simulations of the growth of red spruce (Picea rubra) in conditions of 

nutrient deficiency and O3 stress showed that, in combination, the two stresses acted less than 

additively (Weinstein and Yanai, 1994). Minimal amelioration by nutrient deficiency was 

predicted with Ponderosa pine. 

Plants may also obtain N and S from airborne sources such as NOx, HNO3, NO3 

!

, SO2, 

and SO4

2!

, although, depending upon their concentration, these may also be phytotoxic. 

In various parts of the world, the deposition of N and S in these forms contributes significantly 

to the levels of nutritionally available N and S in soils. Such depositions may, in turn, influence 

the impact of O3 on sensitive species through their roles as nutrients independent of any 

interactions that may occur because of their acidic properties (see Section AX9.3.6.5). For 

example, Takemoto et al. (2001) recently reviewed the situation in southern California’s mixed 

conifer forests and noted that, where N-deposition is appreciable, its combination with O3 is 

causing a shift in Ponderosa pine biomass allocation toward that of deciduous trees, with 

increased needle drop so that only 1- and 2-year needle classes overwinter. Such changes are 

having significant consequences on the balance of the forest ecosystem and are discussed more 

fully in Section AX9.5. 

Of the micronutrient elements, only manganese (Mn) appears to have been studied 

recently. In beans (Phaseolus vulgaris) Mn-deficiency increased O3 toxicity, despite causing 

reduced O3 uptake (through decreased stomatal conductance) and inducing increased levels of 

Mn-SOD (Mehlhorn and Wenzel, 1996). 

AX9-120
In view of the foregoing, it is impossible to generalize about the interactions of soil fertility 

with O3. While this is especially true of the interactions involving soil nitrogen, for which there 

is much conflicting evidence, the interactions with other nutrients need much more thorough 

investigation than has occurred to date, before any clear patterns become apparent. 

AX9.3.6 Interactions with Other Pollutants 

The ambient air may have pollutant gases other than O3 and its photochemical oxidant 

relatives. In particular, industrial, domestic, and automobile emissions and accidents can lead to 

significant atmospheric concentrations of gases such as sulfur dioxide (SO2) and nitric oxide 

(NO) and nitrogen dioxide (NO2), collectively referred to as NOx, both locally and regionally. 

Local releases of gases such as hydrogen fluoride (HF), hydrogen chloride (HCl), and 

chlorine (Cl2) may result from industrial emissions and accidents. Agricultural fertilizer and 

manure usage can lead to significant increases in ambient ammonia (NH3) and ammonium 

sulfate ((NH4)2SO4). The sulfur and nitrogen oxides may undergo reactions in the atmosphere 

leading to the formation of sulphate (SO4

2!

) and nitrate (NO3 

!

) ions and resultant acid deposition. 

The 1996 O3 AQCD (U.S. Environmental Protection Agency, 1996) discounted much of 

the early research on pollutant combinations, because of its lack of resemblance to the ambient 

experience: the concentrations used were unrealistically high or the exposure regimes employed 

almost invariably used gas mixtures, whereas Lefohn et al. (1987) showed that the co-occurrence 

patterns of significant levels of O3 with SO2 or NO2 in the United States were most frequently 

sequential or partially sequential with overlap; only rarely were they entirely concurrent. On the 

other hand, O3 and peroxyacetylnitrate (PAN) frequently co-occur, as both form photochemically 

under similar conditions. 

To the list of reviews mentioned in the 1996 O3 AQCD should be added the more recent 

ones by Barnes and Wellburn (1998), Robinson et al. (1998), and Fangmeier et al. (2002), which 

also explore some of the potential mechanisms underlying pollutant-pollutant interactions. 

AX9.3.6.1 Oxidant Mixtures 

In 1998, Barnes and Wellburn noted that virtually no information existed on the effects on 

plants of concurrent exposures to O3 and other components of photochemical oxidant other than 

PAN. The situation has not changed since their review appeared, and the topic appears to have 

AX9-121
attracted no research interest since before the 1996 O3 AQCD (U.S. Environmental Protection 

Agency, 1996). The continuing conclusion must, therefore, be that, from the limited information 

available, the two gases appear to act antagonistically, with O3 raising the threshold for the 

visible injury response to PAN and PAN reducing the harmful effects of O3. 

AX9.3.6.2 Sulfur Dioxide 

In reviewing O3 × SO2 interactions, Barnes and Wellburn (1998) remarked: “The outcome 

of exposure to this combination of pollutants has probably been the most studied, yet is one of 

the least understood.” More recent studies have only added to the conflicts referred to in the 

1996 criteria document (U.S. Environmental Protection Agency, 1996), rather than resolve them. 

For example, Diaz et al. (1996) reported that, after a year of daily exposures of Aleppo pine 

seedlings to 50 ppb O3 and/or 40 ppb SO2, the combination of pollutants synergistically reduced 

shoot and root growth and impaired mycorrhizal colonization of the roots. With tomato 

(Lycopersicon esculentum), on the other hand, effects on growth ranged from synergistic at low 

exposures (50 ppb) to antagonistic at exposures of 200 ppb of each gas (Khan and Khan, 1994). 

Although various physiological measurements were made in these and earlier studies, it has not 

been possible to determine any consistent mechanism or mechanisms that might account for the 

conflicting results. 

Since the information available about O3 × SO2 interactions appears to be highly dependent 

upon species, the type of response measured, and the experimental protocol used, it would still 

appear prudent to heed the statement of Heagle et al. (1988) in their summary of the studies 

undertaken in 12 field experiments over several years within the NCLAN program: “There were 

no cases where O3 and SO2 interactions significantly affected yield.” (emphasis added.) 

AX9.3.6.3 Nitrogen Oxides, Nitric Acid Vapor, and Ammonia 

The major oxides of nitrogen that occur in ambient air are nitrous oxide (N2O), NO, 

and NO2, of which the latter two (conveniently symbolized as NOx) are particularly important in 

connection with O3, because they are components of the reaction mix that leads to photochemical 

O3 formation and because they can interact with O3-responses. Their reactions in the atmosphere 

can also lead to the occurrence of nitric acid vapor (HNO3) in ambient air. The other major 

AX9-122
N-containing contaminant of ambient air in many parts of the world is NH3, largely released 

through agricultural practices. 

Despite various combinations of O3 and NOx being probably the most common air 

pollutant combinations found in the field, Barnes and Wellburn (1998) noted that they have been 

little studied. Much early work with O3 and NOx focused on O3 × NO2 interactions and can be 

discounted, because of the unrealistic concentrations employed and their use as mixtures rather 

than in types of sequences. The 1996 O3 AQCD (U.S. Environmental Protection Agency, 1996) 

concluded that evidence from studies involving concurrent exposures to both O3 and NOx at 

realistic concentrations was so fragmented and varied that no firm conclusions could be drawn as 

to the likelihood and nature such interactions. However, the few recent investigations taken 

together with the earlier data are now beginning to reveal a pattern of response. 

With regard to NO, Nussbaum et al. (1995a, 2000b) reported their findings with concurrent 

exposures to NO and O3 and observed that, at low O3 levels, NO tended to act similarly to O3 by 

increasing the scale of responses such as growth reductions. However, in ambient air in 

which O3 is a dominant factor, the effects of NO were usually found to be negligible due to 

low levels, although the authors admitted that the effects observed were confounded by the 

inevitable O3-induced oxidation of NO to NO2. 

Two possible mechanisms whereby NO may influence plant response to O3 are suggested 

by recent biochemical studies. First, there is growing evidence for the role of NO as a signaling 

agent in plants that can induce defense responses to a range of biotic and abiotic stressors 

(Beligni and Lamattina, 2001; Neill et al., 2002). Second, a role for NO as an antioxidant 

scavenger of reactive oxygen species has been demonstrated by Beligni and Lamattina (2002) in 

potato leaves and chloroplasts. However, both of these cases concern endogenously synthesized 

NO, and it must be noted that in none of these or other reports of studies of NO signaling have 

the authors considered the potential significance of exogenous NO in ambient air. 

An independent case for O3 × NO interactions comes from Mills et al. (2000). The ANN 

model developed to predict the O3 effects on white clover biomass based on experiments at 

18 locations throughout Europe suggested that the minimum daily NO concentration (at 5 p.m.) 

may have been a contributor to adverse effects. 

Turning to NO2, Maggs and Ashmore (1998) found that, although concurrent but 

intermittent exposures of Bismati rice (Oryza sativa) revealed no significant growth 

AX9-123
interactions, NO2 reduced the rate of O3-induced senescence, an antagonistic response possibly 

related to enhanced N-metabolism. 

With regard to sequential exposures, two studies on gene activation in tobacco revealed 

that NO2 counteracted the effect of O3 in reducing mRNA levels for three genes encoding 

photosynthetic proteins (Bahl and Kahl, 1995) and tended to counteract the O3-induced 

enhancement of defense-protein gene activation (Bahl et al., 1995). However, despite 

compelling evidence for significant interactive effects provided by earlier studies (Bender et al., 

1991; Goodyear and Ormrod, 1988; Runeckles and Palmer, 1987), the only recent investigation 

of growth effects seems to have been that of Mazarura (1997) using sine-wave exposure profiles. 

He found that although 4 weeks of twice daily 3-h exposures to NO2 (120 ppb peak 

concentrations) slightly stimulated growth of radish (Raphanus sativa) and while daily 6-h 

exposures to O3 (120 ppb peak concentration) did not significantly reduce growth, the daily 

sequence, NO2 - O3 - NO2, led to a 13% drop in dry matter production. 

The combined evidence to date, therefore, suggests that, in leguminous species, the effects 

of these sequences are antagonistic with NO2 tending to reduce (or reverse) the negative effects 

of O3 on growth, while the effects are increased in other species. These conclusions differ from 

those of Barnes and Wellburn (1998) who suggested that sequential exposures tended to result in 

antagonistic effects (largely based on the summary by Bender and Weigel (1992), whereas 

simultaneous exposures were likely to lead to synergistic responses. With disagreements both 

among the data and their interpretation, it is not possible to determine the circumstances under 

which specific interactions of O3 and NO2 may occur, but there is no reason to doubt the validity 

of the individual findings of each study. Far more systematic investigation is needed to clarify 

the situation. 

There appear to have been no studies of O3 interacting with HNO3 in the vapor phase. 

However, in the southern California montane forests (Takemoto et al., 2001), in Sweden (Janson 

and Granat, 1999), and elsewhere, significant amounts of N are deposited in this form because of 

the vapor’s high deposition velocity. As a consequence, although much of it ultimately reaches 

the ground through leaching and leaf fall and enters the soil as NO3 

!

, it may also be used as a N 

source by the foliage itself (Garten and Hanson, 1990; Hanson and Garten, 1992; Norby et al., 

1989). This nutritional role is independent of any contribution that HNO3 vapor may make to 

acidic deposition. Indirect interactions with the effects of O3 through N-deposition of NOx, 

AX9-124
HNO3, and NH3 are related to the interactions of O3 with N as a nutrient, and have recently been 

examined in the review by Takemoto et al. (2001). The 1996 O3 AQCD (U.S. Environmental 

Protection Agency, 1996) stated that the evidence available at that time led to estimates of total 

forest dry deposition, including HNO3, ranging from 5.7 to 19.1 kg N ha!1

 year!1 (Taylor et al., 

1988a). However, Takemoto et al. (2001) pointed out that in parts of the mid-elevation forests 

of southern California, dry deposition rates may reach more than 40 kg N ha!1

 year!1. As a 

result, some locations have seen the conversion from N-limited to N-saturated forests. The 

concern for California’s forests is well stated by Takemoto et al.: “As potential modifiers of 

long-term forest health, O3 is a stressor and N deposition is an enhancer of ponderosa/Jeffrey 

pine physiology and growth (Grulke and Balduman, 1999). The progression toward a deciduous 

growth habit, higher shoot:root biomass ratios, increasing depths of litter, tree densification, and 

elevated NO3 

!

 levels in soil and soil solution, all point to the replacement of pine species with 

nitrophilous, shade- and O3-tolerant tree species, such as fir and cedar (Minnich, 1999; Minnich 

et al., 1995).” 

Few studies have been reported of interactions of O3 with NH3. The 1996 criteria 

document made reference to the work on kidney bean by Tonneijck and Van Dijk (1994, 1998). 

Although NH3 alone tended to increase growth and O3 alone to inhibit it, one interaction was 

noted (Tonneijck and Van Dijk, 1994) on the number of injured leaves. Dueck et al. (1998) 

studied the effects of O3 and NH3 on the growth and drought resistance of Scots pine. 

Significant interactions were found for some growth features, but there were no consistent 

patterns of the effects of NH3 on O3 response or vice versa. However, O3 was found to 

ameliorate the enhancement of drought stress caused by NH3 on Scots pine. 

At this time there is insufficient information to offer any general conclusions about the 

interactive effects of O3 and NH3. 

AX9.3.6.4 Hydrogen Fluoride and Other Gaseous Pollutants 

Although HF and other fluorides are important local air pollutants associated with 

aluminum smelting and superphosphate fertilizer manufacture, no studies of possible interactions 

with oxidants appear to have been reported since that of MacLean (1990). He found that HF 

retarded the accelerated senescence and loss of chlorophyll resulting from O3 exposure in corn 

seedlings. However, such an isolated observation cannot be taken to indicate that HF can reduce 

AX9-125
the impact of O3 on other species or even that the effect would ultimately have led to an effect on 

mature plants. 

AX9.3.6.5 Acidic Deposition 

The deposition of acidic species onto vegetation may elicit direct effects on the foliage or 

indirect effects via changes induced in the soil. The 1996 O3 AQCD (U.S. Environmental 

Protection Agency, 1996) included an extensive listing of investigations into the effects of O3 

and acidic deposition (usually in the form of simulated acid rain, SAR) on plant growth and 

physiology. The majority of studies found no effects of SAR or acidic mists or fogs at pH values 

greater than about 3.0 and no interactive effects with O3. (In ambient air, pH values less than 

3.0 have rarely been reported.) In the few reports in which significant interactions were found, 

most were antagonistic and were explained as probably being the result of increased fertility due 

to NO3 

!

 and SO4

2- supplied in the SAR. 

Although numerous reviews have recently appeared (e.g., Bussotti and Ferretti, 1998; 

Flückiger et al., 2002; Fox and Mickler, 1996; Nussbaum et al., 1999; and Sheppard and Cape, 

1999), the shift in interest in air pollution effects away from acid deposition has resulted in little 

new research having been reported over the past 10 or so years. In most of the reported studies, 

no effects due to the O3 exposures, the SAR treatments used, or their combinations were 

observed, e.g., Baker et al. (1994) on loblolly pine; Laurence et al. (1997), and Vann et al. 

(1995) on red spruce; and Laurence et al. (1996) on sugar maple. Branch chamber studies of 

12-year-old Ponderosa pine trees by Momen et al. (1997, 1999) revealed no O3 effects or 

interactions. With red spruce, Sayre and Fahey (1999) noted no effects of O3 on the foliar 

leaching of Ca or Mg, which only became significant with SAR at pH 3.1. Izuta (1998) 

observed no interactions with Nikko fir (Abies homolepis), although SAR at ph 4.0 reduced dry 

matter. Shan et al. (1996) reported adverse effects of O3 but none attributable to SAR on the 

growth of Pinus armandi. 

With herbaceous species, Ashenden et al. (1995) noted significant antagonistic interactions 

of O3 and acid mist in white clover, in which the adverse effect of low pH was countered by O3. 

In contrast, Ashenden et al. (1996) found that, although pH 2.5 mist caused a significant 

stimulation of the growth of ryegrass attributed to a fertilizer effect, and O3 caused reduced 

growth, there was no interaction. Bentgrass (Agrostis capillaris) behaved similarly. 

AX9-126
A study by Bosley et al. (1998) on the germination of spores of the moss, Polytrichum 

commune, and the ferns, Athyrium felix-femina and Onoclea sensibilis, revealed no effect of O3 

on moss spores, while SAR at pH <4.0 was completely inhibitory. With the ferns, germination 

was progressively reduced by both increased O3 and acidity. 

In summary, the few findings of interactions in these recent studies are consistent with the 

previous conclusion regarding the likelihood of such interactions being antagonistic. However, 

the interactions observed were in each case largely the result of the response to the lowest pH 

used, which, in several cases, was below 3.0, and hence may not be relevant to most field 

conditions. 

AX9.3.6.6 Heavy Metals 

As there appears to have been no further research into the interactions of oxidants with 

heavy metal pollutants, our understanding is unchanged from at the time of the 1996 O3 AQCD 

(U.S. Environmental Protection Agency, 1996). As noted therein, the limited data available 

from early studies indicates varying degrees of enhancement of any adverse effects of O3 but 

precludes the development of any response relationships. 

AX9.3.6.7 Mixtures of Ozone with Two or More Pollutants 

In many airsheds, the mixtures that occur, both concurrently and over time, may involve 

three or more pollutants. Very little useful information exists on the effects of O3 with multiple 

pollutants. As the 1996 criteria document and others have pointed out, most of the early studies 

on such combinations can be discounted, because of their use of (1) high and environmentally 

irrelevant exposure concentrations and (2) unrealistic, repetitive exposure profiles (Barnes and 

Wellburn, 1998; U.S. Environmental Protection Agency, 1996). 

The large investment in experimental facilities required to study these complex interactions 

is a major deterrent. So, although the topic has been included in several reviews that have 

appeared in the last decade, there appear to have been only two studies that have provided new 

information on the effects of O3 in combination with more than one other pollutant stress. 

Ashenden et al. (1996) studied the effects of O3 and/or (SO2 + NO2) with four acidities of SAR 

applied to each gas treatment, on white clover and two pasture grasses (Lolium perenne and 

Agrostis capillaris). With each species, the antagonism reported for the O3 × SAR interaction 

AX9-127
(Section AX9.3.4.6.5) tended to be nullified by concurrent exposure to the other gases, while the 

combination of the three gaseous pollutants resulted in the most severe growth inhibition, 

regardless of the acidity of the SAR. 

With such meager evidence, no clear conclusions can be drawn as to the ways in which the 

effects of multiple airborne stressors could influence or be influenced by O3. 

AX9.3.7 Interactions with Agricultural Chemicals 

The review of interactions involving O3, plants, and various agricultural chemicals 

presented in the 1996 O3 AQCD (U.S. Environmental Protection Agency, 1996) remains a valid 

assessment of our limited knowledge of these interrelationships. Our knowledge is largely based 

on the protection against O3 afforded to a range of crop species by applications of various 

chemicals, particularly fungicides, such as benomyl (benlate; methyl-1-[butylcarbamoyl]-2-

benzimidazolecarbamate) and several carbamates and triazoles. A recent report has added 

azoxystrobin (AZO) and epoxyconazole (EPO) to the list (Wu and Tiedemann, 2002). Foliar 

sprays of either AZO or EPO provided 50 to 60% protection against O3 injury to barley 

(Hordeum vulgare) leaves. Both had similar modes of action involving stimulation of the levels 

of antioxidant enzymes such as SOD, ascorbate peroxidase, guaiacol peroxidase, and catalase. 

In contrast, applications of herbicides have yielded variable results ranging from increased 

sensitivity to protection from O3; the nature of the effect is usually species- or cultivardependent. Although of less wide application, some plant growth retardants have also been 

found to provide protection, but no insecticide appears to have been clearly shown to have 

similar properties. 

Despite the attraction of the use of permitted chemicals to provide crop protection, the 

statement in the 1996 O3 AQCD (U.S. Environmental Protection Agency, 1996) is still valid: 

“It is premature to recommend their use specifically for protecting crops from the adverse effects 

of O3, rather than for their primary purpose.” 

AX9.3.8 Factors Associated with Global Climate Change 

During the last decade, interest in the effects of climatic change on vegetation has replaced 

concerns over the purported causes of forest decline and the effects of acidic deposition. Two 

AX9-128
specific components of climate change have been singled out as the foci of most of the research 

activity: 

C the effects of increasing mean global CO2 concentrations in the lower atmosphere, and 

C the effects of increasing levels of surface-level irradiation by UV-B (the result of

stratospheric O3 depletion). 

In spite of the crucial role of temperature as a climatic determinant (Monteith and Elston, 1993), 

the effects of increasing mean global temperatures and their interactions with increasing CO2 

levels in particular have received less attention. 

All of the biotic and chemical interactions with oxidants discussed in the preceding 

sections may be modified by these climatic changes. However, research activities have largely 

focused on the two-way O3 × CO2 interaction. Little if any experimental evidence exists related 

to three-way interactions, such as O3 × CO2 × disease or O3 × CO2 × nutrient availability, 

although such interactions are difficult to predict from the component two-way interactions. 

Numerous reviews have appeared since the 1996 O3 AQCD (U.S. Environmental 

Protection Agency, 1996) dealing with the issues involved. General reviews include 

publications of IPCC (1996, 2001); and UNEP (1993, 1999); the volume by Wellburn (1994); 

the volumes edited by Alscher and Wellburn (1994), De Kok and Stulen (1998), Singh (2000), 

and Yunus and Iqbal (1996); and papers by Idso and Idso (1994), Krupa and Groth (2000), Luo 

et al. (1999), Polle and Pell (1999), Poorter and Pérez-Soba (2001), Runeckles (2002), and 

Weubbles et al. (1999). Effects on agriculture and crop production, growth, and metabolism 

have been reviewed by Groth and Krupa (2000), Rötter and Van De Geijn (1999), and Schnug 

(1998); effects on forests have been reviewed by Bortier et al. (2000a); with focus on insect 

pests, Docherty et al. (1997), Karnosky et al. (2001a,c), McLaughlin and Percy (1999), and Saxe 

et al. (1998). 

As background to the discussion of interactions with O3, it should be noted that the 

increased levels of CO2 experienced since the mid-18th century are such that, without abatement 

of the rates of increase, increased levels of from 540 to 970 ppm have been projected by the year 

2100 (IPCC, 2001). Such increases in the concentration of CO2, the principal GHG released into 

the atmosphere, will inevitably lead to increased global mean temperatures, evidence for which 

is already available from oceanic, icepack, and other records. The latest estimates of the global 

warming are for an increase in the range of 1.4 to 5.8 °C over this century, in contrast to the 

AX9-129
0.6 °C rise experienced since 1900 (IPCC, 2001). However, considerable uncertainty is 

associated with such projections of future increases in global temperature. 

The use of elevated CO2 concentrations has been common practice for many years in the 

production of many greenhouse crops. Much of our early knowledge of the effects of higher 

than ambient CO2 levels on plant growth derives from this application, coupled with research of 

plant physiologists on how CO2 concentrations affect the process of photosynthesis. Information 

available about the effects of increased CO2 levels on photosynthesis and stomatal function, in 

particular, has provided the underlying bases for numerous process models that simulate plant 

growth under stress and in changed climates. 

Although simple O3 × temperature interactions were discussed in Section AX9.3.4.2, the 

close linkage between global CO2 levels and global mean temperatures in the context of climate 

change requires that an assessment of the interactive effects with O3 should focus, as much as 

possible, on interactions involving all three factors. 

AX9.3.8.1 Ozone-Carbon Dioxide-Temperature Interactions 

Idso and Idso (1994) reviewed several hundred reports published between 1982 and 1994 

on the effects of increased CO2 on plant growth and net photosynthesis. Their survey covered a 

wide range of temperate and tropical, herbaceous and perennial species, including coniferous 

trees. They concluded that, for responses to a 300-ppm increase in CO2, somewhat less than a 

doubling of present-day levels, but somewhat greater than the 540 ppm lower limit suggested by 

the IPCC (IPCC, 2001), averaged across all species: 

C light intensity had a negligible effect on net photosynthesis other than at limiting low 

intensities under which the CO2-driven enhancement was increased; 

C increased temperature tended to increase the CO2-driven enhancement of dry matter

accumulation (growth) and net photosynthesis;

C drought conditions tended to increase the CO2-driven enhancements of both growth and 

net photosynthesis, but increased salinity had little effect; 

C mineral nutrient deficiency (especially of nitrogen) tended to increase the CO2-driven 

enhancement of growth; and 

C in the presence of air pollutants (especially SO2 and NOx), the CO2-driven enhancement 

of net photosynthesis tended to be increased. 

AX9-130
It should be noted that the statement that CO2-enhanced growth increased with temperature 

referred to total dry matter accumulation by the whole plant and not to the yield of grain, fruit, or 

seed. Unfortunately, despite the existence of several reports at the time, the summary of 

interactions with air pollutants contained only a single reference to O3, i.e., Pfirrmann and 

Barnes (1993), who reported surprisingly that a doubling of CO2 levels led to a 27% increase in 

dry weight of radish but that the combination with O3 led to a 77% increase. 

The more recent reviews by Rudorff et al. (2000) and Olszyk et al. (2000) have 

addressed CO2 × O3 interactions in detail, with the latter focusing on the implications for 

ecosystems. They concluded that: 

C the effects of both gases on stomatal closure were predominantly additive, with little

evidence of interaction;

C increased photosynthesis resulting from elevated CO2 may be canceled by exposures to 

high O3 levels; 

C foliar O3 injury is reduced by elevated CO2; and 

C interactions between CO2 and O3 can affect storage carbohydrates, leaf free-radical

metabolism, and carbon allocation to shoots and roots. 

Olszyk et al. (2000) also made specific note of the relative lack of information on below-ground 

effects. 

Much of the recently published information on the effects of increased CO2 and O3 levels 

is summarized in Table AX9-12. Note that the table only lists the directions of O3-induced 

effects and any modifications of these effects resulting from elevated CO2, not their magnitudes. 

These directions are usually, but not necessarily, the same as the corollary effects of O3 

on CO2-induced responses. 

The bulk of the available evidence clearly shows that, under the various experimental 

conditions used (which almost exclusively employed abrupt or “step” increases in CO2 

concentration, as discussed below), increased CO2 levels may protect plants from the adverse 

effects of O3 on growth. This protection may be afforded in part by CO2 acting together with O3 

in inducing stomatal closure, thereby reducing O3 uptake, and in part by CO2 reducing the 

negative effects of O3 on Rubisco and its activity in CO2-fixation. Although both CO2-induced 

and O3-induced decreases in stomatal conductance have been observed primarily in short-term 

AX9-131
Table AX9-12. Effects of Increased Carbon Dioxide on Ozone-Induced Responses of Plants at th

Physiological, and Whole-Plant Levels 

AX9-132

O3 Effects: w, Decrease; v, Increase; F, No Significant Effect. 

Co2 Effects: •, Additive/Synergistic; –, Antagonistic/Ameliorative; F, No Significant Effect. 

Plant Response O3 Responsea CO2 Modificationb Species Facilityc Reference 

Biochemical/Metabolic Ascorbate peroxidase w –

w F w F 

Catalase w F 

v F 

Wheat (T. aestivum) CSTR, P Rao et al. (1995)Sugar maple (A. saccharum) CEC, P Niewiadomska et Trembling aspen (P. FACE, G Wustman et al. (2tremuloides) Wheat (T. aestivum) CEC, P McKee et al. (199

CEC, P Niewiadomska et 

w F 

Chlorophyll w –

w	 –

Soybean (G. max) OTC, P Booker et al. (199Wheat (T. aestivum) OTC, G Donnelly et al. (2

(1999) 

Potato (S. tuberosum) OTC, G Donnelly et al. (2

w	 –

Glutathione reductase w –

F F F F 

Soybean (G. max) Heagle et al. (199

(1998); Reid et al

Wheat (T. aestivum) CSTR, P Rao et al. (1995) Sugar maple (A. saccharum) CEC, P Niewiadomska et Aspen (P. tremuloides) FACE, G Wustman et al. (2

Glycolate oxidase w – Soybean (G. max) OTC, P Booker et al. (199Hydroxypyruvate w F reductase 

Rubisco w –

w –

Soybean (G. max) OTC, P Booker et al. (199

Soybean (G. max) OTC, P Reid et al. (1998) 

Wheat (T. aestivum) OTC, G McKee et al. (200

[ v ] – Trembling aspen FACE, G Noormets et al. (2

(P. tremuloides) 

w – Sugar maple (A. saccharum) CEC, P Gaucher et al. (20
Table AX9-12 (cont’d). Effects of Increased Carbon Dioxide on Ozone-Induced Responses of Pl

Metabolic, Physiological, and Whole-Plant Levels 

AX9-133

O3 Effects: w, Decrease; v, Increase; F, No Significant Effect. 

Co2 Effects: •, Additive/Synergistic; –, Antagonistic/Ameliorative; F, No Significant Effect. 

Plant Response O3 Responsea CO2 Modificationb Species Facilityc Reference 

Biochemical/Metabolic (cont’d) Rubisco activity w –

w –w [–] w –

Superoxide dismutase F F 

F F 

Soybean (G. max) OTC, P Reid et al. (1998)

Wheat (T. aestivum) CEC, P McKee et al. (199

Wheat (T. aestivum) OTC, G McKee et al. (200European beech (F. sylvatica) CEC, P Lütz et al. (2000)

Wheat (T. aestivum) CEC, P McKee et al. (199

Sugar maple (A. saccharum) CEC, P Niewiadomska et 

Physiological Stomatal conductance w • Radish (R. sativus) CEC, P Barnes and Pfirrm

w – Soybean (G. max) OTC, G Mulchi et al. (199

w • Soybean (G. max) OTC, P,G Booker et al. (200

w – Bean (P. vulgaris) OTC, P Heagle et al. (200w [•] White clover (T. repens) (O3- CSTR, P Heagle et al. (199

sensitive) 

v - F – - F White clover (T. repens) CSTR, P Heagle et al. (199

(O3-tolerant) 

w • Tomato (L. esculentum) CEC, P Hao et al. (2000) 
Table AX9-12 (cont’d). Effects of Increased Carbon Dioxide on Ozone-Induced Responses of Plants a

Physiological, and Whole-Plant Levels 

AX9-134

O3 Effects: 

w, Decrease; 

v, Increase; 

F, No Significant Effect. 

Co2 Effects: 

•, Additive/Synergistic; 

–, Antagonistic/Ameliorative; 

F, No Significant Effect. 

Plant Response 

O3 Responsea CO2 Modificationb Species Facilityc Reference 

Physiological (cont’d) Stomatal conductance 

w • Potato (S, tuberosum) OTC, G Finnan et al. (200

(cont’d) 

w • Wheat (T. aestivum) CEC, P Balaguer et al. (1

Barnes et al. (199

[ w ] 

• CEC, P McKee et al. (199

w F OTC, G Mulholland et al. w F -

• CEC, P Donnelly et al. (1

v F CEC, P Tiedemann and Fw F Agropyron smithii CEC, P Volin et al. (1998w F Koeleria cristata CEC, P Volin et al. (1998

F F Bouteloua curtipendula CEC, P Volin et al. (1998F F Schizachyrium scoparium CEC, P Volin et al. (1998F F Black cherry (P. serotina) CSTR, P Loats and RebbecF F Green ash (F. pennsylvanica) CSTR, P Loats and Rebbec

[ v ] 

– Yellow poplar (L. tulipifera) CSTR, P Loats and Rebbec

w • Trembling aspen CEC, P Volin et al. (1998

(P. tremuloides) 
Table AX9-12 (cont’d). Effects of Increased Carbon Dioxide on Ozone-Induced Responses of Plants a

Physiological, and Whole-Plant Levels 

AX9-135

O3 Effects: 

w, Decrease; 

v, Increase; 

F, No Significant Effect. 

Co2 Effects: 

•, Additive/Synergistic; 

–, Antagonistic/Ameliorative; 

F, No Significant Effect. 

Plant Response 

O3 Responsea CO2 Modificationb Species Facilityc Reference 

Physiological (cont’d) Stomatal conductance 

F -

w – FACE, G Noormets et al. (2

(cont’d) 

F F Red oak (Q. rubra) CEC, P Volin et al. (1998

[ v ] 

– Durmast oak (Q. petraea) CEC, P Broadmeadow et 

Photosynthesis 

w – Radish (R. sativus) CEC, P Barnes and Pfirrmw – Soybean (G. max) OTC, P Booker et al. (199w – Soybean (G. max) OTC, G,P Mulchi et al. (199w – Bean (P. vulgaris) OTC, P Heagle et al. (200w F Cotton (G. hirsutum) OTC, P Heagle et al. (199w F Tomato (L. esculentum) CEC, P Hao et al. (2000) 

F F Potato (S. tuberosum) OTC, G Donnelly et al. (2F [–] OTC, G Lawson et al. (20

[ w ] 

– Wheat (T. aestivum) CEC, P Barnes et al. (199

w – Wheat (T. aestivum) OTC, G Donnelly et al. (2

Mulholland et al. Reid and Fiscus (

CEC, P Tiedemann and F

w F CEC, P Cardoso-Vilhena 
Table AX9-12 (cont’d). Effects of Increased Carbon Dioxide on Ozone-Induced Responses of Plants a

Physiological, and Whole-Plant Levels 

AX9-136

Co2 Effects: 

O3 Effects: 

w, Decrease; 

v, Increase; 

F, No Significant Effect. 

•, Additive/Synergistic; 

–, Antagonistic/Ameliorative; 

F, No Significant Effect. 

Plant Response 

O3 Responsea CO2 Modificationb Species Facilityc Reference 

Physiological (cont’d) Photosynthesis (cont.) F F Agropyron smithii CEC, P Volin et al. (1998

w – Koeleria cristata CEC, P Volin et al. (1998

F -

w F Bouteloua curtipendula CEC, P Volin et al. (1998

[ w ] 

– Schizachyrium scoparium CEC, P Volin et al. (1998

w – Ponderosa pine 

(P. ponderosa) 

CEC, G Olszyk et al. (200

w – Scots pine (P. sylvestris) OTC, G Kellomäki and W

F F Black cherry (P. serotina) CSTR, P Loats and Rebbec

[ w ] 

– Green ash (F. pennsylvanica) CSTR, P Loats and Rebbec

F F Yellow poplar (L. tulipifera) CSTR, P Loats and Rebbec

w – Trembling aspen (P. 

tremuloides) 

CEC, P Volin et al. (1998

w – European beech (F. sylvatica) CEC, P Grams et al. (199

F F Red oak (Q. rubra) CEC, P Volin et al. (1998

w – Sugar maple (A. saccharum) CEC, P Gaucher et al. (20
Table AX9-12 (cont’d). Effects of Increased Carbon Dioxide on Ozone-Induced Responses of Plants a

Physiological, and Whole-Plant Levels 

AX9-137

O3 Effects: 

w, Decrease; 

v, Increase; 

F, No Significant Effect. 

Co2 Effects: 

•, Additive/Synergistic; 

–, Antagonistic/Ameliorative; 

F, No Significant Effect. 

Plant Response 

O3 Responsea CO2 Modificationb Species Facilityc Reference 

Physiological (cont’d) Photorespiration 

v – Soybean (G. max) OTC, P Booker et al. (199w • Wheat (T. aestivum) CEC, P McKee et al. (199

Growth, Yield Total biomass 

w – Parsley (P. sativum) CEC, P Cardoso-Vilhena w – Bean (P. vulgaris) CEC, P Cardoso-Vilhena w – OTC, P Heagle et al. (200w – Soybean (G. max) 

O

T

C, G Mulchi et al. (199

CSTR, P Reinert et al. (199OTC, P,G Booker et al. (200OTC, P Booker et al. (200

[ w ] F Alfalfa (M. sativa) CEC, P Johnson et al. (19

w – White clover (T. repens) CSTR, P Heagle et al. (199

(O3-sensitive) 

F F White clover (T. repens) CSTR, P Heagle et al. (199

(O3-tolerant) 

w – Tomato (L. esculentum) CEC, P Hao et al. (2000) 

CSTR, P Reinert and Ho (1

F F Potato (S. tuberosum) OTC, G Donnelly et al. (2

Persson et al. (20
Table AX9-12 (cont’d). Effects of Increased Carbon Dioxide on Ozone-Induced Responses of Plants a

Physiological, and Whole-Plant Levels 

AX9-138

Co2 Effects: 

O3 Effects: 

w, Decrease; 

v, Increase; 

F, No Significant Effect. 

•, Additive/Synergistic; 

–, Antagonistic/Ameliorative; 

F, No Significant Effect. 

Plant Response 

O3 Responsea CO2 Modificationb Species Facilityc Reference 

Growth, Yield (cont’d) Total biomass (cont’d) 

w F OTC, G Lawson et al. (20w – Mustard (S. alba) CEC, P Cardoso-Vilhena w – Plantain (P. major) CEC, P Cardoso-Vilhena w – Cotton (G. hirsutum) OTC, P Booker (2000) 

Heagle et al. (199

w – Wheat (T. aestivum) 

C

E

C, P 

OTC, G OTC, P CEC, P OTC, G CSTR, P OTC, G 

Cardoso-Vilhena Fangmeier et al. (Heagle et al. (200McKee et al. (199Pleijel et al. (200Rao et al. (1995) Rudorff et al. (19

[ w ] 

– Wheat (T. aestivum) OTC, G Bender et al. (199

Mulholland et al. 

w F Wheat (T. aestivum) CEC, P Cardoso-Vilhena 

Tiedemann and F

w [–] Wheat (T. aestivum) OTC, G Ewart and Pleijel 

[ w ] 

– Timothy (P. pratense) CEC, P Johnson et al. (19

w – Agropyron smithii CEC, P Volin et al. (1998w – Koeleria cristata CEC, P Volin et al. (1998
Table AX9-12 (cont’d). Effects of Increased Carbon Dioxide on Ozone-Induced Responses of Plants a

Physiological, and Whole-Plant Levels 

AX9-139

O3 Effects: 

w, Decrease; 

v, Increase; 

F, No Significant Effect. 

Co2 Effects: 

•, Additive/Synergistic; 

–, Antagonistic/Ameliorative; 

F, No Significant Effect. 

Plant Response 

O3 Responsea CO2 Modificationb Species Facilityc Reference 

Growth, Yield (cont’d) Total Biomass 

w – Corn (Z. mays) OTC, G Rudorff et al. (19

(cont’d) 

w F Bouteloua curtipendula CEC, P Volin et al. (1998

F F Schizachyrium scoparium CEC, P Volin et al. (1998

w – Ponderosa pine (P. ponderosa) CEC, G Olszyk et al. (200w • Birch (B. pendula) CEC, P Kytöviita et al. (1

[ w ] F Black cherry (P. serotina) CSTR, P Loats and Rebbec

v – Green ash (F. pennsylvanica) CSTR, P Loats and Rebbec

F F European ash (F. excelsior) OTC, G Broadmeadow an

[ v ] 

• Yellow poplar (L. tulipifera) CSTR, P Loats and Rebbec

w – Sugar maple (A. saccharum) CEC, P Gaucher et al. (20w – Trembling aspen CEC, P Volin et al. (1998

(P. tremuloides) OTC, P Dickson et al. (19(O3-tolerant clone} OTC, G Dickson et al. (20

w F Trembling aspen OTC, G Dickson et al. (20

(P. tremuloides) (O3-sensitive clone) 

F F Red oak (Q. rubra) CEC, P Volin et al. (1998

w – Durmast oak (Q. petraea) CEC, P Broadmeadow et 

OTC, G Broadmeadow an
Table AX9-12 (cont’d). Effects of Increased Carbon Dioxide on Ozone-Induced Responses of Plants a

Physiological, and Whole-Plant Levels 

AX9-140

O3 Effects: 

w, Decrease; 

v, Increase; 

F, No Significant Effect. 

Co2 Effects: 

•, Additive/Synergistic; 

–, Antagonistic/Ameliorative; 

F, No Significant Effect. 

Plant Response 

O3 Responsea CO2 Modificationb Species Facilityc Reference 

Growth, Yield (cont’d) Total Biomass F F Aleppo pine (P. halepensis) CEC, P Kytöviita et al. (1(cont’d) 

F F Scots pine (P. sylvestris) OTC, G Broadmeadow an

Seed/grain/fruit/tuber 

w – Soybean (G. max) 

O

T

C, P Fiscus et al. (199

yield OTC, G Mulchi et al. (199

OTC, G Mulchi et al. (199OTC, P,G Booker et al. (200

w – Bean (P. vulgaris) OTC, P Heagle et al. (200w – Tomato (L. esculentum) CSTR, P Reinert and Ho (1w – Potato (S, tuberosum) OTC, G Finnan et al. (200

F F OTC, G Persson et al. (20

w – Wheat (T. aestivum) 

O

T

C, G Bender et al. (199

OTC, G Fangmeier et al. (CEC, P McKee et al. (199OTC, G Mulchi et al. (199OTC, G Mulholland et al. OTC, G Rudorff et al. (19

[ w ] 

– Wheat (T. aestivum) (cont.) OTC, G Fangmeier et al. (

w [–] OTC, G Mulholland et al. 

Mulholland et al. 
Table AX9-12 (cont’d). Effects of Increased Carbon Dioxide on Ozone-Induced Responses of Plants a

Physiological, and Whole-Plant Levels 

AX9-141

O3 Effects: 

w, Decrease; 

v, Increase; 

F, No Significant Effect. 

Co2 Effects: 

•, Additive/Synergistic; 

–, Antagonistic/Ameliorative; 

F, No Significant Effect. 

Plant Response 

O3 Responsea CO2 Modificationb Species Facilityc Reference 

Growth, Yield (cont’d) Relative growth rate 

w – Wheat (T. aestivum) CEC, P Barnes et al. (199w – CEC, P Cardoso-Vilhena w – Agropyron smithii CEC, P Volin et al. (1998w – Koeleria cristata CEC, P Volin et al. (1998

F F Bouteloua curtipendula CEC, P Volin et al. (1998F F Schizachyrium scoparium CEC, P Volin et al. (1998F F European ash (F. excelsior) OTC, G Broadmeadow an

w – Trembling aspen (P. CEC, P Volin et al. (1998

tremuloides) 

[ w ] 

– Red oak (Q. rubra) CEC, P Volin et al. (1998

w – Durmast oak (Q. petraea) OTC, G Broadmeadow an

F F Scots pine (P. sylvestris) OTC, G Broadmeadow an

Specific leaf area-SLA 

w F Radish (R. sativus) CEC, P Barnes and Pfirrmw – Soybean (G. max) OTC, P Reid et al. (1998) 

w – Cotton (G. hirsutum) OTC, P Booker (2000) 

w F OTC, G Mulchi et al. (199
Table AX9-12 (cont’d). Effects of Increased Carbon Dioxide on Ozone-Induced Responses of Plants a

Physiological, and Whole-Plant Levels 

AX9-142

O3 Effects: 

w, Decrease; 

v, Increase; 

F, No Significant Effect. 

Co2 Effects: 

•, Additive/Synergistic; 

–, Antagonistic/Ameliorative; 

F, No Significant Effect. 

Plant Response 

O3 Responsea CO2 Modificationb Species Facilityc Reference 

Growth, Yield (cont’d) Specific leaf area-SLA 

w • White clover (T. repens) CSTR, P Heagle et al. (199

(cont’d) (O3-sensitive) 

v • White clover (T. repens) CSTR, P Heagle et al. (199

(O3-tolerant) 

v • Agropyron smithii CEC, P Volin et al. (1998v • Koeleria cristata CEC, P Volin et al. (1998

F F Bouteloua curtipendula CEC, P Volin et al. (1998

v F Schizachyrium scoparium CEC, P Volin et al. (1998v • Trembling aspen CEC, P Volin et al. (1998

(P. tremuloides) 

v • Red oak (Q. rubra) CEC, P Volin et al. (1998

Root/shoot ratio 

w – Radish (R. sativus) CEC, P Barnes and Pfirrmw • Alfalfa (M. sativa) CEC, P Johnson et al. (19w – White clover (T. repens) CSTR, P Heagle et al. (199

(O3-sensitive) 

F F White clover (T. repens) CSTR, P Heagle et al. (199

(O3-tolerant) 

v • Wheat (T. aestivum) CEC, P McKee et al. (199
Table AX9-12 (cont’d). Effects of Increased Carbon Dioxide on Ozone-Induced Responses of Plants a

Physiological, and Whole-Plant Levels 

AX9-143

O3 Effects: 

w, Decrease; 

v, Increase; 

F, No Significant Effect. 

Co2 Effects: 

•, Additive/Synergistic; 

–, Antagonistic/Ameliorative; 

F, No Significant Effect. 

Plant Response 

O3 Responsea CO2 Modificationb Species Facilityc Reference 

Growth, Yield (cont’d) Root/shoot ratio 

w – Timothy (P. pratense) CEC, P Johnson et al. (19

(cont’d) 

w F Soybean (G. max) OTC, G,P Booker et al. (200

F F Black cherry (P. serotina) CSTR, P Loats and RebbecF F Green ash (F. pennsylvanica) CSTR, P Loats and RebbecF F Yellow poplar (L. tulipifera) CSTR, P Loats and RebbecF F Aspen (P. tremuloides) OTC, G Dickson et al. (20

Foliar injury 

v – Potato (S. tuberosum) OTC, G Donnelly et al. (2

Persson et al. (20

v – Bean (P. vulgaris) OTC, P Heagle et al. (200v – Soybean (G. max) OTC, P Heagle et al. (199v – Cotton (G. hirsutum) OTC, P Heagle et al. (199v – Wheat (T. aestivum) 

C

E

C, P Barnes et al. (199

OTC, G Mulholland et al. 

v – Trembling aspen (P. FACE, G Karnosky et al. (1

tremuloides) Wustman et al. (2

v – European beech (F. sylvatica) CEC, P Grams et al. (199

O3 m # 0.15 ppm. 

b CO2-modifications of O3-effects resulting from ~2× present levels. (Trends are shown in brackets. Pronounced changes with ontoge

 indicated thus: F - –.) 

c Exposure facilities used: CEC: controlled environment chambers; CSTR: continuously stirred tank reactors (Heck et al., 1978); FA enrichment facilities; OTC: open-top chambers. G: plants rooted in the ground; P: plants grown in pots. All species are 

C3 except 

 and Schizachyrium. 
studies, recent data clearly show a long-term and sustained reduction in stomatal conductance 

under elevated CO2 for a number of species (Ainsworth and Long, 2005; Ellsworth et al., 2004; 

Gunderson et al., 2002). Instances of increased stomatal conductance have also been observed in 

response to O3 exposure, suggesting partial stomatal dysfunction after extended periods of 

exposure (Maier-Maercker, 1998). 

At the mechanistic level, Rubisco plays a key role in CO2-assimilation, and while both O3 

and elevated CO2 per se can lead to reduced activity, CO2 can also reverse the O3-induced 

inhibition of Rubisco activity and photosynthesis (Table AX9-12). However, in their review of 

the possible mechanisms involved, Polle and Pell (1999) cautioned that Rubisco should not be 

regarded “as a unique target for the interaction of the two gases.” But it is clear from the bulk of 

the evidence in Table AX9-12 that elevated CO2 levels can ameliorate the inhibition of growth 

caused by O3 in many species, although the precise balance among the mechanisms involved 

may well vary from species to species. Three important caveats must be raised with regard to 

the findings presented in Table AX9-12: 

C the applicability of results from experiments with an abrupt (step) increase in CO2 level 

to understanding the consequences of the gradual increase in CO2 predicted for the 

troposphere over the next hundred years; 

C the validity of the findings in several long-term studies (particularly with tree species) 

conducted using potted plants, because of possible added stressors imposed on their root 

systems relative to trees growing in the field; and 

C the relevance to understanding the effects of climate change of studies focused solely on 

CO2 enrichment at current ambient conditions of temperature and precipitation patterns 

that provide no insights into possible interactive effects as these other climatic variables 

change concurrently with increasing CO2 (Intergovernmental Panel on Climate Change 

[IPCC], 2001). 

The first caveat concerns the distinctly different natures of the exposures to O3 and CO2 

experienced by plants in the field. Changes in the ambient concentrations of these gases have 

very different dynamics. In the context of climate change, CO2 levels increase relatively slowly 

and may change little over several seasons of growth. On the other hand, O3 presents a 

fluctuating stressor with considerable hour-to-hour and day-to-day variability (Polle and Pell, 

1999). Almost all of the evidence presented in Table AX9-12 comes from experimentation 

involving plants grown from the outset in, or subjected to, an abrupt or step increase to a higher 

AX9-144
more or less double), steady CO2 concentration. In contrast, the O3 exposure concentrations 

usually varied from day to day. Luo and Reynolds (1999), Hui et al. (2002), and Luo (2001) 

noted the difficulties in predicting the likely effects of a gradual CO2 increase from experiments 

involving a step increase or those using a range of CO2 concentrations. Indeed, although using 

the much accelerated timescale of an 80-day growing season, Hui et al. (2002) clearly showed 

significant differences between the rates and magnitudes of various physiological and growth 

responses of plantain (Plantago lanceolata) to CO2 between gradual and step increase 

treatments. The authors concluded that, even though there were major differences in most of the 

parameters studied between the gradual and step treatments, “the convergence of the measured 

parameters at the end of the experiment provides some encouragement for the applicability of 

step-type experiments in the field; however, the study suggests caution in interpreting early 

results from short-term studies.” 

In long-term studies, the matter of photosynthetic acclimation to elevated CO2 levels has to 

be considered. Lawlor and Keys (1993) define acclimation in terms of long-term (days, weeks), 

irreversible physiological changes, in contrast to regulation, which relates to more rapid 

(minutes, hours), reversible changes. Each may be positive or negative, but many studies 

indicate that, while positive acclimation to elevated CO2 levels initially led to enhanced 

photosynthesis and growth, negative acclimation ultimately ensued and reduced CO2 

assimilation and growth rates. However, the consensus from recent studies and reviews is that 

such negative acclimation is most likely to occur in situations in which plants are grown under 

some additional stress, induced, e.g., by limitations to growth posed by lack of resources such as 

water or nutrients. The meta-analysis by Curtis (1996) revealed that slow or little negative 

acclimation was noted in studies on unstressed tree species with unhindered opportunities for 

root growth and development, a view originally suggested by Arp and Drake (1991) and largely 

supported in the review by Eamus (1996). A nonwoody perennial, the rhizomatous wetland 

sedge, Scirpus olneyi, grown in its natural environment with no edaphic limitations showed no 

negative acclimation after 4 years; in fact, photosynthetic capacity increased by 31% (Arp and 

Drake, 1991). No negative acclimation of well-watered, field-grown Ponderosa pine trees was 

observed by Tissue et al. (1999) after 6 years of growth at 2×-ambient CO2 levels. Gifford and 

Morison (1993) have summarized the situation thus: “Where the aerial or root environment for a 

plant is restricted (as with inter-plant competition, for example), positive feedback is limited and 

AX9-145
adjustments to the changed resource input balance under high CO2 can include ‘downregulation’ of leaf photosynthesis rate as an integral part of a positive growth response.” 

The influence of other environmental stressors is borne out by several long-term tree 

studies. After 3 years in 565 ppm CO2 in the Duke Forest FACE facility in North Carolina, 

maturing loblolly pine trees showed only a marginal CO2-induced carbon gain if grown on a 

nutritionally moderate site, but zero gain if grown on a nutritionally poor site (Oren et al., 2001). 

This is in sharp contrast to the substantially increased initial growth rates in elevated CO2 

reported by DeLucia et al. (1999), but it is supported by the observations of Tognetti et al. 

(2000) on five Mediterranean tree species growing for many years adjacent to geothermal 

springs releasing CO2 sufficient to provide ambient levels averaging 700 ppm. No significant 

differences in radial growth of the oaks (Quercus cerris, Q. ilex, and Q. pubescens), strawberry 

tree (Arbutus unedo), and flowering ash (Fraxinus ornus) could be detected between trees at the 

naturally enriched site and those at a nearby site exposed to normal ambient CO2 (~350 ppm). 

The authors concluded that limited availability of water and nutrients may have counteracted any 

positive effects of CO2 on growth at the enriched site or that the trees had acclimated to the 

higher CO2 levels. 

Because the ameliorative effects of CO2 on responses to O3 (Table AX9-14) were reported 

mostly in short-term studies involving an abrupt increase in CO2 level, it is appropriate to ask 

whether this amelioration is likely to persist to a time when the ambient CO2 concentration is 

relatively stable at such levels. Regardless of any negative acclimation due to resource 

limitations that may occur in the interim, steadily rising CO2 levels may well lead to natural 

selection and genetic change. Nevertheless, it seems reasonable to expect that the amelioration 

of O3 impact at elevated CO2 levels will be maintained in many situations, but the negative 

acclimation that will probably occur in situations where other resources become limiting will 

reduce the degree of protection. 

Another caveat regarding the validity of some of the observations in Table AX9-12 is 

related to the matter of stress-induced negative acclimation to elevated CO2 and concerns related 

to using potted plants. Although much of the recent information on CO2 effects has come 

from experiments with plants rooted in the ground, more than half of the studies listed in 

Table AX9-12 used potted plants, whether in controlled environment and greenhouse chambers 

or in OTCs. The degree to which pot-based studies resulted in similar patterns of response to 

AX9-146
soil-grown plants appears to depend on the treatment conditions and plant growth conditions 

used in the study. The use of potted plants was a confounding factor in the studies of Taylor 

et al. (2001) of the differences in leaf growth of poplar (Populus) hybrids between plants 

exposed to elevated CO2 in controlled environment chambers (potted plants), OTCs, or a FACE 

facility. Loats and Rebbeck (1999) suggested that their lack of CO2 response in three broadleaf 

species may have resulted from their use of pot-grown plants. In contrast, Heagle et al. (1999a) 

found that the relative enhancement of soybean photosynthesis, growth, and yield by CO2 

enrichment was similar in pots and in the ground. These findings were supported by Booker 

et al. (2005). The recent meta-analysis of data on the effects of elevated CO2 on soybean 

physiology and growth by Ainsworth et al. (2002) revealed a threefold smaller stimulation of 

seed yield in pot-grown than in field-grown plants, even when large (>9 L) pots were used. 

However, Ainsworth et al. (2002) included a wide range of treatment conditions (e.g., CO2 

treatments ranging from 500 to1200 ppm) and plant growth conditions in the meta-analysis, so 

caution is needed when generalizing conclusions about the applicability of pot-based studies. 

Although the majority of the cases cited in Table AX9-12 indicate that O3 and CO2 act 

additively or synergistically in causing stomatal closure, there are numerous exceptions. 

Any reduction in stomatal aperture has consequences other than merely restricting O3 uptake and 

the exchange of other gases. In particular, stomatal closure initially reduces the rate of 

transpiration, although increased leaf temperature and VPD associated with stomatal closure can 

offset decreases in transpiration. In instances where transpiration is reduced, water-use 

efficiency may increase, however decreased transpirational flux may lead to decreased mineral 

uptake, which could adversely impact growth over extended periods. 

Hence, the final caveat regarding Table AX9-12 concerns the interactions of O3 and CO2 

with other climatic variables, especially mean temperature. In light of the key role played 

by temperature in regulating physiological processes and modifying plant response to 

increased CO2 levels (Long, 1991; Morison and Lawlor, 1999) and the knowledge that 

relatively modest increases in temperature may lead to dramatic consequences in terms of 

plant development (Lawlor, 1998), it is unfortunate that much of the large investment in time 

and resources spent on recent studies of the effects of climate change on vegetation have gone 

into investigations limited to increasing our knowledge of the effects of higher levels of CO2 at 

current ambient temperatures. 

AX9-147
Some attention is now being paid to investigating the concurrent effects of CO2 increases 

and warming (recently reviewed by Rowland-Bamford [2000] and Morison and Lawlor [1999]), 

but the observed interactive effects on plant growth are inconsistent. For example, a FACE 

study with ryegrass (Lolium perenne) showed that increased temperatures (provided by infrared 

heaters) reduced the dry matter gain resulting from increased CO2 levels (Nijs et al., 1996). The 

field studies by Shaw et al. (2002) on a California annual grassland dominated by the grasses 

Avena barbata and Bromus hordeaceus and the forbs Geranium dissectum and Erodium botrys 

involved free-air increased CO2 as well as increased temperature, precipitation, and N supply. 

Not only did increased temperature reduce CO2-stimulated net primary productivity (NPP), but 

increased CO2 itself, combined with other factors, was found to be able to reduce NPP. 

There have been several investigations of effects on wheat. Batts et al. (1997) used plastic 

tunnels to create temperature gradients and maintain elevated CO2 levels over field-grown wheat 

and found that, in each of 4 years of study, a temperature rise of ~1.5 °C consistently canceled 

the growth and yield increases caused by a doubling of the CO2 level above ambient. Similar 

findings were reported by Van Oijen et al. (1999) and Van Oijen and Ewart (1999) in OTC field 

studies. Half of the chambers were cooled 1.6 to 2.4 °C below the uncooled chambers, to cancel 

out the normal temperature increase over ambient, due to the so-called “chamber effect” (usually 

a 1 to 3 °C increase above ambient temperature (Heagle et al., 1988). Although temperature had 

no effect on CO2-enhanced assimilation rates, the CO2-enhanced growth and grain yields 

observed in the cooled chambers were effectively canceled out in the warmer chambers. The 

authors attributed this effect to accelerated phenology, a shorter period for grain filling, and a 

lower leaf area index (LAI; total leaf area per unit ground area) in the warmer chambers. 

Wheeler et al. (1996) observed that the benefit to wheat of doubling the CO2 level was offset by 

a mean seasonal increase of only 1 to 1.8 °C. With the continuing use of OTCs for field 

research, Runeckles (2002), has suggested that the temperature rise due to the chamber effect in 

OTCs should be exploited (and measured) as a means of exploring temperature × CO2 as well as 

temperature × CO2 × O3 interactions. 

An indirect affirmation of the importance of temperature as a component of climate change 

on wheat yield was provided by Van Oijen and Ewart (1999) using two simulation models, 

AFRCWHEAT2-O3 and LINTULCC (Ewart et al., 1999). They analyzed data from the 

ESPACE-wheat program, which involved 25 OTC experiments in 1994, 1995, and 1996 at nine 

AX9-148
European locations (Jäger et al., 1999). Both models were able to predict control-treatment grain 

yields closely (5.5 ± 1.2 and 5.8 ± 1.2 t@ha!1, respectively, versus the observed 5.9 ± 1.9 t@ha!1), 

and both indicated a predominantly negative effect of temperature on the yield response to 

increased CO2 (a 3 °C rise reduced the gain in yield from 30 to 14%). However, neither model 

had an R2

 > 0.3, indicating that the models included other sources of variability among the sites 

than the climatic factors. The multiple linear regression developed by Bender et al. (1999) based 

on the same datasets also included temperature as a highly significant covariant. Both studies 

are discussed more fully below. 

Other studies, however, have found positive temperature-related growth effects, as 

suggested by the early Idso and Idso (1994) analysis. In an OTC study using the perennial 

grass Festuca pratensis in which a temperature increase of 3 °C above ambient was combined 

with CO2-enrichment to 700 ppm, both CO2 and temperature caused increases in total aboveground biomass (Hakala and Mela, 1996). Studies with potato (Cao et al., 1994) and soybean 

(Ziska and Bunce, 1997) using potted plants in controlled environment chambers also showed 

temperature-enhanced increases in growth in enriched CO2 atmospheres. Read and Morgan 

(1996) compared the effects of enriched CO2 and temperature on two grasses: cool-season 

Pascopyrum smithii and warm-season Bouteloua gracilis. In the latter (a C4 species), 

750 ppm CO2 resulted in increased dry matter production at daytime temperatures as high as 

35 °C, but in P. smithii (a C3 species), CO2-stimulated growth was greatest at 20 °C. However, 

the stimulation was progressively attenuated by increased temperature, so that at 35 °C, growth 

in 750 ppm was only one third of that in 350 ppm CO2 at 20 °C. 

Although the picture we have of temperature × CO2 interactions is inconsistent, RowlandBamford (2000) has provided persuasive evidence that the nature of the response to temperature 

in the grain yield of crops with as different temperature optima as rice and wheat will depend 

upon whether the change is above or below the temperature optimum. 

But what if we add O3 as another variable? Unfortunately, there have been very few 

studies of the three-way interaction. With the information available on CO2 × O3 interactions 

(Table AX9-12) and the limited information on temperature × O3 interactions (discussed in 

Section AX9.3.4.2) simulation modeling can attempt to provide estimates of O3 × CO2 × 

temperature effects, but experimental observation is still required to validate the models. 

The questions that need to be answered are: if increased temperature can offset the gains in 

AX9-149
productivity afforded by increased CO2 in important species such as wheat, and increased CO2 

can offset the reductions in productivity caused by O3, will increased temperature modify this 

protective effect? And if so, in what manner? 

To date, the only information available appears to consist of the reports by Van Oijen and 

Ewart (1999) and Bender et al. (1999) referred to above. In the former’s simulation studies, the 

overall yield depression of wheat caused by O3 was found to be 7 ± 4% for both 

AFRCWHEAT2-O3 and LINTULCC models versus an observed 9 ± 11%. The enhancements 

due to CO2 were predicted to be 24 ± 9% and 42 ± 11%, respectively, which straddled the 

observed 30 ± 22% gain. Based on the 13 experiments that included all four treatments (±O3, 

±CO2), an actual 10% yield loss due to O3 at ambient CO2 levels was reduced to a 4% loss by the 

elevated CO2. The AFRCWHEAT2-O3 model predicted 7 and 4% losses, and LINTULCC 

model predicted 8 and 5% losses due to the O3 and O3 + CO2 treatments. The actual and 

simulated yield increases in response to CO2 increased further with increasing temperature, but 

although temperature had no discernible effect on the observed depression of yield caused by O3 

alone, both models suggested that the yield reduction was diminished both by higher 

temperatures and higher CO2 levels. 

The analysis of the ESPACE-wheat experiments by Bender et al. (Bender et al., 1999) led 

to the following multiple linear regression: 

. . 2 . 3 . . . Y 2 = 1004 6 * * * + 0 588 * * *[CO ] − 1 908 * *[O ] − 31 230 * * *[T] + 7 309[I ] − 1448 423* * *[H O], (9-5) 

where Y = grain yield, g @ m!2

; [CO2] = ppm CO2; [O3] = ppb O3, 12-h mean; [T] = °C; [I] = light 

intensity, MJ @ m!2

/day; and [H2O] is a dummy variable: well watered = 1; limited water 

supply = 2. (***, p < 0.001; **, p < 0.01; the coefficient for I was not significant). With R2

 = 

0.3983, adjusted for 258 degrees of freedom, a large part of the variability was still unaccounted 

for by the five variables. However, this analysis suggests that CO2, O3, temperature, and 

water-status are important codeterminants of wheat yield, but assumes no interactions. 

Substitution in the model at summer light intensities and with well-watered plants indicates that, 

at 20° C, a doubling of CO2 levels to 700 ppm alone would lead to a 29.5% increase in yield, 

while 50 ppb O3 alone would decrease yield by 10.9%. With both gases at those levels, the yield 

AX9-150
would only increase 20%, but with a concurrent temperature rise of 2 °C, it would shrink to a 

9.6% increase. 

Both studies, therefore, indicate an amelioration of the effects of O3 by CO2, the magnitude 

of which would be reduced at warmer temperatures. However, they relate to a single crop whose 

response to CO2 is temperature-sensitive. Information about other species in which the effects 

of CO2 and temperature are additive are limited. However, Wolf and Van Oijen (2003) recently 

described a model (LPOTCO) simulating the effects of changes in climatic variables, CO2 

and O3 on tuber yield potential of irrigated potato (S. tuberosum cv. Bintje) over locations within 

the European Union ranging from Finland to Italy. They noted that although increased CO2, O3, 

and light intensity were predominant controlling factors, increased temperature also influenced 

potential yields substantially, with increases in northern latitudes (attributed to a longer growing 

season) but decreases in southern latitudes (attributed to decreased assimilate production). 

A clear understanding of the complex interactions of increased CO2 and temperature 

with O3 must await further experimentation or simulations. However, it seems likely that any 

CO2-induced amelioration of the adverse effects of O3 on aspects of growth other than seed or 

grain yield may be lessened or increased by increased temperature, depending upon the 

temperature optima for the species, along the lines suggested by Rowland-Bamford (2000). 

Other crop simulation models which incorporate O3 and some of the various environmental 

factors, including elements of climate change, have been reviewed by Kickert et al. (1999) and 

Rötter and Van De Geijn (1999). However, to date, the applications tend to have focused on 

interactions of O3 with factors such as soil moisture or nutrient availability. 

With forest trees, the situation has the added complexity of a perennial growth form and 

the inevitability, over time, of subjection to additional environmental stresses such as nutrientlimitation. Here, too, although numerous models of tree growth have been described, there 

appear to have been few applications to interactions of O3 and factors of climate change. 

Constable et al. (1996) used TREGRO to model the growth of Ponderosa pine exposed to 

three O3 levels (0.5×, 1.0×, and 2× ambient), two levels of CO2 (ambient and ambient + 

200 ppm CO2), and two temperature regimes (ambient and ambient + 4 °C). Plant growth was 

predicted to be decreased 1, 19, and 39% by the three levels of O3, respectively. Increased CO2 

reduced the loss at the highest O3 level to 7%; however, the combination of elevated CO2 with 

the higher temperature more than overcame the adverse effects of O3, leading to a 4% increase, 

AX9-151
largely attributed to increased fine root mass. The authors suggested that, in relation to the 

baseline conditions used in the simulations (Corvallis, OR), higher concentrations of CO2 and O3 

and a warmer climate will have little impact on total-tree growth, but they noted the importance 

of undertaking multiple stress studies in order to be able to make accurate forecasts of the impact 

of such changes on forest trees. 

More recently, Constable and Friend (2000) compared the capabilities of six published 

process-based models (CARBON, ECOPHYS, PGSM, TRE-BGC, TREGRO, and W91) for 

simulating tree response to elevated CO2, O3, and temperature. They concluded that, although 

these models were capable integrators of the effects of various environmental factors on 

individual processes such as photosynthesis, they were less reliable when extrapolating to 

growth. 

Although most of the research emphasis has been on simple CO2 × O3 interactions, a few 

isolated studies of interactions have involved O3, CO2, and biotic environmental factors. Heagle 

et al. (1994a) observed that both O3 and CO2 tended to be additive in encouraging the growth of 

spider mite (Tetranychus urticae) populations on clover. Infection of wheat with leaf rust 

(Puccinia recondita) sensitized the plants to O3 injury, but its severity was significantly reduced 

in elevated CO2 (Tiedemann and Firsching, 2000). The effects of O3 and CO2 on mycorrhizal 

symbioses was studied by Kytöviita et al. (1999) who found that CO2 did not ameliorate the 

adverse effects of O3 on the root growth of Aleppo pine and European white birch. In another 

study with Aleppo pine, Kytöviita et al. (2001) noted that both O3 and elevated CO2 reduced 

mycorrhiza-induced N-uptake by the roots. In Scots pine, Kasurinen et al. (1999) observed 

transient effects of elevated CO2 and O3 on root symbiosis, but none of the effects persisted over 

the 3 years of the study. 

The soil water × O3 × CO2 interaction was experimentally investigated by Broadmeadow 

and Jackson (2000) in Durmast oak, European ash, and Scots pine. No interactions were noted 

with ash and pine; but with oak, elevated CO2 ameliorated and irrigation exacerbated the effects 

of O3, although the resultant effects were essentially additive. 

Booker (2000) noted that soil nitrogen levels interacted only slightly with O3 and CO2 in 

determining the composition of cotton leaves and roots. Carbon dioxide reversed the inhibition 

of leaf growth caused by O3, but increased N-fertility tended to reduce this reversal. 

AX9-152
Because of the small number of studies of possibly significant interactions of three or more 

environmental factors, it is impossible to draw any sweeping conclusions as to how O3, in the 

context of global climate change, may affect relationships among plants and insects, diseases, 

and symbionts or among plants and nutrients or other air pollutants. The only interaction that 

has some degree of general support is the amelioration of adverse O3 effects by elevated CO2. 

AX9.3.8.2 Ozone-UV-B Interactions 

As noted in the 1996 O3 AQCD (U.S. Environmental Protection Agency, 1996), depletion 

of stratospheric O3 by halofluorocarbons has resulted in increased intensities of UV-B radiation 

(280 to 320 nm wavelengths) at the Earth’s surface. The situation is discussed more fully in 

Chapter 10. 

While stratospheric O3 depletion may result in increased surface UV-B irradiation, 

absorption of UV-B is a property of the O3 molecule regardless of its location; surface UV-B 

flux is, therefore, also reduced by O3 in the troposphere. Although only about 10% of the total 

atmospheric O3 column occurs in the troposphere (Fishman et al., 1990), it contributes a 

disproportionately greater absorption effect than stratospheric O3, because the UV radiation 

penetrating the troposphere becomes increasingly diffuse as it reaches the surface, with a 

consequent increase in mean path length (Brühl and Crutzen, 1989). Any benefits to vegetation 

from reduced ambient O3 stress must, therefore, also be viewed in the context of possible 

adverse effects due to increased UV-B irradiation. There are, thus, two distinct types of possible 

interactions between surface level O3 and UV-B radiation: 

C direct interactions involving simultaneous, sequential, or mixed exposures to O3 and 

UV-B stresses; and 

C effects on responses to UV-B itself resulting from changes in radiation intensity caused 

by changes in surface-level O3 concentrations. 

Only the first type of interaction is discussed here. The second type of interaction has broad 

implications for both health and welfare and focuses on the impacts of UV-B radiation per se. 

This topic is dealt with separately in Chapter 10. 

The most recent reviews specifically addressing the combined effects of tropospheric O3 

and UV-B on plants are by Runeckles and Krupa (1994) and Krupa and Jäger (1996), although 

the topic has also been included in several more general reviews of O3 effects and factors of 

AX9-153
climate change, such as those by Unsworth and Hogsett (1996), Krupa et al. (1998), Posthumus 

(1998), and Krupa and Groth (2000). 

However, little new information has become available since Runeckles and Krupa (1994) 

noted that the scanty knowledge of the effects of UV-B and O3 combinations available at that 

time was derived solely from studies of soybean. Miller et al. (1994) observed no interaction 

and no effect of UV-B on yield, in contrast to a previous report by Teramura et al. (1990) using 

the same cultivar, Essex. More recently, in a study of the saltmarsh grass Elymus athericus 

subjected to reciprocal exposures to O3 and UV-B, Van De Staaij et al. (1997) observed no 

interactive effects and no adverse effects of UV-B following 14-day exposures, even though an 

earlier report showed that longer exposures to UV-B (65 days) could cause a 35% loss of 

biomass (Van De Staaij et al., 1993). However, in a study in which ambient, high-altitude UV-B 

levels were compared with near-zero levels, at ambient or 2×-ambient levels of O3, interactions 

involving the levels of antioxidants in Norway spruce and Scots pine were reported by 

Baumbusch et al. (1998). Schnitzler et al. (1999) subsequently reported that O3-induced injury 

and adverse effects on photosynthesis were more pronounced with near-zero UV-B levels, 

indicating an amelioration of the O3-response. A later study with Scots pine (Zinser et al., 2000) 

revealed O3 × UV-B interactions at the gene expression and biochemical levels. In contrast, 

Ormrod et al. (1995) reported that UV-B predisposed Arabidopsis thalliana to injurious growth 

effects from O3 exposure. 

At various organizational levels, Runeckles and Krupa (1994) identified several similarities 

between plant response to O3 and UV-B, and at the level of gene expression, there have recently 

been several reports of both similarities and distinctions. Willekens et al. (1994) reported similar 

effects of O3, UV-B, and SO2 on the expression of antioxidant genes in Nicotiana 

plumbaginifolia. For parsley (Petroselinum crispum), Eckey-Kaltenbach et al. (1994a) found 

that O3 was a cross-inducer for both the UV-B-induced enhanced biosynthesis of flavonoids 

and the pathogen-induced furanocoumarin phytoalexins, in keeping with the previously 

observed O3-induction of fungal and viral defense reactions. In this regard, Yalpani et al. (1994) 

provided evidence that O3 and UV-B acted similarly in increasing disease-resistance via a 

salicylate-mediated enhancement of defense proteins in tobacco. However, subsequent work 

with tobacco led Thalmair et al. (1996) to conclude that exposure to UV-B did not lead to the 

accumulation of pathogenesis-related proteins. In Scots pine, although O3 is known to induce 

AX9-154
stilbene synthase and cinnamyl alcohol dehydrogenase, UV-B was found to enhance the former 

but suppress the latter, revealing an interaction at the level of gene expression (Zinser et al., 

2000). 

In summary, the present base of information about possible interactions between increased 

UV-B radiation and O3 is insufficient to draw any firm conclusions in terms of gross effects, but 

there is some evidence of similarities in the effects of O3 and UV-B individually and of the 

mechanisms involved at the molecular level. 

AX9.3.8.3 Interactions of Ozone with Multiple Climate Change Factors 

Despite the need for experimental investigations of three-way or more complex 

interactions among O3, CO2, UV-B, temperature, and other climate change factors, few studies 

have been reported, even without O3 as a factor. In an isolated report, using tomato seedlings, 

Hao et al. (2000) employed preexposure to UV-B (±CO2 enrichment) followed by exposure 

to O3 (±CO2 enrichment). They observed that CO2 enrichment more than overcame the 

inhibition of photosynthesis caused by O3, but pretreatment with UV-B reduced the resultant 

increase. 

In view of the unexpected observations made in their grassland study of the combined 

effects of CO2, temperature, precipitation, and N-supply, Shaw et al. (2002) affirmed that 

“Ecosystem responses to realistic combinations of global changes are not necessarily simple 

combinations of the individual factors.” The addition of O3 to the list of variables results in 

further complexity. 

Although computer simulation modeling may ultimately lead to improved understanding of 

these complex issues, to date, no such models appear to have been applied to these interactions, 

possibly because of the scarcity of experimental data for parameterization. 

AX9.3.9 Summary - Environmental Factors 

Although O3 and other photochemical oxidants are phytotoxic, their actions on vegetation 

may be modified by a host of biotic and abiotic factors in the environment; conversely, they may 

modify plant response to these other factors. The extensive review of these biological, physical, 

and chemical factors conducted for the 1996 O3 AQCD (U.S. Environmental Protection Agency, 

1996) concluded with a statement that our understanding was too fragmented to permit drawing 

AX9-155
many general conclusions. With today’s increased awareness of the need for more complete 

information on interactions, it is unfortunate that, in the interval since the 1996 O3 AQCD, 

rigorous, systematic investigations of interactions have been rare, and most of the new 

information is as fragmented as before. This is inevitable, partly in view of the vast scope of the 

possible interactions between O3 and other environmental variables and partly due to the overall 

lack of funding for such research in these areas. 

In the area of biotic interactions, new evidence with regard to insect pests and diseases has 

done little to remove the uncertainties noted in the 1996 criteria document. Most of the large 

number of such interactions that may affect crops, forest trees, and other natural vegetation have 

yet to be studied. The trend suggested previously that O3 increases the likelihood and success of 

insect attack has received some support from recent studies, but only with respect to chewing 

insects. With the economically important group of sucking insects such as the aphids, no clear 

trends have been revealed by the latest studies. Hence, although it seems likely that some insect 

problems could increase as a result of increased O3 levels, we are still far from being able to 

predict the nature of any particular O3 plant insect interaction, its likelihood, or its severity. 

The situation is a little clearer with respect to interactions involving facultative 

necrotrophic plant pathogens with O3, generally leading to increased disease. With obligate 

biotrophic fungal, bacterial, and nematode diseases, there are twice as many reports 

indicating O3-induced inhibitions than enhancements. The frequent reports that infection 

by obligate biotrophs reduces the severity of O3-induced foliar injury should not be interpreted 

as “protection”, because of the negative effects on the host plant of the disease per se. 

With obligate biotrophs, the nature of any interaction with O3 is probably dictated by the unique, 

highly specific biochemical relationships between pathogen and host plant. At this time, 

therefore, although some diseases may become more widespread or severe as a result of 

exposure to O3, it is still not possible to predict which diseases are likely to present the greatest 

risks to crops and forests. 

Several studies have indicated that the functioning of tree root symbioses with mycorrhizae 

may be adversely affected by O3, but there is also evidence that the presence of mycorrhizae may 

overcome root diseases stimulated by O3 and that O3 may encourage the spread of mycorrhizae 

to the roots of uninfected trees. The latest studies, therefore, present no clearer picture of the 

likely nature of simple interactions of O3 and root symbionts, but in view of the importance of 

AX9-156
mycorrhizae as below-ground components of ecosystems, they are discussed more fully in 

Section AX9.5. 

The few recent studies of the impact of O3 on intraspecific plant competition have again 

confirmed that grasses frequently show greater resilience than other types of plants. In grasslegume pastures, the leguminous species suffer greater growth inhibition. And the suppression 

of Ponderosa pine seedling growth by blue wild-rye grass was markedly increased by O3. 

However, we are far from being able to predict the outcome of the impact of O3 on specific 

competitive situations, such as successional plant communities or crop-weed interactions. 

Light, a component of the plant’s physical environment, is an essential “resource” whose 

energy content drives photosynthesis and CO2 assimilation. It has been suggested that increased 

light intensity may increase the sensitivity to O3 of light-tolerant species while decreasing that of 

shade-tolerant species, but this appears to be an oversimplification with many exceptions. 

Temperature affects the rates of all physiological processes based on enzyme-catalysis and 

diffusion, and each process and overall growth (the integral of all processes) has a distinct 

optimal temperature range. Although some recent field studies have indicated that O3 impact 

significantly increases with increased ambient temperature, other studies have revealed little 

effect of temperature. But temperature is unquestionably an important variable affecting plant 

response to O3 in the presence of the elevated CO2 levels contributing to global climate change 

(see below). In contrast, evidence continues to accumulate to indicate that exposure to O3 

sensitizes plants to low temperature stress by reducing below-ground carbohydrate reserves, 

possibly leading to responses in perennial species ranging from rapid demise to impaired growth 

in subsequent seasons. 

Although the relative humidity of the ambient air has generally been found to increase the 

adverse effects of O3 by increasing stomatal conductance and thereby increasing O3 flux, 

abundant evidence indicates that the ready availability of soil moisture results in greater 

sensitivity to O3. The partial “protection” against the adverse effects of O3 afforded by drought 

has been observed in field experiments and modeled in computer simulations. There is also 

compelling evidence that O3 can predispose plants to drought stress. Hence, the response will 

depend to some extent upon the sequence in which the stresses occur, but, even though the 

nature of the response is largely species-specific, successful applications of model simulations 

will lead to larger-scale predictions of the consequences of O3 × drought interactions. However, 

AX9-157
it must be recognized that regardless of the interaction, the net result on growth in the short-term 

is negative, although in the case of tree species, other responses such as increased water use 

efficiency could be a benefit to long-term survival. 

Mineral nutrients in the soil, other gaseous air pollutants, and agricultural chemicals 

constitute chemical factors in the environment. The evidence regarding interactions with 

specific nutrients is still contradictory. Some experimental evidence indicates that low general 

fertility increases sensitivity to O3, while simulation modeling of trees suggests that nutrient 

deficiency and O3 act less than additively; however there are too many example of contrary 

trends to permit any sweeping conclusions. Somewhat analogously with temperature, it appears 

that any shift away from the nutritional optimum may lead to greater sensitivity, but the shift 

would have to be substantial before a significant effect on response to O3 was observed. 

Interactions of O3 with other air pollutants have received relatively little recent attention. 

The situation with SO2 remains inconsistent, but seems unlikely to pose any additional risk to 

those related to the individual pollutants. With NO and NO2, the situation is complicated by 

their nutritional value as N sources. In leguminous species, it appears that NO2 may reduce the 

impact of O3 on growth, with the reverse in other species, but the nature of the exposure pattern, 

i.e., sequential or concurrent, also determines the outcome. Much more investigation is needed 

before we will be able to predict the outcomes of different O3-NO-NO2 scenarios. The latest 

research into O3 × acid rain interactions has confirmed that, at realistic acidities, significant 

interactions are unlikely. A continuing lack of information precludes offering any 

generalizations about interactive effects of O3 with NH3, HF, or heavy metals. More evidence 

has been reported that the application of fungicides affords some protective effects against O3. 

Over the last decade, considerable emphasis has been placed on research into O3 

interactions with the components of global climate change: increased atmospheric CO2, 

increased mean global temperatures, and increased surface-level UV-B radiation. However, 

many of these studies have tended to regard increased CO2 levels and increased mean 

temperatures as unrelated phenomena. Experiments into the effects of doubled CO2 levels at 

today’s mean ambient temperatures may not reveal the impact of climate change on responses to 

O3. For example, the limited experimental evidence and evidence obtained by computer 

simulation suggest that in a 600+ ppm world, although the enriched CO2 would more than offset 

the impact of O3 on responses as varied as wheat yield or the growth of young Ponderosa pine 

AX9-158
trees, the concurrent increase in temperature would reduce, but probably not eliminate, the net 

gain. A similar decrease in the net gain resulting from the complete reversal by CO2 of the 

inhibition of photosynthesis caused by O3 has been reported for increased UV-B irradiation. 

Clearly, additional research is needed in this important area. 

In conclusion, although the increased use of computer simulations may be important in 

suggesting outcomes of the many complex interactions of O3 and various combinations of 

environmental factors, the results obtained will only be as reliable as the input data used for their 

parameterization. The data needed for good simulations can only come from organized, 

systematic study. For predicting the future, ignorance is as good as dependence on poor 

simulations. 

AX9.4 EFFECTS-BASED AIR QUALITY EXPOSURE INDICES 

AX9.4.1 Introduction 

Indices are metrics that relate measured plant damage (i.e., reduced growth) to monitored 

ambient O3 concentration over time. An index is needed to provide a consistent metric for 

reviewing and comparing exposure-response effects obtained from various studies. Such indices 

may also provide a basis for developing a biologically-relevant air quality standard for protecting 

ecological resources. Effects on plant growth and/or yield has been a major focus of the 

characterization of O3 impacts on plants for purposes of the air quality standard setting process 

(U.S. Environmental Protection Agency, 1986, 1996b). The quantifying function over some 

time frame has frequently been referred to as “dose-response” and “exposure-response.” The 

distinction is in how the pollutant concentration is measured: “dose” is the measured pollutant 

concentration absorbed by the leaf over some time period, whereas “exposure” is the ambient air 

concentration measured near the plant over some time period. 

A measure of plant O3 uptake from the ambient air (either rate of uptake or cumulative 

seasonal uptake) is the ideal measure, because without O3 or its reactive product(s) reaching the 

target tissue there is no effect (Tingey and Taylor, 1982). Uptake is controlled in part by stomata 

(see Section AX9.2 for a detailed discussion). An uptake measure should integrate all those 

environmental factors that influence stomatal conductance, e.g., temperature, humidity, soil 

water status. A direct measure of the internal leaf concentration of O3, however, is technically 

AX9-159
difficult and thus uptake values are generally obtained with simulation models that require 

species- and site-specific variables. Because of this, a surrogate for uptake was sought early on 

using statistical summaries of monitored ambient pollutant concentration over some integral of 

time (Lee et al., 1988; Lefohn and Benedict, 1982; O'Gara, 1922; U.S. Environmental Protection 

Agency, 1986, 1992, 1996b). 

An index of exposure then must consider those abiotic and biotic factors known to modify 

the plant response by altering O3 uptake (Hogsett et al., 1988; U.S. Environmental Protection 

Agency, 1996b), including the temporal dynamics of exposure (e.g., concentration, frequency, 

duration), plant phenology (see Section AX9.3), plant defense mechanisms (e.g., antioxidants) 

(see Section AX9.2), and site climate and soil factors (e.g., temperature, VPD, soil moisture) 

(see Section AX9.3). The development of such indices continues to be a challenge (U.S. 

Environmental Protection Agency, 1996b). 

AX9.4.2 Summary of Conclusions from the Previous Criteria Document 

The 1996 O3 AQCD (U.S. Environmental Protection Agency, 1996b) focused on the 

research developing exposure indices to quantify growth and yield effects in crops, perennials, 

and trees (primarily seedlings) and not foliar injury. The indices were various functional and 

statistical summaries of monitored hourly O3 concentrations over designated time periods. The 

testing of the adequacy of these indices to order the measured responses of growth and/or yield 

in crops and tree species as seedlings was accomplished through regression analyses of earlier 

exposure studies. No direct experimental testing of the adequacy of these indices was 

accomplished. Their development focused on consideration and inclusion of some, but not all, 

factors that affect O3 uptake and expression of effects (e.g., Lee et al., 1988). The 1996 

document (U.S. Environmental Protection Agency, 1996b) drew a number of conclusions that 

built on even earlier conclusions (U.S. Environmental Protection Agency, 1992). Based on a 

review of the research published since 1996, those conclusions are still valid. 

Studies prior to 1996 indicated that the components of exposure, including concentrations, 

temporal dynamics (e.g., time of day of peak events), frequency of occurrence, duration, and 

respite time, were integral to developing indices of exposure related to growth response. 

Evidence from the few direct experimental studies of varying exposure components indicate the 

importance of peak concentrations, temporal pattern of occurrence, respite time and the 

AX9-160
importance of cumulating the concentrations over the exposure period (Hogsett et al., 1985a; 

Musselman et al., 1983, 1986, 1994; U.S. Environmental Protection Agency, 1996b). 

Exposure duration influences the degree of plant response. Single season, year-long, or 

multiyear experimental results indicated that greater yield losses occurred when plants were 

exposed for the longer duration and that a cumulative-type index was able to better describe the 

exposure-yield relationship. Indices that do not consider duration, e.g., 7-h seasonal mean 

concentration index, single peak event index, or the index that cumulated all concentrations (i.e., 

SUM00), were unable to adequately describe the relationship. These single event or mean-type 

indices do not consider the role of duration of exposure and either focus only on the peak event 

or put too much focus on the lower hourly average concentrations (U.S. Environmental 

Protection Agency, 1996b). 

Higher hourly average concentrations had a greater effect on plant response. It was 

concluded that cumulative indices that gave greater weight to higher concentrations related well 

with plant response (crops and tree seedlings) and ordered the treatment means in monotonically 

decreasing fashion with increasing exposure, based on studies that applied two or more types of 

exposure regimes with replicate studies of the same species. Examples of these indices, among 

others, were: AOT60 (the seasonal sum of the difference between an hourly concentration above 

the threshold value of 60 ppb, minus the threshold value of 60 ppb), SUM06 (the seasonal sum 

of hourly concentrations at or above the threshold value of 60 ppb), W126 (a sigmoid functional 

weighting of all hourly concentrations for the season), (U.S. Environmental Protection Agency, 

1996b). 

No studies before or after 1996, have enabled a discrimination among the various 

weighted, cumulative indices. Various functional weighting approaches have been used, 

including allometric, sigmoid, and threshold weighting, and compared for best statistical fit of 

the plant growth or yield data; however, no one functional weighting was favored. 

An exposure index that incorporated either the daily or seasonal temporal patterns of 

higher concentration occurrence with the temporal pattern of individual species’ stomatal 

conductance was not reported in the 1996 O3 AQCD (U.S. Environmental Protection Agency, 

1996b). Based on available data, it was unresolved how to proceed with weighting time of day 

or temporal patterns of species conductance. 

AX9-161
The relative importance of cumulative peak concentrations (>0.10 ppm) versus cumulative 

mid-range concentrations (0.05 to 0.099 ppm) was questioned. Although controlled experiments 

had provided important evidence that the higher hourly average concentrations should be given 

greater weight than the mid-level values in developing indices, there was concern that, under 

ambient conditions in the field, the higher concentrations did not occur at the time of maximum 

plant uptake. This coincidence was considered to be the critical factor in determining peak 

concentration impacts on plants. Based on the evidence at that time, it was not possible to 

conclude whether the cumulative effects of mid-range concentrations were of greater importance 

than those of peak hourly average concentrations in determining plant response (U.S. 

Environmental Protection Agency, 1996b). No direct experimental studies, however, had 

addressed this question prior to 1996, nor have any since. 

The composite exposure-response functions for crops and tree seedlings were derived from 

single and multiyear exposure studies that used modified or simulated ambient exposure profiles. 

These profiles were typified by episodic occurrence of a large number of higher O3 

concentrations. This type of pattern is not atypical but is not found in all rural agricultural and 

some forested areas in the United States. Selecting a concentration value from these crop and 

seedling response models may result in an over- or underestimation of growth effects if applied 

to regions of the country where a different type of temporal pattern of occurrence is prevalent 

(U.S. Environmental Protection Agency, 1996b). A multicomponent index was suggested that 

combined the concentration-weighted, cumulative index with the number of occurrences of 

hourly averaged concentrations $0.10 ppm that might reduce the uncertainty associated with 

selecting the exposure value for protection based on NCLAN-type studies (Lefohn and Foley, 

1992; Musselman et al., 1994; U.S. Environmental Protection Agency, 1996b). No direct 

experimental studies addressed this question prior to 1996, nor have any since. 

Since 1996, additional research has focused on the time of day when the higher hourly 

average concentrations occur, the time of day of maximum plant uptake, the diurnal variability 

of plant defense mechanisms, and various suggestions as to including these factors in any one of 

the cumulative, concentration-weighted exposure indices. A much broader literature has focused 

since 1996 on relating O3 flux to plant response and how to use this as an index relating ambient 

concentration to effects. These new developments are discussed in the sections that follow. 

AX9-162
AX9.4.3 Evaluation of Various Exposure Indices for Describing Ambient 

Exposure-Response Relationships 

Mathematical approaches for summarizing ambient air quality information in biologically 

meaningful forms that can serve as surrogates for dose for O3 vegetation effects assessment 

purposes have been explored for more than 80 years (O'Gara, 1922; U.S. Environmental 

Protection Agency, 1996b). Several indices have attempted to incorporate some of the 

biological, environmental, and exposure factors (directly or indirectly) that influence the 

magnitude of the biological response and contribute to observed variability (Hogsett et al., 

1988). In the 1996 O3 AQCD (U.S. Environmental Protection Agency, 1996b), the exposure 

indices were arranged into five categories; (1) One Event, (2) Mean, (3) Cumulative, 

(4) Concentration Weighted, and (5) Multicomponent, and were discussed in detail (Lee et al., 

1989). Figure AX9-16 illustrates how several of the indices weighted concentration and 

accumulate exposure. 

Various components of the exposure-response relationship, including concentration, time 

of day, respite time, frequency of peak occurrence, plant phenology, predisposition, etc., were 

weighted with various functions and evaluated on their ability in ordering the regression of 

exposure versus growth or yield response. The statistical evaluations for each of these indices 

were accomplished using growth/yield response data from many earlier exposure studies (e.g., 

NCLAN). This retrospective approach was necessary, because there were no studies specifically 

designed to test the goodness of fit of the various indices. The regression approach selected 

those indices that most properly ordered and spaced the treatment means to optimize the fit of a 

linear or curvilinear model. This approach provided evidence for the best indices, albeit not as 

defensible as that from studies with experimental designs and analyses that focus on specific 

components of exposure. 

Most of the early retrospective studies reporting regression approaches used data from the 

NCLAN program or data from Corvallis, Oregon or California (Lee et al., 1987; Lee et al., 1988; 

Lefohn et al., 1988; Musselman et al., 1988; U.S. Environmental Protection Agency, 1992; U.S. 

Environmental Protection Agency, 1986). These studies were previously reviewed by the EPA 

(U.S. Environmental Protection Agency, 1992, 1996b) and were in general agreement that the 

best fit of the data were cumulative concentration-weighted exposure indices. Lee et al. (1987) 

suggested that exposure indices that included all the 24-h data performed better than those that 

AX9-163
Figure AX9-16. Diagrammatic representation of several exposure indices, illustrating 

how they weight concentration and accumulate exposure. (a) SUM06: 

the upper graphic illustrates an episodic exposure profile; the shaded 

area under some of the peaks illustrates the concentrations greater than 

or equal to 0.06 ppm that are accumulated in the index. The insert shows 

the concentration weighting (0 or 1) function. The lower portion of the 

graphic illustrates how concentration is accumulated over the exposure 

period. (b) SIGMOID: the upper graphic illustrates an episodic 

exposure profile; the variable shaded area under the peaks illustrates the 

concentration-dependent weights that are accumulated in the index. The 

insert shows the sigmoid concentration weighting function. The midpoint 

of the sigmoid weighting scheme was 0.062 ppm. The lower portion of the 

graphic illustrates how concentration is accumulated over the exposure 

period. (c) 2ndHDM and M-7: the upper graphic illustrates an episodic 

exposure profile. The lower portion of the graphic illustrates that the 

2ndHDM considers only a single exposure peak, while the mean applies a 

constant exposure value over the exposure period. 

Source: Tingey et al. (1991). 

used only 7 h of data; this was consistent with the conclusions of Heagle et al. (1987) that plants 

receiving exposures for an additional 5-h/day showed 10% greater yield loss than those exposed 

for 7-h/day. In an earlier analysis using the NCLAN data, Lee et al. (1988) found the “best” 

exposure index was a phenologically weighted cumulative index, with sigmoid weighting on 

concentration and a gamma weighting function as a surrogate for plant growth stage. This index 

AX9-164
was the best statistical fit, but it depended upon a greater knowledge of species and site 

conditions making specification of weighting functions difficult for general use. 

The next best fits were the several indices which only cumulated and weighted higher 

concentrations (e.g., W126, SUM06, SUM08, AOT40). Amongst this group it was not possible 

to distinguish a single best fit (Heagle et al., 1994b; Lee et al., 1988; Musselman et al., 1988). 

A statistical approach based on profile likelihoods was used to estimate parameters in 

generalized exposure indices similar to the SUM06 and AOT40 indices (Blankenship and 

Stefanski, 2001) using data from experiments conducted during 1993 at eight sites in the eastern 

United States in which O3-sensitive and -tolerant white clover genotypes were grown using 

methods developed by Heagle et al. (Heagle et al., 1994b). The results showed that for the 

SUMX family of indices, where X is a cutoff value, hourly O3 concentrations over ~71 ppb 

contribute the most to yield prediction. For the AOTX family of indices, the parameter was 

54.4 ppb. These values are similar to those used in the SUM06 and AOT40 indices already in 

use. Furthermore, investigation of weighting for time of day confirmed the importance of the 

mid-afternoon hours for this data set, unlike the results found for wheat in Sweden (Danielsson 

et al., 2003; Pleijel et al., 2000a). 

Other factors, including predisposition time (Hogsett et al., 1988; McCool et al., 1988) and 

crop development stage (Heagle et al., 1991; Tingey et al., 2002), contributed to variation in the 

biological response and suggested the need for weighting O3 concentrations to account for 

predisposition time and phenology. However, the roles of predisposition and phenology in plant 

response vary considerably with species and environmental conditions, so specification of a 

weighting function for general use in characterizing plant exposure was not possible. 

European scientists took a similar approach in developing indices describing growth and 

yield loss in crops and tree seedlings, using OTCs with modified ambient exposures, but many 

fewer species and study locations were employed in the European studies. There is evidence 

from some European studies that a lower (Pleijel et al., 1997) or higher (Finnan et al., 1996, 

1997) cutoff value may provide a better statistical fit to the experimental data. Finnan et al. 

(1997) used seven exposure studies of spring wheat to confirm that cumulative exposure indices 

emphasizing higher O3 concentrations were best related to plant response and that cumulative 

exposure indices using weighting functions, including cutoff concentrations, allometric and 

sigmoidal, provided a better fit and that the ordering of these indices differed with different 

AX9-165
linear or Weibull dose-response models. Weighting those concentrations associated with 

sunshine hours in an attempt to incorporate a element of plant uptake did not improve the index 

performance (Finnan et al., 1997). A more recent study using data from several European 

studies of Norway spruce, analyzed the relationship between relative biomass accumulation and 

several cumulative, weighted indices, including the AOT40 and the SUM06 (Skärby et al., 

2004). All the indices performed relatively well in regressing biomass and exposure index, with 

the AOT20 and AOT30 doing slightly better (r2

 = 0.46-0.47). In another comparative study of 

four independent data sets of potato yield and different cumulative uptake indices with different 

cutoff values, a similarly narrow range of r2

 was observed (0.3 -0.4) between the different 

cumulative uptake of O3 indices (Pleijel et al., 2002). 

In both the United States and Europe, the adequacy of these statistical summaries of 

exposure in relating biomass and yield changes have, for the most part, all been evaluated using 

data from studies not necessarily designed to compare one index to another (Lee et al., 1988, 

1989; Skärby et al., 2004). But given the available data, the cumulative, concentration-weighted 

indices perform better than the peak or mean indices. It is not yet possible, however, to 

distinguish differences between the cumulative, concentration-weighted indices with direct 

experimental studies. 

The main conclusions from the 1996 O3 AQCD (U.S. Environmental Protection Agency, 

1996b) regarding an index based on ambient exposure are still valid. No information has come 

forth in the interim to alter those conclusions significantly, and in fact, some recent studies have 

further substantiated them. These key conclusions can be restated as follows: 

• O3 effects in plants are cumulative; 

• higher O3 concentrations appear to be more important than lower concentrations in 

eliciting a response; 

• plant sensitivity to O3 varies with time of day and crop development stage; 

• exposure indices that accumulate the O3 hourly concentrations and preferentially weight 

the higher concentrations have better statistical fits to growth/yield response than do the 

mean and peak indices. 

Following the 1996 criteria review process (U.S. Environmental Protection Agency, 

1996a,b), the EPA proposed an alternative form of the secondary NAAQS for O3 using a 

cumulative, concentration-weighted exposure index to protect vegetation from damage (Federal 

AX9-166
Register, 1997). The EPA considered three specific concentration-weighted indices: the cutoff 

concentration weighted SUM06 and AOT60 and the sigmoid-weighted W126 exposure index 

(U.S. Environmental Protection Agency, 1996a). All three indices performed equally well in 

predicting the exposure-response relationships observed in the crop and tree seedlings studies 

conducted during the previous 20 years (Lee et al., 1989). In a workshop convened to consider 

the science supporting these indices (Heck and Cowling, 1997), the participants agreed that all 

the cumulative concentration- weighted indices considered were equally capable of predicting 

plant response. 

The cutoff concentration-weighted index AOT40 was selected for use in Europe in 

developing exposure-response relationships based on OTC studies of a limited number of crops 

and trees (Grünhage and Jäger, 2003). The United Nations Economic Commission for Europe 

(UNECE, 1988) adopted the critical levels approach for assessment of O3 risk to vegetation 

across Europe. As used by the UNECE, the critical levels are not air quality regulatory standards 

in the U.S. sense, but rather planning targets for reductions in pollutant emissions to protect 

ecological resources. Critical levels for O3 are intended to prevent long-term deleterious effects 

on the most sensitive plant species under the most sensitive environmental conditions, but not to 

quantify O3 effects. A critical level was defined as “the concentration of pollutant in the 

atmosphere above which direct adverse effects on receptors, such as plants, ecosystems, or 

materials may occur according to present knowledge” (UNECE, 1988). The nature of the 

“adverse effects” was not specified in the original definition, which provided for different levels 

for different types of harmful effect (e.g., visible injury or loss of crop yield). There are also 

different levels for crops, forests, and seminatural vegetation. The caveat, “according to present 

knowledge,” is important because critical levels are not rigid; they are revised periodically as 

new scientific information becomes available. For example, the original critical level for O3 

specified concentrations for three averaging times, but further research and debate led to the 

current critical level being stated as the cumulative exposure (concentration × hours) over a 

cutoff concentration of 40 ppb (AOT40) (Fuhrer et al., 1997). The level of 3 ppmAh was 

selected, corresponding to a 5% yield loss in spring wheat as determined from 15 OTC studies. 

The critical level was defined for a 3-month period calculated for daylight hours. This value is 

currently used for all crops, because it is the best supported value and because the limited data 

from other crop species do not provide strong evidence that a more stringent value is required 

AX9-167
(Fuhrer et al., 1997). “Level I” critical levels are currently used to map and identify areas in 

Europe in which the levels are exceeded, and that information is then used to plan optimized and 

effects-based emissions abatement strategies. In the 1990s, areas of exceedance were mapped, 

but analyses of many exposure studies led to the conclusion that the simple, exposure-based 

approach led to the overestimation of effects in some regions and underestimation in others 

(Fuhrer et al., 1997; Kärenlampi and Skärby, 1996). The “Level I” approach did not 

differentiate between plant species, and it did not include modifying site and such 

micrometeorological factors of O3 uptake as VPD, water stress, temperature, and light and 

variation in canopy height. 

A decision was made to work towards a flux-based approach for the critical levels 

(“Level II”), with the goal of modeling O3 flux-effect relationships for three vegetation types: 

crops, forests, and seminatural vegetation (Grünhage and Jäger, 2003). Progress has been made 

in modeling flux (see Section AX9.4.5 (Ashmore et al., 2004a,b) and the Mapping Manual is 

being revised (Ashmore et al., 2004a,b; Grennfelt, 2004; Karlsson et al., 2003a). The revisions 

may include a flux-based approach for three crops: wheat, potatoes, and cotton. However, 

because of a lack of flux-response data, a cumulative, cutoff concentration-based (AOTx) 

exposure index will remain in use for the near future for most crops and for forests and 

seminatural herbaceous vegetation (Ashmore et al., 2004a). 

AX9.4.4 Identifying Exposure Components That Relate to Vegetation Effects 

The efficacy of exposure indices in predicting biological responses requires that 

researchers identify a relationship between measured growth and/or yield effects and exposure 

components and those environmental and site factors that control pollutant uptake by the plant 

effects. A number of these relationships were identified and discussed in the 1996 O3 AQCD 

(U.S. Environmental Protection Agency, 1996a). A significant, but in some instances, 

unquantified role was identified for (1) concentration; (2) duration of exposure; (3) the diurnal 

and seasonal patterns of exposure, e.g., time of day of peak event, season of higher exposures, 

seasons of high precipitation and humidity, the frequency of occurrence of peak events to respite 

time (peak to valley ratios); (4) plant phenology; (5) plant canopy structure; (6) meteorological 

and site factors, e.g., light, humidity; and (7) plant defense mechanisms. 

AX9-168
AX9.4.4.1 Role of Concentration 

A significant role of higher concentrations was established earlier, based on several 

experimental studies (U.S. Environmental Protection Agency, 1996b). Several studies since the 

last review (Nussbaum et al., 1995b; Oksanen and Holopainen, 2001; Yun and Laurence, 1999a) 

have added support for the important role that peak concentrations, as well as the pattern of 

occurrence, plays in plant response to O3. Oksanen and Holopainen (2001) found that the peak 

concentrations and the shape of the O3 exposure (i.e., duration of the event) were important 

determinants of foliar injury in European white birch saplings, but growth reductions were found 

to be more related to total cumulative exposure. Based on air quality data from 10 U.S. cities, 

three 4-week exposure treatments having the same SUM06 value were constructed by Yun and 

Laurence (1999a). They used the regimes to explore effects of treatments with variable versus 

uniform peak occurrence during the exposure period. The authors reported that the variable peak 

exposures were important in causing injury, and that the different exposure treatments, although 

having the same SUM06, resulted in very different patterns of foliar injury. Nussbaum et al. 

(1995b) also found peak concentrations and the pattern of occurrence to be critical in 

determining the measured response. The authors recommended that to describe the effect on 

total forage yield, peak concentrations >0.11 ppm must be emphasized by using an AOT with 

higher threshold concentrations. 

A greater role for higher concentrations affecting plant growth might be inferred based on 

recent air quality analyses for the Southern California area (Lee et al., 2003; Tingey et al., 2004). 

In the late 1960s and 1970s, extremely high O3 concentrations had impacted the San Bernardino 

NF. However, over the past 15 plus years, significant reductions in O3 exposure have occurred 

in the San Bernardino National Forest (Davidson, 1993; Lee et al., 2003; Lefohn and Shadwick, 

2000; Lloyd et al., 1989). An illustration of this improvement in air quality is shown by the 

37-year history of O3 air quality at a site in the San Bernardino Mountains (Figure AX9-17) 

(Lee et al., 2003). The O3 exposure increased from 1963 to 1979 concurrent with increased 

population and vehicular miles, followed by a decline to the present mirroring decreases in 

precursor emissions. The pattern in exposure was evident in various exposure indices including 

the cumulative concentration weighted (SUM06), as well as maximum peak event (1-h 

peak), and the number of days having hourly averaged O3 concentrations $95 ppb (i.e., the 

California O3 standard). The number of days having hourly averaged O3 concentrations $95 ppb 

AX9-169
Figure AX9-17. Trends in May to September 12-h SUM06, peak 1-h O3 concentration 

and number of daily exceedances of 95 ppb for Crestline in 1963 to 1999 

in relation to trends in mean daily maximum temperature for Crestline 

and daily reactive organic gases (ROG) and oxides of nitrogen (NOx) 

for San Bernardino county. Annual ROG and NOx emissions data for 

San Bernardino county were obtained from Alexis et al. (2001) and the 

California Air Resource Board’s emission inventory available at 

http://www/arb/ca.gov/emisinv/emsmain/emsmain.htm. 

Source: Lee et al. (2003). 

AX9-170 
declined significantly from 163 days in 1978 to 103 days in 1997. The changes in ambient O3 air 

quality for the site were reflected in the changes in the frequency and magnitude of the peak 

hourly concentration and the duration of the exposure (Figure AX9-17). Considering the role of 

exposure patterns in determining response, the seasonal and diurnal patterns in hourly O3 

concentration did not vary appreciably from year to year over the 37-year period (Lee et al., 

2003). 

The inference for a role of higher concentrations comes both from results of ground 

measures of tree conditions on established plots and from results of model simulations. Across a 

broad area of the San Bernardino NF, the Forest Pest Management (FPM) method of injury 

assessment indicated an improvement in crown condition from 1974 to 1988; and the area of 

improvement in injury assessment is coincident with an improvement in O3 air quality (Miller 

and Rechel, 1999). A more recent analysis of forest changes in the San Bernardino NF using an 

expanded network of monitoring sites has verified significant changes in growth, mortality rates, 

basal area, and species composition throughout the area since 1974 (Arbaugh et al., 2003). 

A model simulation of Ponderosa pine growth over the 40-year period in the San Bernardino NF 

showed a significant impact of O3 exposure on tree growth and indicates improved growth with 

improving O3 air quality. This area has also experienced elevated N deposition and based on a 

number of environmental indicators, it appears that this area is experiencing N saturation (Fenn 

et al., 1996). To account for this potential interaction, the model simulations were conducted 

under conditions of unlimited soil N, which helps account for these effects. The actual 

interactions are not known. The improvement in growth was assigned to improved O3 air 

quality, but no distinction was made regarding the relative role of mid-range and higher hourly 

concentrations, only that improved growth tracked decreasing SUM06, maximum peak 

concentration, and number of days of hourly O3 $95 ppb (Tingey et al., 2004). A summary of air 

quality data from 1980 to 2000 for the San Bernardino NF area of the number of “mid-range” 

hourly concentrations indicated no dramatic changes over this 20-year period, ranging from 

about 1500 to 2000 hours per year (Figure AX9-18). There was a slow increase in the number of 

mid-range concentrations from 1980 to 1986, which corresponds to the period after 

implementation of the air quality standard. Another sharper increase was observed in the late 

1990s. This pattern of occurrence of mid-range hourly concentrations suggests a lesser role for 

AX9-171
Figure AX9-18. The number of hourly average concentrations between 50 and 89 ppb 

for the period 1980 to 2000 for the Crestline, San Bernardino, CA 

monitoring site. 

Source: U.S. Environmental Protection Agency (2003). 

these concentration ranges compared to the higher values in either of the ground-level tree injury 

observations of the model simulation of growth over the 40-year period. 

AX9.4.4.2 Role of Duration 

Recent studies have called into question the period of time over which concentrations are 

accumulated and the form of the exposure index. Heagle and Stefanski (2000) reported that the 

form of the exposure index was important only for 24-h indices for which SUM00 (cumulated all 

hourly concentrations with no concentration weighting) provided the poorest fit. The authors 

reported that the SUM00, SUM06, W95 (Lefohn and Runeckles, 1987), W126, and AOT40 

produced similarly good fits of the foliage biomass data for 6-, 5-, and 4-h midday accumulating 

periods. The study pooled data from San Bernardino (CA) and Riverside (CA) with data from 

Amherst (MA), Corvallis (OR), Kennedy Space Center (FL), Raleigh (NC), and Blacksburg 

(VA). Ozone exposures were much higher at the two California sites (indicated by high W126, 

AX9-172
SUM06, W95, and AOT40 values) compared to the other locations. Because of the pooling of 

the data, the large number of high hourly average O3 concentrations that occurred at the 

California sites may have resulted in the exposure indices being highly correlated with one 

another and made it difficult to identify one optimal index. 

In another study in California, Arbaugh et al. (1998) reported that the SUM00 exposure 

index performed better for describing visible injury than the SUM06, W126, number of 

hours $0.08 ppm, and the number of days between measurement periods (U.S. Environmental 

Protection Agency, 1996b). These exposure indices were originally developed and tested using 

only growth/yield data, not foliar injury (U.S. Environmental Protection Agency, 1996b). This 

distinction is critical in comparing the efficacy of one index to another. However, for many 

locations in California, a large number of higher hourly average concentrations occur; thus the 

SUM00 could be highly correlated with the frequency of elevated hourly average concentrations 

and could be a good predictor of vegetation effects. 

AX9.4.4.3 Patterns of Exposure 

A significant factor in developing exposure indices is the temporal patterns of O3 

occurrence over a day, a month, and a year, as well as seasonally overlaying the daily and 

seasonal temporal patterns of those influential climatic and site factors. The coincidence of 

peak O3 with maximal stomatal conductance and detoxification processes is key to affecting 

plant growth response (Musselman and Minnick, 2000). 

Daily Patterns 

The diurnal patterns of coincidence of the maximal leaf/needle conductance and 

occurrence of higher ambient concentrations are relevant to the question of which hours during 

the day over a season that hourly concentrations should be cumulated for those indices that 

cumulate and weight concentration. 

A 12-h daylight period for cumulating exposure was proposed following the 1996 O3 

AQCD based primarily on the assumption that most species probably do not have significant 

conductance at night (U.S. Environmental Protection Agency, 1996a). An extensive review of 

the literature, however, reported that a large number of species had varying degrees of nocturnal 

stomatal conductance (Musselman and Minnick, 2000). The role of nighttime stomatal 

AX9-173
conductance and O3 exposure was demonstrated experimentally as well. Grulke et al. (2004) 

showed that the stomatal conductance at night for Ponderosa pine in the San Bernardino NF 

(CA) ranged from one tenth to one fourth that of maximum daytime gas exchange. In June, at 

the high-elevation site, 11% of the total daily O3 uptake of pole-sized trees occurred at night. In 

late summer, however, O3 uptake at night was negligible. Birch seedlings exposed to O3 at night 

show greater reductions in growth than those exposed to O3 in daylight (Matyssek et al., 1995). 

Field mustard (Brassica rapa L.) plants exposed to O3 during the day or night showed little 

significant difference in the amounts of injury or reduced growth response to O3 treatment, 

although the stomatal conductance was 70 to 80% lower at night (Winner et al., 1989). Tissue 

biomass of Ponderosa pine seedlings was significantly reduced when seedlings were exposed to 

either daytime or nighttime episodic profiles (Lee and Hogsett, 1999). However, the biomass 

reductions were much greater with daytime peak concentrations than with nighttime peak 

concentrations. 

Although stomatal conductance was lower at night than during the day for most plants, 

nocturnal conductance could result in some measurable O3 flux into the plants. In addition, 

plants might be more susceptible to O3 exposure at night than during the daytime, because of 

possibly lower plant defenses at night (Musselman and Minnick, 2000). Nocturnal O3 flux also 

depends on the level of turbulence that intermittently occurs at night. Massman (2004) 

suggested that nocturnal stomatal O3 uptake accounted for about 15% of the cumulative daily 

effective O3 dose that was related to predicted injury. Based on a review of the literature 

regarding plant nocturnal stomatal conductance, Musselman and Minnick (2000) recommended 

that any O3 exposure index used to relate air quality to plant response should use the 24-h 

cumulative exposure period for both exposure-response and effective flux models. However, in 

an evaluation of a very large number of indices that described the O3 impact on spring wheat, 

Finnan et al. (1997) did not find any improvement in performance of the cumulative 

concentration-weighted indices by weighting those concentrations occurring during sunlight 

hours. 

Stomatal conductances are species-dependent and linked to both diurnal and seasonal 

meteorological activity as well as to soil/site conditions (e.g., soil moisture). Daily patterns of 

leaf/needle conductance were often highest in midmorning, whereas higher ambient O3 

concentrations generally occurred in early to late afternoon when stomata were often partially 

AX9-174
closed and conductances were lower. Total O3 flux depends on atmospheric and boundary layer 

resistances, both of which exhibit variability throughout the day. Recent experimental studies 

with tree species demonstrated the decoupling of ambient O3 exposure, peak occurrence, and gas 

exchange, particularly in areas of drought (Panek, 2004). Several recent studies have suggested 

that Ponderosa pine trees in the southern and northern Sierra Nevada Mountains may not be as 

susceptible to high O3 concentrations as to lower concentrations, due to reduced needle 

conductance and O3 uptake during the period when the highest concentrations occur (Arbaugh 

et al., 1998; Bauer et al., 2000; Panek et al., 2002; Panek and Goldstein, 2001). Panek et al. 

(2002) compared direct O3 flux measurements into a canopy of Ponderosa pine and demonstrated 

a lack of correlation of daily patterns of conductance and O3 occurrence, especially in the lateseason drought period; they concluded that a consideration of climate or season was essential, 

especially considering the role of soil moisture and conductance/uptake. In contrast, Grulke 

et al. (2002a) reported high conductance when O3 concentrations were high in the same species, 

but under different growing site conditions. The decoupling of conductance and higher ambient 

O3 concentration would hold true for more mesic environments as well as xeric landscapes. The 

longer-term biological responses reported by Miller and Rechel (1999) for Ponderosa pine in the 

same region, and the general reduction in recent years in ambient O3 concentrations, suggest that 

stomatal conductance alone may not be a sufficient indicator of potential vegetation injury or 

damage. 

The generalized models of stomatal conductance may provide a means to link patterns 

of O3 occurrence with climatic and site factors that affect O3 uptake, provided conductance is 

modeled by regions of similar seasonal moisture and by similar canopy structure (Emberson 

et al., 2000a, 2000b) (Grünhage et al., 2000) (Massman, 2004). 

Seasonal Patterns 

Several of the recent studies measuring O3 flux to pine canopies also reported on the 

importance of seasonal patterns in relating exposure to response (Bauer et al., 2000). These 

seasonal patterns can be early- versus late-season occurrence of higher O3 concentrations, 

reflecting climate and precursor availability. The patterns also reflected seasonal drought and 

the role soil moisture played in stomatal conductance and O3 uptake. Recently, studies looked 

directly at this linkage. Panek et al. (2002) compared direct O3 flux measurements into a canopy 

AX9-175
of Ponderosa pine with a number of exposure indices and demonstrated a lack of correlation, 

especially in the late-season drought period; the authors concluded that a consideration of 

climate, especially soil moisture, was essential. They suggested that a better metric for 

seasonally drought-stressed forests would be one that incorporates forest physiological activity, 

through mechanistic modeling, by weighting ambient O3 concentrations by stomatal 

conductance, or by weighting O3 concentrations by site moisture conditions. Panek (2004) 

demonstrated a decoupling of O3 exposure and uptake seasonally as well, via seasonal drought 

influence. Maximum O3 uptake occurred at the beginning of the season and in the winter, 

whereas the pines were nearly dormant during August to September. 

Using TREGRO, a process-based tree growth model, Tingey et al. (2004) simulated longterm growth of Ponderosa pine over a 37-year period. The simulation showed a high degree of 

association between O3 exposure and O3-induced reductions in tree growth (R2

 = 0.56). The 

scatter about the line, however, indicated that other factors beside O3 are required to describe the 

association between exposure and response. Incorporating annual temperature and precipitation 

increased the R2

 to 0.67. In keeping with the observations of Panek (2004) on the decoupling of 

peak O3 occurrence and maximal conductance, the remaining unexplained variation is attributed 

to differences in timing of peak O3 uptake and peak O3 exposure over the years. 

AX9.4.4.4 Frequency of Occurrence of Peak Concentrations 

Several earlier studies demonstrated the greater effect of episodic occurrence of O3 peaks 

compared to daily peak events (U.S. Environmental Protection Agency, 1996b). Since the 

1996 O3 AQCD, a few studies have corroborated the importance of this pattern in growth 

response (Köllner and Krause, 2003; Yun and Laurence, 1999a; Nussbaum et al., 1995b). 

Köllner and Krause (2003) reported that, under equal exposure conditions, the most pronounced 

effects on the yield of sugar beet (Beta vulgaris L.) and soybeans occurred with those regimes 

that emphasized the episodic occurrence of peak events. Similarly, Yun and Laurence (1999b) 

used exposure regimes constructed from 10 U.S. cities to demonstrate that variable peak 

occurrence versus uniform occurrence was important in causing injury in tree seedlings. 

Nussbaum et al. (1995b) compared the effects of different patterns of peak occurrences with 

similar AOT40 values and reported a stronger effect on total forage yield from the episodic 

treatment. 

AX9-176
AX9.4.4.5 Canopy Structure 

Another factor important in either O3 exposure or uptake is how canopy structure affects O3 

concentration in and under forest canopies. There have been several investigations of O3 

concentrations under tree canopies (Enders, 1992; Fontan et al., 1992; Fredericksen et al., 1995; 

Joss and Graber, 1996; Kolb et al., 1997; Lorenzini and Nali, 1995; Neufeld et al., 1992; 

Samuelson and Kelly, 1997). In general, they indicated a reduction in O3 of ~20 to 40% in the 

area below the canopy but above the shrub/herb layers. An essential component in the 

determination of the AOT40 as a critical level was the height at which the O3 concentration was 

measured. The measurement heights are related to the O3 concentration measured at the top of 

the canopy, i.e., upper surface boundary of the (quasi-) laminar layer (Grünhage and Jäger, 

2003). This location is presumably more related to stomatal uptake. Weighting the O3 

concentration at this location takes into account stomatal opening and, if weighted with the 

Jarvis-Steward factors for radiation, temperature, and soil moisture, the “toxicologically” 

effective AOT40 is obtained (Grünhage and Jäger, 2003). A question exists however as to 

whether this “canopy” O3 concentration is clearly connected to stomatal O3 uptake. During site 

conditions that limit stomatal conductance (e.g., low soil moisture, high VPD) at the top of the 

canopy, high concentrations of O3 can occur with minimal risk. 

In a study that considered those factors important in O3 uptake that are also spatially 

distributed as a result of canopy structure, Davison et al. (2003) reported that the variation in 

visible injury in coneflower (Rudbeckia laciniata var. digitata) populations in Great Smoky 

Mountains National Park was unlikely to be due to differences in O3 flux and more likely due to 

variation in PAR. At a height of 50 cm above ground, PAR was reduced by almost 90%, 

whereas the O3 varied from about 15 to 90% of ambient. Ozone injury was not solely related to 

O3 flux. Although there have been studies of the effects of different light levels on O3 response, 

there have been few at the very low levels that occur in canopies of tall herbaceous stands or in 

the ground layer of forests. Davison et al. (2003) reported that conductance was not related to 

diurnal changes in light. The O3 levels were still about 90% of the O3 concentration above the 

canopy when light was less than 5%. Light intensity dropped to 1.5% of open at 130 cm from 

the edge of the canopy, while O3 dropped to only 42%. The study, although reporting on the 

adequacy of visible foliar injury as an indicator of O3 effects, suggested that consideration of 

AX9-177
other factors such as light were important in predicting response. How this might be included in 

developing exposure-response indices was not considered. 

AX9.4.4.6 Site and Climate Factors 

Soil moisture is a critical factor in controlling O3 uptake through its effect on plant water 

status and stomatal conductance. In an attempt to relate uptake, soil moisture, and ambient air 

quality to identify areas of potential risk, available O3 monitoring data for 1983 to 1990 were 

used along with literature-based seedling exposure-response data from regions within the 

southern Appalachian Mountains that might have experienced O3 exposures sufficient to inhibit 

growth (Lefohn et al., 1997). In a small number of areas within the region, O3 exposures and 

soil moisture availability were sufficient to possibly cause growth reductions in some O3­

sensitive species (e.g., black cherry). The conclusions were limited, however, because of the 

uncertainty in interpolating O3 exposures in many of the areas and because the hydrologic index 

used might not reflect actual water stress. 

AX9.4.4.7 Plant Defense Mechanism - Detoxification 

The non-stomatal component of plant defenses are the most difficult to quantify, but some 

studies are available (Barnes et al., 2002; Chen et al., 1998; Massman and Grantz, 1995; Plöchl 

et al., 2000), and a larger discussion can be found in Section AX9.3. Massman et al. (2000) 

developed a conceptual model of a dose-based index to determine how plant injury response 

to O3 relates to the traditional exposure-based parameters. The index used time-varyingweighted fluxes to account for the fact that flux was not necessarily correlated with plant injury 

or damage. Their model applied only to plant foliar injury and suggested that application of 

flux-based models for determining plant damage (yield or biomass) would require a better 

understanding and quantification of the injury and damage relationship. 

AX9.4.5 Ozone Uptake or Effective Dose as an Index 

Another approach in developing an index that relates growth response to ambient O3 is 

based on determining the O3 concentration going from the atmosphere into the leaf, or flux. 

Interest has been increasing in recent years, particularly in Europe, in using mathematically 

tractable flux models for O3 assessments at the regional and national scale (Emberson et al., 

AX9-178
2000a,b). Reducing uncertainties in flux estimates for areas with diverse surface or terrain 

conditions to within ±50% requires “very careful application of dry deposition models, some 

model development, and support by experimental observations” (Wesely and Hicks, 2000). As 

an example, the annual average deposition velocity of O3 among three nearby sites in similar 

vegetation was found to vary by ±10%, presumably due to terrain (Brook et al., 1997). 

Moreover, the authors stated that the actual variation was even greater, because stomatal uptake 

was unrealistically assumed to be the same among all sites, and flux is strongly influenced by 

stomatal conductance (Brook et al., 1997). This uptake-based approach to quantify the 

vegetation impact of O3 requires inclusion of those factors that control the diurnal and 

seasonal O3 flux to vegetation (e.g., climate patterns and species and/or vegetation-type factors 

and site-specific factors). The models have to distinguish between stomatal and non-stomatal 

components of O3 deposition to adequately estimate actual concentration reaching the target 

tissue of a plant to elicit a response. Determining this O3 uptake via canopy and stomatal 

conductance by necessity relies on models to predict flux and ultimately the “effective” flux 

(Grünhage et al., 2004; Massman et al., 2000; Massman, 2004). “Effective flux” has been 

defined as the balance between the O3 flux and the detoxification process (Dämmgen et al., 

1993; Grünhage and Haenel, 1997; Musselman and Massman, 1999). The time-integrated 

“effective flux” is termed “effective dose”. The uptake mechanisms and the resistances in this 

process, including stomatal conductance and biochemical defense mechanisms, are discussed in 

the previous Section AX9.3. The flux-based index is the goal for the “Level II” critical level for 

assessment of ozone risk to vegetation and ecosystems across Europe (Ashmore et al., 2004a). 

AX9.4.5.1 Models of Stomatal Conductance 

Only a limited number of studies have measured O3 concentration or its reaction products 

within the leaf (e.g., Moldau and Bichele (2002); see Section AX9.4.3). Altimir et al. (2002) 

described an enclosure technique for measuring O3 flux to foliage at the shoot level that allowed 

determination of partitioning and seasonality of the removal pathways on the foliage. The loss 

of O3 to the wall material of the chamber was great and required a correction when the stomatal 

activity was low. Only a few instances of direct measures of O3 flux to foliage in the field have 

been reported. Most measures of O3 flux are from canopy measurements made with 

micrometeorological techniques, but a number of assumptions are necessary and there are 

AX9-179
limitations due to landscapes (Grünhage et al., 2000; Wesely and Hicks, 2000). Comparison of 

simulated and measured O3 flux densities show good agreement in the mean (Grünhage et al., 

2000). Comparison, however, of continuous O3 concentrations and fluxes measured over a 

5-year period by the gradient method (Fowler et al., 1989) in a 30-year-old Norway spruce stand 

demonstrated a correlation over 5 years but were not correlated on a diurnal or seasonal basis. 

The correlation was based on two uncoupled processes inside and outside the stomata, i.e., the 

destruction of O3 outside the stomata in the canopy was influenced by those same factors 

(temperature, light, humidity) that control the diurnal opening and closing of the stomata. 

A similar lack of correlation of measured concentration and estimated flux, daily and seasonally, 

into Norway spruce and cembran pine (Pinus cembra) at six sites was due mostly to the control 

of stomatal conductance by those same microenvironmental factors (temperature, humidity, 

irradiance) (Emberson et al., 2000b). Seasonal variation of flux was attributed to the 

temperature course. During the growing season, the leaf-air VPD was the environmental factor 

controlling stomatal conductance and O3 flux into the needles. 

Given the limitations of actual measures of flux and the lack of correlation between 

measured concentrations and flux, a key goal is to develop an uptake or flux-based response 

index using models that consider site, climatic, meteorological, and species-specific (e.g., 

detoxification reactions) factors. Models of O3 conductance into plant tissue are available 

(Grünhage et al., 1997; Massman, 1993; Wesely, 1989). The European Monitoring and 

Evaluation Program (EMEP) developed an O3 deposition model for application across Europe in 

conjunction with the EMEP photochemical model as a tool for the critical levels program 

(Emberson et al., 2000a). The model was developed to estimate vegetation type-specific O3 

deposition and stomatal flux, calculated according to a standard three-resistance formulation 

incorporating atmospheric, boundary layer, and stomatal resistances (Emberson et al., 2000a). 

The model used a multiplicative algorithm of the stomatal conductance of O3 (Jarvis, 1976) and 

has been parameterized for 10 European tree species, seven crop species, and one type of 

seminatural vegetation. The model calculates conductance as a function of leaf phenology, 

temperature, photosynthetic flux density (PFD), VPD, and soil moisture deficit (SMD). The 

environmental variables are site-specific (or regionally-specific). The most important factors 

limiting O3 with this model were VPD, SMD, and phenology (Emberson et al., 2000a). These 

AX9-180
factors demonstrate the critical linkage of high VPD and stomatal closure, which typically cooccur with high O3 concentrations. 

A number of recent model-based studies have investigated the relationship of flux and 

plant growth response in several crop and forest tree species (Karlsson et al., 2004a,b; Pleijel 

et al., 2004; Altimir et al., 2004; Bassin et al., 2004; Elvira et al., 2004; Emberson et al., 2000b; 

Gerosa et al., 2004; Matyssek et al., 2004; Mikkelsen et al., 2004; Soja et al., 2004; Tuovinen 

et al., 2004; Wieser and Emberson, 2004). The studies used earlier exposure experiments as well 

as explicitly designed field studies, but no clear associations emerged to provide a basis for a 

flux-based index. Grünhage and Jäger (2003) emphasized the need for chamber-less 

experiments to develop flux-effect relationship based on flux estimates at canopy height. 

The complexity of using flux as an index of O3 exposure for growth response is shown in 

field studies that measured O3 flux into Norway spruce and cembran pine (Emberson et al., 

2000b). They demonstrated that stomatal conductance was the main limiting factor for O3 

uptake and showed the dependence of that measure on crown position, needle age, and altitude. 

Consideration of the role of climate illustrates the importance of a flux measure. Pleijel et al. 

(2000b) reported the improved relationship of yield in spring and winter wheat grown in OTCs 

in many areas across Europe when it was related to the cumulative stomatal O3 uptake during the 

grain-filling period. Compared to the AOT40, the cumulative uptake index estimated larger 

yield losses in the relatively humid parts of western and northern Europe, while smaller yield 

loss was estimated for the dry summer climates in southern and central Europe. 

Danielsson et al. (2003) compared the ability of two different stomatal models to relate 

grain yield in field-grown spring wheat to cumulated O3 uptake and an exposure index of AOT40 

and found that the cumulated O3 uptake determined with either model performed better in 

relating exposure to yield than did the cumulative exposure index of AOT40. 

Cumulative O3 uptake (CU) was modeled for three deciduous and two coniferous species 

growing at different sites and elevations and compared with the AOT40 exposure measure at 

these sites (Matyssek et al., 2004). A general linearity was demonstrated between the two 

measures of O3 exposure, and, at any given AOT40, there was a 25 ± 11% variation in CU. 

Although no correlation of growth alterations was observed with either the exposure or the 

uptake measure, the modeled cumulative uptake was able to describe the variation in tree size 

and site location, making for a better measure in risk assessment of O3 (Matyssek et al., 2004). 

AX9-181
Karlsson et al. (2004b) compared the biomass-response relationship in young trees at seven 

experimental sites across Europe using modeled cumulative O3 uptake and AOT40. A weaker 

dose-response relationship was reported for the cumulative uptake metric compared to the 

AOT40 (Karlsson et al., 2004b). 

Concern about the complexity of the stomatal models and the data needed to model O3 

uptake has led some researchers to offer modified accumulated exposure indices that consider 

the meteorological factors controlling uptake (Gerosa et al., 2004; Karlsson et al., 2004a). In a 

study of subterranean clover in Austria, Belgium, and southern Sweden, Karlsson et al. (2004a) 

reported on the performance of a modified accumulated exposure over the threshold (mAOT) 

which was based on solar radiation and VPD. This index improved the relationship for observed 

visible injury. But when modeled uptake of O3 was derived from a simple stomatal conductance 

model considering solar radiation, VPD, and air temperature, this index gave an even greater 

improvement in the relationship to visible injury than did the ambient exposure index of AOT40 

(Karlsson et al., 2004a). The added value of the mAOT was worthwhile, as was its lower degree 

of complexity and data requirements compared to simulating O3 uptake with stomatal models. 

Based on a study of O3 fluxes over a barley (Hordeum vulgare L.) field in Italy, a similar 

modified exposure index was reported and referred to as “effective exposure” (Gerosa et al., 

2004). Their approach was similar in its consideration of physiological aspects in conjunction 

with monitored O3 concentrations. It also addressed the shortcomings of the data needs for 

modeled O3 uptake. 

Models that partition O3 uptake into stomatal and non-stomatal components are also now 

available and predict a significant non-stomatal component in calculating O3 flux (Altimir et al., 

2004; Bassin et al., 2004; Mikkelsen et al., 2004; Nikolov and Zeller, 2003; Nussbaum et al., 

2003; Zeller and Nikolov, 2000). Altimir et al. (2004) compared the relative contributions of 

stomatal and non-stomatal sinks at the shoot level for Scots pine. Using the EMEP model with a 

revised parameterization for Scots pine, they demonstrated that a major removal of O3 was due 

to the non-stomatal component; when a non-stomatal term was introduced dependent on ambient 

relative humidity, the non-stomatal contribution to the total conductance was about 50%. Zeller 

and Nikolov (2000) demonstrated a large non-stomatal O3 uptake (41% of the total annual flux) 

in subalpine fir at a site in southern Wyoming using the biophysical model FORFLUX. In a 

5-year study of measured O3 flux to a Norway spruce canopy, Mikkelsen et al. (2004) showed 

AX9-182
monthly patterns of non-stomatal and stomatal deposition as part of total deposition to the 

canopy. Their study demonstrated that daily means of O3 concentration and fluxes averaged 

over 5 years correlated well, but the correlation was based on two different noncoupled 

processes outside and inside the stomata. The destruction of O3 in the canopy was influenced by 

temperature, light, and humidity, and these same factors influence stomatal opening, e.g., 

midday and night closure. Consequently, the diurnal O3 concentration and O3 flux do not 

correlate at all during the growing season. The study estimated yearly stomatal uptake to be a 

minimum of 21% of total deposition (i.e., non-stomatal uptake was as high as 80% of total). 

The stomatal uptake was highest May to August (30 to 33%) and lowest November to 

February (4 to 9%). 

AX9.4.5.2 Nonlinear Response and Developing Flux Indices 

If O3 flux were used as the only metric to predict vegetation injury or damage, the 

prediction might be overestimated, because of nonlinear relationships between O3 and plant 

response (Amiro et al., 1984; Amiro and Gillespie, 1985; Bennett, 1979, 1996b, 1986; U.S. 

Environmental Protection Agency, 1978). The nonlinearity in the response surface suggests the 

existence of a biochemical threshold. Musselman and Massman (1999) suggested that those 

species having high amounts of detoxification potential might show less of a relationship 

between O3 stomatal uptake and plant response. More recently, nonlinear relationships 

between O3 flux and yield were shown for potato (Pleijel et al., 2002) and spring wheat 

(Danielsson et al., 2003). The relationship between O3 flux and potato yield led to the use of 

an instantaneous flux threshold to overcome the nonlinear relationship (Pleijel et al., 2002). 

However, the authors did not report a substantial improvement in the mathematical fitting of the 

model when applying the threshold. Most of the flux was accumulated below 0.06 ppm. 

However, Danielsson et al. (2003) showed an improved relationship between O3 uptake and 

yield of spring wheat using a threshold of 5 nmoles m-2 sec-1 (0.24 mg m-2 sec-1). These results 

suggest not all O3 entering the stomata contribute to a reduction in yield, which depends to some 

degree on the amount of internal detoxification occurring for each particular species (see Section 

AX9.2). The cellular detoxification reactions and repair processes which both detoxify oxidants 

as well as play central roles in the carbon economy of the plant are another level of resistance 

to O3 reaching the target tissue (see Section AX9.2). The magnitude of the response is 

AX9-183
determined by the amount of the pollutant reaching the target site and the ability of the plant to 

reestablish homeostatic equilibrium. Thus, one would expect to observe a decoupling of O3 

uptake with vegetation effects, which would manifest as a nonlinear relationship between O3 flux 

and injury or damage. 

Additional factors for inclusion in flux-based models to predict vegetation effects would be 

the defense and repair mechanisms. However, the fact that the defense and repair mechanisms 

vary diurnally as well as seasonally may make it extremely difficult to apply a mathematically 

determined threshold to instantaneous flux measurements to calculate cumulative flux. The 

threshold models do not allow for the temporal (i.e., daily and seasonal) variability of defense 

mechanisms. Specifically, the relationship between conductance, O3 concentration, and 

defense/repair mechanisms needs to be included. Recently, Massman (2004) illustrated that the 

combination of stomatal conductance, O3 concentration, and diurnal variation of defense 

mechanisms showed the daily maximum potential for plant injury (based on effective dose) 

coincided with the daily peak in O3 mixing ratio. Massman et al. (2000) stressed that the product 

of the overlapping mathematical relationships of conductance, concentration, and defense 

mechanisms results in a much different picture of potential impact to vegetation than just the use 

of conductance and concentration in predicting vegetation effects. 

AX9.4.5.3 Simulation Models 

Another approach for determining O3 uptake and relating growth response to ambient O3 

exposure may be the use of physiologically-based simulation models. Several of these have 

been used in various contexts, comparing O3 response in a number of tree species with varying 

climate and site factors (e.g., soil moisture) (Hogsett et al., 1997; Laurence et al., 2001; Ollinger 

et al., 1997, 1998; Weinstein et al., 2001, 2002). These process-based models provide for an 

integration of species, climate, and site factors controlling O3 uptake with long-term growth. 

One of the important considerations in applying simulation modeling is to carefully assess the 

uncertainties associated with the modeling predictions. Further efforts need to be made to 

exercise the models so that they predict past growth losses associated with changes in O3 

exposures that can be verified with on-the-ground surveys. 

AX9-184
AX9.4.6 Summary 

A large number of studies pertinent to the development of exposure indices have been 

published since 1996, and these are predominantly focused on the development of a flux-based 

index to relate ambient O3 to effects. There were only a few such studies prior to 1996 and these 

were reviewed in the 1996 O3 AQCD (U.S. Environmental Protection Agency, 1996b). The few 

studies published since 1996 on the role of O3 exposure components (including concentration, 

duration, and exposure patterns) in describing growth response to O3 exposures have 

substantiated earlier conclusions of the importance of higher concentration, shape of the peak, 

and the episodicity of peak occurrence in the plant response to O3 exposure. An inferred role of 

peak concentrations is possible from consideration of improved O3 air quality in regions such as 

the San Bernardino Mountains in southern California. Studies provide the basis for focusing on 

the higher O3 concentrations, while including the lower levels, when estimating the effects of 

emission reductions on vegetation. 

A few studies have demonstrated the potential disconnection of the temporal patterns of 

peak events and maximal stomatal conductance. In addition, a few other studies have 

demonstrated the uptake of O3 during nighttime hours, suggesting the need to cumulate O3 

exposure 24 h per day and not just during daylight hours. 

Several studies since 1996 have demonstrated another critical concern in developing an 

index for exposure. The concern is that peak O3 events and maximum stomatal conductance 

may be temporally separate. This disconnection introduces uncertainty in assessing O3 impact 

when using the current ambient exposure based cumulative, concentration-weighted indices. 

If stomatal conductance is relatively low, as in the late afternoon in arid climates, and that is the 

same time as the peak O3 concentrations, then use of an exposure index that does not consider 

this disconnect will overestimate the effect of the exposure. This concern is especially apparent 

when assessing the impact of O3 across all the varied climatic regions of the United States or 

Europe. Some studies use stomatal models to predict uptake (Ashmore et al., 2004a) or 

physiological process-based models (Laurence et al., 2001) to integrate those species, climate, 

and site factors that drive this temporal pattern of stomatal conductance and exposure, and thus 

reduce some of the uncertainty in regional and national assessments of effects. These 

approaches, however, are still limited by being species-dependent. 

AX9-185
The results of these studies and reviews indicate the need to continue to develop indices 

that are more physiologically and meteorologically connected to the actual dose of O3 the plant 

receives. The cumulative concentration-weighted exposure indices are acknowledged surrogates 

for effective dose and are simple conceptually and easy to measure. They do not fully 

characterize the potential for plant uptake and resulting effects associated with O3, because the 

indices, being measures of ambient concentration, do not include the physical, biological, and 

meteorological processes controlling the transfer of O3 from the atmosphere through the leaf and 

into the leaf interior (U.S. Environmental Protection Agency, 1996b). Use of such indices is 

especially limited in spatial risk characterizations, because of the lack of linkage between 

meteorology and species- and site-specific factors influencing O3 uptake. The flux-based 

approach should provide an opportunity to improve upon the concentration-based (i.e., exposure 

indices) approach. A cautionary argument was advanced in a few publications centered around 

the nonlinear relationship between O3 uptake and plant injury (not growth alteration) response. 

The concern was that not all O3 stomatal uptake results in a reduction in yield, which depends to 

some degree on the amount of internal detoxification occurring with each particular species; 

species having high amounts of detoxification potential may show less of a relationship 

between O3 stomatal uptake and plant response. 

The European approach and acceptance of flux-based critical values is a recognition of this 

problem; a concerted research effort is needed to develop the necessary experimental data and 

modeling tools that will provide the scientific basis for such critical levels for O3 (Dämmgen 

et al., 1994; Fuhrer et al., 1997; Grünhage et al., 2004). 

At this time, based on the current state of knowledge, exposure indices that differentially 

weight the higher hourly average O3 concentrations but include the mid-level values represent 

the best approach for relating vegetation effects to O3 exposure in the United States. A large 

database exists that has been used for establishing exposure-response relationships. Such a 

database does not yet exist for relating O3 flux to growth response. The pattern disconnects 

between period of uptake and peak occurrence, as well as the potential for nocturnal uptake, 

should be considered by adding some weighting functions into the currently used exposure 

indices. Of particular consideration would be their inclusion in regional-to-national estimations 

of O3 impacts on vegetation. Another useful approach to regional assessment for certain species 

AX9-186
is to simulate growth effects with process-based models that account for seasonal climate and 

site factors that control conductance. 

It is anticipated that, as the overlapping mathematical relationships of conductance, 

concentration, and defense mechanisms are better defined, O3-flux-based models may be able to 

predict vegetation injury and/or damage at least for some categories of canopy-types with more 

accuracy than the exposure-response models. 

AX9.5 OZONE EXPOSURE-PLANT RESPONSE RELATIONSHIPS 

AX9.5.1 Introduction 

Ambient O3 concentrations have long been known to cause visible symptoms, decreases in 

photosynthetic rates, decreases in plant growth, and decreases in the yield of marketable organs 

(U.S. Environmental Protection Agency, 1978, 1986, 1996). Yet, despite considerable research 

in the U.S. and other countries during the past three decades, quantifying the effects of ambient 

O3 exposure on vegetation remains a challenge. Numerous studies have related O3 exposure to 

plant responses, with most effort focused on the yield of crops and the growth of tree seedlings. 

Most experiments exposed individual plants grown in pots or soil under controlled conditions to 

known concentrations of O3 for a segment of daylight hours for some portion of the plant’s life 

span (Section AX9.1). The response of a plant species or variety to O3 exposure depends upon 

many factors discussed in previous sections, including genetic characteristics (Section AX9.3.2), 

biochemical and physiological status (Section AX9.3), and previous and current exposure to 

other stressors (Sections AX9.3, AX9.4). Section AX9.3 describes how O3 moves from the 

atmosphere into the leaf and the subsequent biochemical and physiological responses of plants. 

The current section focuses on the quantitative responses of plants to seasonal or multiyear 

exposures to known amounts of O3. Quantitative responses include foliar symptoms and 

decreased growth of whole plants or decreased harvestable portions of them. Because of the 

available information, most of this section focuses on the response of individual plants, 

especially crop plants and tree seedlings, with limited discussion of mixtures of herbaceous 

species. The responses of natural ecosystems are discussed in Section AX9.6. 

This section will pay particular attention to studies conducted since the publication of the 

1996 AQCD (U.S. Environmental Protection Agency, 1996). However, because much O3 

AX9-187
research was conducted prior to the 1996 AQCD, the present discussion of vegetation response 

to O3 exposure is largely based on the conclusions of the 1978, 1986, and 1996 criteria 

documents (U.S. Environmental Protection Agency, 1978, 1986, 1996). To provide a context for 

the discussion of recent research, the key findings and conclusions of those three documents are 

summarized below. 

AX9.5.2 Summary of Key Findings/Conclusions from Previous 

Criteria Documents 

Experimental data reviewed in the 1978 and 1986 O3 AQCDs dealt primarily with the 

effects of O3 on agricultural crop species (U.S. Environmental Protection Agency, 1978, 1986). 

The chapter on vegetation effects in the 1978 O3 AQCD (U.S. Environmental Protection Agency, 

1978) emphasized foliar symptoms and growth effects, but not those effects that affected yield, 

an emphasis dictated by the kind of data available at the time. The 1986 O3 AQCD reviewed a 

substantial new body of evidence based on OTC experiments (see Section AX9.1) showing that 

ambient O3 exposures reduced the growth and yield of herbaceous plants, again with a focus on 

major crop species. In the 1986 and 1996 O3 AQCDs, data were presented from regression 

studies conducted to develop exposure-response functions for estimating yield loss of major crop 

species in different regions of the United States. The 1996 O3 AQCD included results from 

additional herbaceous crop species as well as shrub and tree species. For a number of tree 

species, biomass growth of seedlings was related to growing season O3 exposures to produce 

response functions for estimating O3 exposures that reduce growth by 10 or 30%. Also, in the 

1986 and 1996 O3 AQCDs, data from studies using EDU as a protectant were reviewed. The 

1978, 1986, and 1996 O3 AQCDs also reviewed data on the response to O3 exposures of forest 

ecosystems in the San Bernardino Mountains of southern California (U.S. Environmental 

Protection Agency, 1978, 1986, 1996). Because this region is exposed to high concentrations of 

O3 and has shown evidence of ecosystem-level changes, it remains an important study area (see 

Section AX9.6). 

Ozone can cause a range of effects, beginning with individual cells, leaves, and plants, and 

proceeding to plant populations and communities. These effects may be classified as either 

“injury” or “damage”. Injury encompasses all plant reactions, such as reversible changes in 

plant metabolism (e.g., altered photosynthetic rate), altered plant quality, or reduced growth that 

AX9-188
does not impair yield or the intended use or value of the plant (Guderian, 1977). In contrast, 

damage includes all effects that reduce or impair the intended use or value of the plant. Damage 

includes reductions in aesthetic values as well as losses in terms of weight, number, or size of the 

plant part that is harvested (yield loss). Yield loss also may include changes in crop quality, i.e., 

physical appearance, chemical composition, or the ability to withstand storage. Losses in 

aesthetic values are difficult to quantify. Although foliar symptoms cannot always be classified 

as damage, their occurrence indicates that phytotoxic concentrations of O3 are present, and, 

therefore, studies should be conducted to assess the risk to vegetation. 

Visible symptoms due to O3 exposures reduce the market value of certain crops and 

ornamentals for which leaves are the product, e.g., spinach, petunia, geranium, and poinsettia. 

The concept of limiting values used to summarize foliar symptoms in the 1978 O3 AQCD (U.S. 

Environmental Protection Agency, 1978) was also considered valid in the 1986 O3 AQCD (U.S. 

Environmental Protection Agency, 1986). Jacobson (1977) developed limiting values by 

assessing the available scientific literature and identifying the lowest exposure 

concentration/duration reported to cause foliar symptoms in a variety of plant species. Graphical 

analyses presented in those documents indicated that the limit for reduced plant performance was 

an exposure to 50 ppb for several hours per day for more than 16 days. Decreasing the exposure 

period to 10 days increased the concentration required to cause symptoms to 100 ppb; and a 

short, 6-day exposure further increased the concentration required to cause symptoms to 

300 ppb. These limiting values established in 1978 were still deemed appropriate in the 1986 

and 1996 O3 AQCDs. Such foliar symptoms are caused by O3 concentrations that occur in the 

United States as shown in Table AX9-13 (adapted from U.S. Environmental Protection Agency, 

(U.S. Environmental Protection Agency, 1996). 

The 1986 O3 AQCD emphasized that, although foliar symptoms on vegetation are often an 

early and obvious manifestation of O3 exposure, O3 effects are not limited to foliar symptoms. 

Other effects include reduced growth of many organs (including roots), changes in crop quality, 

and alterations in plant susceptibility to biotic stressors and sensitivity to abiotic stressors. The 

1986 O3 AQCD also emphasized that O3 exerts phytotoxic effects only if a sufficient amount 

of O3 reaches sensitive sites within the leaf (Section AX9.2). Ozone injury will not occur if the 

rate of O3 uptake is low enough that the plant can detoxify or metabolize O3 or its metabolites 

or if the plant is able to repair or compensate for the effects (Tingey and Taylor, 1982; U.S. 

AX9-189
Table AX9-13. Summary of Ozone Exposure Indices Calculated for 3- or 5-Month 

Growing Seasons from 1982 to 1991a 

3-Month Growing Season (June-August) 

2HDMc M7 SUM00 SUM06 SIGMOID 

No. of ppm ppm ppmAh ppmAh ppmAh 

Year Sitesb Mean CVd Mean CV Mean CV Mean CV Mean CV 

1982 99 0.114 23.7% 0.05 18.7% 82.9 19.1% 26.8 68.8% 26.3 56.7% 

1983 102 0.125 24.9% 0.06 21.9% 86.1 22.1% 34.5 58.1% 33.0 52.3% 

1984 104 0.117 24.6% 0.05 18.2% 84.1 19.9% 27.7 58.4% 27.4 47.9% 

1985 117 0.117 24.6% 0.05 17.1% 84.6 18.0% 27.4 59.6% 27.4 47.6% 

1986 123 0.115 21.8% 0.05 19.1% 85.3 18.0% 27.7 65.0% 27.7 51.8% 

1987 121 0.119 22.9% 0.06 17.6% 86.9 17.3% 31.2 56.4% 30.4 46.8% 

1988 139 0.129 21.3% 0.06 17.8% 97.6 19.6% 45.2 46.8% 42.9 42.4% 

1989 171 0.105 23.1% 0.05 17.5% 86.4 19.9% 24.8 78.7% 25.8 59.4% 

1990 188 0.105 21.6% 0.05 18.3% 85.7 21.0% 25.8 76.2% 26.6 59.2% 

1991 199 0.106 22.0% 0.05 18.4% 87.7 21.3% 28.3 74.2% 28.9 59.5% 

Among Years 0.113 11.1% 0.05 10.0% 87.0 9.9% 29.5 42.1% 29.4 31.0% 

5-Month Growing Season (May-September) 

M7 SUM00 SUM06 SIGMOID 

No. of ppm ppmAh ppmAh ppmAh 

Year Sites Mean CV Mean CV Mean CV Mean CV 

1982 88 0.048 20.6% 122.9 22.3% 37.3 70.9% 37.1 57.8% 

1983 87 0.051 22.1% 129.6 24.4% 44.4 61.9% 43.8 52.7% 

1984 95 0.048 18.0% 126.2 19.1% 36.7 60.8% 37.6 46.9% 

1985 114 0.048 18.4% 124.5 19.4% 36.2 63.8% 37.0 50.3% 

1986 118 0.048 20.3% 123.3 21.4% 34.9 70.7% 35.6 55.7% 

1987 116 0.050 20.3% 128.7 20.4% 42.2 62.0% 41.8 50.3% 

1988 134 0.054 18.7% 141.7 22.0% 58.0 50.5% 55.6 45.0% 

1989 158 0.047 18.6% 127.8 22.5% 32.7 87.8% 35.2 64.1% 

1990 172 0.049 19.8% 129.4 22.7% 34.6 82.7% 37.0 62.1% 

1991 190 0.050 19.8% 130.6 23.6% 36.8 80.7% 38.8 62.9% 

Among Years 0.049 9.8% 129.0 9.9% 38.7 42.5% 39.6 29.8% 

a Updated and additional years from data given in Table III of Tingey et al. (1991), where the spatial and temporal variation

 in ambient O3 exposures is expressed in terms of several exposure indices. b Indicates the number of separate monitoring sites included in the analysis; fewer sites had 5 months of available data than

 had 3 months of available data. 

c The 2HDM index is calculated for sites with at least 3 months of available data. SUM00, SUM06, M7, SIGMOID, and 

 2HDM are the cumulative sum above 0.0 ppm, the cumulative sum above 0.06 ppm, the 7-h seasonal mean, the sigmoid

 weighted summed concentration, and the second highest daily maximum 1-h concentration, respectively. d CV = coefficient of variation. 

Source: Table 5-30 from U.S. Environmental Protection Agency (1996) based on Tingey et al. (1991). 

AX9-190 
Environmental Protection Agency, 1986). Cellular disturbances that are not repaired or 

compensated for are ultimately expressed as foliar symptoms, reduced root growth, or reduced 

yield of fruits or seeds. 

Beginning in the 1986 O3 AQCD and continuing in the 1996 O3 AQCD, OTC studies that 

better quantified the relationship between O3 exposure and effects on crop species were 

reviewed, with a focus on yield loss. 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 O3 exposure to plant response, and (2) studies that 

compared the effects of discrete treatment level(s) to a control. The advantage of the regression 

approach is that exposure-response models can be used to interpolate results between treatment 

levels. 

Discrete treatment experiments were designed to test whether specific O3 treatments were 

different from the control rather than to develop exposure-response equations, and the data were 

analyzed using analyses of variance. When summarizing these studies using discrete treatment 

levels, the lowest O3 concentration that significantly reduced yield was determined from 

analyses done by the original authors. Often, the lowest concentration used in a study was the 

lowest concentration reported to reduce yield; hence, it was not always possible to estimate a noeffect exposure concentration. In general, the data indicated that 100 ppb O3 (frequently the 

lowest concentration used in the studies) for a few hours per day for several days to several 

weeks usually caused significant yield reductions of 10 to 50%. 

By the time the 1986 O3 AQCD was prepared, much new information concerning the 

effects of O3 on the yield of crop plants had become available through EPA’s NCLAN research 

program and through research funded by other agencies. The NCLAN project was initiated by 

the EPA in 1980 primarily to improve estimates of yield loss under field conditions and to 

estimate the magnitude of crop losses caused by O3 throughout the United States (Heck et al., 

1982; Heck et al., 1991). The cultural conditions used in the NCLAN studies approximated 

typical agronomic practices. The primary objectives were: 

(1) to define relationships between yields of major agricultural crops and O3 exposure 

as required to provide data necessary for economic assessments and development 

of O3 NAAQS; 

(2) to assess the national economic consequences resulting from O3 exposure of major 

agricultural crops; and 

AX9-191
(3) to advance understanding of cause-and-effect relationships that determine crop 

responses to pollutant exposures. 

Using NCLAN data, the O3 concentrations predicted to cause 10 or 30% yield loss were 

estimated using linear or Weibull response functions. The data in Table AX9-14 are from the 

1996 document and were based on yield-response functions for 38 species or cultivars developed 

from studies using OTCs of the type developed by Heagle et al. (1973) (see Section AX9.1). 

Composite exposure-response functions for both crops and tree seedlings as a function of O3 

exposure expressed as SUM06 are shown in Figure AX9-19. Review of these data indicate that 

10% yield reductions could be predicted for more than 50% of experimental cases when: 

(1) 12-h SUM06 values exceeded 24.4 ppmAh, (2) SIGMOID values exceeded 21.5 ppmAh, or 

(3) 7-h seasonal mean concentrations were 50 ppb. The SIGMOID index is very similar to the 

W126 index (see Section AX9.4 for further information about O3 indices). Much lower values 

are required for each index to protect 75% of experimental cases (Table AX9-14). Grain crops 

were generally found to be less sensitive than other crops. The data summarized in the 1996 

criteria document also indicated that the variation in sensitivity within species may be as great as 

differences between species. 

The chemical protectant, EDU, was also used to provide estimates of yield loss. The 

impact of O3 on yield was determined by comparing the yield data from plots treated with EDU 

versus untreated plots. Studies indicated that yields were reduced by 18 to 41% when daytime 

ambient O3 concentrations exceeded 80 ppb for 5 to 18 days over the growing season. For this 

approach to be credible, the effects of EDU itself on a particular species must be preestablished 

under conditions without O3 exposure (Kostka-Rick and Manning, 1992a). 

The 1996 O3 AQCD reviewed several experiments demonstrating that the seedlings of 

some tree species such as poplars (Populus spp.) and black cherry are as sensitive to O3 as are 

annual plants, in spite of the fact that trees are longer-lived and generally have lower rates of gas 

exchange, and, therefore, a lower uptake of O3. The 1996 document also reviewed data showing 

that O3 exposures that occur at present in the United States are sufficient to affect the growth of a 

number of trees species. For example, exposure-response functions for 51 cases of tree seedling 

responses to O3, including 11 species representing deciduous and evergreen growth habits, 

suggest that a SUM06 exposure for 5 months of 31.5 ppmAh would protect hardwoods from a 

10% growth loss in 50% of the cases studied (Table AX9-15). Similarly, a SUM06 exposure of 

AX9-192
Table AX9-14. Ozone Exposure Levels (Using Various Indices) Estimated to Cause at 

Least 10% Crop Loss in 50 and 75% of Experimental Casesa 

50th PERCENTILEb SUM06 SEc SIGMOID SE M7 SE 2HDM SE 

NCLAN Data (n = 49; wet and dry)d 24.4 3.4 21.5 2.0 0.049 0.003 0.094 0.006 

NCLAN Data (n = 39; wet only) 22.3 1.0 19.4 2.3 0.046 0.003 0.090 0.010 

NCLAN Data (n = 54; wet and dry)e 26.4 3.2 23.5 2.4 NA NA 0.099 0.011 

NCLAN Data (n = 42; wet only)e 23.4 3.1 22.9 4.7 NA NA 0.089 0.008 

NCLAN Data (n = 10; wet) 25.9 4.5 23.4 3.2 0.041 0.001 0.110 0.042 

NCLAN Data (n = 10; dry) 45.7 23.3 40.6 0.1 0.059 0.014 0.119 0.017 

Cotton Data (n = 5) 23.6 2.3 19.3 2.3 0.041 0.001 0.066 0.032 

Soybean Data (n = 13) 26.2 5.4 22.6 3.6 0.044 0.005 0.085 0.013 

Wheat Data (n = 6) 21.3 15.2 19.3 12.7 0.061 0.018 0.098 0.059 

Cotton Data (n = 5)e 30.0 12.7 27.2 12.8 NA NA 0.075 0.012 

Soybean Data (n = 15)e 23.9 6.5 22.0 8.0 NA NA 0.088 0.008 

Wheat Data (n = 7)e 25.9 10.5 21.4 9.4 NA NA 0.097 0.028 

75th PERCENTILEb 

NCLAN Data (n = 49; wet and dry) 14.2 4.2 11.9 5.6 0.040 0.007 0.051 0.010 

NCLAN Data (n = 39; wet only) 14.3 2.7 12.6 2.3 0.039 0.005 0.056 0.006 

NCLAN Data (n = 54; wet and dry)e 16.5 4.3 14.5 3.2 NA NA 0.073 0.006 

NCLAN Data (n = 42; wet only)e 17.2 3.0 14.7 2.4 NA NA 0.070 0.006 

NCLAN Data (n = 10; wet) 16.4 3.7 13.7 3.2 0.040 0.001 0.080 0.032 

NCLAN Data (n = 10; dry) 24.0 0.8 22.3 0.1 0.053 0.022 0.093 0.003 

Cotton Data (n = 5) 21.8 5.0 17.5 2.8 0.041 0.001 0.065 0.014 

Soybean Data (n = 13) 14.2 0.1 12.4 0.1 0.041 0.006 0.069 0.004 

Wheat Data (n = 6) 11.7 2.5 10.9 2.4 0.054 0.032 0.062 0.035 

Cotton Data (n = 5)e 21.1 6.0 16.7 5.7 NA NA 0.070 0.034 

Soybean Data (n = 15)e 15.3 4.1 13.4 4.1 NA NA 0.078 0.007 

Wheat Data (n = 7)e 5.1 2.6 8.5 3.4 NA NA 0.054 0.027 

a

See Appendix A for abbreviations and acronyms. 

b

The numbers in parentheses are the number of cases used in deriving the various exposure levels. 

c

Standard error (SE).

d

NCLAN data refers to studies conducted as part of the NCLAN project. Wet and dry refer to watering regimes

 used in the studies, wet being well-watered, and dry meaning some level of drought stress was imposed. 

e

24-h exposure statistics reported in Lee et al. (1994b). Relative yield loss for 2HDM is relative to yield at 40 ppb rather 

than 0 ppb as was used in Tingey et al. (1991). 

Source: U.S. Environmental Protection Agency (1996) modified from Tingey et al. (1991). 

AX9-193
Figure AX9-19. Distribution of biomass loss predictions from Weibull and linear 

exposure-response models that relate biomass to O3 exposure. 

Exposure is characterized with the 24-h SUM06 statistic using data 

from (a) 31 crop studies from National Crop Loss Assessment 

Network (NCLAN) and (b) 26 tree seedling studies conducted at U.S. 

Environmental Protection Agency’s Environmental Research Laboratory 

in Corvallis, OR; Smoky Mountains National Park, TN; Houghton, 

Michigan; and Delaware, Ohio. Separate regressions were calculated 

for studies with multiple harvests or cultivars, resulting in a total of 

54 individual equations from the 31 NCLAN studies and 56 equations 

from the 26 seedling studies. Each equation was used to calculate the 

predicted relative yield or biomass loss at 10, 20, 30, 40, 50, and 60 ppmAh, 

and the distributions of the resulting loss were plotted. The solid line is 

the calculated Weibull fit at the 50th percentile. 

Source: U.S. Environmental Protection Agency (1996); Hogsett et al. (1995). 

AX9-194 
Table AX9-15. SUM06 Levels Associated with 10 and 20% Total Biomass 

Loss for 50 and 75% of the Seedling Studies 

(The SUM06 value is adjusted to an exposure length of 92 days per year.)a 

Weibull Equations (all 51 seedling studies): 

50th Percentile PRYLb

 = 1 - exp(-[SUM06/176.342]**1.34962) 

75th Percentile PRYL = 1 - exp(-[SUM06/104.281]**1.46719) 

Weibull Equations (27 fast-growing seedling studies): 

50th Percentile PRYL = 1 - exp(-[SUM06/150.636]**1.43220) 

75th Percentile PRYL = 1 - exp(-[SUM06/89.983]**1.49261) 

Weibull Equations (24 slow-to-moderate growing seedling studies): 

50th Percentile PRYL = 1 - exp(-[SUM06/190.900]**1.49986) 

75th Percentile PRYL = 1 - exp(-[SUM06/172.443]**1.14634) 

Weibull Equations (28 deciduous seedling studies): 

50th Percentile PRYL = 1 - exp(-[SUM06/142.709]**1.48845) 

75th Percentile PRYL = 1 - exp(-[SUM06/87.724]**1.53324) 

Weibull Equations (23 evergreen seedling studies): 

50th Percentile PRYL = 1 - exp(-[SUM06/262.911]**1.23673) 

75th Percentile PRYL = 1 - exp(-[SUM06/201.372]**1.01470) 

Levels Associated with Prevention of a 10 and 20% Total Biomass Loss 

for 50 and 75% of the Seedlings 

All 51 Seedling Cases 

Percent of Seedlings 

50% 75% 

Relative 10% 33.3 22.5 

Biomass Loss 20% 58.0 37.5 

27 Fast-Growing Seedling Cases 

Percent of Seedlings 

Relative 10%

Biomass Loss 20%

50% 75% 

31.3 19.4 

52.9 32.4 

AX9-195 
Table AX9-15 (cont’d). SUM06 Levels Associated with 10 and 20% Total 

Biomass Loss for 50 and 75% of the Seedling Studies 

(The SUM06 value is adjusted to an exposure length of 92 days per year.)a 

24 Slow-to-Moderate-Growth Seedling Cases 

Percent of Seedlings 

50% 75% 

Relative 10% 42.6 24.2 

Biomass Loss 20% 70.2 46.6 

28 Deciduous Seedling Cases 

Percent of Seedlings 

50% 75% 

Relative 10% 31.5 20.2 

Biomass Loss 20% 52.1 33 

23 Evergreen Seedling Cases 

Percent of Seedlings 

50% 75% 

Relative 10% 42.6 21.9 

Biomass Loss 20% 78.2 45.9 

a

See Appendix A for abbreviations and acronyms.

b PRYL = predicted relative yield (biomass) loss

Source: U.S. Environmental Protection Agency (1996), based on Hogsett et al. (1995). 

42.6 ppmAh should provide the same level of protection for evergreen seedlings. However, these 

results do not take into the account the possibility of effects on growth in subsequent years. 

Because multiple-year exposures may cause a cumulative effect on the growth of some trees 

(Simini et al., 1992; Temple et al., 1992), it is likely that a number of species are currently being 

affected even at ambient exposures (Table AX9-13). 

In 1986, the EPA (U.S. Environmental Protection Agency, 1986) established that 7-h per 

day growing season mean exposures to O3 concentrations above 50 ppb were likely to cause 

measurable yield loss in agricultural crops. At that time, few conclusions could be drawn about 

the response of deciduous or evergreen trees or shrubs, due to the lack of information about 

AX9-196
response of such plants to season-long exposures to O3 concentrations of 40 to 60 ppb and above. 

However, the 1978 and 1986 O3 AQCDs (U.S. Environmental Protection Agency, 1978, 1986) 

indicated that the limiting value for foliar symptoms on trees and shrubs was 60 to 100 ppb for 

4 h. From 1986 to 1996, extensive research was conducted, establishing the sensitivity of many 

tree species. Based on research published since the 1986 O3 AQCD (U.S. Environmental 

Protection Agency, 1986), a number of conclusions were drawn in 1996 O3 AQCD (U.S. 

Environmental Protection Agency, 1996): 

(1) An analysis of 10 years of monitoring data from more than 80 to almost 

200 nonurban sites in the United States established ambient 7-h growing season 

average concentrations of O3 for 3 or 5 months of 51 to 60 ppb and 47 to 54 ppb, 

respectively. The SUM06 exposures ranged (a) from 24.8 to 45.2 ppmAh for 

3 months and (b) from 32.7 to 58.0 ppmAh for 5 months (Tingey et al. (1991), 

Table AX9-13). 

(2) The results of OTC studies that compared yields at ambient O3 exposures with those 

in filtered air and retrospective analyses of crop data (Table AX9-14) established that 

ambient O3 concentrations were sufficient to reduce the yield of major crops in the 

United States. Research results since 1978 did not invalidate EPA conclusions 

(U.S. Environmental Protection Agency, 1978, 1986) that foliar symptoms due to O3 

exposures reduce the market value of certain crops and ornamentals where leaves are 

the product (such as spinach, petunia, geranium, and poinsettia) and that such 

damage occurs at ambient O3 concentrations observed in the United States. 

(3) A 3-month SUM06 exposure of 24.4 ppmAh, corresponding to a 7-h mean of 49 ppb 

and a 2HDM of 94 ppb O3 may prevent a 10% loss in 50% of the 49 experimental 

cases analyzed by Tingey et al. (1991). A 12-h growing season mean of 0.045 ppb 

should restrict yield losses to 10% in major crop species (Lesser et al., 1990). 

(4) Depending on duration, concentrations of O3 and SUM06 exposures currently in the 

United States are sufficient to affect the growth of a number of tree species. Given 

the fact that multiple-year exposures may cause a cumulative effect on the growth of 

some trees (Simini et al., 1992; Temple et al., 1992), it is likely that a number of 

species currently are being impacted, even at ambient O3 exposures (Tables AX9-13 

and AX9-20). 

(5) Exposure-response functions for 51 cases of seedling response to O3 (Hogsett et al., 

1995), including 11 species representing deciduous and evergreen growth habits, 

suggest that a SUM06 exposure for 5 months of 31.5 ppmAh would protect 

hardwoods from a 10% growth loss in 50% of the cases studied. A SUM06 exposure 

of 42.6 ppmAh should provide the same level of protection for evergreen seedlings. 

Note that these conclusions do not take into the account the possibility of effects on 

growth in subsequent years, an important consideration in the case of long-lived 

species. 

AX9-197
(6) Studies of the response of trees to O3 have established that, in some cases (for 

instance, poplars and black cherry), trees are as sensitive to O3 as are annual plants, 

in spite of the fact that trees are longer-lived and generally have lower gas exchange 

rates, and, therefore, lower O3 uptake. 

(7) Use of the chemical protectant, EDU, is of value in estimating O3-related losses in 

crop yield and tree growth, provided that care is exercised in establishing appropriate 

EDU dosages to protect the plants without affecting growth. 

The major question to be addressed in the remainder of this section is whether new 

information supports or alters the 1996 criteria document conclusions summarized above. 

In particular, this section evaluates whether the response of plants to experimental treatments at 

or near O3 concentrations characteristic of ambient levels in many areas of the United States 

(Tables AX9-13 and AX9-20) can be compared to a control or reduced O3 treatment to establish 

a potential adverse effect. Before evaluating new information from the literature on O3 effects 

on vegetation, O3 exposure indices used in O3 studies and trends in O3 exposure patterns during 

the past two decades are briefly reviewed. 

AX9.5.3 Ozone Indices and Ambient Exposure 

As recognized in both the 1986 and the 1996 criteria documents, the characterization and 

representation of the exposure of vegetation to O3 is problematic, because the specific aspects of 

pollutant exposure that cause injury or damage are difficult to quantify. This issue was 

addressed in Section AX9.4, and only a few points will be discussed here in order to provide a 

context for interpreting data on exposure-response relationships. The most important effects 

of O3 on vegetation occur due to uptake of O3 through stomata, with subsequent oxidative injury 

that appears to be rather nonspecific (Section AX9.2). As has been discussed by numerous 

authors during the last three decades, from a toxicological and physiological view, it is much 

more realistic to relate effects to internal (absorbed) O3 dose rather than to exposure near the leaf 

or canopy (Fowler and Cape, 1982; Fuhrer et al., 1992; Grünhage et al., 1993, 1999; Legge et al., 

1995; Massman et al., 2000; Musselman and Massman, 1999; Pleijel et al., 1995b; Runeckles, 

1974; Taylor et al., 1982; Tingey and Taylor, 1982) (see also Section AX9.4). Theoretically, 

flux estimates should improve the assessment of O3 effects, but despite recent attention to this 

topic, particularly in Europe, it remains difficult to estimate flux in the field outside of 

AX9-198
experimental sites where continuous measurements of wind speed and other environmental 

conditions are made. This topic is discussed further below in Section AX9.5.4.5. 

No simple exposure index can accurately represent all of the numerous factors operating at 

different timescales that affect O3 flux into plants and subsequent plant response (Section 

AX9.4). Indices of peaks, such as the 2HDM, are not well suited for discerning exposureresponse relationships, because they do not capture the effects of lower O3 concentrations nor the 

cumulative effects of O3 on vegetation (Heck and Cowling, 1997; U.S. Environmental Protection 

Agency, 1996). Therefore, peak indices have not been used in recent decades to develop 

exposure-response relationships for vegetation. Fortunately, other simple indices have shown 

substantial correlation with responses such as crop yield under experimental conditions. During 

the 1980s, the most commonly used indices for expressing O3 exposure were 7-, 8-, or 12-h 

daytime average values over the duration of O3 exposure, which was often 3 months or 

somewhat less for experimental studies with crops. These indices perform reasonably well for 

interpreting experimental data on the response of vegetation to O3, particularly for individual 

experiments, although they do not explain all of the variation among experiments in 

retrospective analysis of multiple experiments (Lesser et al., 1990). 

Since the 1980s, cumulative indices such as the SUM06, AOT40, or W126 that 

preferentially weight higher concentrations have been used in conjunction with mean indices for 

developing exposure-response relationships (Tables AX9-14 and AX9-15, and Figure AX9-18). 

Such indices are often more suitable than mean values, because they are cumulative and because 

they preferentially weight higher concentrations. Thus, these indices generally provide 

somewhat better fits to experimental data than do mean indices, especially in retrospective 

analyses of multiple experiments on multiple species (Lee et al., 1994a; Lee et al., 1994b; 

Lee and Hogsett, 1999; Tingey et al., 1991). Unfortunately, no single index has been used 

consistently even in the recent literature, making it difficult to compare results among and 

between experiments and with ambient exposure data. However, Tables AX9-13 and AX9-20 

provide summaries of ambient exposure data for several indices that can be compared to the 

experimental results reviewed in the remainder of this section. Of the cumulative indices that 

preferentially weight higher concentrations, the SUM06 index has been used most commonly in 

the U.S. literature, and it was selected in a meeting of scientific experts on O3 effects on 

vegetation as suitable for a secondary standard to protect vegetation (Heck and Cowling, 1997). 

AX9-199
However, it should be noted that the W126 index has been selected for use in protecting 

vegetation in Class 1 areas (Federal Land Manager's Air Quality Related Values Workgroup 

(FLAG), 2000). Even in recent studies, O3 data are often presented using only a seasonal 

mean index value, and so mean values are frequently presented in this section. Such reporting of 

mean indices should not be interpreted as a preference for them, but rather as a limitation in the 

data reported in the literature. Additional information about O3 exposure for individual 

experiments, including the number and type of O3 treatments (addition of a constant 

concentration of O3 or an amount proportional to ambient levels), and duration, are reported in 

Tables AX9-16 through AX9-19. 

Since the 1996 O3 AQCD, the use of the AOT40 index has become quite common in 

Europe for identifying and mapping areas of exceedance, but it has not been used much in the 

United States. Thus, studies reporting O3 exposure only as AOT40 values are presented in tables 

summarizing effects on annual, herbaceous perennial, and woody vegetation. However, such 

studies are not as commonly cited in the text of this section, because AOT40 summary data 

on O3 exposures in the United States are rarely available. This lack makes it difficult to compare 

experimentally derived exposure-response data expressed as AOT40 to ambient U.S. O3 

exposures. The development of critical levels in Europe has been based primarily on the AOT40 

index, so this index is discussed in that context. 

In addition to peak weighting, there is also evidence that the timing of exposure during 

plant growth is important. For example, the greatest effects on grain yield are due to exposure 

during grain filling, rather than earlier or later in the growing season (Lee et al., 1988; Pleijel 

et al., 1998; Soja et al., 2000; U.S. Environmental Protection Agency, 1996; Younglove et al., 

1994). A recent study grew bush bean in OTCs with CF or above-ambient O3 using exposure 

dynamics typical of the Midwestern United States for three time periods: (1) the entire season, 

(2) the period prior to anthesis, (3) during pod filling and maturation (Tingey et al., 2002). 

Ozone exposure prior to anthesis reduced growth by less than 1% per ppmAh (SUM06) while 

exposure during pod filling and maturation reduced growth by 4 to 7%. A meta-analysis of 

53 studies of O3 effects on soybean found that O3 had greater effects with increases in 

developmental stage, with the greatest effect during seed filling (Morgan et al., 2003). 

The importance of respite times was discussed in the previous criteria documents (U.S. 

Environmental Protection Agency, 1978, 1986, 1996) but remains difficult to quantify 

AX9-200
Table AX9-16. Summary of Selected Studies of Ozone Effects on Annual Species 

O3 Concentration Response 

(units are ppb unless Variable (Decrease from 

Species Facility Location otherwise specified)1 Duration lowest, %) 

Bean, cv. OTC The Netherlands CF to CF75: 62 days Green pod 29 at 9-h mean = Pros 9-h mean = 3-70, yield (AOT40 = 3.6 pp

AOT40 = 0 to 17.7 ppm•h 

Bean, cv. Lit OTC Germany CF, NF, CF-1×, CF-2×: 3 months Pod yield 56 (CF, 2×) 

mean = 1, 14, 15, 32 

Bean , cv. OTC CF, +O3: 63-65 days Pod dry 51, 57 (early and Bush Blue Corvallis, OR SUM06 = 0.0, 75.7 or 68.4 weight season experimeLake 290 ppm•h; AOT40 = 0.0, 50.9 or 

46.4 ppm•h; 7-h mean = 7, 89 or 85 (early and late season experiments) 

Bean , cvs. OTC Raleigh, NC CF, 1.4×: 1 year Pod dry n.s. for TenderetTenderette, 12-h mean = 23, 72 weight for S156 S156 Bean , cvs. OTC Raleigh, NC CF, NF, AA: 2 years Pod dry n.s. for R123 n.R123, 12-h mean = 31, 51, 49 in weight Oregon-91 in 20Oregon-91, year 2000; 25, 46, 47 in year 27 in 2001; 21, 4S156 2001 S156 Corn OTC Beltsville, MD CF, +40: 7-h mean = 20, 70 1 year Grain yield 13 Cotton, cv. OTC Raleigh, NC CF, 1.5×: 1 year Seed-cotton 22 Deltapine 12-h mean = 21, 71 weight Cotton, cv. OTC Raleigh, NC CF, NF, 1.5×, 1 year Seed-cotton 21, 49 (NF, 1.5×Deltapine 12-h mean = 24, 51, 78 weight 

Oat, cv. Vital OTC Ostad, Sweden CF, NF: 7-h mean = 12, 27 1 year Grain yield +2 (n.s.) 

AX9-201
Table AX9-16 (cont’d). Summary of Selected Studies of Ozone Effects on Annual Speci

O3 Concentration Response 

(units are ppb unless Variable (Decrease from 

Species Facility Location otherwise specified)1 Duration lowest, %) 

Potato, cvs. OTC Raleigh, NC CF, NF, 1.5×: 1 year Tuber yield 15, 31 for NorlanSuperior, 12-h mean = 15, 45, 80 NF and 1.5x; 11 Dark Red Superior in 1.5x Norland Potato2 OTC 6 sites N. Europe AOT40 = 6-27 ppm•h 2 years (1 year Tuber yield 4% average for 

at 2 sites) experiments 

Rape, oilseed Open Air Northumberland, AA, +O3: 7-h mean for 17 17 days in fall, Seed yield 14 

UK days Aug.-Sept = 30, 77, for overwinter, 32 

32 days in days in spring May-June = 31, 80 

Rice, cvs. OTC Japan CF, 1×, 1.5×, 2×, 2.75×: 7-h 3 years Grain yield 3 to 10 at 40 ppb Koshi-hikari, mean = 13.5-93.4 Nippon-bare Soybean OTC Beltsville, MD CF, +40: 7-h mean = 25, 72 2 years Seed yield 25 Soybean, cv. OTC Raleigh, NC CF, 1.5×: 12-h mean for 3 3 years Seed yield 41 Essex years = 23, 82 Soybean, cvs. OTC Maryland CF, +40: 7-h mean = 24 and 2 years Seed yield 10, 32 (2 cvs.) Forrest, Essex 24, 63 and 62 for each year Soybean, cv. OTC Raleigh, NC CF, NF, 1.5×: 1 year Seed yield 16, 37 (NF, 1.5×Essex 12-h mean = 20, 50, 79 

Soybean, cv. OTC Raleigh, NC CF, NF, 1.5×: 1 year Seed yield 15, 40 (NF, 1.5×Essex 12-h mean = 18, 42, 69 

Soybean, cv. OTC Raleigh, NC CF, NF, 1.5×: 1 year Seed yield 22, 36 (NF, 1.5×Holladay 12-h mean = 18, 42, 69 

Soybean, cv. OTC Raleigh, NC CF, NF, 1.5×: 1 year Seed yield +46, +4 (NF, 1.5NK-6955 12-h mean = 18, 42, 69 

Soybean, 3 OTC Raleigh, NC CF, NF, 1.5×: 3 months Seed yield At ambient = +1cvs. 12-h mean = 14, 36, 64 16 for 3 cvs. 

AX9-202
Table AX9-16 (cont’d). Summary of Selected Studies of Ozone Effects on Annual Speci

AX9-203

O3 Concentration Response 

Species Facility Location 

(units are ppb unless 

otherwise specified)1 Duration Variable 

(Decrease from lowest, %) 

Soybean, 3 OTC Raleigh, NC CF, NF, 1.5×: 12-h mean = 4 months Seed yield At ambient = 17cvs. 24, 49, 83 18 (3 cvs.) Soybean, cv. OTC Raleigh, NC CF, NF, 1.5×: 4 months Seed yield 11, 22 (amb., 1.5Essex 12-h mean = 20, 50, 79 

Soybean, cvs, OTC Beltsville, MD CF, NF+: 7-h mean = 24, 58 134 days Seed yield Essex = +11 (n.sEssex, Forrest Forrest = 21 Soybean, cv. OTC Raleigh, NC CF, 1.5×: 164 d (1999), Seed yield 24 (1999), 39 Essex 12-h mean = 24, 75 (1999); 149 d (2000) (pots in 2000), 

22, 67 (2000) 41 (ground in 20

Soybean, cv. FACE Urbana- AA, +O3 in 2002 2 expts of 1 year Seed yield 15, 25 (2002, 2093B15 Champaign, IL 8-h mean = 62, 75; 

SUM06 = 36, 70 ppm•h; AOT40 = 22, 42 ppm•h; in 2003 8-h mean = 50, 63; SUM06 = 22, 53 ppm•h; AOT40 = 14, 34 ppm•h 

Timothy OTC Sweden AOT40 = 10, 20, 340; 1 year Biomass 58 

12-h mean = 20, 152 

Watermelon OTC Spain CF (O3 = 0), NF in 1988 2 expts of 1 year Fruit yield 19, 39 (2 expts) 

AOT40 = 5.96 ppm•h, SUM06 = 0.29 ppm•h, in 1989 AOT40 = 18.92 ppm•h, SUM06 = 4.95 ppm•h 

Wheat1, cv. OTC 8 sites in 12-h mean (SD) low = 26.3 13 studies of 1 Grain yield 13 (n.s.) 

Minaret N Europe (12.2), 12-h mean (SD)-high year each 

= 51.4 (18.3) AOT40 mean (SD) low = 6.18 (8.54) ppm•h, AOT40 mean (SD) high = 28.23 (23.05) ppm•h 
Table AX9-16 (cont’d). Summary of Selected Studies of Ozone Effects on Annual Speci

AX9-204

O3 Concentration 

Species Facility Location 

(units are ppb unless 

otherwise specified)1 Duration Variable 

Response (Decrfrom lowest, %

Wheat1 OTC Sweden AOT40 0 to 15 ppm•h 7 years Grain yield 23 at AOT40 = 

ppm•h 

Wheat, cv. OTC SE Ireland CF, +50: 3 h/day, Grain yield 53 Promessa 12-h total = 5.6, 32.6 ppm•h 5 days/week, 

7 weeks 

Wheat, cv. OTC SE Ireland CF, +25: 6 h/day, Grain yield +17 Promessa 12-h total = 6.2, 33.4 ppm•h 5 days/week, 

7 weeks 

Wheat, cv. OTC SE Ireland CF, +25, +50: +25 = 6 h/day, Grain yield Amb + 25 = 3 (nPromessa 12-h total = 6.7, 34, 34 5 days/week, Amb + 50 = 17 

ppm•h +50 = 3 h/day, 

5 days/week, both 7 weeks 

Wheat, cvs. OTC Beltsville, MD CF, +40: 2 years Grain yield 20 Massey, 7-h mean = 19, 20 and 61, 65 Saluda (2 years) Wheat, cv. OTC Germany 8-h mean = 5.9, 61.2, 92.5 1 year Grain yield 14, 40 Turbo (mid, high O3) 

Wheat, cv. OTC Germany 8-h mean = 4.7, 86.4 1 year Grain yield 20 Turbo Wheat, cv. OTC Germany 7-h mean = 5, 41, 73 1 year Grain yield 35 Turbo Wheat, OTC Raleigh, NC 12-h mean = 27, 47, 90 2 months Grain yield 5 (n.s.) winter, 8 cvs. Wheat, OTC Raleigh, NC 12-h mean = 22, 38, 74 2 months Grain yield 16 (n.s.) winter, 8 cvs 
Table AX9-16 (cont’d). Summary of Selected Studies of Ozone Effects on Annual Speci

O3 Concentration 

Species Facility Location 

(units are ppb unless 

otherwise specified)1 Duration Variable 

Response (Decrfrom lowest, %

Wheat, cv. OTC Finland 1992: 12-h mean = 14, 30, 2 years Grain yield At highest O3 = Drabant 61; AOT40 = 16.3, 34.8, each year 

54.6 ppm•h. 1993: 

12-h mean = 9, 21, 45; AOT40 = 10.2, 24.8, 40.6 ppm•h 

Wheat, cv. Open Air Northumberland, AOT40 for Mar to 1 year, Grain yield 13 Riband UK Aug 93 = 3.5, 6.2 ppm•h overwinter 

1 Values for ambient or NF treatments are indicated in bold. 2 Bold indicates that multiple experiments (more than just 2 years at a single site) were included in the analysis. 

AX9-205
Table AX9-17. Summary of Selected Studies of the Effects of Ozone on Perennial Herbaceou

AX9-206

O3 Concentration 

Species Facility Location 

(units are ppb unless otherwise 

specified)1 Duration Variable 

Response (from lowe

Alfalfa, cvs. OTC Quebec, 12-h mean: 1991 = 6, 39, 49, 110; 3 months Biomass For NF: AApica, Team Canada 1992 = 0, 34, 42, 94 in each of 21; Team = 

2 years (n.s.) 

Bahia grass OTC Auburn, AL 12-h mean = 22, 45, 91 24 weeks Biomass at ambient O3 for 34, 29 (n.s.

1st, 2nd cutting of early 9 (n.s.) and late season plantings. 

Bent grass OTC United AOT40 = 0.8-15.0 ppm•h 8 h/day for Biomass, in competition 8 (uncut), +(Capillaris sp.) Kingdom 3 months with 3 other spp. Total 

biomass in uncut pots, aboveground biomass in cut pots (cut every 14 days). 

Blackberry Large Alabama 1994: AOT40 = 2-112 ppm•h, 7 months in Percent canopy cover +124 for c

OTC SUM06 = 1-162 ppm•h, 1994, 6 (grown in old field for biomas

1995: AOT40 = 3-83 ppm•h, months in community), biomass ripe ripe fruit nu

SUM06 = 0-132 ppm•h 1995 fruit number. sig. chamb

Clover, white Ambient MA, OR, NC, SUM06 for 6-h/day = 10.2-39.4 2 growing Biomass ratio 4 at 6-h SU

air CA (2 sites) ppm•h, AOT40 for 12-h/day = seasons (sensitive/resistant) 39.4 ppm•h

and VA 0.6-50.1 ppm•h AOT40 = 5

Clover, white Ambient 14 European AOT40 for 28 d = 0-12 ppm•h 3 growing Biomass ratio 5 at AOT40 

air sites seasons sensitive/resistant) days = 0.9-

Clover, white OTC United AOT40 = 0.8-15.0 ppm•h 8 h/day for Biomass, in competition 18 (uncut), 

Kingdom 3 months 3 other spp. Total biomass 

in uncut pots, aboveground biomass in cut pots (cut every 14 days). 

Clover, white OTC Switzerland CF, NF, NF+. NF++: 3.5 months/ Biomass, in managed 24, 26, 52 and red 12-h mean = 21, 39, 47, 65 year for pasture 

2 years 

Clover, white, OTC Italy CF, NF: AOT40 = 0.1; 2 months Biomass 20 cv. Menna 8.9 ppm•h, 7-h mean = 24, 53 
Table AX9-17 (cont’d). Summary of Selected Studies of the Effects of Ozone on Perennial Herba

AX9-207

O3 Concentration 

Species Facility Location 

(units are ppb unless otherwise 

specified)1 Duration Variable 

Response (from lowe

Fescue, red OTC United AOT40 = 0.8-15.0 ppm•h 8 h/day for Biomass, in competition + 30 (uncu

Kingdom 3 months with 3 other spp. Total (cut) 

biomass in uncut pots, aboveground biomass in cut pots (cut every 14 days). 

Lespedeza, OTC Auburn, AL CF, NF, 2×: 12-h mean = 23, 40, 10 weeks Biomass n.s. Sericea 83, SUM06 = 0.2, 9.1, 61.0, 

AOT40 = 0.6, 7.0, 39.8 

Little bluestem OTC Auburn, AL CF, NF, 2×: 12-h mean = 23, 40, 10 weeks Biomass n.s. 

83 ppb, SUM06 = 0.2, 9.1, 61.0, AOT40 = 0.6, 7.0, 39.8 

Phleum alpinum OTC Sweden AOT40 = 0.01, 0.02, 0.34 ppm•h; 1 year biomass 87 

12-h mean = 20, 152 

Speedwell, OTC United AOT40 = 0.8-15.0 ppm•h 8 h/day for Biomass, in competition 14 (uncut), Germander Kingdom 3 months with 3 other spp. Total 

biomass in uncut pots, aboveground biomass in cut pots (cut every 14 days). 

Strawberry OTC United 8-h mean = 27, 92; 69 days Fruit size, yield Size = 14, 

Kingdom AOT40 for +O3 = 24.59 ppm•h yield = (n.s

Sumac, winged OTC Alabama SUM06 = 0 to 132 ppm•h 6 months Percent canopy cover 95 

(grown in old field community) 

Timothy OTC Sweden CF, NF, CF+: 1 year Biomass n.s. in NF, 

AOT40 = 0.0, 1.3, 20.3 ppm•h; 

12-h mean = 20, 68, 152 

1 Values for ambient or NF treatments are indicated in bold.2 Bold indicates that multiple experiments (more than just 2 years at a single site) were included in the analysis.
Table AX9-18. Summary of Selected Studies of Ozone Effects on Deciduous Trees and Sh

Species Age Facility Location 

O3 Concentration 

(units are ppb unless 

otherwise specified)1 Duration Variable 

Respo(Decrlowes

Ash, European 

OTC Hampshire, UK NF, NF+: Mean = 17.7, 

44.1; AOT40 for 24h = 1.9, 59.9 ppm•h 

3 years for day 100 to day 162 

Growth and biomass of organs 

n.s. 

Ash, European 

Seedling OTC Switzerland 0.5×, 0.85×, 1×, 0.5×+30: 

AOT40 = 0.1, 3.4, 7.1, 19.7 ppm•h 

5 months Biomass 26 at 

0.5× + 

Aspen Cutting, 

Seedling 

OTC Michigan CF, 1×, 2×: 

3 months 7-h mean for 1990 =7, 43, 63 ; for 1991 (square wave exposure) = 11, 45, 66 

98 days Total 

biomass of 3 clones and seedlings 

For 19at 1x f(mean for seeFor 19at 1x f(mean 

Aspen Cutting OTC Michigan CF, 1×, 2×: 

SUM00 = 11, 58, 71 ppm•h 

98 days Total 

biomass 

25-38 

Aspen Cutting FACE Wisconsin Ambient, 1.5×: 4-year 

ambient 12-h mean = 34.6, 36.9, 36.0, 36.6; 4-year 1.5× 12-h mean = 54.5, 51.1, 48.9, 52.8 

7 years (only 4 years of O3 

data reported) 

Volume (d2*h) 

21 aft14 aftat 1.5x 

Aspen Cutting Large OTC New York 1×, 1.7×, 3×: 

SUM06 = 1, 20, 62 ppm•h; 

9-h mean = 40, 74, 124 

92 days Shoot 

biomass 

14, 25 clones 

Aspen First year OTC Pennsylvania 8-h mean = 39, 73 11 weeks Biomass 14-30 

N trea

Beech, European 

OTC Switzerland 0.5×, 0.85×, 1×, 0.5×+30: 

AOT40 = 0.1, 3.4, 7.1, 19.7 ppm•h 

5 months Biomass 6 at 1×

0.5×+

Beech, European 

Seedling OTC Belgium CF, NF, +30: 

8-h mean = 5. 29, 33; 

AOT40 = 0. 4.06, 

8.88 ppm•h 

23 April ­30 Sept 

Growth No eff

AX9-208
Table AX9-18 (cont’d). Summary of Selected Studies of Ozone Effects on Deciduous Trees a

AX9-209

Species Age Facility Location 

O3 Concentration 

(units are ppb unless 

otherwise specified)1 Duration Variable 

Re(Dlow

Beech, European Seedling Growth chamber Belgium CF, CF+40, CF+100: 

Sum0 = 0.48, 8.93, 25.14 ppm•h; AOT40 of 

NF+100 = 13.91 ppm•h; 

uptake = 159, 2965, 7095 mol m2 

7 episodes of 5 days 

Biomass, diameter 

No 

Beech, European 0-3 years OTC Switzerland AOT40 for 24 h/days = 

4-73 ppm•h 

1-3 years Total 

biomass 

20 fopp

Beech, Japanese 4 years Growth chamber Japan CF, +60 ppb for 7 h/day 156 days Total 

biomass 

19 

Birch, silver Sapling FACE Finland AOT40 = 1, 15 ppm•h; 

7-h mean = 26, 40 

5 years Biomass 34 

fo

Birch, silver Sapling OTC Sweden NF, NF+, NF++, daylight 

mean 1997 = 29, 37, 54; 1998 = 25, 42, 71 ppb; 

AOT40 1997 = 2.4, 6.9, 

35.1; 1998 = 0.6, 19.6, 

2 years Total 

biomass 

Ton.s22 ro30 

74.7 ppm•h 

Birch, [B. pubescens] 

Seedling Chamber in 

glasshouse 

Norway AOT40 = 0.1, 2.5, 7.1, 

7.4, 17.8, 19.8 ppm•h 

40 days Biomass Si

in AOpp7.

Birch, paper Sapling FACE Wisconsin Ambient, 1.5x: 4 y 

Ambient 12-h mean = 34.6, 36.9, 36.0, 36.6; 4 y 1.5x 12-h mean = 54.5, 

7 years (only 4 years of O3 

data reported) 

Volume (d2*h) 

No 

51.1, 48.9, 52.8 

Cherry, black 2 years OTC Norris, TN CF, 1×, 2×: 

7-h mean = 21, 50, 97 

April to August Biomass No 
Table AX9-18 (cont’d). Summary of Selected Studies of Ozone Effects on Deciduous Trees a

Species Age Facility Location 

O3 Concentration 

(units are ppb unless 

otherwise specified)1 Duration Variable 

Re(Dlow

Cherry, black Seedling OTC GSMNP 1 CF, 1×, 1.5×, 2×: 

SUM06 = 0-40.6 ppm•h, 

AOT40 = 0.03-28.3 

76 days Biomass n.s

1.5

ppm•h 

Cherry, black Seedling OTC GSMNP 1 CF, 0.5×, 1×, 1.5×, 2×: 

SUM06 = 0-53.7 ppm•h; 

AOT40 = 0-40.4 ppm•h 

140 days Biomass n.s

1.5

Cherry, black 1 year OTC Delaware, OH CF, 0.5, 1, 1.5, 2×: 

SUM00 in 1990 = 17-107 ppm•h, in 1991 = 31-197 

ppm•h 

2 years (in 1990 for 3.5 months, 1991 for 4 months) 

Total biomass 

no an2× 

Cherry, black Seedling OTC Pennsylvania CF, 0.75×, 0.97×: 

7-h mean = 39 to 46, SUM06 = 0-10.34 ppm•h 

3 years for 17 weeks 

Total biomass 

6 a0.9

Cottonwood, Eastern 

Cutting Ambient, in buried 

pots with irrigation 

In and within 100 km of New York City, NY 

12 h mean = 23-49 ppb 2 months each 

year, 3 10-years experiments 

Total biomass 

3338 co23 

Grape 3 years OTC Austria CF, 1×, +30, +50: 

(AOT40 = 0-50 ppm•h 

2 years (preflowering, past harvest) 

Fruit yield Ca

AO27 

Oak Seedling OTC Hampshire, UK NF, NF+: 

Mean = 17.7, 44, AOT40 for 24h = 1.9, 59.9 ppm•h 

3 years for day 100 to day 162 

Biomass of organs 

30 bio

Oak, red Seedling OTC Norris, TN SUM06 for 3 years = 0, 

29, 326 ppm•h; SUM00 

for 3 years = 147, 255, and 507 ppm•h 

3 years Total 

biomass 

n.s

AX9-210
Table AX9-18 (cont’d). Summary of Selected Studies of Ozone Effects on Deciduous Trees a

AX9-211

O3 Concentration Re

Species Age Facility Location 

(units are ppb unless 

otherwise specified)1 Duration Variable 

(Dlow

Oak, red 30 years OTC Norris, TN SUM06 for 3 years = 0, 3 years Stem n.s

29, 326 ppm•h; SUM00 increment redfor 3 years = 147, 255 and ph507 ppm•h 

Maple, red 2 years OTC Norris, TN CF, 1×, 2× : April to August Biomass No 

7-h mean = 21, 50, 97 

ppm•h 

Maple, sugar 1 year OTC Delaware, OH CF, 0.5, 1.5, 2×: 2 years (in 1990 Total n.s

UM00 in 1990 = 17 to 107 ppm•h, in 1991 = 31 to 

for 3.5 months, 1991 for 4 

biomass tre

197 ppm•h months) 

Maple, sugar Seedling Large OTC Ithaca, NY CF, 1×, 1.5×, 2×: 3 years 3 years for 134, Biomass No 

SUM00 = 148 to 591 128, 109 days ppm•h; daytime mean = 19.7 to 40.7 

Maple, sugar Seedling Large OTC Ithaca, NY 1×, 1.7×, 3×: 3 years for 109, Total Fo

3 years 12-h mean = 38, 143, 116 days biomass 3×

69, 117 low 

anlig

Maple, sugar Sapling FACE Wisconsin Ambient, 1.5×: 4 y 7 years (only 4 Volume 18 

Ambient 12-h mean = y of O3 data (d2*h) 34.6, 36.9, 36.0, 36.6; 4y reported) 1.5x 12-h mean = 54.5, 51.1, 48.9, 52.8 

Plum, Casselman Sapling Large OTC Fresno, CA CF, 1×, +O3: 4 years Stem Fr

12-h mean = 31, 48, 91 increment, 1×

fruit yield +O

Poplar, black Seedling OTC Belgium CF, NF, +30: 23 April - 30 Diameter, 29 

8-h mean = 5, 29, 33; Sept height in 

AOT40 = 0, 4, 8.9 ppm•h eff
Table AX9-18 (cont’d). Summary of Selected Studies of Ozone Effects on Deciduous Trees a

AX9-212

O3 Concentration Re

Species Age Facility Location 

(units are ppb unless 

otherwise specified)1 Duration Variable 

(Dlow

Poplar, hybrid 0 year FACE Finland AOT40 = 0.07, 1.6 2 months Biomass, n.s(P. tremuloides × ppm•h; 7-h mean = 30, 38 height 6 fP. tremula) Poplar, hybrid Cutting OTC Michigan CF, CF+100: 60 days Total 46 

12, 48 ppm•h biomass of 

Poplar, hybrid Cutting OTC Michigan CF, CF+100: 60 days Total 46 

12, 48 ppm•h biomass of 

Yellow-poplar 1 year OTC Delaware, OH CF, 0.5×, 1.5×, 2×: 2 years (in 1990 Total No 

SUM00 in 1990 = 17 to 107 ppm•h, in 1991 = 31 

for 3.5 months, 1991 for 4 

biomass 

to 197 ppm•h months) 

Yellow-poplar 1- 7 year Large OTC Delaware, OH CF, 0.5×, 1.5×,: 5 years Total No 

SUM00= 145, 583, 861 ppm•h ; SUM06 = 0.3, 

biomass 

228.7, 661.8 ppm•h over 5 years 

1 Values for ambient or NF treatments are indicated in bold. 2 Bold indicates that multiple experiments (more than just 2 years at a single site) were included in the analysis.
Table AX9-19. Summary of Selected Studies of Ozone Effects on Evergreen Trees and Sh

O3 Concentration Response 

(units are ppb unless (decrease from 

Species Age Facility Location otherwise specified)1 Duration Variable lowest, %) 

Fir, Seedling Open air British 12 trts: 1988 = 92 Second flush Calculated 55 Douglas Columbia 12-h mean 1988 = 18-41; days; 1989 = biomass highest expos

1989 = 27-66 101 days 

Hemlock, Seedling OTC GSMNP1 , CF to 2×: 3 years Biomass No effect eastern TN SUM06 = 0.2-108.1 ppm•h, 

AOT40 = 0.2-63.9 ppm•h 

Pine, 12 weeks OTC Oak Ridge, CF to 2×: 3 months Biomass 14 in 1× (avg floblolly TN 24-h summer = 74, 137, 169, families) 

206, 284 ppm•h 

Pine, 1 year OTC Alabama 1994: 2 years, April Dry weight, n.s. loblolly AOT40 = 2-112 ppm•h, to October height, 

SUM06 = 10-162 ppm•h, diameter 1995: AOT40 = 3-83 ppm•h, SUM06 = 0-132 ppm•h 

Pine, 3 years OTC Raleigh, NC Ambient, CF, NF, 1.5×., 5 months Height, No effect on stloblolly 2.5×: 12-h mean = 54, 29, 47, diameter, height or diam

76, 98 needle length decrease in ne

length 

Pine, 4 weeks Large Auburn, AL CF, 1×, 2×: 2 12-week Shoot biomass, Shoot = 15 in loblolly OTC 12-h mean = 13, 47,, 98 experiments root biomass, in 2x in both y

ppm•h in 1998; 12, 44, 97 foliar injury Root = 26 in 2

both years; Fosig. greater in 1999 and 2x in years. 

Pine, Seedling OTC Corvallis, For CF 12-h SUM06 = 0 3 years: 16 Total biomass No effect withponderosa OR ppm•h; for +03 12-h weeks, 16 grass, 25 with 

SUMO6 = 22, 27, 31 ppm•h weeks, 14 present for 3 years weeks 

AX9-213
Table AX9-19 (cont’d). Summary of Selected Studies of Ozone Effects on Evergreen Trees an

O3 Concentration Response 

(units are ppb unless (decrease from 

Species Age Facility Location otherwise specified)1 Duration Variable lowest, %) 

Pine, 39 to 45 Ambient CA 24-h mean for 3 weeks late Ambient Fine and 85 at most polponderosa years gradient July and early August for gradient medium root site. 

1993 and 1994 = 70-90 ppb growth 

Pine, Seedling OTC CA CF, 1×, 2×: Total biomass n.s. ponderosa 24-h mean 

approx. 20, 60, 120 

Pine, OTC Hampshire, NF, NF+: 3 years for 62 Total biomass 15 Scots UK Mean = 17.7, 44.1; 24-h days 

AOT40 = 1.9, 59.9 ppm•h 

Pine, 3-6 years Free air Finland Amb, +O3: 3 years Biomass No effect Scots AOT40 = 0-1, 2-13 ppm•h 

Pine, Seedling OTC Switerzland 0.5×, 0.85×, 1×, 0.5×+30: 5 months Biomass 14 at 1×, 22 at Scots AOT40 = 0.1, 3.4, 7.1, 0.5×+30 

19.7 ppm•h 

Pine, 3 years OTC Finland CF, 1×, +O3: 2 years (4 Biomass No effect Scots 24 h AOT40 for months each) 

2 years = 0.5, 6, 73 ppm•h 

Pine, 3 years Free air Finland 1×, +O3: 3 years Root and shoot 32 only for rooScots 24 h AOT40 for 2 years = 2, (3-4 months biomass biomass in hig

37 ppm•h each) treatment 

Pine, Seedling OTC GSMNP 2 , CF to 2×: 3 years Biomass Slight decrease Table TN SUM06 = 0.2-116.4 ppm•h, older needle mMountain AOT40 = 0.2-71.7 ppm•h only 

Pine, Seedling OTC GSMNP 2 , CF to 2×: 1-2 years (3 Biomass No effect Virginia TN SUM06 = 0.1-32.8, 47.9, expts) 

56.2 ppm•h; AOT40 = 0.1-19.3, 27.1, 34.4 ppm•h 

AX9-214
Table AX9-19 (cont’d). Summary of Selected Studies of Ozone Effects on Evergreen Trees an

AX9-215

O3 Concentration Response 

Species Age Facility Location 

(units are ppb unless 

otherwise specified)1 Duration Variable 

(decrease from lowest, %) 

Sequoia, 125 years Branch California 0.25×, 1×, 2×, 3×: 61 days Branch growth No effect giant chamber 24-h SUM00 approx. 10, 85, 

180, 560 ppm•h 

Spruce, 4-7 years Open air Finland Amb, +O3: 3 years Biomass No effect Norway AOT40 = 0 1, 2-13 ppm•h 

Spruce, 3-7 years OTC Sweden CF, 1.5×: 4 years Total biomass 8 Norway 12-h mean for 4 years = 12, 

44; AOT40 = 2, 23 ppm•h 

Spruce, Seedling OTC Switerzland 0.5×, 0.85×, 1×, 0.5×+30: 5 months Biomass n.s. Norway AOT40 = 0.1, 3.4, 7.1, 

19.7 ppm•h 

Spruce, 0-3 years OTC Switzerland AOT40 for 24 h for 1 to 1-3 years Total biomass n.s. Norway 3 years = 22 to 63 ppm•h 

Spruce, 3 to 7 OTC Sweden CF, 1×, 1.5×: 4 years Biomass 5.3 at 1.5x Norway years AOT40 daylight for 

4 years = 1, 16, 79 ppm•h 

Spruce, Sapling Large Ithaca, NY CF, 1×, 1.5×, 2×: 4 years: Biomass No effect red OTC total for 4 years = 211 to 98-124 

569 ppm•h; days/year daytime mean = 21-71 

1 Values for ambient or NF treatments are indicated in bold. 2 Great Smoky Mountains National Park. 
(Section AX9.4). Even when some of these aspects of O3 exposure can be elucidated, it is 

difficult to apply this knowledge to developing exposure-response relationships based on data in 

the scientific literature, because O3 exposure is often reported only in the form of a summary 

index such as a 12- or 24-h mean, SUM06, or AOT40. 

Table AX9-13 presents summaries of ambient O3 exposure patterns in the United States for 

1982 to 1991 for several indices including the 7-h mean and SUM06. More recent summaries 

for the entire United States for these indices are not available, but Table AX9-20 summarizes 

more recent data for the central and eastern United States. As shown in Table AX9-20, from 

1989 to 1995, mean 12-h 3-month SUM06 values (in ppmAh) at 41 rural sites in the Clean Air 

Status and Trends Network were 31.5 for the Midwest, 18.9 for the Upper Midwest, 33.2 for the 

Northeast, 13.2 for the Upper Northeast (New Hampshire, Maine), 34.5 for the South-Central, 

and 19.2 for the Southern Peripheral subregions (Baumgardner and Edgerton, 1998). These 

results are important because these sites were selected to represent rural areas, while many other 

monitoring sites represent urban or suburban areas. For these same subregions, W126 values 

ranged from 12.8 to 25.6 ppmAh. From 1989 to 1995, O3 concentrations decreased about 5% for 

daily and 7% for weekly values for most of these sites, after adjusting for meteorological 

conditions (Holland et al., 1999). These trends were statistically significant at about 50% of the 

sites (p # 0.05). However, because the trend analysis was intended to examine the efficacy of O3 

emissions controls, the trends were adjusted for meteorological conditions. Thus, they do not 

reflect the actual trends in O3 exposure over time. 

AX9.5.4 Effects of Ozone on Annual and Biennial Species 

Much of the research on short-lived species during the last decade has been conducted in 

Europe. Several European studies have focused on wheat with an emphasis on developing 

critical levels as discussed in Section AX9.4.3 and reviewed briefly below in Section AX9.5.4.5. 

An extensive search of the literature was performed using several electronic databases to 

identify scientific articles containing quantitative information on both the amount of O3 exposure 

and its effects on vegetation. Greater emphasis is placed on studies with longer duration with O3 

exposure concentrations and environmental conditions that were as similar as possible to 

ambient conditions. Many of the studies reviewed herein were conducted in OTCs. In the 

United States, nearly all of such studies have used the type of OTC developed by Heagle et al. 

AX9-216
Table AX9-20. Ozone Exposures at 35 Rural Sites in the Clean Air Status and Trends 

Network in the Central and Eastern United States From 1989 to 1995 

SUM06 SUM06 W126 W126 Max. Max. 

Subregion 

12-h, 3-Month 

Mean 

12-h, 

3-Month SD 

3-Month 

Mean 

3-Month 

SD 

8 h > 80 ppb(n) 

Mean 

8 h > 80 ppb(n) 

SD 

Midwest 31.5 10.2 25.1 7.7 13.8 10.6 

Upper 18.9 8.5 16.0 5.9 5.6 5.6 

Midwest 

Northeast 33.2 11.9 26.6 9.5 15.8 12.2 

Upper 13.2 8.6 12.8 6.5 3.3 5.5 

Northeast 

South 34.5 16.6 25.6 11.5 7.1 10.0 

Central 

Southern 19.2 7.6 15.2 5.4 1.9 1.6 

Periphery 

Units for SUM06 and W126 are ppmAh.

Source: Baumgardner and Edgerton (1998).

(1973). For the few studies in the U.S. that used other types of OTCs, they are described briefly 

in the text. In Europe, a wide variety of styles of OTCs have been used. See Section AX9.1 for 

further information about the use of OTCs. The emphasis in this subsection is on quantifying 

exposure-response relationships for annual plants, with a focus on the response of above-ground 

biomass and yield of species grown as crops or occurring as native or naturalized species in the 

United States. Emphasis is placed on studies not included in the 1996 AQCD (U.S. 

Environmental Protection Agency, 1996), including a few studies published prior to 1996. 

However, an attempt is made to compare the results of these more recent studies of individual 

species to those reviewed in the 1996 AQCD. 

AX9.5.4.1 Effects on Growth, Biomass, and Yield of Individual Species 

Most research on the effects of O3 on herbaceous species has evaluated growth, biomass, or 

yield of commercial portions of crop or forage species. It is well established that reproductive 

organs such as seeds may be particularly sensitive to injury or biomass reductions due to O3, as 

reviewed by (Black et al., 2000). As discussed in Section AX9.3, numerous analyses of 

AX9-217
experiments conducted in OTCs and with naturally occurring gradients demonstrate that the 

effects of O3 exposure vary depending on the growth stage of the plant. Plants grown for seed or 

grain are often most sensitive to exposure during the seed or grain-filling period (Lee et al., 

1988; Pleijel et al., 1998; Soja et al., 2000; Younglove et al., 1994), whereas plants grown for 

biomass production, such as alfalfa, may be sensitive throughout the growth period (Younglove 

et al., 1994). However, because different species are sensitive during different periods of their 

growth and, because planting or germination dates vary throughout large regions even for a 

single species, no single phenological weighting scheme can appropriately and practically 

represent all vegetation in all locations in the United States. For natural populations, reductions 

in seed yield might be particularly important if subsequent seedling establishment is 

compromised by O3. 

Green beans (cv. Pros) were grown in pots in OTCs in the Netherlands for 62 days and 

exposed to 6 treatments consisting of constant O3 additions to CF chambers (see Section AX9.1) 

for 9 h/day (Tonneijck and Van Dijk, 1998). Bean yield response to O3 was nonlinear, with an 

apparent threshold near the CF30 (charcoal filtered with a constant addition of 30 ppb O3) 

treatment with a 9-h mean O3 concentration of 28 ppb and an AOT40 value of 0.1 ppm"h). Yield 

was reduced by 29% at a 9-h mean value of 44 ppb corresponding to an AOT40 value of 3.6 

ppmAh (Table AX9-16). Beans were grown in pots in OTCs for 3 months with the following O3 

treatments: CF, nonfiltered (NF), CF with O3 added up to ambient, and CF with 2×-ambient O3 

(Brunschon-Harti et al., 1995). Ozone reduced pod mass by 56% with a mean concentration of 

32 ppb in the 2× ambient treatment as compared with 1 ppb in the CF treatment (daily averaging 

time not reported). A second treatment factor in this experiment was addition of EDU, as 

discussed below under the heading “Studies Using Ethylene Diurea as a Protectant.” Bush beans 

(cv Bush Blue Lake 290) were grown in pots in OTCs in Corvallis Oregon for 63 or 65 days in 

two experiments, one from May to July and one from August to October (Tingey et al., 2002). 

Plants were exposed to either CF air or CF air with above-ambient O3 with temporal frequency 

and exposure dynamics typical of the Midwestern United States, with SUM06 values of 0.0 for 

the CF treatment and 75.7 or 68.4 ppmAh for the two experiments. Ozone exposure reduced pod 

dry weight by 51 and 57% in the two experiments. The sensitive cultivar S156 and the more 

resistant cultivar Tenderette were grown in pots in OTCs in Raleigh, NC at either CF (12-h mean 

of 23 ppb) or 1.4× (12-h mean of 72 ppb) (Heagle et al., 2002). At final harvest, the dry weight 

AX9-218
of the sensitive cultivar was reduced 90% by the O3 treatment, but the more resistant cultivar 

was not reduced. In two additional OTC experiments during 2000 and 2001 in Raleigh, NC, the 

sensitive cultivar S156, the moderately sensitive cultivar Oregon-91, and the more resistant 

cultivar R123 were grown in CF, NF, and ambient air treatments (Fiscus et al., 2005). For the 

NF treatment of the two experiments, the yield of S156 was reduced 21% and 45%, that of 

Oregon-91 by 27% only in 2001, while R123 was not significantly reduced. These yield 

reductions are greater than those previously reported in four similar studies summarized in the 

1996 AQCD (Table 5-25 of U.S. Environmental Protection Agency, 1996). Greater sensitivity 

in the more recent experiments may be due to cultivar differences or other differences in 

experimental protocols. 

In a study with OTCs on silty loam soil in Beltsville, MD, corn yield was reduced by 13% 

with exposure to a 7-h mean concentration of 70 ppb O3 compared to a CF treatment with a 7-h 

mean concentration of 20 ppb (Mulchi et al., 1995; Rudorff et al., 1996a,c). In this study, 

different amounts of O3 were added above ambient levels for 5 days follows: 20, 30, 40, 50, 

60 ppb, except that O3 was not added to exceed a total concentration of 120 ppb (Rudorff et al., 

1996a,c). 

In two studies conducted in Raleigh, NC, cotton (cv. Deltapine 51) was grown in pots and 

exposed to CF and 1.5× (nonfiltered, see Section AX9.1) O3 in one year, and CF, NF, and 

1.5×-ambient O3 in the second year, with ambient and elevated CO2 concentrations (Heagle 

et al., 1999b) (Table AX9-16). In the first year, yield decreased by 22% with 1.5×-ambient O3 

(12-h mean value of 71 ppb). In the second year, yield decreased by 21% and 49% with 

exposure to ambient or 1.5× ambient O3 (12-h mean values of 51 and 78 ppb, Table AX9-16). 

Increased CO2 levels prevented or reduced this yield suppression (Heagle et al., 1999b). These 

yield reductions are similar to those reported previously in four similar studies summarized in 

the 1996 AQCD (Table 5-25 of U.S. Environmental Protection Agency, 1996). 

In a study of oats in OTCs in southern Sweden, exposure to ambient (NF) O3 did not affect 

grain yield (Pleijel et al., 1994a). Ambient O3 concentration expressed as a 7-h mean was 

27 ppb, with only 1 h greater than 80 ppb and none above 90 ppb. 

The interactive effects of elevated O3 and CO2 additions on potato yield (cv Bintje) were 

studied in OTCs at 6 sites in northern Europe (Craigon et al., 2002). Ozone was added to a 

target daily average value of 60 ppb, and AOT40 values across all years and experiments ranged 

AX9-219
from ~6 to 27 ppmAh. The O3 treatment reduced total tuber yield an average of 4.8% with 

elevated O3 treatment across all experiments (Craigon et al., 2002). This total effect was 

statistically significant even though the effects of individual experiments generally were not 

(Craigon et al., 2002), due to the increased statistical power of the pooled analysis. Several 

publications report other aspects of the “CHIP” experiments or present results of individual 

experiments (De Temmerman et al., 2002a,b; Donnelly et al., 2001a,b; Fangmeier et al., 2002; 

Finnan et al., 2002; Hacour et al., 2002; Lawson et al., 2002; Pleijel et al., 2002; Vandermeiren 

et al., 2002; Vorne et al., 2002). Resistant and susceptible cultivars of potato (Superior and Dark 

Red Norland, respectively) were grown for one season in Raleigh, NC and exposed to CF, NF, 

and 1.5× ambient O3 treatments with 12-h mean values of 15, 45, 80 (Heagle et al., 2003). 

Tuber yield was decreased by 15 and 31% for Dark Red Norland in NF and 1.5x treatments, 

but by only 11% in only the 1.5x treatment for Superior. 

The effect of an intermittent constant addition of O3 using a free air exposure system in 

Northumberland, UK was investigated with the oilseed rape cultivar Eurol (Ollerenshaw et al., 

1999). Ozone was added for 6 h/day for 17 days. The ambient treatment had a mean value of 

30 ppb and the O3 addition treatment had a mean of 77 ppb. After overwintering, O3 was added 

for 32 days for 7 h/day between May and June (mean values of 31 and 80 ppb). Yield was 

reduced by 14% despite the lack of any foliar symptoms. 

Field fumigation chambers ventilated with fans on both ends were used to assess effects of 

five O3 treatments on rice over 3 years in Japan (Kobayashi et al., 1994, 1995). All O3 

treatments used CF air, and O3 was added to the 0.5, 1.0, 1.5, 2.0, or 2.75× ambient 

concentration for 7 h/day. Based on a linear regression for the 3 years, yield decreased by 3 to 

10% at a 7-h mean concentration of 40 ppb (Table AX9-16). This decrease is greater than that 

found for rice in earlier studies in California (Kats et al., 1985b), although whether this 

difference is due to differences in cultivars, experimental treatment, or environmental factors 

cannot be determined. 

During 3 years in Beltsville, MD, the soybean cultivars Essex and Forrest were exposed to 

CF air and NF air in OTCs (Chernikova et al., 2000; Robinson and Britz, 2000) with O3 added as 

described for experiments with corn and wheat at Beltsville (Mulchi et al., 1995; Rudorff et al., 

1996c). During 1994 and 1995, as previously found for these cultivars, Essex was less sensitive 

than Forrest, with yield decreases of 10% (n.s.: p > 0.1) compared to 32% for Forrest (p < 0.01) 

AX9-220
(Chernikova et al., 2000). There was no evidence of water stress in this experiment. In 1997, 

the two O3 treatments were CF (7-h mean = 24 ppb) and NF with a constant addition of O3 (7-h 

mean = 58 ppb) (Robinson and Britz, 2000). The yield of Essex was not significantly affected, 

while the yield of Forrest was decreased by 21% (Table AX9-16). 

In a study in Raleigh, NC, the soybean cultivar Essex was grown in pots and exposed to CF 

and 1.5× ambient O3 concentrations during three growing seasons (Fiscus et al., 1997). Over the 

3 years, exposure to an average 12-h mean O3 concentration of 82 ppb reduced soybean yield by 

41% (Table AX9-16). In similar studies also in Raleigh, NC, Essex was exposed to CF, NF, and 

1.5× ambient O3 for two seasons (Heagle et al., 1998b). Yield decreased by 16% and 15% in the 

2 years by ambient O3 (12-h mean values of 50 and 42 ppb), and decreased by 37 and 40% with 

exposure to 1.5× ambient O3 (12-h mean values of 79 and 69 ppb, Table AX9-16). In this same 

experiment in the second year, similar yield reductions were observed for the cultivar Holladay, 

while the growth of cultivar NK-6955 was increased substantially by ambient O3 exposure. 

All three cultivars were grown in the same chambers in this experiment, and the authors 

suggested that NK-6955 plants may have shaded the other cultivars to some extent. 

In a 2-year study using OTCs in Raleigh, NC, the soybean cultivars Coker 6955, Essex, 

and S53-34 were exposed to CF, NF, and 1.5× ambient O3 treatments (Miller et al., 1994). 

Seasonal mean 12-h O3 concentrations ranged from 14 to 83 ppb. As compared to the CF 

treatment, ambient O3 exposure (NF treatment) reduced seed yield by 11 to 18% except for 

Coker 6955 in the first year (1989), which showed a yield increase of 14%. The 1.5× ambient O3 

treatment reduced yield by 32 to 56% in all cultivars in both years. In a similar subsequent 

experiment with the cultivar Essex, exposure to a 12-h mean ambient O3 concentration of 50 ppb 

reduced yield by 11%, while exposure to 79 ppb reduced yield by 22% (Table AX9-16) (Miller 

et al., 1998). In similar experiments with Essex during 1999 and 2000, plants were exposed to 

1.5× ambient O3 in OTCs either in pots or planted in the ground (Booker et al., 2005). Exposure 

to a 12-h mean ambient O3 concentration of 75 ppb reduced yield by 27% in pots and 24% in the 

ground in 1999, and exposure to a 12-h mean ambient O3 concentration of 67 ppb reduced yield 

by 41% in pots and 39% in the ground in 2000 (differences between results in pots and in the 

ground were not statistically significant). 

In a study using a free air exposure system near Urbana-Champaign, IL, the soybean 

cultivar 93B15 was grown for two seasons (2002, 2003) and exposed to ambient ozone and 

AX9-221
elevated ozone approximately 1.23 × ambient (Morgan et al., in press). The maximum 8-hour 

average ozone concentration was 62 and 50 ppb in the control plots for 2002 and 2003 

respectively, and 75 and 63 ppb in the elevated ozone plots. Thus, the effective increase in 

ozone concentration in the elevated plots was 21 and 25% for 2002 and 2003, respectively. In 

2003, a severe hailstorm in July removed many of the leaves in all plots early in the season, but 

the plants recovered and produced new foliage. Compared to the ambient treatment, the elevated 

treatment reduced yield by 15% in 2002 and 25% in 2003. The authors ascribe the greater 

proportional yield decrease seen in the second year despite lower O3 exposure to the interactive 

effects of hail damage and O3 exposure. Because free air exposure systems can only add O3 

above ambient concentrations, it should be noted that yield also might have been decreased in 

the ambient plots, and that the reduction in yield in the elevated O3 treatment might have been 

greater in comparison to a lower background O3 exposure regime. This FACE facility study is 

important, because it confirmed yield reductions reported previously with soybean plants grown 

in OTCs. 

These yield reductions for soybean are generally similar to those reported previously in 

13 similar studies summarized in the 1996 AQCD (Table 5-23 of U.S. Environmental Protection 

Agency, 1996). A meta-analysis of 53 studies of O3 effects on soybean found that at an average 

O3 exposure of 45 ppb, seed yield was decreased by 10% compared to the CF treatment, while at 

70 ppb, seed yield was decreased by 24% (Morgan et al., 2003). The 95% confidence limits of 

these responses based on a bootstrap method did not include a value of zero yield loss. These 

results suggest that seasonal O3 concentration patterns that occur in some years throughout many 

parts of the U.S. can reduce soybean seed yield. 

A reanalysis of 7 years of data from OTC experiments with wheat in Ostad, Sweden 

showed that relative yield linearly decreased with increasing O3, with a maximum yield loss of 

23% at an AOT40 value of 15 ppmAh (Danielsson et al., 2003). A very similar response was 

found using the flux (stomatal conductance) model of Emberson et al. (2000b) and a similar 

amount of the variance was explained by the flux model (for AOT40 model, r2

 = 0.34 and for the 

Emberson flux model, r2

 = 0.39). A modified flux model developed and calibrated for this site 

also had a similar linear response equation, but explained much more of the variance (r2

 = 0.90). 

During the 1990s, a major European research program investigated the combined effects of 

CO2, O3, and other physiological stresses on wheat (Bender et al., 1999; Hertstein et al., 1996, 

AX9-222
1999; Jäger et al., 1999). The ESPACE-wheat program included 13 experiments in OTCs at 

eight sites in northern Europe over 3 years. Low- and high-O3 exposures in these experiments 

had the following values: 12-h mean (SD) low = 26.3 ppb (12.2), 12-h mean (SD) high = 51.37 

(18.3) ppb, AOT40 mean (SD) low = 6.2 (8,5) ppmAh, AOT40 mean (SD) high = 28.3 (23.0) 

ppmAh, as calculated from data presented in Table 3 of Hertstein et al. (1999). An analysis of all 

13 experiments showed that high O3 at ambient CO2 reduced yield by 13% on average (Bender 

et al., 1999). However, this reduction was not statistically significant based on an ANOVA, and 

the authors concluded that the wheat cultivar Minaret may be relatively tolerant to O3 (Bender 

et al., 1999). Results of some individual studies within this program have been reported 

previously (Donnelly et al., 1999; Fangmeier et al., 1996, 1999; Mulholland et al., 1997a, 

1998a,b; Pleijel et al., 2000b). 

In a study with OTCs on silty loam soil in Beltsville, MD, wheat yield was reduced by 

20% on average over 2 years with 7-h mean concentrations of 61 and 65 ppb O3, compared with 

CF treatment with a 7-h mean concentration of 20 ppb (Mulchi et al., 1995; Rudorff et al., 

1996a,c). In the above study, different amounts of O3 were added above ambient levels for 

5 days as follows: 20, 30, 40, 50, 60 ppb, except that O3 was not added to exceed a total 

concentration of 120 ppb (Rudorff et al., 1996c). Wheat grown in pots in OTCs was exposed to 

elevated O3 and water stress in Germany, and yield was decreased by 35% in the 2×-ambient 

treatment with a 7-h mean O3 concentration of 71 ppb, statistically significant effects were not 

seen in the 1×-ambient treatment (Fangmeier et al., 1994). In two studies conducted in Raleigh, 

NC, soft red winter wheat was grown in pots and exposed to CF, NF, and 1.5×-ambient O3, with 

ambient and elevated CO2 concentrations (Table AX9-16) (Heagle et al., 2000). In the first 

experiment, eight cultivars were exposed to 12-h mean O3 concentrations of 27, 47, and 90 ppb, 

and in the second experiment two of these cultivars were exposed to 22, 38, and 74 ppb. There 

was a trend toward decreased yield in both experiments, but these trends were not statistically 

significant. The wheat cultivar Drabant was exposed to CF, NF, and a constant addition of 35 

ppb during 1992 and 1993 in Finland (Ojanpera et al., 1998), using the Heagle-type OTCs 

(Heagle et al., 1973). The following 12-h mean O3 exposures were observed in 1992: 14, 30, 

61 ppb. In 1993, the values were 9, 21, and 45 ppb, (see Table AX9-16 for AOT40 values and 

other information). Yield was reduced 13% in each year by the added O3 treatment. 

AX9-223
The effect of an intermittent constant addition of O3 using a free air exposure system was 

investigated with the winter wheat cultivar Riband in Northumberland, UK (Ollerenshaw and 

Lyons, 1999). Ozone exposures expressed as AOT40 values for September and October 1992 

were 0.14 and 3.5 ppmAh; while for April to August 1993, values were 3.5 and 6.2 ppmAh. Yield 

was reduced by 13%. 

These results provide an additional line of evidence supporting the OTC-studies that 

demonstrated yield reductions in wheat due to O3 exposures that occur in the United States. 

These yield reductions for wheat are generally similar to those reported previously in 

22 comparable studies summarized in the 1996 AQCD (Table 5-25 of U.S. Environmental 

Protection Agency, 1996). 

AX9.5.4.2 Effects on Plant Quality 

In addition to reductions in biomass or crop yield, O3 may also reduce the quality or 

nutritive value of annual species. Many studies have shown effects of O3 on various measures of 

plant organs that affect quality, with most studies focusing on characteristics important for food 

or fodder (U.S. Environmental Protection Agency, 1996). 

The effect of a continuous intermittent addition of O3 using a free air exposure system in 

Northumberland, UK was investigated with the oilseed rape cultivar Eurol as discussed above 

(Ollerenshaw et al., 1999). Ozone exposures expressed as AOT40 values for August to October 

1991 were 0.2 and 3.8 ppmAh; for June 1992 they were 0.7 and 8.1 ppmAh. Yield quality 

measured as crude protein and oil content was decreased significantly. Because the price of the 

product is reduced in direct proportion to the oil content, such a decrease represents a substantial 

loss to growers (Ollerenshaw et al., 1999). 

Two wheat cultivars, Massey and Saluda, were each grown for one year each in Beltsville, 

MD (Table AX9-16) and exposed to either CF or an addition of 40 ppb for 7 h/day for 

5 days/week (Mulchi et al., 1995; Rudorff et al., 1996a,c). Milling and baking quality scores and 

flour protein were not significantly affected by elevated O3 exposure, but the softness equivalent 

was increased slightly (2.4%) in both experiments (Rudorff et al., 1996b). The authors 

concluded that these changes, along with other slight changes due to an increased CO2 treatment, 

suggested that O3 and CO2 had only minor effects on wheat grain quality. In wheat grown in 

AX9-224
Sweden, the harvest index was significantly decreased and the protein content increased due to 

exposure to a 12-h mean of 48 ppb (Gelang et al., 2000). In an analysis of 16 experiments 

conducted with spring wheat and either O3 or CO2 exposures in four Nordic countries, a negative 

linear relationship was found between grain yield and grain protein content (y = !0.38× +138.6, 

expressed as percentages of the NF treatment (Pleijel et al., 1999a). 

For three soybean cultivars grown in Raleigh, NC, O3 significantly decreased oleic acid 

content, although the authors stated that the reduction was not large enough to be economically 

important (Heagle et al., 1998b). 

In a UK study, potato exposed during 1998 to an AOT40 value of 12.5 ppmAh in OTCs 

(in Nottingham) resulted in the paste from tubers being more viscous (Donnelly et al., 2001b). 

In this study, an AOT40 exposure of 27.11 ppmAh in 1999 caused starch granules to be less 

resistant to swelling, and total glycoalkaloid content was increased due to an increase in asolanine (Donnelly et al., 2001b). Such increases in glycoalkaloid content have been observed 

previously in potato (Pell and Pearson, 1984) and may be important, because glycoalkaloids 

cause bitter flavors and, at higher concentrations, toxicity. The authors indicated that levels 

found in this study approached those that may cause bitterness, but not those of concern for 

toxicity (Donnelly et al., 2001b). 

In the CHIP program the effects of O3 were studied using OTCs at six sites in northern 

Europe, and yield decreases were observed as described above. The reducing sugar and starch 

content of tubers decreased linearly due to O3 exposure, while the ascorbic acid concentration 

increased linearly (Vorne et al., 2002). Compared to the CF treatment, exposure to an AOT40 

value of 14 ppmAh decreased starch concentrations by 2%, decreased reducing sugar 

concentration by 30%, and increased ascorbic acid concentration by 20%. While the changes in 

reducing sugars and ascorbic acid increase tuber quality, the reduction in starch concentration 

decreases tuber quality. 

In two 1-year studies using OTCs in commercial fields in Spain, the soluble solids content 

of watermelon was decreased 4 to 8% due to seasonal O3 exposures as follows: AOT40 = 5.96 

ppmAh and SUM06 = 0.295 ppmAh in one year; and AOT40 = 18.9, SUM06 = 4.95 in the second 

year (Gimeno et al., 1999). 

AX9-225
AX9.5.4.3 Effects on Foliar Symptoms 

For most annual crop species, the most important effects of O3 are on yield of the 

commercially important part of the crop, expressed as the mass of the harvested portion. 

However, for some crops, foliar symptoms are important if they reduce the marketability of the 

crop. This is why efforts have been made to identify O3 exposures associated with foliar 

symptoms. In Europe, Level I critical levels have been determined for such effects based on 

observations from experiments conducted in 15 countries under the auspices of the United 

Nations Economic Commission for Europe International Cooperative Programme on effects of 

air pollution and other stresses on crops and non-woody plants (UN ECE ICP-Vegetation; 

formerly ICP-Crops), as well as on observations of symptoms in commercial fields from 1993 to 

1996 (Benton et al., 1995; Benton et al., 2000). Because the occurrence of symptoms increased 

with greater humidity, these levels took into account the VPD. Two short-term critical levels 

were derived from 1995 data: an AOT40 value of 0.2 ppmAh over 5 days when mean VPD is 

below 1.5 KPa (0930 - 1630 h), and a value of 0.5 ppmAh when the mean VPD is above 1.5 Kpa 

(Benton et al., 1996). The 1996 data supported the critical levels in 83% of observations, 

although symptoms occurred on three occasions when the AOT40 was less than 0.05 ppmAh and 

the VPD was very low — less than 0.6 Kpa. The authors concluded that these critical levels are 

good indicators of the likelihood of foliar symptoms, but that further refinement may be 

required, such as including factors that modify O3 uptake by stomata. 

In a more recent study in Germany, 25 native herbaceous species were exposed to 

several square-wave O3 exposures in CF OTCs (Bergmann et al., 1999). Six of the 25 species 

showed O3-specific symptoms, and five species responded to single-day peaks. The most 

sensitive species exhibiting O3-specific symptoms were Cirsium arvense and Sonchus asper, 

both of which responded to AOT40 values < 1.5 ppmAh (Bergmann et al., 1999). 

AX9.5.4.4 Other Effects 

Several studies during recent decades have demonstrated O3 effects on different stages of 

reproduction. Effects of O3 have been observed on pollen germination, pollen tube growth, 

fertilization, and abortion of reproductive structures, as reviewed by Black et al. (2000). This 

issue is not addressed here, because reproductive effects will culminate for seed-bearing plants 

in seed production, and the substantial body of evidence relating O3 exposure and reduced seed 

AX9-226
production was discussed above. However, one example of a native species will be presented, 

because of its implications for extrapolating exposure-response data to noncommercial species. 

Spreading dogbane (Apocynum androsaemifolium L.) has been identified as a useful species for 

O3 biomonitoring, because of O3-induced diagnostic symptoms (Kohut et al., 2000). A study in 

Massachusetts found that exposure to O3 in NF OTCs or ambient plots for 103 days produced 

significantly fewer flowers and that fewer of these flowers survived to produce mature fruits 

(Bergweiler and Manning, 1999). Because foliar symptoms were not common, the authors 

concluded they are not required for effects on reproduction to occur. Genotoxic effects and 

effects on population genetics were discussed in Section AX9.3. 

AX9.5.4.5 Scaling Experimental Data to Field Conditions 

Substantial effort has been invested in the design of OTCs for assessing the effects of air 

pollutants on vegetation under near-ambient conditions. The design, construction, and 

performance of many types of chambers has been reviewed extensively (Hogsett et al., 1987a,b). 

Despite such design efforts, the influence of experimental chambers on exposure-response 

functions has been debated for many years (e.g., Manning and Krupa, 1992), because several 

factors differ between OTC studies and actual fields. This issue was addressed in Sections 

AX9.1.2.2 and AX9.1.2.4, and only a few comments about the implications of chamber artifacts 

for interpreting exposure-response relationships are presented here. 

While it is clear that chambers can alter some aspects of plant growth (for example, 

(Bytnerowicz et al., 2004; Elagöz and Manning, 2005b), the more important issue is whether 

they alter the response of plants to O3. A review of such chamber studies done in California 

found that plants responded similarly to O3 whether OTCs, closed-top chambers, or air exclusion 

systems were used; differences were found for fewer than 10% of growth parameters (Olszyk 

et al., 1986a). In another review of literature about Heagle-type OTCs (Heagle et al., 1988), the 

authors concluded that “Although chamber effects on yield are common, there are no results 

showing that this will result in a changed yield response to O3.” A more recent study of chamber 

effects examined the responses of tolerant and sensitive white clover clones to ambient O3 in 

greenhouse, OTCs, and ambient plots (Heagle et al., 1996). For individual harvests, O3 reduced 

the forage weight of the sensitive clone 7 to 23% more in the greenhouse than in OTCs. 

However, the response in OTCs was the same as in ambient plots. A similar study with these 

AX9-227
white clover clones near Naples, Italy also found no significant difference between O3 effects 

measured in OTCs versus those measured by comparing the ratio of sensitive and resistant 

clones in ambient air (Fagnano et al., 2004). Several studies have shown very similar yield 

responses to O3 for plants grown in pots or in the ground, suggesting that even such a significant 

change in environment does not alter the proportional response to O3, at least as long as the 

plants are well- watered (Heagle, 1979; Heagle et al., 1983). As discussed in Section AX9.1, 

results from recent FACE studies are similar to those obtained from OTC studies, providing 

another line of evidence that chamber effects in OTCs do not substantially alter O3 

exposure-response relationships (Morgan et al., in press). 

Most experiments investigating O3 effects on annual vegetation provide adequate water to 

avoid substantive drought stress. Because drought stress has generally been shown to reduce the 

effect of O3 on annual vegetation, such experiments may tend to overestimate O3 effects on crops 

and especially on unmanaged or seminatural vegetation. 

As mentioned above, the use of O3 flux, rather than exposure, is theoretically more 

realistic, and such an approach would also address the vertical gradient issue (Section 3.3.2, 

Section AX9.1). A number of investigators have suggested that modeling O3 flux can improve 

estimates of O3 effects on vegetation. Models of O3 flux can reduce the variation in the response 

to O3 that is sometimes observed between years in an experiment (Emberson et al., 2000a, b; 

Fuhrer et al., 1992; Grünhage et al., 1993; Grünhage and Haenel, 1997; Pleijel et al., 2000a). In 

a study of O3 deposition to an oat crop in OTCs, O3 flux in the chamber was estimated to be up 

to twice that in an adjacent field based on a K-theory approach and measurements of stomatal 

conductance and environmental conditions (Pleijel et al., 1994b). These measurements were 

made for 2 hours on 5 days when the canopy was physiologically active and wind speeds were 

moderate. However, the O3 flux in a plant-less chamber was nearly as high as that in the open 

field. The authors conclude that O3 uptake in the chamber was between 100 and 200% of that in 

the field. These models of flux have a sound biological and meteorological basis and are useful 

for interpreting experimental data. Flux models have been successfully applied at intensive 

study sites with detailed site-specific data on stomatal conductance and micrometeorological 

conditions (e.g., Fredericksen et al., 1996b; Grünhage et al., 1993, 1994)). Yet even at a single 

well-studied site, different methods can provide different estimates of O3 flux. For example, at a 

site in a vineyard in California, an evapotranspiration-based method overestimated the O3 flux as 

AX9-228
compared to an eddy covariance approach by 20 to 26% (Massman and Grantz, 1995). At a site 

in a nearby cotton field the evapotranspiration-based approach overestimated the eddycovariance method by 8 to 38%. Flux-response relationships have received substantial attention 

in Europe during the past decade, as part of an attempt to move beyond the exposure-based 

Level 1 critical levels to flux-based Level II critical levels (Section AX9.4). However, the 

database for flux-response relationships is very limited (Grünhage et al., 2004), and these 

approaches do not always explain greater amounts of response variation than do exposure-based 

approaches (Karlsson et al., 2004b). The critical level approach is discussed briefly below, 

because it has been used to extrapolate from field studies to landscapes, countries, and regions. 

There has been criticism that the Level I critical level for crops overestimates O3 effects in 

the Mediterranean countries, because it was developed based on studies in Northern Europe (De 

Santis, 1999; De Santis, 2000). However, there is evidence of substantial crop loss due to O3 in 

some southern European counties, such as the Po valley in Northern Italy (Fumagalli et al., 

2001). In these studies, Heagle-type OTCs were used. Losses in NF chambers as compared to 

CF chambers over several years at two sites ranged from 11.2 to 22.8% for barley and wheat, 

from 0.3 to 31.5% for other crop species, and from 4.1 to 19.8% for forage species (Fumagalli 

et al., 2001). Surprisingly, the least effect was observed for soybean, despite AOT40 values of 

9.32 ppmAh, 3× the Level I critical level. Similarly, a review of studies in Northern Italy found 

that ambient O3 episodes have been reported to cause foliar symptoms on 24 agricultural and 

horticultural crops in commercial fields (Fumagalli et al., 2001). Ambient O3 has also been 

reported to cause yield losses in several crop species, although no data on O3 exposure were 

presented by the reviewers (Fumagalli et al., 2001). 

The Level I approach has also been criticized for focusing only on a single annual species 

(wheat) and a single woody perennial species (beech). However, this species focus is 

appropriate, because the goal was to determine an exposure-response relationship for a sensitive 

species based on available data. In support of standards in Germany, an effort was made to 

combine data from different species, and consisted of a meta-analysis of studies conducted in 

both closed and OTCs (Grünhage et al., 2001). In this study, experiments published between 

1989 and 1999 were included and analyzed if they met the following conditions: (1) a 

significant O3 effect was determined; (2) exposure conditions were well defined; (3) foliar 

symptoms, growth, or yield was measured; and (4) plant species were relevant to Europe 

AX9-229
(Grünhage et al., 2001). Despite the focus on European species, many of the species studies also 

occur in the United States. Separate regressions for herbaceous plants and for tree species were 

created as a function of duration of exposure at a given level of O3 exposure at the top of the 

plant canopy. These regression equations, with confidence limits and with correction for the 

vertical gradient in O3 from the top of the quasi-laminar boundary layer, can be used to define 

whether effects are unlikely (below the lower confidence limit), probable (between the 

confidence limits), or highly likely (above the upper confidence limit) to occur near a given O3­

monitoring station. 

A further concern about the Level 1 approach is that foliar symptoms, rather than biomass, 

may be an important endpoint, because foliar symptoms may be more sensitive (VanderHeyden 

et al., 2001). In an OTC study in southern Switzerland, it was shown that a number of tree 

species show foliar symptoms at AOT40 values lower than the Level 1 value of 10 ppmAh 

(VanderHeyden et al., 2001). 

Exposure-response relationships developed primarily from OTC experiments, with 

confirming evidence from other approaches such as resistant and sensitive clover clones exposed 

in ambient air and FACE experiments, are useful for estimating the effects of ambient O3 on 

vegetation in the U.S. However, despite the substantial number of experimental studies listed in 

Table AX9-16 and in previous Air Quality Criteria Documents for O3 (for example, U.S. 

Environmental Protection Agency, 1996), most studies have been conducted in only a few 

locations with only a few species of economically important crops. While these studies provide 

strong evidence that ambient O3 in the U.S. is likely reducing crop yields and plant growth 

significantly in many regions in many years, it remains difficult to extrapolate to all landscapes 

and all plant species and to determine whether exposure-response relationships based on existing 

studies will protect all species in all locations from significant deleterious effects of O3. 

AX9.5.4.6 Summary of Effects on Short-Lived Species 

For annual vegetation, the data summarized in Table AX9-16 show a range of growth and 

yield responses both within species and among species. Nearly all of these data were derived 

from studies in OTCs, with only two studies using open-air systems in the UK (Ollerenshaw 

et al., 1999; Ollerenshaw and Lyons, 1999). It is difficult to compare studies that report O3 

exposure in different indices such as AOT40, SUM06, or 7-h or 12-h mean values. However, 

AX9-230
when such comparisons can be made, the results of this more recent body of research confirm 

the earlier results summarized in the 1996 O3 AQCD (U.S. Environmental Protection Agency, 

1996). A summary of earlier literature concluded that a 7-h, 3-month mean of 49 ppb 

corresponding to a SUM06 exposure of 24.4 ppmAh would cause 10% loss in 50% of 49 

experimental cases (Tingey et al., 1991). A similar study using a 24-h, rather than 7-h, 

averaging period found that a SUM06 exposure of 26.4 ppb would cause 10% loss in 50% of 

54 experimental cases (Lee et al., 1994a,b). Recent data summarized in Table AX9-16 support 

this conclusion, including data from a FACE study which confirmed results found previously 

using OTCs (Morgan et al., in press). These values represent ambient exposure patterns that 

occur in some years over large portions of the United States. Some annual species such as 

soybean are more sensitive, and greater losses may be expected (Table AX9-16). Thus, the 

recent scientific literature supports the conclusions of the 1996 AQCD that ambient O3 

concentrations are reducing the yield of major crops in the United States. 

Much research in Europe has used the AOT40 exposure statistic, and substantial effort has 

gone into developing Level-1 values for vegetation. Based on regression analysis of 15 OTC 

studies of spring wheat, including one study from the United States and 14 from locations 

ranging from southern Sweden to Switzerland, an AOT40 value of 5.7 ppmAh was found to 

correspond to a 10% yield loss, and a value of 2.8 ppmAh corresponded to a 5% yield loss 

(Fuhrer et al., 1997). Because a 4 to 5% decrease could be detected with a 99% confidence 

level, a critical level of an AOT40 value of 3 ppmAh was selected in 1996 (Kärenlampi and 

Skärby, 1996). 

In addition to reductions in crop yield, O3 may also reduce the quality or nutritive value of 

annual species. Many studies have shown effects of O3 on various measures of plant organs that 

affect quality, with most studies focusing on characteristics important for food or fodder. These 

studies indicate that there may be economically important effects of ambient O3 on the quality of 

crop and forage species. Previous criteria documents have concluded that foliar symptoms on 

marketable portions of crops and ornamental plants can occur with seasonal 7-h mean O3 

exposures of 40 to 100 ppb (U.S. Environmental Protection Agency, 1978, 1986, 1996). The 

recent scientific literature does not refute this conclusion. 

The use of OTCs may reverse the usual vertical gradient in O3 that occurs within a few 

meters above the ground surface (Section 3.3.2, Section AX9.1). Such a reversal suggests that 

AX9-231
OTC studies may overestimate, to some degree, the effects of an O3 concentration measured 

several meters above the ground, because the O3 concentration is slightly lower at the canopy 

height than at the height at which O3 is monitored. For example, as shown in Figure 3.6 

(Section 3.3.2), under unstable atmospheric conditions that are typical during daylight in 

agricultural fields, there was an average decrease of 7% in ozone concentration from 4 m to 

0.5 m above the surface. However, such considerations do not invalidate the conclusion of the 

1996 AQCD (U.S. Environmental Protection Agency, 1996) that ambient O3 exposures 

(Tables AX9-13 and AX9-21) are sufficient to reduce the yield of major crops in the United 

States. Recent studies using OTCs confirmed previous results that ambient ozone exposures can 

significantly reduce the growth of annual species and the yield of commercially important crop 

species. Additionally, a two-year study using a free air exposure system with soybean confirmed 

yield reductions found previously using OTCs (Morgan et al, in press). 

AX9.5.5 Effects of Ozone on Long-Lived (Perennial) Species 

Although there has been considerable research in Europe on annual species during the past 

10 years, much research in the United States has focused on perennial species. In Europe, and in 

a few studies in the United States, effects of O3 on mixtures of annual and perennial herbaceous 

species have been investigated using growth chambers, greenhouses, and OTCs. Section 

AX9.5.5.1 reviews such studies, with an emphasis on studies using OTCs. 

AX9.5.5.1 Herbaceous Perennial Species 

Two alfalfa cultivars were grown in pots and exposed to CF, NF, 1.5×-ambient and 

2×-ambient O3 concentrations in two 1-year studies in Quebec, Canada (Renaud et al., 1997). 

One cultivar, Apica, is commonly grown in the region, and another, Team, is normally grown 

farther south and is more tolerant to O3. For Apica in both years and for Team in 1991, O3 

exposure caused a linear reduction in above-ground biomass. In the NF treatment, growth of 

Apica was decreased by 31 and 21% in the 2 years, while the growth of Team was reduced by 

14% in 1991, but not reduced in 1992. The authors suggested that the differing effects on Team 

could be due to different progenies and propagation methods in the 2 years or to more rapid 

growth in 1991 along with higher O3 peak values in 1991. In 1991, O3 maxima exceeded 60 ppb 

in 15 days, whereas in 1992 there were only three such days. At the end of the growing season, 

AX9-232
total starch reserves in roots were decreased by O3, due primarily to a decrease in root mass, that 

the authors suggested could accelerate decline in alfalfa yields. These yield reductions are 

generally similar to those reported previously in five similar studies summarized in the 1996 

AQCD (Table 9-25 of U.S. Environmental Protection Agency, 1996). 

A study in Alabama exposed early- and late-season-planted Bahia grass (Paspalum 

notatum cultivar Pensacola) in OTCs to CF, NF, or 2×-ambient O3 treatments (Muntifering et al., 

2000). Ozone exposures expressed as 12-h mean values over the 24-week experiment were 22, 

45, and 91 ppb, and the highest ambient O3 concentrations were recorded in late June, late July, 

late August and mid-September at approximately 90 ppb. Above-ground biomass growth was 

reduced by the NF treatment for the first and second harvest by 34% and 29% for the earlyseason planting, but statistically significant effects were not observed in the late-season planting 

(Table AX9-17). The 2×-ambient treatment did not cause further significant reductions in 

biomass. The authors suggested that the lack of a significant O3 effect in the late planting may 

have been due to the shorter total O3 exposure time as well as to the lower O3-exposure 

concentrations during the weeks immediately preceding harvest. These results are important, 

because this is an economically important species and because previous studies have focused on 

grass species that use the C3, rather than C4, metabolic pathway. 

An investigation of the use of different O3 indices and averaging times on the correlation 

with growth effects was undertaken with the North Carolina clover system (Heagle et al., 1995). 

For 2 years of data at six sites in Massachusetts, Oregon, North Carolina, California (2 sites), 

and Virginia, averaging time was found to be more important than the choice of the type of 

index including mean, SUM06, and AOT40 (Heagle and Stefanski, 2000). The best correlation 

between O3 exposure and the ratio of sensitive-to-tolerant clover types was found for the 6-h 

period from 1000 to 1600 h. For this period, very similar r2

 values (0.91 to 0.94) were found for 

SUM06, W126, and AOT40 (Heagle and Stefanski, 2000). For the above indices, a linear 

relationship was found, with no effect in Corvallis, OR with exposure to a SUM06 value of 

10.2 ppmAh and a ratio of 0.53 (sensitive/tolerant) at San Bernardino with a SUM06 exposure of 

39.4 ppmAh (Heagle and Stefanski, 2000). 

In a study of the biomass ratio of O3-sensitive versus O3-insensitive clover at 14 sites in 

Europe during 1996 to 1998, a model was developed using ANN techniques (see Section AX9.1) 

that had r2

 values for the training data of 0.84 and for unseen validation data of 0.71 (Mills et al., 

AX9-233
2000). The predictive factors in the model were AOT40, 24-h mean O3, daylight mean 

temperature, and 24-h mean temperature. This model was selected after a thorough investigation 

of a number of models using many more or fewer parameters using both ANN and multiple 

linear regression techniques. This model predicted that a 5% reduction in biomass ratio was 

associated with AOT40 values in the range 0.9 to 1.7 ppmAh accumulated over 28 days, with 

plants being most sensitive under conditions of low NOx, moderate temperature, and high 24-h 

mean O3 concentration. 

Exposure to a square-wave 8-h mean O3 concentration of 92 ppb for 62 days in an 

experiment in OTCs the UK did not significantly reduce the total yield of strawberry fruits, but 

did decrease the average size of the fruits by 14% (Drogoudi and Ashmore, 2000). This 

contrasts with an increase in total yield (fruit weight) found in a previous study in California 

(Takemoto et al., 1988). 

When timothy was exposed in OTCs in Sweden to NF, CF, and CF+O3 treatments, there 

was no effect of a 12-h mean O3 exposure of 68 ppb (NF treatment), but a 12-h mean exposure 

of 152 ppb decreased yield by 58% (Danielsson et al., 1999). A similar lack of effect of 

exposure to a 12-h mean O3 exposure of 62 ppb was found in a previous study in the United 

States (Kohut et al., 1988). 

Although most investigations of O3-response relationships focus on growth or yield of 

marketable portions of plants, some studies also investigate effects on plant quality. In the study 

of bahia grass in Alabama discussed above, in addition to the effects on yield, there were 

significant effects on quality for ruminant nutrition (Muntifering et al., 2000). Concentrations of 

neutral detergent fiber (NDF) were higher in primary-growth and regrowth forages from the 

early-season planting when exposed to 2×-ambient O3 than when exposed to the NF treatment. 

The concentration of acid detergent fiber was higher in the 2×-ambient treatment than in NF 

treatment regrowth, whereas acid detergent lignin concentration was higher in 2×-ambient than 

in NF primary-growth forage. Crude protein concentrations were lower in CF-exposed than in 

NF-exposed regrowth forage from the early planting and in CF- than in NF-exposed primarygrowth forage from the initial harvest of the late-season planting. No differences were observed 

among treatments in concentrations of total phenolics in primary-growth or regrowth forages 

from either planting, although concentrations of total phenolics tended to be higher in 

CF-exposed than in NF-exposed primary-growth forage from the late-season planting. The 

AX9-234
authors concluded that the alterations in quality of primary-growth and vegetative regrowth 

forages were of sufficient magnitude to have nutritional and possibly economic implications to 

their use for ruminant animal feed. 

Sericea lespedeza and little bluestem were exposed to CF, NF, and 2×-ambient O3 in OTCs 

in Alabama for 10 weeks (Powell et al., 2003). Ozone treatments expressed as 12-h mean 

concentrations were 23, 40, and 83 ppb and expressed as seasonal SUM06 values were 0.2, 9.1, 

and 61.0 ppmAh. Although there were few statistically significant effects of O3 on yield (the 

yield of only the 2×-ambient compared to NF for Sericea lespedeza in the last of six harvests), 

plant quality as feed for ruminants was reduced. The nutritive quality of Sericea lespedeza was 

decreased by 7% and that of little bluestem by 2% as a result of increased cell wall constituents 

and decreased in vitro digestibility. 

For some annual species, particularly crops, the endpoint for an assessment of the risk of 

O3 exposure can be defined as yield or growth; e.g., production of grain. For plants grown in 

mixtures such as hayfields, and natural or seminatural grasslands (including native 

nonagricultural species), endpoints other than production of biomass may be important. Such 

endpoints include biodiversity or species composition and measures of plant quality such as total 

protein and effects may result from competitive interactions among plants in mixed-species 

communities. Most of the available data on non-crop herbaceous species are for grasslands. 

In a study of two perennial grasses (bent grass and red fescue) and two forbs (white clover, 

Germander speedwell [Veronica chamaedrys L.]) grown in pots in OTCs, O3 effects differed 

among species and cutting treatments (Ashmore and Ainsworth, 1995) (see also Table AX9-17). 

Fescue biomass increased with higher O3 treatments both in pots that were not cut during the 

growing season (mid-June to mid-September) and those that were cut every two weeks. 

However, bent grass biomass decreased with higher O3 exposure in the uncut treatment and 

increased in the cut treatment. White clover and Germander speedwell biomass decreased 

substantially with higher O3 exposure with and without cutting, with greater decreases in the cut 

treatment. The authors cautioned that the experiment did not replicate field circumstances. The 

plants were all cut to 1 cm above the ground, which does not simulate grazing, and there may 

have been effects due to growing the plants in pots. However, two key results of this study 

likely apply to mixtures of species growing in hay or forage fields or seminatural and natural 

communities. First, O3 exposure increased the growth of O3-tolerant species, while exacerbating 

AX9-235
the growth decrease of O3 sensitive species. Second, the total biomass of the mixed-species 

community was unaffected by O3 exposure due to the differential effects on O3-sensitive and 

O3-tolerant species. 

In a 2-year study using OTCs placed over managed pasture in Switzerland, the aboveground biomass of clover (red and white) was reduced linearly in response to increased O3 

exposure (Table AX9-17) (Fuhrer et al., 1994). Exposure to a 12-h mean concentration of 

39 ppb O3 reduced biomass by 24% as compared to the CF treatment with a 12-h mean 

concentration of 21 ppb O3. There was a trend toward increased above-ground biomass of 

grasses (primarily orchard grass), but this trend was not statistically significant. As often found 

in other studies of mixtures of species, by O3 exposure did not significantly affect total aboveground biomass O3 exposure. 

A field-grown grass/clover mixture was exposed to CF, NF, and two O3 addition treatments 

for two growing seasons in OTCs in southern Sweden (Pleijel et al., 1996). The mixture 

consisted of 15% (by seed weight) red clover cv. Fanny, 60% timothy cv. Alexander, and 25% 

fescue cv. Svalofs Sena. Ozone concentrations expressed as AOT40 ranged from 0 to 

approximately 47 ppmAh and expressed as 7-h mean from 11 to 62 ppb. Over this range, a slight, 

but statistically significant, linear decrease of 4% in total above-ground biomass was seen 

growth over six harvests. No significant decrease was seen in the proportion of clover, and the 

authors ascribed this lack of effect to the relatively higher O3 sensitivity of timothy and lower 

sensitivity of this clover cultivar as compared to previously published results for other 

grass/clover mixtures (e.g., Fuhrer et al., 1994). 

A mixture of species in an old farm field in Alabama was exposed to O3 for two growing 

seasons in large OTCs (4.8 m high and 4.5 m diameter); and a similar lack of effect of O3 was 

found on total plant community growth measured as both canopy cover and vertical canopy 

density (Barbo et al., 1998). Of the 40 species in the plots, O3 effects were examined only on the 

most common species: blackberry, broomsedge bluestem, bahia grass, Panicum spp., and 

winged sumac (second year only). Of these species, a 2×-ambient O3 treatment increased the 

percent canopy cover of blackberry over 2 years by 124%, while that of winged sumac was 

decreased by 95% (Table AX9-17). Blackberry showed no significant effect on biomass, but 

ripe fruit mass was decreased by 28% (Chappelka, 2002). However, there was a significant 

chamber effect for this latter response. Biomass was not reported for other species in this study 

AX9-236
due to a hurricane. Effects on loblolly pine grown in this experiment are discussed subsequently 

in Section AX9.5.5.5. 

In summary, results of studies on perennial herbaceous species conducted since the 1996 

criteria document was prepared are presented in Table AX9-17. As for single-season 

agricultural crops, yields of multiple-year forage crops are reduced at O3 exposures that occur in 

some years over large areas of the United States (Tables AX9-13, AX9-21). This result confirms 

that reported in the 1996 AQCD (U.S. Environmental Protection Agency, 1996). When species 

are grown in mixtures, O3 exposure can increase the growth of O3-tolerant species while 

exacerbating the growth decrease of O3-sensitive species (e.g., Ashmore and Ainsworth, 1995; 

Fuhrer et al., 1994). Because of this competitive interaction, the total growth of the mixedspecies community may not be affected by O3 exposure (Ashmore and Ainsworth, 1995; Barbo 

et al., 1998; Fuhrer et al., 1994). However, in some cases mixtures of grasses and clover species 

have shown significant decreases in total biomass growth in response to O3 exposure in studies 

in the United States (Heagle et al., 1989; Kohut et al., 1988) and in Sweden (Pleijel et al., 1996). 

In Europe, a provisional critical level for perennials of an AOT40 value of 7 ppmAh over 6 

months has been proposed to protect sensitive plant species from the adverse effects of O3. 

AX9.5.5.2 Deciduous Woody Species 

It is extremely difficult and costly to study entire mature trees under controlled conditions 

such as those in OTCs, with the possible exception of some species managed for fruit or nut 

production. For this reason, the great majority of investigations have been of seedlings in 

growth chambers, greenhouses, or OTCs. A few investigations have been carried out on 

saplings or more mature trees using free air exposure systems (Haeberle et al., 1999; Isebrands 

et al., 2000, 2001; Werner and Fabian, 2002). Exposure-response functions based on 28 

experimental cases of seedling response to O3 suggest that a SUM06 exposure for 3 months of 

31.5 ppmAh would protect hardwoods from a 10% growth loss in 50% of the cases (Table 

AX9-18). However, there is a substantial range in sensitivity among species. A risk analysis 

was undertaken to predict tree biomass growth reductions due to O3 based on exposure-response 

equations for seedlings of individual species combined with the species’ spatial distribution 

across the eastern United States and interpolated O3 exposure expressed as SUM06 (Hogsett 

et al., 1997). The growth of sensitive species such as aspen and black cherry was predicted to be 

AX9-237
reduced by at least 20% across 50% of their ranges in a high O3 year and approximately 10% in a 

lower-than-average O3 year (Hogsett et al., 1997). 

A few investigations reported since the last criteria document was prepared have examined 

saplings or mature trees, notably of oak species in the southern Appalachian Mountains and pine 

species in California. Most of these studies have been of natural (uncontrolled) O3 exposures. 

Additional studies have examined foliar symptoms on mature trees, and in recent years such 

surveys have become more common and with greater attention to the standardization of methods 

and the use of reliable indicator species (Campbell et al., 2000; Smith et al., 2003). Previous 

criteria documents have noted the difficulty in relating foliar symptoms to effects on individual 

tree growth, stand growth, or ecosystem characteristics (U.S. Environmental Protection Agency, 

1996). This difficulty still remains to the present day. 

Some investigators have suggested that a comprehensive risk assessment of the effects of 

O3 on mature tree species might best be accomplished by extrapolating measured effects of O3 

on seedlings to effects on forests using models based on tree physiology and forest stand 

dynamics (Chappelka and Samuelson, 1998; Laurence et al., 2000, 2001). Several such efforts 

are discussed in Sections AX9.3 and AX9.6. 

In this subsection, emphasis will be placed on experimental evidence of O3 effects on the 

growth of woody species under controlled conditions with some information from observational 

studies under ambient conditions in forests. Experimental results are summarized for deciduous 

species in Table AX9-18; the species are discussed below in the order in which they appear in 

this table. 

A series of studies in Michigan and Wisconsin during the 1990s on clones of trembling 

aspen previously demonstrated that they differ in their O3 sensitivity (Coleman et al., 1995a,b, 

1996; Dickson et al., 2001; Isebrands et al., 2000, 2001; Karnosky et al., 1996, 1998, 1999; King 

et al., 2001). Several of those studies were undertaken with plants in pots or in the ground in 

OTCs and additional studies were undertaken at three sites selected to differ primarily in O3 

exposure (Karnosky et al., 1999). An ongoing study was undertaken using a FACE carbon 

dioxide and O3 enrichment facility in Rhinelander, WI (Isebrands et al., 2000, 2001). These 

studies showed that O3-symptom expression was generally similar in OTCs, FACE, and gradient 

sites, supporting the previously observed variation among aspen clones (Karnosky et al., 1999). 

In the Michigan OTC study, plants were grown in pots and exposed to CF, 0.5× ambient, 1× 

AX9-238
ambient, 1.5× ambient, 2×-ambient O3 treatments for 98 days (Karnosky et al., 1996). Ozone 

concentrations expressed as 3-month, 7-h mean values were 7 ppb (CF), 43 ppb (1×) and 63 ppb 

(2×). Ozone decreased total plant biomass between 2 and 22% for three clones previously 

selected to represent high, intermediate and low O3 tolerance based on previous studies of larger 

populations(clones 216, 271 and 259; Karnosky et al., 1996). Seedlings produced from 15 

parent trees responded similarly as did the 3 clones, with an average biomass reduction of 14% 

in the 1×-ambient treatment for the seedlings compared to 16% for the clones. In a second 

experiment using square wave exposures, biomass reduction for the clones was ranged from 23 

to 39% (mean = 31%) at a 7-h mean O3 concentration of only 45 ppb, which is similar to the 

response to the 2×-ambient treatment in the previous experiment at a 7-h mean O3 concentration 

of 66 ppb. 

The FACE study evaluated the effects of multiple years of exposure to combinations of 

elevated CO2 and 1.5× ambient O3 on growth responses in mixture of five trembling aspen 

clones (Isebrands et al., 2000, 2001). Height, diameter, and stem volume (diameter2

 × height) 

were decreased by elevated O3. On average for all clones, stem volume was decreased by 20% 

over the first 3 years in the elevated O3 treatment as compared with the 1×-ambient treatment. 

However, one clone showed increased growth in response to O3. Ozone concentrations were not 

reported. Over the first 7 years of the study, average stem volume was decreased by 14% with 

12-h mean O3 concentrations between 49 and 55 ppb as compared with effects at ambient O3 

concentrations with 12-h mean values of 35 to 37 ppb (O3 exposure data are for the first 4 years, 

as they have not been reported for subsequent years) (Karnosky et al., 2003b, 2005). This FACE 

facility study is important, because it confirmed responses reported previously with these clones 

grown in pots or soil in OTCs, without the alterations of microclimate induced by chambers. 

Currently, this is the only U.S. study using this technology to have examined the effects of O3 

under these conditions. This study is also significant, because the elevated O3-exposure pattern 

used was intended to reproduce the 6-year average pattern from Washtenaw County, Michigan 

(Karnosky et al., 1999). 

Rooted cuttings of two aspen clones from Acadia National Park in Maine were exposed to 

1×-ambient, 1.7×-ambient, and 3×-ambient O3 concentrations in large OTCs in Ithaca, NY for 

much of one growing season (15 June to 15 September) (Yun and Laurence, 1999a). Both 

circular (4.7 m diameter, 3.7 m height) and rectangular (7.4 m × 2.75 m × 3.7 m height) 

AX9-239
chambers were used (Mandl et al., 1989). Exposure to 1.7×-ambient O3 (SUM06 = 20 ppmAh, 

9-h mean = 74 ppb) reduced shoot growth by 14 and 25% compared to ambient O3 for the two 

clones (Yun and Laurence, 1999a). Total dry weight was reduced by 55 and 35% in the two 

clones by the 3×-ambient treatment (SUM06 = 62 ppmAh, 9-h mean = 124 ppb) compared to the 

ambient O3 treatment. 

When black poplar cuttings in OTCs in Belgium were exposed to 8-h mean O3 

concentrations of 5, 29, and 33 ppb, diameter growth decreased by 29% in the highest O3 

treatment, but height growth was unaffected (Bortier et al., 2000b). A 2-month study of hybrid 

poplar (Populus tremuloides × P. tremula) in a free air exposure system in Finland with 7-h 

mean O3 concentrations of 30 and 38 ppb found a 6% decrease in height with no effect on 

biomass (Oksanen et al., 2001). Eastern cottonwood cuttings in pots buried in the ground with 

drip irrigation were exposed to ambient O3 at several sites in and near New York City in three 

2-month experiments during three summers (Gregg et al., 2003). Ozone concentrations were 

lower at urban sites than at rural sites within 100 km of the urban sites. Total biomass growth 

was greater in urban than rural sites, with a strong linear decrease in biomass with increasing O3 

across all sites and years (r2 = 93). Total biomass decreased 33% with 12-h mean O3 levels of 38 

ppb compared to 23 ppb. Multiple regression analysis showed no significant temperature effect 

on biomass. Therefore, the authors suggested that O3 exposures were the most likely explanation 

for the reduced biomass in rural areas. The overall growth reductions and the variation among 

genotypes seen on all of the above aspen and poplar studies is similar to those previously 

reported in three OTCs studies summarized in the 1996 O3 AQCD (Table 9-26 of U.S. 

Environmental Protection Agency, 1996). 

For paper birch, over the first 7 years of the Wisconsin FACE study, average stem volume 

was unaffected by 12-h mean O3 concentrations between 49 and 55 ppb as compared to effects at 

ambient O3 concentrations with 12-h mean values of 35 to 37 ppb (based on O3 exposure data for 

the first 4 years, as they have not been reported for subsequent years) (Karnosky et al., 2003b, 

2005). In contrast, significant effects were found in this study for aspen and sugar maple, so 

these results indicate that paper birch is relatively insensitive to O3 compared to these other 

species. 

Black cherry seedlings grown in pots were exposed in OTCs in the Great Smoky Mountain 

National Park in Tennessee to O3 treatments ranging from CF to 2×-ambient in two experiments 

AX9-240
during 1989 and 1992 (Neufeld et al., 1995). Ozone exposure, expressed as SUM06, ranged 

from 0 to 40.6 ppmAh in 1989 and from 0 to 53.7 ppmAh in 1992. Corresponding AOT40 values 

were 0.0 to 28.3 ppmAh in 1989 and 0 to 40.4 ppmAh in 1992. In 1989, total biomass was 

decreased in the 1.5×-ambient treatment by 18% and in the 2×-ambient treatment by 38%. 

In 1992, total biomass was decreased in the 1.5×-ambient treatment by 27%, and in the 

2×-ambient treatment by 59%. In this study, SUM06 and AOT40 provided better fits than did 

SUM00 with Weibull regression analyses to log-transformed biomass data. Although a Weibull 

model was used, responses to O3 expressed as SUM06 and AOT40 appeared to be linear. The O3 

exposures in the 1.5×-ambient and 2×-ambient treatments were reported to be similar to those 

for a site near Charlotte, NC in a high-O3 year (1988). In a 2-year experiment in OTCs in Ohio, 

seedlings of black cherry, sugar maple, and yellow poplar were exposed to O3 treatments with 

SUM00 values ranging from 16 to 107 ppmAh in 1990 and 31 to 197 ppmAh in 1991 (Rebbeck, 

1996). After two seasons of exposure, only black cherry showed a growth decrease: total 

biomass was reduced by 32% in the 2×-ambient O3 treatment compared to the CF treatment; root 

biomass was decreased by 39%. These results contrast with those of a previous study with black 

cherry seedlings in which significant biomass reductions with exposures up to 2×-ambient were 

not observed (7-h mean = 21 to 97 ppb), perhaps because of the small sample size (3 seedlings 

per chamber (Samuelson, 1994) in the earlier study. 

A multiyear study of effects of O3 on both seedling and mature (30-year-old) red oak trees 

was conducted in Norris, TN in large OTCs with three replicates per O3 treatment. Trees were 

exposed for 3 years to CF, 1×-ambient, and 2×-ambient treatments, with the following O3 

exposures: SUM06 for 3 years = 0, 29, 326 ppmAh; SUM00 for 3 years = 147, 255, and 507 

ppmAh. The net photosynthetic rate in mature trees was reduced by 25% in the ambient 

treatment and by 50% in the 2×-ambient treatment (Samuelson and Edwards, 1993; Hanson 

et al., 1994; Wullschleger et al., 1996). Despite these large decreases, no significant effects on 

stem increment at the base, stem increment in the canopy, or leaf mass were observed for the 

mature trees (Samuelson et al., 1996). Similarly, seedling biomass was not significantly reduced 

by O3 exposure. The difficulty in replicating experiments with mature trees makes it difficult to 

detect changes in growth or biomass. However, the mean values of the stem increment at the 

base and within the canopy in the ambient treatment were larger than those in the CF treatment, 

although those in the 2×-ambient treatment were lower. Therefore, this study of mature trees 

AX9-241
does not provide evidence that these ambient concentrations reduced above-ground tree growth, 

even after 4-years exposure. 

Sugar maple seedlings were exposed for 3 years to ambient, 1.7×-ambient, and 3×-ambient 

O3 treatments at both high-light (35% of ambient) and low-light levels (15% of ambient) 

(Topa et al., 2001). This experiment was conducted in large OTCs near Ithaca, NY. Over the 

3 years, O3 exposures expressed as SUM00 for the three treatments were 88, 126, and 

185 ppmAh. After 3 years, total seedling biomass in the 3×-ambient treatment was reduced by 

64% and 41% in the low- and high-light treatments, respectively (compared to the 1×-ambient 

treatment). The larger reduction of biomass under low-light conditions suggests that seedlings 

growing under closed canopies may be substantially more sensitive to O3 than are seedlings 

exposed to higher-light levels in gaps or clearings. These results differ from other studies in 

which seedling biomass was unaffected by exposure to SUM00 values of 304 ppmAh over 2 years 

(Rebbeck, 1996) or 591 ppmAh over 3 years (daytime mean of 40.7) (Laurence et al., 1996). 

However the latter two studies used much higher light levels, which may have reduced the O3 

effect, based on the results of Topa et al. (2001). Over the first 7 years of the Wisconsin FACE 

study, average stem volume of sugar maple was decreased by 14% with 12-h mean O3 

concentrations between 49 and 55 ppb as compared with effects at ambient O3 concentrations 

with 12-h mean values of 35 to 37 ppb (based on O3 exposure data for the first 4 years, as they 

have not been reported for subsequent years) (Karnosky et al., 2003b, 2005). These growth 

effects were not statistically significant for the first 3 years, but became significant subsequently. 

These results are important because they demonstrate that 3 years of exposure may not be long 

enough to evaluate effects of O3 on the growth of tree species. 

Although most studies demonstrate that O3 decreases biomass growth, occasional results 

indicate that O3 can increase growth of some portions of woody perennials. When Casselman 

plum trees near Fresno, CA were exposed to O3 in large, rectangular OTCs to three O3 treatments 

(CF, 1×-ambient, and an above-ambient O3 treatment) for 4 years (12-h mean = 31, 48, 91 

ppmAh), stem increment increased 14% in the highest O3 treatment compared to the CF 

treatment; and this effect was statistically significant (Retzlaff et al., 1997). However, fruit yield 

decreased in this treatment by 42% and also decreased by 16% in the 1×-ambient-O3 treatment. 

Root growth was not measured in this study. Hence, the increase in stem diameter may have 

been at the expense of other organs. However, in a fifth year, all plants were exposed to 

AX9-242
1×-ambient O3, and there were no differences in fruit yield, suggesting that trees were able to 

recover to some extent from the effects of O3 exposure in prior years. 

When potted yellow poplar seedlings were exposed to O3 concentrations up to SUM00 

values of 107 ppmAh in one year and 197 ppmAh in a second year, no effects on biomass were 

observed (Rebbeck, 1996). In a study at the same location with seedlings planted in the ground 

and exposed to O3 concentrations with SUM06 values of 0.3, 228.7, and 661.8 ppmAh over 

5 years, no effects on biomass were found (Rebbeck and Scherzer, 2002). 

Many studies have demonstrated that root growth is more sensitive to O3 exposure than is 

stem growth. For example, in a study with black cherry seedlings exposed in OTCs in 

Tennessee in 1989, root biomass in a 2×-ambient treatment was decreased by 42%, while stem 

biomass was decreased by only 24%. However, in a second experiment in 1992, root and stem 

growth reductions in the 2×-ambient treatment were similar (65% versus 60%) (Neufeld et al., 

1995). In Finland, reduced root growth was found for a number of clones of silver birch 

(Oksanen and Saleem, 1999). After 5 years, root growth was decreased by 33%, but shoot 

growth was not affected by O3 exposures of a 7-h mean of 15 ppmAh over 5 years in a FACE 

system (Oksanen, 2001). When first-year poplar seedlings (P. tremuloides) were exposed in 

OTCs to two O3 concentrations and six N concentrations, the root/shoot ratio was decreased soon 

after exposure to O3 in all N treatments, even though O3 effects on total biomass were not 

detected in the low-N and very high-N treatments (Pell et al., 1995). These results suggest that 

effects on the root/shoot ratio occur before significant growth effects arise. In a series of OTC 

experiments lasting 1 to 3 years at 3 different elevations in Switzerland, fine root growth in 

European beech was found to be more sensitive to O3 than was shoot or total biomass (Braun and 

Fluckiger, 1995). Although the estimated effect of O3 on fine root biomass was similar to that 

for total biomass, fine root biomass was significantly decreased at AOT40 (24-h) values of only 

3 ppmAh, while total biomass was not significantly decreased until AOT40 values reached 30 to 

40 ppmAh. 

AX9.5.5.3 European Critical Levels 

In Europe, a Level I critical level has been set for forest trees based on OTC studies of 

saplings. This level is discussed here because it was based on a deciduous tree species. For 

consistency with the approach used for crops, an AOT40 index value was selected. A few 

AX9-243
studies have shown that O3 can be taken up by tree species at nighttime, e.g., young birch trees 

(Matyssek et al., 1995). However, because most evidence suggests that O3 deposition at 

nighttime is low (Coe et al., 1995; Rondon et al., 1993), a value for only daylight hours was 

selected in Europe (Fuhrer et al., 1997; Kärenlampi and Skärby, 1996). European beech was 

selected for development of a Level I critical level, because data from several studies were 

available for this species and because deciduous tree species were judged to be more sensitive to 

O3 compared to evergreen tree species (Fuhrer et al., 1997; Kärenlampi and Skärby, 1996). 

A critical level was defined as an AOT40 value of 10 ppmAh for daylight hours for a 6-month 

growing season (Kärenlampi and Skärby, 1996). However, other studies have shown that other 

species such as silver birch may be more sensitive to O3 than beech (Pääkkönen et al., 1996). 

Level I critical values are not designed for making quantitative estimates of the O3 effects on 

vegetation at the regional scale, instead a so-called Level II critical value is required for this 

purpose. For long-lived perennials, additional problems complicate extrapolation. As discussed 

below (Section AX9.5.5.7), considerable scaling is required to extrapolate from experiments 

conducted with tree seedlings to estimate effects on mature trees in forests. Because of these 

scaling issues, there is greater uncertainty in estimating effects on forest trees than on annual 

plants such as crops. While some information is available for addressing issues such as scaling 

from seedlings to mature trees and estimating O3 uptake, this information may still be 

insufficient for developing a Level II approach that can provide quantitative estimates of forest 

growth losses due to O3 (Broadmeadow, 1998). 

AX9.5.5.4 Summary of Effects on Deciduous Woody Species 

Recent evidence from free air exposure systems and OTCs supports results observed 

previously in OTC studies (Table AX9-15, Figure AX9-18). Specifically, a series of FACE 

studies undertaken in Rhinelander, WI (Isebrands et al., 2000, 2001) showed that O3-symptom 

expression was generally similar in OTCs, FACE, and ambient-O3 gradient sites, supporting the 

previously observed variation among aspen clones using OTCs (Karnosky et al., 1999). This 

study also found no effects on sugar maple growth after 3 years, but in years 4 to 7 found 

significant growth reduction due to O3 (Karnosky et al., 2005). These results are important, 

because they indicate results obtained from OTCs are supported by results from free air exposure 

systems and also that more than 3 years may be required to adequately investigate the effects of 

AX9-244
O3 on the growth of tree species. Finally, this study found that competition may alter the effect 

of O3, depending on environmental conditions and genotype (McDonald et al., 2002). New 

evidence is also available comparing various aspects of O3 sensitivity between seedlings and 

mature trees of some species, notably red oak. As has been observed in previous O3 criteria 

documents, root growth is often found to be the most sensitive indicator in terms of biomass 

response to O3. 

Results since 1996 support the conclusions of the 1996 AQCD (U.S. Environmental 

Protection Agency, 1996) that individual deciduous trees are generally less sensitive to O3 than 

are most annual plants, with the exception of a few genera such as Populus, which are highly 

sensitive. However, the data presented in Table AX9-18 suggest that ambient exposures that 

occur in different regions of the United States can sometimes reduce the growth of seedlings of 

deciduous species. Results from multiple-year studies sometimes show a pattern of increasing 

effects in subsequent years. Although, in some cases, growth decreases due to O3 become less 

significant or even disappear over time. While some mature trees show greater O3 sensitivity in 

physiological parameters such as net photosynthetic rate compared to seedlings, these effects 

may not translate into measurable reductions in biomass growth. Because even multiple-year 

experiments do not expose trees to O3 for more than a small fraction of their life span and 

because competition may, in some cases, exacerbate the effects of O3 on individual species, 

determining the effects on mature trees remains a significant challenge. Effects on mature trees 

under natural conditions are discussed after the review of evergreen species below and more 

fully in Section AX9.6, in the context of extrapolating from controlled studies to forest 

ecosystems. 

AX9.5.5.5 Evergreen Woody Species 

Most investigations have shown evergreen tree species to be less sensitive to O3 compared 

to deciduous species (U.S. Environmental Protection Agency, 1996). For example, exposureresponse functions based on 23 experimental cases of seedling response to O3, suggest that a 

SUM06 exposure for 3 months of 42.6 ppmAh would protect evergreen species from a 10% 

growth loss in 50% of the cases (Table AX9-15). For deciduous species, the corresponding 

SUM06 value was 31.5 ppmAh (Table AX9-15). As another example, experiments in the Great 

Smoky Mountains National Park found black cherry seedlings to demonstrate substantial 

AX9-245
decreases in biomass, as discussed above and shown in Table AX9-18 (Neufeld et al., 1995). 

However, exposure for up to three growing seasons did not decrease the biomass of eastern 

hemlock, Table Mountain pine, and Virginia pine seedlings exposed to O3 under similar 

conditions in this location, as shown in Table AX9-19 (Neufeld et al., 2000). 

As for deciduous species, there is a substantial range in sensitivity among evergreen 

species. As discussed above for deciduous species, a risk analysis was undertaken to predict tree 

biomass growth reductions due to O3 based on exposure-response equations for tree seedlings 

combined with species distribution across the eastern United States and interpolated O3 exposure 

(Hogsett et al., 1997). While some species such as Virginia pine were predicted to be affected 

only slightly even in a high O3 year, the growth of sensitive evergreen species such white pine 

was predicted to be reduced by 5% in a lower-than-average O3 year and 10% in a high O3 year 

across 50% of its range (Andersen et al., 1997). The remainder of this section discusses 

experimental results for evergreen species in the order shown in Table AX9-19. 

Douglas fir seedlings were exposed to elevated O3 concentrations in a free air zonal air 

pollution system in British Columbia, Canada for two growing seasons with 12-h mean values in 

1988 of 18 to 41 ppb and in 1989 of 27 to 66 ppb (Runeckles and Wright, 1996). Although 

substantial variation was seen in effects among the different treatments, there was a significant 

decrease in the growth of the second flush weight in the second year, with reductions of 55% at 

the highest O3 exposure, based on a linear regression. This result contrasts with the lack of 

effect seen in a previous study with seedlings of this species grown in pots for 134 days and 

exposed to 7-h mean O3 concentrations up to 71 ppb (Table 9-30 in U.S. Environmental 

Protection Agency, 1996). 

First-year loblolly pine seedlings of 53 open-pollinated families were exposed to 

1×-ambient O3 in OTCs for a single growing season, and average growth volume was decreased 

by 14% compared to a CF treatment (McLaughlin et al., 1994). The 1×-ambient O3 exposure in 

this study expressed as 24-h SUM00 was 137 ppmAh, and the CF treatment reduced O3 by 47%. 

In this study, the root-to-shoot ratio was decreased significantly in six of the nine families 

examined. Exposure to O3 with SUM06 values up to 162 ppmAh and 132 ppmAh in 2 successive 

years in OTCs had no effect on seedlings grown in competition with various species of grasses 

and forbs (Barbo et al., 2002). Exposure of 3-year-old seedlings to O3 exposures of up to 

AX9-246
2.5×-ambient (12-h mean of 98 ppb) also had no significant effect (Anttonen et al., 1996). 

Four-week-old loblolly pine seedlings were grown in large OTCs in Alabama and exposed to 

CF, 1× ambient, 2× ambient O3 treatments in 2 one year experiments, with seasonal 12-h 

mean O3 concentrations of 13, 47, 98 ppmAh in 1998 and 12, 44, 97 ppmAh in 1999 (Estes et al., 

2004). Shoot biomass was decreased 15% in the 1× treatment and 22% in 2× in both years, 

while root biomass was decreased by 26% in the 2× treatment in both years. Foliar symptoms 

were significantly greater in the 1× treatment iin1999 and in the 2× treatment in both years. 

Information summarized in the 1996 AQCD (U.S. Environmental Protection Agency, 1996), 

indicated that significant effects on seedling growth were observed in several studies of 

seedlings exposed to elevated O3 concentrations for one or more years. Several studies, 

including that of McLaughlin et al. (1994), demonstrate considerable variation in O3 sensitivity 

among different genotypes of loblolly pine. 

For Ponderosa pine seedlings, the 1996 AQCD reviewed a number of studies with 

exposures to elevated O3 concentrations for one to three growing seasons (U.S. Environmental 

Protection Agency, 1996). More recent similar studies support the earlier results (Table 

AX9-19) (Andersen et al., 2001; Takemoto et al., 1997). The 1996 criteria document also 

discussed at some length the ongoing work examining effects of O3 across a naturally-occurring 

O3 gradient in the San Bernardino Mountains in California. Since that time, much research on 

ponderosa pine has focused on interactive effects of additional stresses such as nitrogen and on 

effects of O3 on physiological parameters (Sections AX9.3, AX9.6). Effort has also been 

focused on the effects of O3 on root growth because such effects could alter sensitivity to 

drought or nutrient stress. Ecosystem level effects of O3 are discussed further in Section AX9.6, 

but some information relevant to exposure-response relationships is discussed below. 

For several tree species, O3 has been shown in experimental studies with seedlings to 

reduce root growth more than shoot growth (U.S. Environmental Protection Agency, 1996). 

Ponderosa pine has been shown to be sensitive to O3, and studies with seedlings have shown 

reduced root growth, decreases root-to-shoot ratios, and decreased allocation to roots (Andersen 

et al., 1991, 1997; Andersen and Rygiewicz, 1995; Andersen and Scagel, 1997; Tingey et al., 

1976b). Data from a long-studied pollution gradient in the San Bernardino Mountains of 

southern California suggests that O3 substantially reduces root growth in natural stands of 

AX9-247
ponderosa pine. Root growth in mature trees was decreased at least 87% in a high pollution site 

as compared to a low pollution site (Grulke et al., 1998a), and a similar pattern was found in a 

separate study with whole tree harvest along this gradient (Grulke and Balduman, 1999). 

Because other potential influences on root growth, including shading by competing trees, soil 

temperature, soil moisture, phenology, were not correlated with the observed pattern of reduced 

root growth, the authors conclude that O3 was the cause of the observed decline in root growth. 

Further results of field investigations with ponderosa pine and other pine species native to 

California are discussed below under the heading “Scaling experimental data on long-lived 

species to field conditions” as well as in Section AX9.6. 

Table Mountain pine, Virginia pine, and eastern hemlock seedlings were exposed to 

various levels of O3 (CF to 2× ambient) in OTCs for in a series of experiments two or three years 

in Great Smoky Mountains National Park in Tennessee (Neufeld et al., 2000). There were no 

statistically significant effects of O3 exposure on stem or root biomass, and only slight effects on 

the biomass of the oldest needles in Table Mountain pine in the 2× ambient treatment. 

As reviewed in the 1996 criteria document, studies of the response of red spruce to O3 

exposures generally have not found effects on growth of seedlings or saplings, even after 

exposure to high concentrations (12-h mean of 90 ppb) for up to 4 years. A report since that 

time confirms that this slow-growing species is O3 insensitive for at least several years (Laurence 

et al., 1997). 

For perennial vegetation, cumulative effects over more than one growing season may be 

important. For 3-year-old Norway spruce in Sweden, exposure to elevated O3 for three growing 

seasons decreased total biomass by 18% and stem biomass by 28% (Karlsson et al., 1995). 

However, after a fourth season of exposure, total biomass decreased significantly by only 8% 

(Karlsson et al., 2002). In this experiment, the O3 exposures expressed as 12-h mean values 

averaged over four growing seasons were 12 and 44 ppb for the CF and 1.5×-ambient treatments, 

respectively; and AOT40 values were 2 and 23 ppmAh, respectively. Despite 4 years of 

exposure, this experiment did not demonstrate a consistent trend in the O3 effect on biomass that 

would suggest a significant carry-over effect. However, a study of 3- to 7-year-old Norway 

spruce in OTCs in Finland found a 5.3% decrease in total plant biomass after 7 years, with an 

elevated O3 AOT40 exposure value of 79 ppmAh (Ottosson et al., 2003; Wallin et al., 2002). 

AX9-248
AX9.5.5.6 Summary of Effects on Evergreen Woody Species 

In summary, the O3 sensitivity of different genotypes within species and between species 

of evergreen vegetation varies widely. Based on studies with evergreen seedlings in OTCs, 

major species in the United States are generally less sensitive than are most deciduous trees, and 

slower-growing species are less sensitive than faster-growing ones. However, exposure to 

ambient O3 may reduce the growth of seedlings of commonly occurring species. Because tree 

species are long-lived, most experiments have only covered a very small portion of the life span 

of a tree, making estimation of any effect on mature trees difficult. Considerations for scaling 

the results of seedling studies to mature forest trees as well as additional information from field 

surveys and studies of mature trees under natural conditions are discussed below and in 

Section AX9.6. 

AX9.5.5.7 Scaling Experimental Data to Mature Trees 

As compared with annual crop species, it is much more difficult to define appropriate 

exposure-response relationships for tree species. For annual species, an experiment may cover 

the whole life span of the plant, but it is difficult and expensive to provide controlled-exposure 

conditions for long-lived plants for any significant portion of their life spans, although a few 

FACE studies have demonstrated that it is feasible. However, FACE studies cannot investigate 

the effects of ambient O3 exposures, because lower-than-ambient O3 treatments cannot be 

applied. Most studies have used small seedlings, because they are manageable under 

experimental conditions; but seedlings and mature trees may have different sensitivities to O3. 

For perennial species, effects of O3 may accumulate over more than 1 year, and may interact 

with other stresses such as drought stress over multiple growing seasons. As for annual species 

(Section AX9.3.2), substantial variability occurs among evergreen genotypes and this variation 

may interact with other stress responses differently in different landscapes and regions. Despite 

these difficulties, investigators have addressed some of these issues since the 1996 AQCD (U.S. 

Environmental Protection Agency, 1996). New information is available on the response of 

mature evergreen trees to O3 under field conditions, and models based on tree physiology and 

stand dynamics have been used to predict O3 effects on forest stands and regions. The following 

issues are reviewed briefly below: (1) interaction of drought and O3 stress, (2) scaling data from 

AX9-249
seedlings to mature-tree studies. Two additional scaling issues are addressed in Section AX9.6: 

(1) scaling data to forest stands, and (2) scaling data to ecosystems and regions. 

AX9.5.5.7.1 Interactive Effects of Drought and Ozone 

Many interacting factors may influence the effect of O3 on vegetation. For crop plants, 

environmental conditions are often managed such that nutrients and water are not strongly 

limiting; but for native vegetation, including most perennial species, such factors are likely to 

limit growth. The effects of interacting stresses on vegetation were reviewed in Section AX9.3. 

However, because drought is common in many forests, and because there is substantial evidence 

that it alters the response of trees to O3, it is discussed in this section in the context of 

determining exposure-response relationships for trees. 

Controlled experiments with seedlings provide direct evidence that drought can reduce the 

impact of O3. For example, for 3-year-old Norway spruce in Sweden, exposure to elevated O3 

for three growing seasons decreased total biomass by 18% and stem biomass by 28% (Karlsson 

et al., 1995). However, for droughted trees, both total and stem biomass decreased only 5%, 

with a statistically significant interaction with O3 for stem biomass. Yet after a fourth season of 

exposure, there was no longer any interaction between drought and O3, while there was a 

significant decrease of 8% in the biomass when both drought and well-watered data were 

combined (Karlsson et al., 2002). In this study, seedlings were grown in sand in 120-L pots and 

for the drought treatment, water was withheld for 4 weeks during the first year and for 7 to 8 

weeks during each of the last 3 years. In this experiment, the O3 exposures expressed as 12-h 

seasonal daylight mean averaged over four growing seasons were 12 and 44 ppb for the CF and 

1.5×-ambient treatments, respectively. Over this period, the AOT40 values for the treatments 

averaged 2 and 23 ppmAh respectively. Despite 4 years of exposures, this experiment did not 

demonstrate a consistent trend in drought O3 interactions. The difference in effects seen between 

the third and fourth season suggest that scaling drought-O3 interactions from seedlings to mature 

trees may be difficult. However, evidence from biomonitoring surveys supports an interaction 

between drought and O3 effects, at least for foliar symptoms. In systematic surveys of foliar 

symptoms on species selected as biomonitors throughout much of the eastern United States, 

symptoms were more common and more severe in areas with high O3 concentrations (Smith 

AX9-250
et al., 2003). However, in very dry years, such as 1999, the occurrence and severity of 

symptoms was greatly reduced, even in areas with high ambient O3 concentrations. 

AX9.5.5.7.2 Scaling from Seedlings to Mature Trees 

Because most experiments are conducted with seedlings, various methods are required to 

scale experimental data on seedlings to mature trees. An overview of physiological differences 

between young and old plants, and the consequences of these differences for O3 sensitivity, was 

provided in Section AX9.3.5.3. The discussion below focuses on information relevant to 

developing exposure-response relationships for mature trees. Information from a few 

experimental studies, as well as scaling efforts based on physiological characteristics 

incorporated into models, are discussed in Section AX9.6. 

Although most studies continue to examine the effects of O3 on seedlings, during the 1990s 

some studies examined the effects of O3 on the response of mature trees. Studies of mature trees 

demonstrate differences in some aspects of O3 sensitivity between seedlings and mature trees. 

For some species, such as red oak, seedlings are less sensitive to O3 than are mature trees 

(Hanson et al., 1994; Samuelson and Edwards, 1993; Wullschleger et al., 1996). Both red oak 

seedlings and genetically related mature trees were exposed to CF, 1×-ambient, or 2×-ambient 

O3 exposures in OTCs in Tennessee for two growing seasons (Hanson et al., 1994). Nine large 

chambers (4.6 × 8.2 m) were used to enclose individual mature trees and standard EPA-style 

OTCs were used for potted seedlings. Ozone exposures expressed as a 24-h SUM00 were 34, 

79, and 147 ppmAh in 1992 and 37, 95, 188 ppmAh in 1993 for the sub-ambient, and 2×-ambient 

treatments. Mature trees had a greater light-saturated net photosynthetic rate and stomatal 

conductance compared to seedling foliage at physiological maturity. By the end of the growing 

season, exposure to 1×-ambient and 1×-ambient O3 reduced the light-saturated net 

photosynthetic rate and stomatal conductance of mature trees by 25 and 50%, respectively, 

compared with the CF treatment (35 ppmAh). In seedlings, however, light-saturated net 

photosynthetic rate and stomatal conductance were less affected by O3 exposure. The authors 

concluded that extrapolations of the results of seedling-exposure studies to foliar responses of 

mature forests without considering differences in foliar anatomy and stomatal response between 

juvenile and mature foliage may introduce large errors into projections of the O3 responses of 

mature trees. 

AX9-251
In a study of ponderosa pine in California, seedlings and branches of mature trees (in 

branch chambers) were exposed to O3 concentrations of 0.5-, 1-, and 2×-ambient O3 

concentrations (Momen et al., 1997). Net photosynthetic rate of 1-year-old, but not current-year, 

foliage was reduced in mature trees but not significantly reduced in seedlings. This effect was 

not due to alteration of stomatal conductance by O3. This result contrasts with those with earlier 

studies of red spruce (Rebbeck et al., 1993). 

In contrast to the findings for red oak and Ponderosa pine, giant sequoia seedlings had 

higher rates of stomatal conductance, CO2-exchange rate, and dark respiration than did mature 

trees (Grulke and Miller, 1994). As compared to older trees, stomatal conductance was more 

than 7-fold greater in current-year, and 4-fold greater in 2-year-old, seedlings (Grulke and 

Miller, 1994). The authors concluded that giant sequoia seedlings are sensitive to atmospheric 

O3 until ~5 years of age. Low conductance, high water use efficiency, and compact mesophyll 

all contribute to a natural O3 tolerance, or to O3 defense, or to both, in the foliage of older trees. 

Similarly, lower stomatal conductance was found in mature Norway spruce in Austria 

compared to seedlings grown with optimal water and nutrients in a growth chamber (Wieser, 

1997). In this study, net photosynthetic rate was less sensitive to added O3 in mature trees 

compared to seedlings. In a related study, the average rate of O3 uptake of 17-year-old trees 

~0.6 nmol m!2 s

!1

, decreasing linearly in older trees, such that rates were only ~0.1 m!2 s!1 in 

216-year-old trees (Wieser et al., 1999). 

Based on a review of studies of stomatal conductance in both seedlings and mature trees, 

Samuelson and Kelly (2001) concluded that O3 uptake in oak species, black cherry, sugar maple, 

and American beech averaged 47% lower in potted seedlings than in mature trees. For evergreen 

species, they concluded that O3 uptake in seedlings averaged 26% higher than in mature trees. 

They also suggested that artifacts introduced by growth in pots confound these differences that 

exposure-response functions derived from seedlings grown in situ are more applicable to mature 

trees than are studies of seedlings grown in pots (Samuelson and Kelly, 2001). 

As discussed above for annual vegetation, it has long been noted that internal O3 dose is 

more appropriate than external O3 exposure for assessing the effects of O3 on vegetation, because 

effects occur primarily via the uptake of O3 through the stomata (Section AX9.2.2). However, 

external O3 exposure sometimes has been shown to explain O3 effects as well or better than 

calculated internal O3 dose. For ponderosa pine, Grulke and others (2002b) found little 

AX9-252
difference in the response of net photosynthetic rate and stomatal conductance to O3 exposure as 

compared to calculated O3 uptake; and estimated O3 uptake by ponderosa pine and O3 exposure 

at several sites were highly correlated (r2

 = 0.92). For red oak, Hanson and others (Hanson et al., 

1994) found that SUM00 explained 83% of the variance in the response of light-saturated 

photosynthetic rate to O3 levels, while estimated internal dose explained only 76% of the 

variance. In this same study, SUM06 explained only 49% of the variance. Due to genetic 

variation or other factors, individual mature trees will vary in their response to similar O3 

exposures. For example, in 125-year-old giant sequoia trees exposed to ~230 ppmAh of O3 in 

branch chambers, O3 uptake in one individual was ~5 mmol m!2

, while in another it was 

~9.5 mmol m!2 (Grulke et al., 1996). 

Based on these results, stomatal conductance, O3 uptake, and O3 effects cannot be assumed 

to be equivalent in seedlings and mature trees. In general, mature deciduous trees are likely to 

be more sensitive to O3 compared to seedlings, while mature evergreen trees are likely to be less 

sensitive than seedlings. However, even when differences in physiological traits occur, 

concomitant effects on stem growth may not be detected in the field. Additionally, complex 

interactions may occur between environmental conditions and O3 responses; and artifacts may 

occur for seedling studies, especially for seedlings grown in pots. Finally, competition between 

species or genotypes within a species can either exacerbate or ameliorate the effects of O3. Such 

effects are predicted by models of the growth of mixed species forests, as discussed in Section 

AX9.6, and various interactions between competitive ability and O3 effects have been found for 

aspen clones in the Wisconsin FACE study (McDonald et al., 2002). Section AX9.6 further 

discusses issues that must be addressed when scaling data from individual mature trees to forests 

and regions. 

AX9.5.6 Studies With the Chemical EDU 

The chemical EDU has been used with the goal of protecting plants from O3 effects 

without controlling O3 exposure (Table AX9-21) (U.S. Environmental Protection Agency, 1986, 

1996). As discussed in Section AX9.1.3.3, the use of EDU has the potential to be a low-cost, 

practical method of evaluating ambient O3 exposures on plants grown under natural conditions 

without the limitations imposed by methodologies such as OTCs (Section AX9.1.3.3). However, 

because EDU is phytotoxic, and may have effects on plants other than antioxidant protection, 

AX9-253
Table AX9-21. Ethylene Diurea Effects on Vegetation Responses to Ozone 

Species Description EDU Application Ozone Exposure Effects of EDU Bean, cv. Lit 10-cm pots in OTCs in Soil drench 200 mL of CF, NF, CF-1×, CF-2×: 

O3 reduced pod, shoot, and 

Germany 150 ppm solution per mean = 1, 14, 15, 32 ppb root mass. EDU increased 

plant every 14 days root, leaf, and shoot mass, 

but a significant interaction with O3 occurred only for 

root weight. 

Bean, cv. BBL-290 2 expts in 5.5 L pots Soil drench every 14 2 expts with CF, NF and Visible injury and reduced 

in OTCs with 4 O3 days, in expt 1 = 0.14, 2 constant additions of total biomass or yield, even 

treatments 28, 56, 120 mg/L potting 

O3. 7-h mean O3 (ppb) CF treatment. Within an O

medium; expt 2 = 0, 8, for Expt 1 = 34, 70, 95, treatment, sometimes 16, 32 mg/L 121; Expt 2 = 19, 42. increased yield (Expt 2 only

74, 106 

Bean, cv. Bush Blue Field-grown in fine Foliar spray at 300 ppm Ambient, with 181 and EDU increased final aboveLake 290, Bush Blue sandy loam in every 7 days between 141 h > 40 ppb, 74 and ground biomass in S156 in Lake 274, lines S156, Massachusetts, in full expansion of primary 95 h> 60 ppb, and 23 h > 2001 and 2002, but R123 2 one-year experiments. leaves and pod 80 ppb in 2001 and 2002 significantly decreased 

senescence. respectively above-ground biomass in 

R123 in both years and BBL 274 in 2002. 

Bean Pots with potting mix Soil drench of 200 mL of AOT40 = 0.4-1.8 ppm-h 0 to 50% increase in pod 

at 3 locations in Spain increasing concentrations mass, but did not restore (2 years at 1 site) of 100, 150, 200, 250 yield at sites with higher O

ppm every 14 days (4-10 mg l:1 soil) 

Bean, cv. Lit Pots with potting mix at Soil drench of 200 mL of AOT40 = 0.64-0.98, Average 20% yield increas

4 sites in the Netherlands increasing concentrations 7-h mean = 49-55 ppb at all sites. 

of 100, 150, 200, 250 ppm every 14 days (4-10 mg l:1 soil) 

Bean, cv. Lit Pots with potting mix at Soil drench of 200 mL of AOT40 = 0.81 ppm-h 16% yield increase. 

1 site in Belgium increasing concentrations 

of 100, 150, 200, 250 ppm every 14 days (4-10 mg l:1 soil) 

AX9-254
Table AX9-21 (cont’d). Ethylene Diurea Effects on Vegetation Responses to Ozone 

Species Description EDU Application Ozone Exposure Effects of EDU Clover, subterranean Plants in 10-cm pots at 100 ml of 150 ppm AOT40 = 0-0.56 ppm-h Injury, but not leaf biomass 

4 rural sites in the solution as soil drench for 4-week periods was affected by EDU and ONetherlands for 3 years every 14 days for exposure. 

2 months 

Clover, white 15-cm pots in field, well 100 mL of 150 ppm AOT40 (28 days) = 0-20 Change in biomass ratio, 

watered, 12 locations throughout Europe, 

solution as soil drench every 14 days for 

ppm-h weak linear relationship 

(r2 = 0.16) stronger 

3 years 3 months relationship using ANN 

and climatic factors 

Clover, white, 2 expts,10-cm pots in 100 mL of150 ppm AOT40 = 15.5, n.s. cv. Menna field in Italy, see also solution as soil drench 12.1 ppm•h; 

companion OTC expt every 14 days for 7-h mean = 69, 60 

2 months 

Clover, white, 2 expts,10-cm pots in Soil drench 100 mL of CF, NF, NF+ 25, No effect of EDU despite cv. Menna OTCs in Denmark 150 ppm solution every NF+50 ppb, highly significant effect of 

14 days 

O3 exposure not reported on above-ground biomass 

Poplar, hybrid Stem injections, field, Approx. 125 or 250 1991: No effect on biomass; 6%, 

cuttings, 1 or 2 years mg/leaf (low, high EDU 7-h mean = 56, 12% more severely O3 

treatments) AOT40 = 23; damaged leaves in high 5 times every 14 days 1992: EDU for 2 years 

7-h mean = 59, AOT40 = 27 

Pine, loblolly 1 year old half-sib 150, 300, 450 ppm every 1995, 1996, 1997, For EDU 450 trt, above- 

seedlings in field in 14 days no. h > 40 ppb = 1723, ground biomass increased TX for 3 years 2297, 2052; approx 46% (n.s. in other 

no. h > 60 ppb = 378, treatments 584, 528; peak = 113, 102, 118 

Radish, cv. Cherry Belle Plants in pots in potting Soil drench containing 24-h mean = 31 ppb, 24% increase in hypocotyl 

mix exposed for 5 weeks 20 mg EDU applied 7-h mean = 36 ppb, mass, 18% increase in in southern Sweden. 2 times, 14 days apart AOT40 = 1.3 ppm-h. shoot mass 

AX9-255
it is crucial that the correct dosage for protection from O3 be determined without the direct 

effects of EDU. For example, a study in Massachusetts applied EDU to foliage of field-grown 

bush beans of two lines (S156, R123) and two cultivars (Bush Blue Lake 290, Bush Blue Lake 

274) (Elagöz and Manning, 2005a). EDU increased the above-ground biomass of one line but 

decreased the above-ground biomass in the other line and in one of the cultivars. Other studies 

have shown that EDU does not always have greater effects at higher O3 exposures (Ribas and 

Penuelas, 2000; Tonneijck and Van Dijk, 1997). Such results suggest that it may be difficult to 

quantify ambient O3 effects using EDU, because the amount of plant growth or yield expected at 

a low (background) O3 concentration cannot be inferred from EDU-treated plants grown at 

locations with higher O3 exposures. Unfortunately, although many studies with EDU have been 

conducted in recent decades, very few have used multiple EDU application levels along with 

multiple O3 exposures to characterize the EDU system for a given plant species. Therefore, the 

text of this section focuses on how data from existing studies can be used for developing or 

validating exposure-response relationships, rather than reviewing results of all individual studies. 

Data from individual studies on O3 exposure, EDU application rates, and the effects of EDU are 

presented in Table AX9-21. In addition to EDU, sodium erythorbate has been used in a few 

studies as a protectant chemical. Since very few published studies have used sodium erythorbate 

and attempts to establish appropriate doses for individual species are even more limited, the use 

of this chemical is not reviewed here. 

In summary, it is difficult to use data from existing EDU studies to develop exposureresponse relationships or to quantify the effects of ambient O3 exposure. Despite these 

limitations, the EDU studies reviewed in previous criteria documents (U.S. Environmental 

Protection Agency, 1986, 1996) and the more recent studies summarized in Table AX9-21 

provide another line of evidence that ambient O3 exposures occurring in many regions of the 

United States are reducing the growth of crops and trees. 

AX9.5.7 Summary 

Data published during the last decade support the conclusions of previous criteria 

documents that there is strong evidence that ambient concentrations of O3 cause injury and 

damage to numerous common and economically valuable plant species. For annual vegetation, 

the data summarized in Table AX9-16 show a range of growth and yield responses both within 

AX9-256
and among species. Nearly all of these data were derived from studies in OTCs, with only two 

studies using open-air systems in the UK (Ollerenshaw et al., 1999; Ollerenshaw and Lyons, 

1999). It is difficult to compare studies that report O3 exposure using different indices, such as 

AOT40, SUM06, or 7-h or 12-h mean values. However, when such comparisons can be made, 

the results of more recent research confirm earlier results summarized in the 1996 O3 AQCD 

(U.S. Environmental Protection Agency, 1996). A summary of earlier literature concluded that a 

7-h, 3-month mean of 49 ppb corresponding to a SUM06 exposure of 26 ppmAh would cause 

10% loss in 50% of 49 experimental cases (Tingey et al., 1991). More recent data summarized 

in Table AX9-16 support this conclusion, and more generally indicate that ambient O3 exposures 

can reduce the growth and yield of annual species. Some annual species such as soybean are 

more sensitive, and greater losses may be expected (Table AX9-16). A two-year study using a 

free-air exposure system with soybean confirmed yield reductions found previously using OTCs 

(Morgan et al., in press). Thus, the more recent scientific literature supports the conclusions of 

the 1996 O3 AQCD (U.S. Environmental Protection Agency, 1996) that ambient O3 

concentrations are probably reducing the yield of major crops in the United States. 

Much research in Europe has used the AOT40 exposure statistic, and substantial effort has 

gone into developing Level-1 critical levels for vegetation using this index. Based on regression 

analysis of 15 OTC studies of spring wheat, including one study from the United States and 

14 from locations ranging from southern Sweden to Switzerland, an AOT40 value of 5.7 ppmAh 

was found to correspond to a 10% yield loss, and a value of 2.8 ppmAh corresponded to a 5% 

yield loss (Fuhrer et al., 1997). Because a 4 to 5% decrease could be detected with a 99% 

confidence level, a critical level of an AOT40 value of 3 ppmAh was selected in 1996 

(Kärenlampi and Skärby, 1996). 

In addition to likely reductions in crop yield, O3 may also reduce the quality or nutritive 

value of annual species. Many studies have shown effects of O3 on various measures of plant 

organs that affect quality, with most studies focusing on characteristics important for food or 

fodder. These studies indicate that there may be economically important effects of ambient O3 

on the quality of crop and forage species. Previous O3 criteria documents have concluded that 

visible symptoms on marketable portions of crops and ornamental plants can occur with seasonal 

7-h mean O3 exposures of 40 to 100 ppb (U.S. Environmental Protection Agency, 1978, 1986, 

1996). The more recent scientific literature does not refute this conclusion. 

AX9-257
The use of OTCs may reverse the usual vertical gradient in O3 that occurs within a few 

meters above the ground surface (Section 3.3.2, Section AX9.1). This reversal suggests that 

OTC studies may somewhat overestimate the effects of an O3 concentration measured several 

meters above the ground. However, such considerations do not invalidate the conclusion of the 

1996 AQCD (U.S. Environmental Protection Agency, 1996) that ambient O3 exposures 

(Tables AX9-13 and AX9-21) are sufficient to reduce the yield of major crops in the United 

States. Recent studies using OTCs confirmed previous results that ambient ozone exposures can 

significantly reduce the growth of annual species and the yield of commercially important crop 

species. Additionally, a two-year study using a free air exposure system with soybean confirmed 

yield reductions found previously using OTCs (Morgan et al, in press). 

As for single-season agricultural crops, yields of multiple-year forage crops are reduced 

at O3 exposures that occur over large areas of the United States. This result is similar to that 

reported in the 1996 O3 AQCD (U.S. Environmental Protection Agency, 1996). When species 

are grown in mixtures, O3 exposure can increase the growth of O3-tolerant species and 

exacerbate the growth decrease of O3-sensitive species (e.g., Ashmore and Ainsworth, 1995; 

Fuhrer et al., 1994). Because of this competitive interaction, the total growth of the mixedspecies community may not be affected by O3 exposure (Ashmore and Ainsworth, 1995; Barbo 

et al., 1998; Fuhrer et al., 1994). However, in some cases, mixtures of grasses and clover species 

have shown significant decreases in total biomass growth in response to O3 exposure in studies 

in the United States (Heagle et al., 1989; Kohut et al., 1988) and in Sweden (Pleijel et al., 1996). 

In Europe, a provisional critical level for herbaceous perennials of an AOT40 value of 7 ppmAh 

over 6 months has been proposed to protect sensitive plant species from adverse effects of O3. 

For deciduous tree species, recent evidence from free air exposure systems and OTCs 

supports results observed previously in OTC studies. For example, a series of FACE studies 

undertaken in Rhinelander, WI (Isebrands et al., 2000, 2001) showed that O3-symptom 

expression was generally similar in OTCs, FACE, and also at sites along an ambient O3 gradient, 

supporting the previously observed variation among aspen clones using OTCs (Karnosky et al., 

1999). This study also found no effects on sugar maple growth after 3 years, but in years 4 to 

7 found significant growth reduction due to O3 (Karnosky et al., 2005). These results are 

important because, they indicate results obtained from OTCs are supported by results from free 

air exposure systems and also that more than 3 years may be required to adequately investigate 

AX9-258
effects of O3 on the growth of tree species. As has been observed in previous criteria documents, 

root growth often is found to be the most sensitive biomass response indicator to O3. 

Results reported since 1996 support the conclusion of the 1996 O3 AQCD (U.S. 

Environmental Protection Agency, 1996) that deciduous trees are generally less sensitive to O3 

than are most annual plants, with the exception of a few very sensitive genera such as Populus 

and sensitive species such as black cherry. However, the data presented in Table AX9-18 

suggest that ambient O3 exposures that occur in the United States can potentially reduce the 

growth of seedlings of deciduous species. Results from multiple-year studies sometimes show a 

pattern of increased effects in subsequent years. In some cases, however, growth decreases due 

to O3 may become less significant or even disappear over time. While some mature trees show 

greater O3 sensitivity in physiological parameters such as net photosynthetic rate than do 

seedlings, these effects may not translate into measurable reductions in biomass growth. 

However, because even multiple-year experiments do not expose trees to O3 for more than a 

small fraction of their life span, and because competition may in some cases exacerbate the 

effects of O3 on individual species, determining effects on mature trees remains a significant 

challenge. 

In Europe, a Level I critical level has been set for forest trees based on OTC studies of 

European beech seedlings. A critical level was defined as an AOT40 value of 10 ppmAh for 

daylight hours for a 6-month growing season (Kärenlampi and Skärby, 1996). However, other 

studies show that some species such as silver birch may be more sensitive to O3 compared to 

beech (Pääkkönen et al., 1996). 

For evergreen tree species, as for other tree species, the O3 sensitivity of different 

genotypes and different species varies widely. Based on studies with seedlings in OTCs, major 

species in the United States are generally less sensitive than are most deciduous trees, and 

slower-growing species are less sensitive than are faster-growing species. Interacting stresses 

such as competition stress may increase the sensitivity of trees to O3. As for deciduous species, 

most experiments with evergreen species have only covered a small portion of the life span of a 

tree and have been conducted with seedlings, making estimating effects on mature trees difficult. 

For all types of perennial vegetation, cumulative effects over more than one growing 

season may be important; furthermore, studies for only a single season may underestimate 

effects. Mature trees may be more or less sensitive to O3 than are seedlings, depending on the 

AX9-259
species, but information on physiological traits may be used to predict some such differences. In 

some cases, mature trees may be more sensitive to O3 than seedlings due to differences in their 

gas exchange rates, growth rates, greater cumulative exposures, or due to the interaction of O3 

stress with other stresses. 

AX9.6 EFFECTS OF OZONE EXPOSURE ON NATURAL ECOSYSTEMS 

AX9.6.1 Introduction 

The preceding section on species-level responses (AX9.5) provides a lead-in to address the 

response of ecosystems to O3. The conclusion of the 1996 O3 AQCD (U.S. Environmental 

Protection Agency, 1996) was that aside from the results from the San Bernardino NF, there was 

no direct evidence that O3 is altering natural ecosystems in the United States. This conclusion is 

generally valid today, except that our understanding of the effects of O3 in the San Bernardino 

NF has been tempered by additional understanding of the complicating role that N deposition 

plays in this system. Despite the lack of any new, direct information linking O3 with ecosystem 

changes, numerous publications since 1996 have highlighted ways in which O3 may affect 

ecosystem structure and/or function. This section addresses new and (where appropriate) older 

literature in order to illustrate possible shifts in energy or material flow through ecosystems as a 

result of O3 exposure. 

An ecosystem is defined as comprising all of the organisms in a given area interacting with 

the physical environment, so that a flow of energy leads to a clearly defined trophic structure, 

biotic diversity, and cycling of materials between living and nonliving parts (Odum, 1963). 

Individuals within a species and populations of species are the building blocks from which 

communities and ecosystems are constructed. Classes of natural ecosystems, e.g., tundra, 

wetland, deciduous forest, and conifer forest, are distinguished by their dominant vegetation 

forms. Ecosystems boundaries are delineated when an integral unit is formed by their physical 

and biological parts. Defined pathways for material transport and cycling and for the flow of 

energy are contained within a given integrated unit. 

Each level of organization within an ecosystem has functional and structural 

characteristics. At the ecosystem level, functional characteristics include, but are not limited, to 

energy flow; nutrient, hydrologic, and biogeochemical cycling; and maintenance of food chains. 

AX9-260
The sum of the functions carried out by ecosystem components provides many benefits to 

mankind, as in the case of forest ecosystems (Smith, 1992). These include food, fiber 

production, aesthetics, genetic diversity, and energy exchange. 

Ecosystems are functionally highly integrated. Changes in one part of an ecosystem, such 

as the primary producer component, may have consequences for connected parts, such as the 

consumer and decomposer components. For example, when needles are shed prematurely as a 

result of O3 exposure, successional development of phyllosphere fungi inhabiting the surface of 

Ponderosa pine needles may be truncated (Bruhn, 1980). In addition, decomposer populations in 

the litter layer may be capable of higher rates of decomposition, due to the higher N content of 

the younger age classes of needles falling from O3-damaged pines (Fenn and Dunn, 1989). 

Because ecological systems integrate the effects of many influences, the results of O3 exposure 

may depend on co-occurring influences that predispose an ecosystem to stress (Colls and 

Unsworth, 1992). One important change in our thinking since the 1996 O3 AQCD is that, at the 

high levels of O3 exposure that are known to result in detectable plant responses (>250 ppm h 

accumulated over a growing season), N deposition must also be considered as a concurrent 

stressor. Both O3 exposure and increased N deposition can cause changes in N cycling and 

compartmentalization within ecosystems. 

The vast majority of O3-effects literature addresses individual species responses (see 

Section AX9.6.4.3), as was also true in 1996 (U.S. Environmental Protection Agency, 1996). 

This section differs from the preceding one in that the physiological stress of individual species 

is considered only within the context of its natural ecosystem. Changes in function at the 

individual level propagate through the higher levels of organization, resulting in changes in 

ecosystem structure and function. However, since ecosystem-level responses result from the 

interaction of organisms with one another and with their physical environment, it takes longer 

for a change to develop to a level of prominence at which it can be identified and measured. The 

paucity of scientific literature on O3 effects at the ecosystem level is a result of both the 

complexity of ecological systems, and long response times. In addition, “indirect” effects of O3 

on plants (e.g., effects that alter the plants’ ability to integrate environmental stresses) may be 

more important than the direct effects on photosynthesis and respiration at the leaf level 

(Johnson and Taylor, 1989). 

AX9-261
A conceptual framework (see Table AX9-22) for discussing the effects of O3 on 

ecosystems was developed by the EPA Science Advisory Board (Young and Sanzone, 2002). 

Their six essential ecological attributes (EEAs) include landscape condition, biotic condition, 

organism condition, ecological processes, hydrological and geomorphological processes, and 

natural disturbance regimes (see Table AX9-22). The major ecological effects of O3, and gaps in 

our knowledge of O3 exposure effects at the ecosystem level, are summarized at the end of this 

chapter. While the main focus is O3 effects newly described since the 1996 O3 AQCD (U.S. 

Environmental Protection Agency, 1996), many key historical papers are cited to demonstrate 

ecosystem response, particularly where they remain the only examples in the literature. 

Although the vast majority of published studies focus on individuals, six case studies (five 

field examples and one FACE experiment) have measured several ecosystem components 

simultaneously to better understand the overall ecosystem response to O3. We provide an 

overview of these six studies up-front to provide a context for the subsequent discussion on 

possible ecosystem effects on an EEA basis. 

AX9.6.2 Case Studies 

AX9.6.2.1 Valley of Mexico 

The first evidence of air pollution impacts on vegetation in the Valley of Mexico (Mexico 

City Air Basin) were observations of foliar injury symptoms in bioindicator plants attributed 

to O3, PAN, SO2, and possibly other pollutants (De Bauer, 1972). Subsequently, O3 injury to 

foliage and crowns of pine trees were reported in forests to the south and southwest of Mexico 

City (De Bauer and Hernández-Tejeda, 1986; De Bauer and Krupa, 1990). Ozone is considered 

to be the pollutant with the most severe impacts on vegetation within the Mexico City urban 

zone and in forests downwind of the city. Pinus hartwegii is the most O3-sensitive pine species 

and is severely impacted by high O3 exposures encountered to the south/southwest of the Mexico 

City metropolitan area (Miller et al., 1994). The potential for O3 injury is particularly high in 

this area, because O3 levels are high during the summer rainy season when soil moisture 

availability and stomatal conductance are greatest; these factors enhance O3 uptake and injury. 

Decline of Abies religiosa (oyamel) in the Desierto de los Leones NP is a well-known example 

of dramatic dieback and mortality of entire forest stands due primarily to air pollution stress 

(Alvarado et al., 1993). Other factors, such as a lack of stand thinning, also contribute to forest 

AX9-262
Table AX9-22. Essential Ecological Attributes for Natural Ecosystems Affected by O3 

Category Species Condition Measures References 

Landscape Condition 

C Habitat Types 

Biotic Condition 

C Ecosystems and 

Communities 

Mixed conifer forest Community composition, 

Stand structure 

Miller et al. (1989) 

Miller and McBride (1999a) 

Community Extent

 and Composition 

Pinus ponderosa Relative abundance Miller (1973); 

Arbaugh et al. (2003) 

Grassland communities Species composition Ashmore et al. (1995); 

Ashmore and Ainsworth (1995) 

Coastal sage scrub Species cover, richness,

 equitability 

Westman (1979, 1981) 

Early successional plant

 community 

Species richness, diversity,

 evenness 

Barbo et al. (1998) 

Populus tremuloides

 and Betula papyrifera 

Soil microbial community Phillips et al. (2002) 

Pinus ponderosa Soil microbial community Scagel and Andersen (1997) 

Pinus taeda Fungal morphotypes Edwards and Kelly (1992); 

Qui et al. (1993) 

Trophic Interactions 

Insects Pinus ponderosa Bark beetle severity Cobb et al. (1968) 

Pinus ponderosa Bark beetle productivity and

 predator/parasitoid density 

Dahlsten et al. (1997) 

Populus tremuloides Blotch leaf miner

 performance 

Kopper and Lindroth (2003a) 

Populus tremuloides Aphid/natural enemy

 abundance 

Percy et al. (2002) 

Populus tremuloides Forest tent caterpillar/

 paratisoid performance 

Percy et al. (2002); 

Holton et al. (2003) 

Diseases Populus hybrids Septoria occurrence Woodbury et al. (1994) 

Populus hybrids Rust occurrence Beare et al. (1999a) 

Populus tremuloides Rust occurrence Karnosky et al. (2002) 

Picea abies and 

Picea sitchensis 

Needle fungi Magan et al. (1995) 

Pinus ponderosa Root disease × O3 interactions Fenn et al. (1990) 

Pinus taeda Canker dimensions Carey and Kelley (1994) 

Pinus sylvestris/

 mycorrhizae 

Disease susceptibility Bonello et al. (1993) 

AX9-263
Table AX9-22 (cont’d). Essential Ecological Attributes for Natural Ecosystems 

Affected by O3 

Category Species Condition Measures References 

Community Dynamics Pinus ponderosa/Abies

 concolor/Calocedrus

 decurrens 

Populus tremuloides 

Pinus ponderosa/

 Elymus glaucus 

Pinus taeda/diverse

 community 

C Species and Populations

 Population Size Pinus strobus

Pinus ponderosa

Genetic Diversity/ Lupinus bicolor

Population Structure 

Populus tremuloides 

Trifolium repens 

Plantago major 

Population Dynamics Trifolium repens

Plantago major

Abundance 

Competitive status 

O3 sensitivity 

Tree growth 

Mortality 

Mortality 

% population sensitive 

% population sensitive 

% population sensitive 

% population sensitive 

Adaptation 

Population changes over time 

Minnich et al. (1995)

McDonald et al. (2002)

Andersen et al. (2001)

Barbo et al. (2002)

Karnosky (1981)

Carroll et al. (2003)

Dunn (1959)

Berrang et al. (1986, 1989, 1991)

Heagle et al. (1991)

Davison and Reiling (1995);

Reiling and Davison (1992b);

Lyons et al. (1997)

Heagle et al. (1991)

Davison and Reiling (1995)

Organism Condition 

C Visible Symptoms Pinus ponderosa 

Pinus jeffreyi 

Prunus serotina 

Liriodendron tulipfera 

Sassafras albidum 

Foliar symptoms 

Foliar symptoms 

Foliar symptoms 

Foliar symptoms 

Foliar symptoms 

Grulke and Lee (1997);

Arbaugh et al. (1998);

Salardino and Carroll (1998);

Temple et al. (1992)

Grulke et al. (2003)

Patterson and Rundel (1995);

Salardino and Carroll (1998);

Fredericksen et al. (1995, 1996b);

Chappelka et al. (1997, 1999a,b)

Hildebrand et al. (1996);

Ghosh et al. (1998); Lee et al.

(1999); Ferdinand et al. (2000);

Schaub et al. (2003)

Yuska et al. (2003); Somers et al.

(1998); Hildebrand et al. (1996) 

Chappelka et al. (1999a,b)

AX9-264 
Table AX9-22 (cont’d). Essential Ecological Attributes for Natural Ecosystems 

Affected by O3 

Category Species Condition Measures References 

Organism Condition (cont’d) 

C Visible Symptoms Populus nigra, Foliar symptoms Chappelka et al. (1999a); 

(cont’d) Fraxinus excelsior and

 Prunus avium 

Novak et al. (2003) 

Fagus sylvatica Foliar symptoms Gerosa et al. (2003); 

Vollenweider et al. (2003b) 

Fraxinus americana Foliar symptoms Schaub et al. (2003) 

Grassland species Foliar symptoms Bungener et al. (1999a) 

Herbaceous species Foliar symptoms Bergmann et al. (1999) 

Asclepias exaltata Foliar symptoms Chappelka et al. (1997) 

Rudbeckia laciniata and Foliar symptoms Chappelka et al. (2003) 

Verbesina occidentalis 

Asclepias incarnata Foliar symptoms Orendovici et al. (2003) 

C Physiological Status Pinus halepensis Allometry Wellburn and Wellburn (1994) 

Populus tremuloides Crown architecture Dickson et al. (2001) 

Betula pendula Crown architecture Kull et al. (2003) 

Fagus sylvatica Crown architecture Stribley and Ashmore (2002) 

Populus tremuloides Root dry weight Coleman et al. (1996) 

Pinus ponderosa Root/shoot ratio Grulke et al. (1998a); 

Grulke and Balduman (1999) 

Festiva ovina Root/shoot ratio Warwick and Taylor (1995) 

Betula pubescens Root/shoot ratio Mortensen (1998) 

Populus tremuloides × Root/shoot ratio Landolt et al. (2000); 

Populus tremula Paludan-Müller et al. (1999) 

Populus tremuloides Leaf area index Oksanen et al. (2001) 

Pinus ponderosa Carbon allocation to Neufeld et al. (1995); 

 mycorrhizae Wiltshire et al. (1996); 

Andersen and Rygiewicz 

(1995a,b) 

Betula pendula Decreased winter bud Karnosky et al. (2003a) 

 formation 

Betula pendula Delayed bud break Oksanen (2003a,b) 

Acer saccharum Early bud break Prozherina et al. (2003); 

Bertrand et al. (1999) 

AX9-265
Table AX9-22 (cont’d). Essential Ecological Attributes for Natural Ecosystems 

Affected by O3 

Category Species Condition Measures References 

C Reproductive Status Apocynun

 androsaemifolium 

Flowering time Bergweiler and Manning (1999) 

Buddleia davidii Flowering time Findley et al. (1997) 

Rubus cuneifolius Pollen germination Chappelka (2002) 

Plantago major Pollen tube elongation Stewart (1998) 

Fragaria × ananassa Fruit yield Drogoudi and Ashmore 

(2000, 2001) 

Plantago major Seed yield Lyons and Barnes (1998); 

Pearson et al. (1996); 

Reiling and Davison (1992a); 

Whitfield et al. (1997) 

Understory herbs Seed yield Harward and Treshow (1975) 

Ecological Processes 

C Energy Flow

 Primary Production Pinus ponderosa Photosynthesis Miller et al. (1969); Clark et al. 

(1995); Takemoto et al. (1997); 

Grulke et al. (2002b) 

Pinus ponderosa Needle retention Temple et al. (1993) 

Populus tremuloides Photosynthesis Coleman et al. (1995ab); 

Noormets et al. (2001a,b); 

Sharma et al. (2003); 

Karnosky et al. (2003a); 

Oksanen (2003a,b) 

Betula pendula Photosynthesis/conductance Matyssek et al. (2002) 

Betula pendula Stem respiration and radial

 growth 

Kelting et al. (1995) 

Quercus rubra Root turnover Coleman et al. (1996) 

Populus tremuloides Soil respiration King et al. (2001); 

Andersen and Scagel (1997); 

Coleman et al. (1995a) 

Pinus ponderosa Soil respiration Scagel and Andersen (1997); 

Andersen (2000); Samuelson 

and Kelly (1996) 

Quercus rubra Carbon partitioning and

 allocation 

Andersen et al. (1997); 

Grulke et al. (1998a) 

Populus tre muloides Carbon allocation Grulke and Balduman (1999) 

AX9-266
Table AX9-22 (cont’d). Essential Ecological Attributes for Natural Ecosystems 

Affected by O3 

Category Species Condition Measures References

 Primary Production 

(cont’d) 

Pinus ponderosa 

Betula pendula 

Carbon allocation 

Carbon allocation 

Grulke et al. (2001) 

Karlsson et al. (2003b); 

Oksanen and Saleem (1999); 

Saleem et al. (2001) 

Fragaria vesca Carbon allocation Manninen et al. (2003) 

Pinus taeda Root respiration Edwards (1991) 

Lespedeza cuneata and 

Schizacbyrium

 scoparium 

Yield Powell et al. (2003) 

Liriodendron tulipfera Radial growth Somers et al. (1998); 

Vollenweider et al. (2003a) 

Prunus serotina Radial growth Somers et al. (1998) 

Pinus jeffreyi Radial growth Peterson et al. (1987) 

Pinus ponderosa Radial growth (no effect) Peterson et al. (1993) 

Pinus strobus Radial growth Bartholomay et al. (1997) 

Pinus taeda Radial growth McLaughlin and Downing 

(1995; 1996) 

Fagus sylvatica Stem volume Braun et al. (1999) 

Picea abies Stem volume Wallin et al. (2002) 

Populus tremuloides Volume growth Isebrands et al. (2001) 

Pinus ponderosa Root growth Andersen et al. (1991)

 Net Ecosystem

 Production 

Northern hardwoods 

Northern hardwoods 

NPP estimates 

Biomass estimates 

Laurence et al. (2000) 

Hogsett et al. (1997)

 Growth Efficiency Plantago major Relative growth rate Davison and Reiling (1995); 

Lyons et al. (1997); Reiling and 

Davison (1992b); Davison and 

Barnes (1998) 

Grassland species Relative growth rate Bungener et al. (1999b) 

Native herbs Relative growth rate Warwick and Taylor (1995) 

Grasses and herbs Relative growth rate Pleijel and Danielsson (1997) 

Populus tremuloides Relative growth rate Yun and Laurence (1999a) 

AX9-267
Table AX9-22 (cont’d). Essential Ecological Attributes for Natural Ecosystems 

Affected by O3 

Category Species Condition Measures References

 Growth Efficiency

 (Cont’d) 

Fagus sylvatica 

Picea abies 

Relative growth rate 

Relative growth rate 

Bortier et al. (2000c) 

Karlsson et al. (2002) 

Prunus serotina Relative growth rate Lee et al. (2002) 

C Material Flow

 Organic Carbon

 Cycling 

Populus tremuloides 

 and Betula papyrifera 

Altered foliar C:N ratio and

 N resorption efficiency 

Lindroth et al. (2001) 

Andropogon virginicus

 and Rubus cuneifolius 

Litter decomposition rate Kim et al. (1998) 

Liriodendron tulipera Litter decomposition rate Scherzer et al. (1998) 

Populus deltoides Litter decomposition rate Findlay and Jones (1990) 

Pinus ponderosa Litter decomposition rate Fenn and Dunn (1989) 

Pinus sylvestris Litter decomposition 

(no effect) 

Kainulainen et al. (2003)

 Nitrogen Cycling Pinus ponderosa Altered foliar N Momen and Helms (1996) 

Pinus ponderosa Foliar N and O3 exposure

 no effects 

Bytnerowicz et al. (1990) 

Pinus taeda Altered foliar N metabolism Manderscheid et al. (1992) 

Prunus serotina and

 Liriodendron

 tulipfera 

Altered foliar N Boerner and Rebbeck (1995)

 Other Nutrient

 Cycling 

Picea sitchensis and

 Pinus sylvestris 

Foliar leaching no effect Skeffington and Sutherland 

(1995) 

Pinus ponderosa Nutrient availability and O3 Bytnerowicz et al. (1990) 

Hydrology and Geomorphology 

C Water Budget Picea rubens Water-use efficiency Laurence et al. (1997) 

no effect 

Pinus armandi Water-use efficiency Shan et al. (1996) 

Pinus jeffreyi Canopy transpiration Grulke et al. (2003a) 

Picea abies Transpiration Maier-Maercker (1997) 

 (xylem sap flow) 

Fraxinus excelsior Water stem flow Wiltshire et al. (1994) 

Betula pendula Water-use efficiency Maurer and Matyssek (1997) 

Populus hybrids Water-use efficiency Reich and Lassoie (1984) 

AX9-268
Table AX9-22 (cont’d). Essential Ecological Attributes for Natural Ecosystems 

Affected by O3 

Category Species Condition Measures References 

Natural Disturbance Regimes 

C Frequency Pinus ponderosa Frequency of fire McBride and Laven (1976); 

Minnich et al. (1995); 

Miller and McBride (1999) 

Pinus ponderosa Occurrence of bark beetle

 outbreaks 

Pronos et al. (1999); 

Dahlsten et al. (1997) 

C Intensity Picea sitchensis Winter damage Lucas et al. (1988) 

Pinus halepensis Reduced winter damage Wellburn and Wellburn (1994) 

Picea rubens Freezing tolerance Waite et al. (1994) 

Fagus sylvatica Drought stress Pearson and Mansfield 

(1993, 1994) 

Picea abies Drought stress Maier-Maercker (1998); 

Maier-Maercker and 

Koch (1992) 

Pinus ponderosa Fire intensity Miller and McBride (1999a) 

C Extent Pinus ponderosa Extent of bark beetle attack Minnich et al. (1995) 

C Duration Pinus ponderosa Duration of bark beetle attack Minnich et al. (1995) 

decline. Lead in automobile gasoline was phased out in 1990, and foliar concentrations of heavy 

metals in forest species are not now at phytotoxic levels (Fenn and De Bauer, 1999). Sulfur 

dioxide concentrations decreased in the early 1990s as a result of regulatory mandates limiting 

their emissions. Sensitive plants in the northeast and northwest sectors of the Mexico City urban 

zone where concentrations are highest may still be impacted by exposure to ambient SO2 levels. 

Deposition of ionic forms of N and S are high in forested areas southwest of Mexico City. 

However, the effects of these chronic nutrient inputs to the forest are only beginning to be 

investigated and understood. 

The ecological perturbations caused by severe air pollution exposures in forests located 

downwind of Mexico City are expected to continue for the near future (the next 5 to 10 years), 

largely as a result of high O3 concentrations as well as emissions of N oxides. The longer-term 

response is more uncertain and depends largely on the effectiveness of regulatory emissions 

control strategies. Currently, pollutant levels are declining. Forest responses to this trend will 

AX9-269
depend on how long it takes to reduce levels sufficiently to allow sensitive species to recover. 

Some of the change to the ecosystem is probably irreversible, such as the loss of lichen diversity 

and the loss of other O3-sensitive species (Zambrano and Nash, 2000). 

AX9.6.2.2 San Bernardino Mountains 

The San Bernardino Mountains lie east of the Los Angeles Air Basin in California, and 

significant levels of pollution have been transported into the mountain range, including into a 

Class I wilderness area. The effects of O3 exposure on the mixed conifer forest of the 

San Bernardino Mountains is perhaps the longest and most thoroughly documented O3 

ecological effects evaluation (Miller and McBride, 1999a). In this classic case study linking 

tropospheric O3 exposure to damage to an entire forest ecosystem (U.S. Environmental 

Protection Agency, 1996) (Table AX9-23), Miller et al. (1963) first identified the unique foliar 

chlorotic mottle that was occurring on two of the dominant tree species, Pinus ponderosa and 

P. jeffreyi. Levels of O3 averaging 100 to 120 ppb over 24 h with 1-h peaks well into the 

200 ppb range were common in the region in the 1960s and 1970s (Miller and McBride, 1999a). 

Single-hour peak values have declined in recent years due to heavily regulated pollution control 

(Arbaugh et al., 1998; Lee et al., 2003; Takemoto et al., 2001). 

Since the 1996 O3 AQCD (U.S. Environmental Protection Agency, 1996) was written, 

the concurrent role of N deposition in modifying ecosystem response to O3 exposure in the 

San Bernardino Mountains has been further elucidated (Arbaugh et al., 2003; Bytnerowicz et al., 

1999; Bytnerowicz, 2002; Fenn et al., 1996; Takemoto et al., 2001). Both O3 exposure and N 

deposition reduce foliar retention (Grulke and Balduman, 1999) and alter tissue chemistry of 

both needles and litter (Poth and Fenn, 1998). In addition, confounding factors such as drought 

and fire suppression add to the complexity of ecosystem response (Arbaugh et al., 2003; 

Minnich et al., 1995; Takemoto et al., 2001). Extensive crown injury measurements have 

also been made, linking ambient O3 exposure data to chlorotic mottle and fascicle retention 

(Arbaugh et al., 1998). Ozone exposure and N deposition reduce carbon allocation to stems and 

roots (Grulke et al., 1998a, 2001), further predisposing trees to drought stress, windthrow, root 

diseases, and insect infestation (Takemoto et al., 2001). Increased mortality of susceptible tree 

species (Ponderosa and Jeffrey pine) has shifted community composition toward white fir and 

incense-cedar (Abies concolor, Calocedrus decurrens) and has altered forest stand structure 

AX9-270
Table AX9-23. Case Studies Demonstrating the Ecological Effects of O3 

Keystone Period 

Study Species Study Type Studied Key Ecological Findings 

Valley of Pinus hartwegii, Field transects 35 yrs • Significant foliar injury 

Mexico Abies religiosa (De Bauer, 1972; De Bauer and 

Hernández-Tejeda, 1986; De Bauer and 

Krupa, 1990) 

• Community composition changes 

(Alvarado et al., 1993) 

• Species richness changes (Zambrano and 

Nash, 2000) 

San Bernardino Pinus ponderosa, Field transects 40 yrs • Community composition changes 

Mountains P. jeffreyi (Arbaugh et al., 2003; Miller, 1973; 

Minnich et al., 1995) 

• Population changes 

(McBride and Laven, 1999) 

• O3-pine-bark beetle interaction 

(Pronos et al., 1999) 

• Altered C flows 

(Arbaugh et al., 1999; Grulke et al., 

1998a, 2001, 2002b; Grulke and 

Balduman, 1999) 

• Interaction of O3, drought, N deposition 

(Fenn et al., 1996; Grulke, 1999; 

Takemoto et al., 2001) 

• Altered carbon cycling 

(Arbaugh et al., 1999) 

Sierra Nevada Pinus ponderosa, Field 35 yrs • Wide-scale nature of effects 

Mountains P. jeffreyi (Edinger et al., 1972; Miller and 

Millecan, 1971) 

• Link to decreased growth 

(Peterson et al., 1987, 1991, 1995) 

• Quantification of O3 flux 

(Bauer et al., 2000; Goldstein et al., 2003; 

Panek et al., 2002) 

• Cumulative O3 effects 

(Takemoto et al., 1997) 

• Canopy level responses 

(Grulke et al., 2003a,b) 

• Population changes (Carroll et al., 2003) 

Appalachian Fraxinus Field 25 yrs C Link of visible symptoms to growth 

Mountains americana, decreases (McLaughlin et al., 1982; 

Liriodendron Somers et al., 1998) 

tulipfera, • Wide-scale nature of effects (Chappelka 

Pinus strobus, et al., 1999a; Hildebrand et al., 1996) 

Prunus serotina 

AX9-271
Table AX9-23 (cont’d). Case Studies Demonstrating the Ecological Effects of O3 

Keystone Period 

Study Species Study Type Studied Key Ecological Findings 

Aspen FACE Acer saccharum, Open-air O3 6 yrs • Competitive interactions 

 Betula exposure (McDonald et al., 2002) 

papyrifera, • O3-aspen-rust interaction 

Populus (Karnosky et al., 2002) 

tremuloides • Plant-insect interactions 

• (Holton et al., 2003; Percy et al., 2002) 

• C and N cycling 

• (King et al., 2001; Lindroth et al., 2001) 

• Moderation of CO2 responses by O3 

(Isebrands et al., 2001; Karnosky et al., 

2003b; McDonald et al., 2002; 

Wustman et al., 2001) 

Plantago Plantago major Field 20 yrs • Population structure 

• (Davison and Reiling, 1995; 

• Lyons et al., 1997) 

• O3 resistance 

• (Reiling and Davison, 1992c) 

• Adaptation (Davison and Reiling, 1995) 

Carpathian Pinus sylvestris, Field 15 yrs • Significant foliar injury 

Mountains Picea abies • Community composition changes 

• Species diversity changes 

(Miller et al., 1989) (Table AX9-25). Ozone exposure is implicated in projected changes in 

stand composition (McBride and Laven, 1999) toward a predominance of oaks, rather than 

mixed conifer forests. Forest understory species have also been affected (Temple, 1999). These 

individual species responses collectively have affected trophic structure and food web dynamics 

(Dahlsten et al., 1997; Pronos et al., 1999), as well as C and N cycling (Arbaugh et al., 2003) 

(Table AX9-24). 

AX9.6.2.3 Sierra Nevada Mountains 

Like the San Bernardino Mountains, the western slope of the Sierra Nevada Mountains in 

central and southern California has also been exposed to elevated O3 for a long time, although 

the effects have been much less.. Symptoms of O3 injury have been found on Ponderosa and 

Jeffrey pines in all of the Sierra Nevada national forests and parks (Carroll et al., 2003). First 

identified as a problem in the 1970s (Miller et al., 1972), elevated O3 with daytime means of 

60-80 ppb are common (Bauer et al., 2000; Böhm et al., 1995; Bytnerowicz et al., 2002c; Panek 

AX9-272
Table AX9-24. The Most Comprehensively Studied Effects of O3 on Natural Ecosystem 

are for the San Bernardino Mountain Forest Ecosystem. Citations Focus on Research 

Published Since U.S. EPA (U.S. Environmental Protection Agency, 1996). 

Pollutant Occurrence Reference 

O3 exposure and N deposition 

Cellular, Biochemical 

Foliar pigments 

Antioxidants 

Foliar Responses 

Foliar Symptoms 

Gas Exchange 

Photosynthesis and Conductance

 O3 flux

 Foliar nutrients 

Whole Organism 

Growth/Biomass 

• Above-ground 

• Below-ground 

• Root/shoot ratio 

• Carbon allocation 

• Crown vigor 

Ecosystem

 Community dynamics/succession

 Simulations 

 Understory vegetation 

Pest interactions 

• Bark beetle/predators 

• Disease occurrence 

• Litter decomposition 

Disturbance 

• Bark beetle occurrence 

• Fire frequency 

Fenn et al. (1996; 2000); Grulke et al. (1998a,

2003a); Bytnerowicz et al. (1999)

Grulke and Balduman (1999); Grulke and Lee

(1997); Tausz et al. (1999a,b,c, 2001)

Arbaugh et al. (1998); Grulke and Lee (1997);

Miller and Rechel (1999)

Grulke (1999); Grulke et al. (2002a,b);

Grulke and Retzlaff (2001) 

Grulke and Balduman (1999)

Grulke et al. (1998a); Grulke and Balduman

(1999); Grulke and Balduman (1999);

Grulke et al. (2001); Arbaugh et al. (1998);

Miller and Rechel (1999)

Arbaugh et al. (2003); Arbaugh et al. (1999);

McBride and Laven (1999)

Temple (1999)

Dahlsten et al. (1997); Pronos et al. (1999)

Miller and Rechel (1999); Pronos et al. (1999)

Miller and McBride (1999b); Minnich et al.

(1995); Minnich (1999)

AX9-273
Table AX9-25. Effects of Ozone, Ozone and N Deposition, and Ozone and Drought 

Stress on Pinus ponderosa and Pinus jeffreyi in the Sierra Nevada and the 

San Bernardino Mountains, California. Citations are Focused on Research 

Published since U.S. EPA (1996). 

O3 O3 + N deposition O3 + Drought References 

Foliar Biochemistry and 

Tissue Chemistry 

Total ascorbate 

Dehydroascorbate 

Total glutathione 

Oxidized glutathione 

" Carotenoids 

Foliar nitrogen 

C:N ratio of foliage2 

Starch 

Chlorophyll content 

Gas Exchange 

Amax lower canopy 

Amax whole canopy 

Amax seedlings 

Stomatal limitation 

Stomatal conductance 

Foliar respiration 

O3 flux 

d1 d i Grulke et al. (2003b); 

Tausz et al. (2001) 

i n.d. d Grulke et al. (2003b) 

d i d Tausz et al. (2001) 

i i d Tausz et al. (2001) 

i n.d. d Grulke et al. (2003b) 

d i d Grulke et al. (1998a); Grulke and 

Lee (1997); Poth and Fenn (1998) 

i n.d. d Poth and Fenn (1998) 

Grulke et al. (2003b) 

n.d. d i Grulke et al. (2001) 

d id d Grulke et al. (1998b, 2003b); 

Grulke and Lee (1997); 

Takemoto et al. (1997)(Grulke 

(1999); Tausz et al. (2001) 

n.d. i d Grulke (1999); Grulke et al. 

(2002b); Grulke and Retzlaff 

(2001); Panek (2004) 

d n.d. d Grulke et al. (2003b); 

Panek and Goldstein (2001) 

di n.d. n.d. Grulke and Retzlaff (2001) 

n.d. n.d. i Panek and Goldstein (2001) 

d di d Grulke (1999); Grulke et al. 

(2003a); Panek (2004) 

n.s. i d Grulke (1999); 

Grulke et al. (2002a) 

d n.s. d Panek et al. (2002, 2003); 

Panek and Goldstein (2001) 

Grulke et al. (2002a, 2004) 

AX9-274 
Table AX9-25 (cont’d). Effects of Ozone, Ozone and N Deposition, and Ozone and 

Drought Stress on Pinus ponderosa and Pinus jeffreyi in the Sierra Nevada and the 

San Bernardino Mountains, California. Citations are Focused on Research 

Published Since U.S. EPA (1996). 

O3 O3 + N deposition O3 + Drought References 

Growth and Productivity 

Foliar biomass n.d. i d Grulke and Balduman (1999) 

Height growth n.d. i d Grulke and Balduman (1999) 

Bole diameter growth d i d Grulke and Balduman (1999) 

Fine root biomass d d i Grulke et al. (1998a) 

Leaf Surfaces 

Stomatal occlusion i n.d. n.d. Bytnerowicz et al. (1999); 

Bytnerowicz and Turunen (1994) 

Trophic Interactions 

Bark beetle n.s. i i Pronos et al. (1999) 

Ecosystem Level 

Competitive indices n.d. d i Miller and Rechel (1999) 

1

Responses are shown as significant increases (i), significant decreases (d), both significant decreases and

 increases reported (di), nonsignificant effects (n.s.), and no data (n.d.) compared to trees or seedlings at field sites

 with lower O3, drought stress, or lack of significant N deposition (<10 kg ha!1 yr!1). Frequently n.d. was used

 for lack of a control site without compounding high N deposition. Foliar analyses and leaf surface properties

 were largely determined from previous year needles. Gas exchange data were generally from previous year

 needles at peak growing season, prior to late summer drought (mid- to late July). 

2

Abbreviations: C = carbon; N = nitrogen; Amax = maximum photosynthesis rate. 

et al., 2002). The west-slope Sierra Nevada forests are also exposed to a wide range of 

additional gaseous and particulate pollutants, including various S and N compounds 

(Bytnerowicz et al., 1999; Fenn et al., 2003b; Takemoto et al., 2001), but at levels much lower 

than in the San Bernardino Mountains. Typical O3-induced visible foliar symptoms, including 

chlorotic mottle, chlorophyll degradation, and premature senescence, are commonly found 

on O3-sensitive genotypes of Ponderosa pine (Arbaugh et al., 1998; Peterson et al., 1991; Staszak 

et al., 2004) and Jeffery pine (Arbaugh et al., 1998; Grulke et al., 2003b; Patterson and Rundel, 

1995; Peterson et al., 1987). Other important conifers in the region, such as giant sequoia, 

appear to be relatively O3-tolerant (Grulke et al., 1996). The symptoms of foliar injury and 

AX9-275
growth reductions have been verified on seedlings in O3 exposure chambers (Momen et al., 

2002; Momen and Helms, 1996; Temple, 1988). 

Ozone foliar injury of dominant pine species in the Sierra Nevada Mountains is correlated 

to decreased radial growth in both Ponderosa pine (Peterson et al., 1991) and Jeffrey pine 

(Patterson and Rundel, 1995; Peterson et al., 1987). Because of the large amount of intraspecific 

variation in O3 sensitivity in these two species, O3 exposure may be a selective agent (Patterson 

and Rundel, 1995), with differential mortality rates for sensitive individuals (Carroll et al., 

2003). The region’s forests may also be experiencing subtle changes in species composition and 

community structure (Patterson and Rundel, 1995; Takemoto et al., 2001). 

Based on fire scar dating, reconstructions of stand age classes, historical records, and 

present stand structure, fire has been largely excluded in western forests for the last 75 to 100 

years (Minnich, 1999; Minnich and Padgett, 2003). Fire exclusion has resulted in fewer large 

stand-replacing fires rather than a mosaic of smaller low-intensity fires. The change in fire 

intensity may have selectively altered stand structure, fitness and competitiveness of component 

species, along with their susceptibility to atmospheric pollutants and other stressors (Minnich, 

1999). Short-lived (50 to 80 years) species such as knobcone (Pinus attenuata) and Coulter pine 

(Pinus coulteri), which occur at the interface of the chaparral and the mixed conifer forest, may 

already have been selected for O3 tolerance by seedling establishment (the most sensitive tree 

age class in conifers) after large fires in the 1950s (Minnich, 1999). Strong measures to suppress 

fires have largely kept chaparral fires from invading the mixed conifer forests, and stand 

densification in the mixed conifer zone has increased. High stand density, in turn, may weaken 

the younger cohorts and increase sensitivity to atmospheric pollution (Minnich, 1999). 

Other disturbances that play a potential role in sensitivity to atmospheric pollution include 

cycles of drought stress. Nearly every decade is marked by one or more years of very low 

precipitation (Graumlich, 1993). During extended periods of drought, foliar injury is lower than 

in subsequent years with higher average precipitation (Carroll et al., 2003). In the first several 

years (1975 to 1977) of a Sierran-wide assessment of O3 injury to pines, O3 injury increased, 

because of greater water availability due to greater stomatal conductance and, presumably, 

greater O3 uptake. Trees instrumented with monitors to directly measure canopy transpiration 

had 20% greater stomatal conductance in mesic microsites (riparian areas, mid-slope seeps) than 

trees in xeric microsites (rock outcrops) (Grulke et al., 2003a). Although the Sierra Nevada 

AX9-276
experienced a prolonged drought between 1987 and 1993, it was less severe than other droughts 

and O3 injury did not significantly decrease (Carroll et al., 2003). The same plots showed only a 

slight increase in O3 injury between 1993 and 2000. While drought stress may make trees more 

susceptible to insect and pathogen infestation, serious outbreaks of insect infestation are believed 

to be indicators, not a cause, of existing stress in the forest (Wickman, 1992). 

AX9.6.2.4 Appalachian Mountains 

The southern Appalachian Mountain region experiences some of the highest O3 exposures 

of any natural areas in the eastern United States (Chappelka et al., 1997; Hildebrand et al., 1996; 

Mueller, 1994; Samuelson and Kelly, 1997). Since the region is the home of the Shenandoah 

and Great Smoky Mountains NPs, which have Class I air quality designations by the 1977 

Clean Air Act, there has been considerable study of the region’s dominant forest species to 

determine O3 effects. Visible foliar symptoms of O3 have been found in natural ecosystems 

consisting of sassafras (Sassafras albidum) (Chappelka et al., 1999b), black cherry (Prunus 

serotina) (Chappelka et al., 1997, 1999a; Hildebrand et al., 1996; Samuelson and Kelly, 1997), 

yellow poplar and white ash (Fraxinus americana) (Chappelka et al., 1999b; Hildebrand et al., 

1996). Visible foliar symptoms induced by O3 have been recreated on the same species in 

chamber studies (Chappelka et al., 1985; Chappelka and Chevone, 1986; Duchelle et al., 1982; 

Fredericksen et al., 1995; Samuelson, 1994). No response to O3 exposure has been found for 

other hardwood trees, nor for the three conifer species tested (Neufeld et al., 2000). 

Long-term foliar injury symptoms have been correlated with decreased radial growth in 

yellow poplar and black cherry (Somers et al., 1998) and with decreased biomass in cherry 

(Neufeld et al., 1995). Although climatic conditions (drought) largely explained radial growth 

reductions, O3 exposure may have also contributed (McLaughlin and Downing , 1996). Ozone 

exposure may also be affecting the understory vegetation in the region (Chappelka et al., 1997; 

2003; Davison et al., 2003; Duchelle et al., 1983; Duchelle and Skelly, 1981) and community 

composition (Barbo et al., 1998), through impacts on both growth and reproduction (Chappelka, 

2002). Foliar litter from trees exposed to elevated O3 have lower decomposition rates (Kim 

et al., 1998). Other air pollutants are likely to be found in this ecosystem but not at the high 

deposition values found in the California studies. A decline in forest health in the northern 

Appalachians has been primarily attributed to the effects of acidic fog and rain on soil 

AX9-277
acidification, lower Ca2+ availability, reduction in fine root biomass, and modification of 

cuticular wax. However, fog- and O3-exposed red spruce forests in New England also show 

winter injury (Percy et al., 1992). 

AX9.6.2.5 Plantago Studies in the United Kingdom 

One of the most well-documented studies of population and community response to O3 

effects are the long-term studies of common plantain (Plantago major) in native plant 

communities in the United Kingdom (Davison and Reiling, 1995; Lyons et al., 1997; Reiling and 

Davison, 1992c). Sensitive populations of P. major had significant growth decreases in elevated 

O3 (Pearson et al., 1996; Reiling and Davison, 1992a,b; Whitfield et al., 1997) and reduced 

fitness as determined by decreased reproductive success (Pearson et al., 1996; Reiling and 

Davison, 1992a). While spatial comparisons of population responses to O3 are complicated by 

other environmental factors, rapid changes in O3 resistance were imposed by ambient levels and 

variations in O3 exposure (Davison and Reiling, 1995). Molecular patterns of genetic variation 

suggest that a change in O3 resistance over time probably resulted from natural selection in 

genotypes already present in local populations, rather than through an influx of new P. major 

germplasm (Wolff et al., 2000). At the site of common plantain seed collection the highest 

correlations occurred between O3 resistance and ambient O3 concentrations occurred (Lyons 

et al., 1997), rather than between O3 resistance and other climatic variables, as found for aspen 

(Berrang et al., 1991). 

AX9.6.2.6 Forest Health in the Carpathian Mountains 

The Carpathian Mountains cross five countries (the Czech Republic, the Slovak Republic, 

Poland, Romania, and the Ukraine) and contain many national parks and several biosphere 

reserves. The forests were largely cleared in the 15th century and were replanted with Norway 

spruce. As elevation increases, beech (Fagus sylvatica) or beech-fir (Abies alba) forests grade 

into Norway spruce or spruce-fir forests. Near the treeline, Norway spruce mixes with dwarf 

mountain pine (Pinus mugo). Dwarf mountain pine forms an almost pure stand just below the 

alpine vegetation. 

The forests of the Carpathian Mountains have been subjected to anthropogenic stressors 

(e.g., shepherding, metal mining, wood harvest for structures and paper) for hundreds of years, 

AX9-278
as described for the Tatra Mountains in the southern Carpathians (Wezyk and Guzik, 2002). The 

Carpathians have been subjected to regional air pollution stressors since industrialization. Most 

of the effects of air pollution on forest health degradation were due to (1) heavy metal 

deposition, (2) soil acidification by acid deposition, and (3) subsequent pest outbreaks, the 

combination of which led to the forest decline and dieback between 1970 and 1989 (Dunajski, 

2002). Industrial pollutants such as SO2 and heavy metals have significantly declined since the 

1980s, but O3 exposure has continued to increase (Bytnerowicz et al., 2002a; Bytnerowicz et al., 

2004). The increased ownership and use of private cars in Central Europe, as well as the 

long-range transport of O3 from western Europe, are believed to be responsible for the continued 

increase in photooxidants. In 1995, drought resulted in significant forest mortality, as well as an 

epidemic of bark beetle infestation in subsequent years. A network of air quality monitoring 

sites was installed across Europe in the late 1980s as part of the International Cooperative 

Programme on Assessment and Monitoring of Air Pollutant Effects on Forests (ICP Forests). 

Mean defoliation rates for six important forest species across Europe have increased or remained 

unchanged from 1989 to 1999 (Percy et al., 2002). Ozone concentrations experienced in the 

Tatra Mountains, especially along the southern slopes, occasionally reach 190 to 200 ppb as 

2-week-long averages, with the highest values experienced in early summer at elevations of 

1700 to 2300 m (Bytnerowicz et al., 2004). In other parts of the Carpathian Mountains, peak 

2-week average O3 concentrations were lower, at 160 ppb (Bytnerowicz et al., 2002b). For all 

trees inventoried, about 13% exhibited greater than 25% defoliation during 1997 to 2000. There 

was no difference between extent of damage for broadleaves or conifers. Trees in Poland and 

the Czech Republic were the most affected by air pollution, while the least damaged forests were 

in Romania (Bytnerowicz et al., 2002b). 

The extent to which O3 exposure affects forest health degradation, and slows forest 

degradation, is still unknown in Europe. In many of the published studies, the response to a 

known O3 gradient is largely confounded by other pollutants and/or climatic gradients (Szaro 

et al., 2002; Widacki et al., 2002). Current levels of ambient O3 are believed to be high enough 

to reduce bole radial growth (Percy et al., 2002). Although average O3 concentration alone was 

not related to bole growth, the peak hourly O3 concentration was negatively correlated to bole 

growth (Muzika et al., 2004). Recent evidence indicates that canopy health of European white 

AX9-279
oak (Quercus robur), Norway spruce, maritime pine (Pinus pinaster), and beech has 

significantly declined (Huttunen et al., 2001). However, the canopy health of Scots pine has 

improved. The network of air quality monitoring stations and forest plots is extensive and 

active. Subsequent correlative analyses including both meteorological and air quality attributes 

throughout the European Union (EU) will help to determine the specific role of O3 exposure in 

forest decline. Historical effects of anthropogenic disturbance may still be confounding. 

AX9.6.2.7 Field Exposure System (FACE), Rhinelander, Wisconsin 

The Aspen Free-Air CO2 Enrichment facility was designed to examine the effects of both 

elevated CO2 and O3 on aspen (Populus tremuloides), birch (Betula papyrifera), and sugar maple 

in a simple reconstructed plantation characteristic of Great Lakes Aspen-dominated forests 

(Karnosky et al., 2003b; Karnosky et al., 1999). Instead of using chambers to expose the plants 

to desired gas concentrations, the gas is piped up vertical delivery tubes in the open air. The 

vertical delivery pipes surround a 30-m diameter circular plot with five different aspen clones in 

half of the plot, one quarter of the plot planted in aspen and birch, and one quarter in aspen and 

maple. The O3 treatment for the first 5 years was 1.5× ambient, with ambient O3 exposures 

averaging 35 to 37 ppb (12 h daytime average over the growing season) compared to elevated O3 

rings averaging 49 to 55 ppb for the same time period (Karnosky et al., 2003b). 

Elevated CO2, elevated O3, and elevated CO2 + O3 have had effects on most system 

components being measured in the study (Table AX9-26) (Karnosky et al., 2003b). One 

interesting finding of the project has been the nearly complete offset by elevated O3 of the 

enhancements induced by elevated atmospheric CO2 for the pioneer keystone species aspen 

(Isebrands et al., 2001) and birch (Percy et al., 2002) even though O3 exposure alone did not 

always result in a significant response when compared to controls. They also found evidence 

that the effects on above- and below-ground growth and physiological processes have cascaded 

through the ecosystem, even affecting microbial communities (Larson et al., 2002; Phillips et al., 

2002). This study also confirmed earlier observations of changes in trophic interactions 

involving keystone tree species, as well as important insect pests and their natural enemies 

(Table AX9-26) (Awmack et al., 2003; Holton et al., 2003; Percy et al., 2002). 

AX9-280
Table AX9-26. Summary of Responses of Populus tremuloides to Elevated CO2 

(+200 μmol mol-1), O3 (1.5 × ambient), or CO2+O3 Compared with Control During 

3 Years of Treatments at the Aspen FACE Project 

(Modified from Karnosky et al. (2003b) 

CO2 O3 CO2 + O3 Reference 

Foliar Gene Expression 

and Biochemistry 

Rubisco; RbcS2

 transcripts d1 d dd Noormets et al. (2001a); Wustman et al. (2001) 

PAL transcripts d I d Wustman et al. (2001) 

Acc oxidase, catalase d I d Wustman et al. (2001) 

Ascorbate peroxidase d n.s. d Wustman et al. (2001) 

Glutathione reductase d I d Wustman et al. (2001) 

Phenolic glycosides I d n.s. Kopper and Lindroth (2003a,b); Lindroth et al. 

(2001) 

Tannins n.s. i i Kopper and Lindroth (2003a,b); Lindroth et al. 

(2001) 

Foliar nitrogen d n.s. d Kopper and Lindroth (2003a,b); Lindroth et al. 

(2001) 

C:N ratio of foliage i n.s. ii Lindroth et al. (2001) 

Starch d d n.s. Wustman et al. (2001) 

Gas Exchange 

Amax lower canopy n.s. dd id Noormets et al. (2001b); Takeuchi et al. (2001) 

Amax whole canopy ii dd n.s. Noormets et al. (2001a); Sharma et al. (2003) 

Stomatal limitation d n.s. d Noormets et al. (2001b) 

Stomatal conductance d di d Noormets et al. (2001b) 

Foliar respiration n.s. i n.s. Noormets et al. (2001a); Takeuchi et al. (2001) 

Soil respiration ii d n.s. King et al. (2001) 

Microbial respiration ii n.s. n.s. Phillips et al. (2002) 

Stomatal density n.s. n.s. n.s. Percy et al. (2002) 

Chlorophyll content d d d Wustman et al. (2001) 

Chloroplast structure i d d Oksanen et al. (2001); Takeuchi et al. (2001); 

Wustman et al. (2001) 

O3 flux d ii i Noormets et al. (2001b) 

Growth and Productivity 

Leaf thickness i n.s. n.s. Oksanen et al. (2001) 

Leaf size i d d Wustman et al. (2001) 

Leaf area i d n.s. Noormets et al. (2001a) 

1

Responses are shown as significant increases p < 0.05 (I), significant increases p < 0.01 (ii), significant decreases 

p < 0.05, significant decreases p < 0.01 (dd), both significant increases and decreases reported (id), nonsignificant 

effects (n.s.), and no data (n.d.). 

AX9-281 
AX9.6.3 Landscape Condition 

In the SAB framework (Figure AX9-20), landscape condition is assessed using the areal 

extent, composition of component landscape ecosystems or habitat types, and the pattern or 

structure of component ecosystems or habitat types (including biocorridors). To date, no 

publications exist on the impacts of O3 exposure on landscape condition. The effects of O3 

exposure have only been reported at the community or stand level (see Biotic Conditions, 

below). The following is a description of current discussions by land stewards and of how 

difficult it will be to quantitatively assess the effect of O3 exposure on landscape condition. 

Landscapes are identified and preserved, such as national parks, Class I wilderness areas, 

etc., so that they are protected from the effects of O3 exposure by law. Efforts to determine 

whether landscapes have been affected by certain levels of exposure rely on valuation of 

landscape and ecosystem components. Several different approaches of valuation have been 

used, including pathological (visible symptoms), biomass and allocation, and biogeochemical. 

In the pathological approach, a “critical levels” concept is developed, with varying levels 

of impact viewed as acceptable, interim targets, or as unacceptable. As an example, land 

managers of Class I wilderness areas may consider a level acceptable if it resulted in no 

visible O3 symptoms to sensitive species. In concrete terms, sensitive species may respond to 

peak O3 exposures of 60 ppb (e.g., coneflower, in Great Smoky National Mountains NP; 

[Davison et al., 2003]), and so the critical exposure level would be < 60 ppb for any hourly value 

during the growing season. An interim target would be that less than 5% of the sensitive plants 

would have visible symptoms on <15% of the leaf surface. An unacceptable level of O3 

exposure would be any result more pronounced than the interim target. The advantage of the 

foliar injury approach is that large crews with relatively simple training can assess individual 

species within the landscape and “see” the effect of the oxidant exposure. There are several 

disadvantages, however. Some species (e.g., white fir) exhibit no foliar injury but do have shifts 

in biomass allocation in response to oxidant exposure (Retzlaff et al., 2000). Other species have 

shown significant decreases in foliar injury due to needle loss, with retranslocation of nutrients 

to remaining foliage, and subsequent increased photosynthetic rate (Beyers et al., 1992). 

In addition, the development of foliar symptoms within a species is related to sunlight and 

microclimate (Davison et al., 2003). 

AX9-282
Figure AX9-20. A conceptual diagram of processes and storage pools in sources and sinks 

that are affected by O3 exposure. A plus (+) denotes an increase in 

process rate or pool size, a minus (!) denotes a decrease in process rate or 

pool size, and a plus-minus (±) denotes that both increases and decreases 

have been reported in response to O3. Primary effects in the shoots (1°) 

are distinguished from secondary effects in roots (2°) since the primary 

site of O3 action occurs in the leaves. 

Source: Andersen (2003). 

In the biomass approach, O3 exposure resulting in a measurable decline in biomass (usually 

of a target, sensitive species) is used to evaluate landscape condition. The bulk of the 

information available is from seedling responses to controlled chamber exposures, reviewed in 

the previous section. Some information exists for species in natural environments, but teasing 

out concurrent stressors and finding adequate controls may be intractable. For example, in a 

long-term gradient of O3 exposure, N deposition, and drought, the site with the highest O3 

exposure had the greatest whole tree biomass (pole-sized trees) due to growth stimulation by 

N deposition (Grulke and Balduman, 1999). 

AX9-283
In the biogeochemical approach, changes in biogeochemical cycling are used to assess 

landscape condition. Ozone-sensitive species have well known responses to O3 exposure, 

including altered C allocation to below- and above-ground tissues, and altered rates of leaf 

production, turnover, and decomposition. Changes in turnover rates of ephemeral tissues 

(leaves, fine roots) also affect nutritional status of the remaining tissue. These shifts can affect 

overall C and N loss from the ecosystem in terms of respired C, and leached aqueous dissolved 

organic and inorganic C and N. Instability in C and N pools and their dynamics can affect 

landscape-level nutrient dynamics even without significant inputs of N deposition. The endpoint 

assessment is based on changes in water quality from or in the landscape, correlated to a defined 

oxidant exposure level. These approaches are linkable: visible injury at a particular level could 

be related to reduction in photosynthate, which would reduce whole plant biomass (and carbon 

dynamics). If O3-sensitive species dominate the landscape, then changes in C and N dynamics 

over time would be expected to alter biogeochemical cycles. Examples of forest types that 

contain geographically extensive, O3-sensitive species that could be used in assessing 

landscape-level changes include Ponderosa pine in the western United States, yellow poplar or 

loblolly pine in the eastern United States deciduous forests, and Norway spruce in the Carpathian 

Mountains of eastern Europe. 

Water quantity may also be affected by O3 exposure at the landscape level. Moderately 

high O3 exposure may affect the mechanism of stomatal opening (McAinsh et al., 2002), 

resulting in sluggish stomatal opening and closing (Reich and Lassoie, 1984). During 

moderately high O3 exposure in a drought year, canopy transpiration was greater for yellow 

poplar than on adjacent days with lower O3 exposure, which could alter water use at the 

landscape level. Oxidant exposure (O3 and NOx) may decrease the ability of exposed plants to 

close stomata at night (Grulke et al., 2004), thus increasing water loss from the landscape. 

Ecosystem models should aid in interpreting O3-exposure effects at the landscape level. 

AX9.6.4 Biotic Condition 

AX9.6.4.1 Ecosystems and Communities 

The SAB framework described by Young and Sanzone (2002) identifies community extent, 

community composition, trophic structure, community dynamics, and physical structure as EEAs 

for assessing ecosystem health. 

AX9-284
COMMUNITY EXTENT 

Ecosystem function is dependent on areal extent, constituent species composition, trophic 

structure and its dynamics, and community physical structure. Genetic variation within species, 

and the dynamics of the interactions that exist among different species and their biotic and 

abiotic environment, are also involved (Agrawal and Agrawal, 2000). There are no reports of O3 

exposure altering community distribution or extent. 

COMMUNITY COMPOSITION 

Significant changes in plant community composition resulting directly from O3 exposure 

has been demonstrated in two forested areas: the mixed conifer forest of the San Bernardino 

Mountains, CA and the mixed conifer forest of the Valley of Mexico near Mexico City. It is 

also likely that community composition has changed in response to O3 exposure in the 

coniferous forests of the Carpathian Mountains, but this has not yet been definitively shown. 

The first forest communities shown to be affected by O3 were the Pinus ponderosadominated stands of the San Bernardino Mountains in southern California (Miller, 1973). Miller 

suggested that mixed forests of P. ponderosa, Pinus jeffreyi, and Abies concolor were changing 

to predominantly A. concolor because of the greater O3 sensitivity of the pines. Significantly 

greater mortality of young mature trees (50 to 99 years old) occurred in sites that also showed 

higher foliar injury relative to sites that showed slight foliar injury (McBride and Laven, 1999). 

For P. ponderosa, 33% of the trees in the high foliar injury sites died versus 7% of the trees in 

the low foliar injury sites over the decade-long census. In contrast, 24% of Abies concolor died 

in high foliar injury sites, whereas no trees died in slight injury sites. The authors suggested that 

certain age classes were especially sensitive to O3 exposure, because they are emerging into the 

canopy where higher O3 concentrations are encountered. Future projections based on past 

changes in community composition have been conducted for 2024 and 2074 (McBride and 

Laven, 1999). In their projections, the population of Ponderosa pine nearly disappears in all tree 

age classes, and the community is dominated by California black oak (Quercus kelloggii) in all 

tree age classes, followed by Incense-cedar (Calocedrus decurrens) and sugar pine (Pinus 

lambertiana) by the year 2074. Their projections do not account for potential changes in genetic 

structure of the more O3-sensitive species. 

AX9-285
In the Valley of Mexico, the closed forest structure changed to a woodland from high 

pollutant exposure (Fiscus et al., 2002). Cryptogamic community diversity also significantly 

declined in response to prolonged, extreme O3 exposure (Zambrano and Nash, 2000). Together, 

these two examples illustrate the potential for shifts in community composition in response to O3 

stress. 

TROPHIC STRUCTURE 

Above-Ground 

One of the first reports of trophic level interactions in natural communities was 

the O3-induced predisposition of Ponderosa pine to attack by bark beetles (Cobb et al., 

1968; Stark et al., 1968; Stark and Cobb, 1969). Trees exposed to oxidant injury had lower resin 

production, flow, and exudation pressure. Also, several attributes associated with tree defense 

against beetle attack including sapwood and phloem moisture content and phloem thickness 

were compromised by oxidant exposure (Pronos et al., 1999). Another trophic level has been 

implicated, in that O3-injured Ponderosa pine had the same rate of bark beetle infection, but 

healthy trees had greater numbers of bark beetle predators and parasitoids (Dahlsten et al., 1997). 

This suggests that O3 damage rendered the pines inhospitable for the natural enemies of the bark 

beetles. Similar findings were presented by Percy et al. (2002) for aphids whose abundance was 

increased in young Populus tremuloides stands exposed to elevated O3. In that study, the levels 

of natural enemies of aphids (ladybirds, lacewings, spiders, and parasitoids) were significantly 

decreased under elevated O3. 

Below-Ground 

Processing of plant-derived carbon compounds by soil organisms comprising the soil food 

web is a fundamental property of a functional and stable below-ground ecosystem (De Ruiter 

et al., 1998; Wolters, 1998). Soil food web organisms are responsible for recycling nutrients and 

for development of soil properties such as porosity, aggregate structure, water-holding capacity, 

and cation exchange capacity. A shift in food-web species diversity or functional complexity in 

response to O3 stress may alter ecosystem processes. 

Evidence that soil organisms are affected by O3 indicates the potential for changes in soil 

food-web structure and function. Since O3 does not penetrate the soil beyond a few centimeters, 

AX9-286
the proposed mechanism by which O3 alters soil biota is through a change in C input to soils 

(Andersen, 2003). Ozone can alter C inputs to soil and hence soil processes through four 

different pathways: (1) leaf-litter quality and quantity (see Material Cycling, below), (2) C 

allocation to roots (see Physiological Status, below), (3) interactions among root symbionts, and 

(4) rhizodeposition. The complex nature of the effects of O3 on trophic interactions and food 

webs calls for additional basic research and modeling. 

There have been no comprehensive studies on the effects of O3 on structural components of 

soil food webs; however, studies have shown that O3 affects free-living soil organisms of food 

webs. In the few cases where soil microbial communities have been examined, O3 has led to 

changes in bacterial and fungal biomass, and, in some cases, changes in soil enzyme activity. 

Phillips et al. (2002) examined the effects of elevated CO2 and O3 on C flow through 

heterotrophic microbial communities in soils collected from a FACE study in Wisconsin. Ozone 

decreased abundance of fungal phospholipid fatty acids in aspen and birch-aspen plots but had 

few other direct effects on measured soil parameters. The greatest effect of O3 was to eliminate 

significant increases in microbial respiration resulting from elevated CO2, suggesting an 

important role for O3 in altering C flow through soils. Shafer (1988) found that O3 tended to 

increase the number of fungal propagules and bacteria exhibiting phosphatase activity in the 

rhizosphere of sorghum. Ozone in combination with simulated acid rain stimulated soil 

arylsufatase activity (Reddy et al., 1991). The response was observed at low concentrations, but 

was reversed at high concentrations, suggesting a threshold level of O3, possibly involving 

different mechanisms. Ozone significantly decreased soil microbial biomass in the fall after one 

season of exposure in a wheat and soybean system (Islam et al., 2000). Other studies have 

shown shifts in microbial and fungal biomass in response to O3 stress, but responses were 

variable (Scagel and Andersen, 1997; Yoshida et al., 2001). 

Decreased allocation to roots associated with O3 exposure alters N fixation in legumes and 

actinorrhizal species. Ozone exposure was found to decrease nodulation in a number of species 

(Manning et al., 1971; Tingey and Blum, 1973). In alder (Alnus serrulata), host root cells of 

nodules showed cytoplasmic breakdown and lacked organelles when seedlings were exposed 

to O3 for 27 days (Greitner and Winner, 1989). 

Ozone has been shown to affect mycorrhizal colonization (Adams and O'Neill, 1991; 

Edwards and Kelly, 1992; Ho and Trappe, 1984; McCool et al., 1982; Simmons and Kelly, 1989; 

AX9-287
Smith and Read, 1997). Although short term in nature, several studies have found enhanced 

mycorrhizal short-root formation under O3 stress. White pine (Pinus strobus) (Stroo et al., 

1988), Norway spruce (Rantanen et al., 1994), Northern red oak (Quercus rubra) (Reich et al., 

1985), Douglas fir (Pseudotsuga menziesii) (Gorissen et al., 1991a), European silver fir (Abies 

alba) (Wollmer and Kottke, 1990), and Scots pine (Kasurinen et al., 1999) all showed some 

increase in mycorrhizal presence when exposed to O3. Others have shown minimal or no effects 

of O3 on mycorrhizas (Kainulainen et al., 2000b; Mahoney et al., 1985; Meier et al., 1990). 

Stroo et al. (1988) found that percent infection increased from 0.02 to 0.06 ppm O3, then 

decreased from 0.06 to 0.14 ppm; the total number of short roots were unaffected, however. In 

cases where stimulation was observed, the response was often noted shortly after initiation of 

exposure, often at relatively low concentrations. Good examples of this transitory response can 

be found in results with Norway spruce and Scots pine (Kasurinen et al., 1999; Rantanen et al., 

1994). In these studies, O3 increased mycorrhizal short roots initially but differences were not 

evident by the end of the experiment. 

Evidence suggests that decreased below-ground allocation associated with O3 stress alters 

mycorrhizal host-symbiont compatibility. Edwards and Kelly (1992) found a shift in fungal 

morphotypes present on loblolly pine roots, even though the number of mycorrhizal short roots 

per gram fine root was not significantly affected by O3. Qui et al. (1993) found increased 

numbers of morphotypes present in O3-sensitive loblolly pine seedlings exposed to O3. Roth and 

Fahey (1998) found an interaction between O3 and acid precipitation treatments on the 

composition of fungi forming ectomycorrhizae on red spruce saplings, possibly driven by 

nutrient availability. Carbohydrate requirements vary among fungal species (Bidartondo et al., 

2001), and O3 may affect species composition by altering carbohydrate availability in roots. 

A shift in species dominance could lead to a change in successional patterns of mycorrhizal 

communities. 

In the few studies that examined root exudation in response to O3 exposure, O3 was found 

to alter the quantity and quality of root exudates. McCool and Menge (1983) found a significant 

decrease in exudation of amino acids in tomato exposed to 300 ppb O3. McCrady and Andersen 

(2000) observed increased root exudation in nonmycorrhizal wheat. No apparent change in root 

exudation was found in labeling studies of ECM Ponderosa pine (Andersen and Rygiewicz, 

1995a). Inconsistency in the literature probably results from species differences and 

AX9-288
experimental protocols; however, these examples illustrate the potential effects of O3 on 

rhizosphere C flux. 

Decreased C allocation to roots of O3-exposed plants may reduce root longevity and 

accelerate root turnover, increasing rhizodeposition of C and N. Fine root turnover decreased in 

mature northern red oak exposed to elevated O3 (seasonal exposure ranging from 152 to 

189 ppmAh), whereas seedlings did not show any reduction in turnover (Kelting et al., 1995). 

King et al. (2001) found a trend toward decreased live root biomass and increased dead root 

biomass in aspen exposed to O3 in a FACE study, suggesting possible changes in both 

production and longevity. 

Other studies also suggest that O3 alters C flux to soils, resulting in changes in CO2 efflux 

from soils. Both root respiration and soil CO2 efflux decreased from loblolly pine seedlings 

exposed to O3 (Edwards, 1991). Soil CO2 efflux increased in response to O3 in Ponderosa pine 

seedlings (Andersen and Scagel, 1997). No direct assessments of hyphal growth and turnover in 

response to O3 stress have been conducted. Ozone decreased C allocation to extrametrical 

hyphae of a Ponderosa pine mycorrhiza, which might be expected to decrease growth and 

increase hyphal turnover (Andersen and Rygiewicz, 1995b). 

COMMUNITY DYNAMICS, PHYSICAL STRUCTURE 

One of the best-documented examples of change in long-term forest community dynamics, 

in dominant overstory trees, occurred in the San Bernardino Mountains between 1968 and 1974 

(as reported by Miller [1973] and Miller et al. [1989]). Plots were recently re-inventoried 

~25 years after establishment (Arbaugh et al., 2003). Of the six codominant canopy species, 

white fir showed the greatest change, increasing in both numbers and bole growth for a 286% 

change in basal area/ha in the San Bernardino Mountains. Sugar pine basal area also increased 

significantly (by 334%), but this species represents only a small portion (1%) of the total basal 

area of the forest sampled. The most O3-sensitive species (Miller et al., 1982), Ponderosa and 

Jeffrey pine, had the lowest increase in basal area/ha (76 and 62%, respectively). These two 

species represented 72% of the basal area/ha of all stands inventoried. Ponderosa pine had the 

greatest mortality rate of all canopy species inventoried (46%), followed by white fir and black 

oak (35% and 33%), Jeffrey pine (29%) and Incense cedar and sugar pine (both 7%). In moist 

sites (at the western end of the San Bernardino Mountains), there was significant recruitment of 

AX9-289
Incense cedar, white fir, and sugar pine. Only one study directly attributed tree mortality to O3 

exposure: it accounted for 7% of mortality in the Sierra and Sequoia NFs (Carroll et al., 2003). 

Species diversity in the understory can be quite large, making studies of O3 effects on 

understory community dynamics very challenging. However, there have been some attempts to 

quantify understory responses, ranging from describing relative sensitivity to their visible 

symptoms (Temple, 1999; Treshow and Stewart, 1973) to very complex measures of community 

structure and composition (Westman, 1979, 1981). The lowest percentage cover and lowest 

species diversity in California coastal sage scrub was correlated with the highest O3 exposures as 

estimated by extrapolation from the closest air monitoring stations (Westman, 1979). The 

understory also has the potential to influence responses to O3 of dominant keystone species, as 

has been shown in controlled experiments with both Ponderosa pine (Andersen et al., 2001) and 

Loblloly pine (Barbo et al., 2002). Barbo et al. (1998) exposed an early successional forest 

community to ambient air, charcoal-filtered air, non-filtered air, and 2× ambient in the 

Shenandoah NP. They found changes in species performance, canopy structure, species 

richness, and diversity index consistent with the view that O3 can induce a shift in vegetation 

dominance and community structure. 

There have been few studies evaluating the effect of O3 exposure on the physical structure 

of natural ecosystems. Despite an extensive array of allometric equations for conifers in the 

western United States (Ter-Mikaelian and Korzukhin, 1997), none appear to predict individual 

tree shape in a site of moderate O3 exposure, suggesting that O3 may effect allometry (Grulke 

et al., 2003a). Canopy structural changes are also implied by the measure of canopy 

transparency used in the USDA Forest Service’s Forest Health Monitoring (FHM) assessment. 

The loss of epiphytic lichens within the canopy is a clear example of plant community structural 

change occurring along an O3 gradient (Nash and Sigal, 1999; Zambrano and Nash, 2000). 

As of yet, there have been no comprehensive studies on the effects of O3 on structural or 

functional below-ground components (Andersen, 2003). Phillips et al. (2002) found evidence 

for changes in the bacterial and fungal biomass below aspen and aspen/paper birch stands 

exposed to elevated O3. Subsequent study showed that O3 exposure decreased cellobiohydrolase 

activity in the soil microorganisms, driving the change in the microbial community (Larson 

et al., 2002). 

AX9-290
AX9.6.4.2 Species and Populations 

Ozone can affect species and populations of species found in ecosystems through changes 

in population size, genetic diversity, population structure and/or dynamics, and habitat suitability 

(Young and Sanzone, 2002). For example, if individuals of a species are lost due to O3 

exposure, population size declines. Often very young (e.g., conifer seedlings, see Section 

AX9.6.4.3 below) and old individuals differ in their sensitivity, so that population structure also 

will be altered by O3 exposure. If resource allocation to reproductive output is altered by O3 

exposure, population dynamics will be altered. Communities dominated by O3 sensitive species 

in the canopy or understory may be altered sufficiently for the habitat to become unsuitable for 

other species. Genetic selection acts on the individual plant, which represents a certain 

proportion of the populations’ genetic variation. If an O3-sensitive individual succumbs through 

multiple stresses, including O3 stress, the genetic variation represented in the population 

generally declines, unless sensitive individuals have low inherent genetic variability (e.g., 

Staszak et al., 2004). 

While the concept of natural selection induced by O3 exposure and related changes in 

natural plant communities has been around for a long time (Dunn, 1959; Miller et al., 1972), the 

concept of evolution of O3 tolerance is still not widely accepted. The unequivocal demonstration 

that considerable genetic variation in O3 resistance exists within and between plant populations, 

and that ambient levels of O3 may differentially affect fitness-related traits (i.e., growth, survival, 

and fecundity), suggests that O3 may potentially drive the natural selection for resistant 

individuals. Dunn (1959) presented circumstantial evidence that ambient O3 in the Los Angeles 

area was high enough to drive the selection of O3-resistant Lupinus bicolor genotypes. Since 

Dunn’s (Dunn, 1959) research on O3-induced population changes, researchers have 

demonstrated differences in O3 tolerance among other plant populations. In the devastating 

forest decline southeast of Mexico City, the remaining trees (primarily Abies religiosa in the 

“cemetery forests”) appear to be less affected by foliar injury than trees that were lost, despite 

continued high-O3 exposures (Alvarado et al., 1993). However, even the most convincing work 

in this field (Berrang et al., 1986, 1989, 1991), with Populus tremuloides, where a strong 

correlation between visible foliar injury after O3 exposure and maximum O3 concentration at the 

origin of the population was shown (Berrang et al., 1991), a change in gene frequency at any one 

site over time has not yet been demonstrated (Bell et al., 1991; Reiling and Davison, 1992a). 

AX9-291
Furthermore, the selection intensity of O3 has been questioned (Bell et al., 1991; Taylor et al., 

1994; Taylor and Pitelka, 1992) and the emergence of O3 exposure since the 1950s as an 

environmental stressor may not have been long enough to affect tree populations with long 

generation times (Barrett and Bush, 1991). 

The loss of O3-sensitive individuals results in natural selection favoring O3-tolerant species 

(Bradshaw and McNeilly, 1991). Increased levels of mortality of O3-sensitive individuals have 

occurred for Pinus jeffreyi and P. ponderosa exposed to ambient O3 along the western slope of 

the Sierra Nevadas (Miller et al., 1998; Peterson et al., 1987), for Pinus strobus exposed to 

ambient O3 in southern Wisconsin (Karnosky, 1981), and for P. ponderosa in the San Bernardino 

Mountains (Carroll et al., 2003). In these examples, individuals that consistently had greater 

foliar injury and lower needle retention were lost in repeated surveys. Ozone-induced loss of all 

individuals except the most tolerant and breeding among the surviving individuals to yield more 

more tolerant populations has not yet been demonstrated for plants exposed to O3, except for the 

relatively short-term (2 years) adaptation exhibited in Trifolium repens (Heagle et al., 1991) and 

Plantago major (Davison and Reiling, 1995). Heagle et al. (1991) were able to show the 

adaptation of a Trifolium repens population to elevated O3 in just two growing seasons. 

Similarly, Davison and Reiling (1995) compared the O3 resistance of P. major populations 

grown from seed collected from the same sites over a period of increasing O3. The two 

independent populations studied exhibited increased O3 resistance, consistent with the idea of 

selection for O3 tolerance. Using random amplified polymorphic DNA primers, this team also 

showed that the later populations are subsets of the earlier ones, consistent with in-situ evolution 

rather than with catastrophic loss and replacement of the populations (Wolff et al., 2000). The 

problem is determining whether spatial patterns in O3 resistance and changes in time are casually 

related to O3, because there were very strong correlations with other factors (Davison and 

Barnes, 1998; Reiling and Davison, 1992c). The potential for the evolution of O3 resistance has 

been clearly demonstrated by Whitfield et al. (1997) in their study of O3 selection of common 

plantain, where they showed that within a matter of a few generations, it was possible to 

increase O3 resistance in an initially O3-sensitive population. Wild radish (Raphanus sativus) 

developed O3 resistance after only one generation of exposure to O3 (Gillespie and Winner, 

1989). 

AX9-292
A third independent line of research suggesting O3 may be affecting the genetic diversity of 

wild plant populations was presented by Paludan-Müller et al. (1999) who showed that northwest 

European provenances of European beech were more sensitive to O3 than were southeast 

European provenances that had experienced higher O3 levels. Recent research on the genetic 

structure of 50-year-old Ponderosa pines in the San Bernardino Mountains suggests that distinct 

differences in frequency of some alleles and genotypes occurred, with the O3-tolerant trees being 

more heterozygous (Staszak et al., 2004). While both of these studies were only correlational, 

they are consistent with previous studies of this type, suggesting O3-induced population changes. 

Again, other environmental stressors besides O3 exposure could have been involved in effecting 

change within these populations. 

Natural selection for O3 tolerance can also be facilitated by reductions in fitness related to 

lower seed yields of O3-sensitive species or individuals. The impacts of O3 on reproductive 

development, recently reviewed by Black et al. (2000) can occur by influencing (1) age of 

flowering, particularly in long-lived trees that often have long juvenile periods of early growth 

without flower and seed production; (2) flower bud initiation and development; (3) pollen 

germination and pollen tube growth; and (4) seed, fruit, or cone yields and seed quality (Table 

AX9-24). In addition, vegetatively propagated species can have lower numbers of propagules 

under elevated O3 conditions (Table AX9-24). 

Several studies suggest that reproductive structures are clearly sensitive to O3 and that O3 

can affect fitness of plants by affecting either the sporophytic or gametic generations. Decreased 

numbers of flower spikes and seed capsules per plant were found for plantain growing under 

elevated O3 (Lyons and Barnes, 1998; Pearson et al., 1996; Reiling and Davison, 1992a). 

Similar responses were seen for Brassica campestris plants exposed to a single dose of 100 ppb 

O3 for 6 h. Stewart et al. (1996) and Bosac et al. (1998) reported an increase of flower bud 

abortion for oilseed rape (Brassica napus L.) similarly exposed to a short duration of elevated 

O3. Floral initiation period can be delayed in O3-sensitive plants, as was described for dogbane 

(Apocynum androsaemifolium) grown under ambient O3 in the eastern United States. In one of 

the few comparisons of whole plant O3 sensitivities with that of male gametophytes, Hormaza 

et al. (1996) found a high correlation of relative O3 sensitivity of pollen tube elongation with that 

of O3 effects on net photosynthesis and relative growth rates for 6 species of fruit trees. 

AX9-293
Clearly, the concept of O3 -induced genetic change is an area that needs additional research 

attention. Repeated collections over time from wild populations receiving high O3 exposures to 

examine population responses and relative sensitivity changes, the sampling of genetic diversity 

along known O3 gradients, and the use of modern biotechnological approaches to characterize 

and quantify genetic diversity are useful approaches to test for O3-induced impacts on diversity 

in natural ecosystems. 

AX9.6.5 Organism Condition 

PHYSIOLOGICAL STATUS 

The generalized effects of O3 exposure on plants are well known, and have been reviewed 

from several viewpoints over the last decade (Darrall, 1989; De Kok and Tausz, 2001; Heath and 

Taylor, 1997; Matyssek et al., 1995; Pell et al., 1997; Reich, 1987; Schraudner et al., 1997; U.S. 

Environmental Protection Agency, 1996). The topic of individual species response and 

modification of response by other factors has been addressed thoroughly in Sections AX9.6.3 

and AX9.6.4 of this chapter. Here, this subsection will describe the physiological changes in 

response to O3 that have been hypothesized to lead to changes in ecosystem structure or function. 

Above-Ground Responses 

The first critical step leading to O3 response is uptake of O3 by the leaves, leading to 

changes in C and nutrient relations that are thought to alter plant growth and competitiveness 

(see Section AX9.3). Ozone enters leaves through stomata, reacts with cell walls or membranes, 

and starts a series of adverse reactions. Cuticular uptake of O3 is believed to be negligible (Coe 

et al., 1995; Kerstiens and Lendzian, 1989). Once inside the leaf, O3 and its byproducts lead to 

membrane disruption, chlorophyll breakdown, and decreased Rubisco levels (Schweizer and 

Arndt, 1990). In turn, photosynthesis is decreased, as is stomatal conductance (Weber et al., 

1993). Also, O3 often leads to increased maintenance respiration, decreased foliar nutrient 

content, and imbalances in tissue nutrient content and retention. When photosynthetic pigments 

have been damaged, the pigment must be fully broken down (and/or new N and Mg must be 

taken up and transported to the leaf) for the pigment to be regenerated (Bjorkman and DemmigAdams, 1995). Ozone exposure alters within-plant priorities for resources: less C is available 

for allocation to roots and spring regrowth, and less foliar biomass is retained. At the wholeAX9-294
organism level, O3 exposure decreases root mass (Grulke et al., 1998a) and radial bole growth 

(Muzika et al., 2004; Peterson et al., 1991) with little impact on height growth. Visible 

symptoms of O3 injury vary between species and genotypes but often include upper leaf surface 

stipple, chlorotic mottle, or large bifacial blotches of necrotic tissue. Premature senescence is 

typical of almost all O3-induced foliar damage. All of these changes can alter the plant’s ability 

to function in a broader ecosystem context. 

The underlying mechanisms of O3-injury response in conifers, broadleaf deciduous trees, 

and herbaceous species are assumed to be similar. However, several differences in long-lived 

species are important at the ecosystem level. Most of the research on O3 effects has been 

conducted on herbaceous species (i.e., crops). Although a number of native herbaceous species 

have been identified as O3-sensitive, there are no published physiological studies on the effect 

of O3 exposure on herbaceous or shrub species in situ. In natural ecosystems, the majority of 

species are not annuals, unless the system is highly disturbed. Nonetheless, response of crop 

species to elevated O3 may be used as an analog for native annual response: phenological 

staging is accelerated (soybeans; Booker et al. [2004]), thus “avoiding” additional O3 exposure. 

Conifers have roughly half the stomatal conductance of deciduous broadleaf trees (Reich, 

1987), leading to proportionally less O3 uptake at the same O3 exposure level. Yet, except for 

species of larch, individual conifer needles are longer-lived and active over a greater portion of 

the year. Therefore, needle longevity can also work against the tree by increasing cumulative O3 

exposure and exposure to other stressors. Increased needle longevity is not always a 

disadvantage, for example, conifers are physiologically active in early spring and late fall, during 

times of lower oxidant concentrations. These periods can contribute significantly to a net 

positive annual C balance and from the standpoint of nutrient storage, are important in reparation 

responses to pollutants. Patterson and Rundel (1995) reported that Jeffrey pine had significant 

stomatal opening (one third that of a typical summer day) in midwinter with snow on the ground. 

At least pole-sized and larger trees can mitigate reductions in C acquisition due to oxidant 

exposure in the summer with C assimilation on favorable days in the winter. The interaction of 

environmental factors, plant phenology (the timing of growth events; birth and mortality of plant 

parts), physiological status (nutritional or moisture status; dormant or active growth within the 

year), and tree age (interannual differences in resource acquisition and allocation) all contribute 

to the complexity of long-lived species (and hence ecosystem) response to O3 exposure. 

AX9-295
One widely observed response to O3 exposure is premature leaf loss. As noted above, 

premature leaf loss may reduce O3 uptake during high-O3 years, but it has several negative 

consequences. Early leaf loss results in reduced C uptake through photosynthesis. Premature 

needle loss also results in less N retranslocation compared to normally senescing leaves, which 

reduces whole plant N balance (Fenn and Dunn, 1989) and carbohydrate availability for 

overwinter storage (Grulke et al., 2001). Because of such effects accumulated over several years 

of O3 exposure, subsequent-year C and N reserves can be affected (Andersen et al., 1991). 

Conversely, a series of drought years can decrease O3 uptake, as well as reduce C and 

nutrient acquisition, altering resource allocation to defenses (e.g., antioxidants) (Grulke et al., 

2003b) (or resins) against insect infestation, rendering the tree more susceptible to O3 injury. 

Conifers have thicker cuticles than either broadleaf deciduous or herbaceous species. 

Continued O3 exposure may compromise cuticular integrity (Bytnerowicz and Turunen, 1994). 

Once cuticular integrity is breached, individual leaves (needles) are likely to be excised, thus 

contributing generally to defoliation and reduced C acquisition. 

With the exception of the extensive research conducted on mature tree response to O3 

exposure in California forests (Arbaugh et al., 1998; Grulke et al., 1996, 1998b, 2001, 2002b, 

2003a,b, 2004; Grulke, 1999; Grulke and Balduman, 1999; Peterson et al., 1987, 1991, 1995; 

Wieser et al., 2002), the vast majority of studies of O3 effects on forest trees have been 

conducted on young seedlings (Chappelka and Samuelson, 1998) and little is known about 

acclimation to O3 (Skärby et al., 1998). Chamber exposure studies can be used to document 

foliar symptoms and develop response variables for the whole plant. These response variables 

can then be field tested on mature trees using correlative analyses (e.g., Grulke and Lee [1997]; 

Grulke et al. [2003b]). Without the initial work in chamber exposure studies, field responses to 

O3 exposure would be difficult to verify and distinguish from other concurrent stressors. 

Predicting mature tree responses to O3 solely from seedling response studies is complex, 

because seedlings or saplings do not necessarily respond to O3 in the same way as mature trees 

(Karnosky et al., 2003a; Norby et al., 1999). Although species dependent, O3 has been found to 

have stronger effects on leaf function in younger rather than older trees (Kolb and Matyssek, 

2001). Each component physiological attribute “matures” at a different rate. Gas exchange 

patterns differ between seedlings and mature trees. For example, leaf respiration of juvenile 

Ponderosa pine was greater than that of mature trees (Momen et al., 1996). In the conifers 

AX9-296
tested, the highest gas exchange rates (and by inference stomatal uptake of O3) are found in 

seedlings (e.g., in scions of red spruce, Rebbeck et al. [1993]; giant sequoia, Grulke and Miller 

[1994]; Ponderosa pine, Grulke and Retzlaff [2001]; and Norway spruce, Wieser et al. [2002]). 

Patterns of biomass (Grulke and Balduman, 1999) and carbohydrate allocation (Grulke et al., 

2001) differ between immature and mature trees. Pole-sized trees had greater reduction in root, 

foliar, and bole carbohydrate concentrations than did old growth trees. Antioxidant defenses 

vary with both tree age and needle age (Tegischer et al., 2002). Based on all attributes measured 

in both Ponderosa pine and giant sequoia, the youngest tree age considered representative of 

mature trees was 20 years old (Grulke et al., 1996; Grulke and Retzlaff, 2001). In some 

broadleaf deciduous tree species, seedlings are more conservative, and mature trees have greater 

gas exchange rates, as is the case for Quercus rubra (Edwards et al., 1994; Kelting et al., 1995; 

Samuelson and Kelly, 1996, 1997) and Fagus sylvatica (Braun et al., 1999). In another 

broadleaf deciduous tree species (black cherry), gas-exchange rates of seedlings were faster, but 

total O3 flux to leaves of seedlings was lower than that of mature trees due to differences in leaf 

ontogeny (Fredericksen et al., 1995, 1996b). 

Nitrogen deposition modifies the effects of oxidant exposure through several offsetting 

physiological mechanisms (see Section AX9.4.4). Nitrogen deposition, in wet or dry particulate 

form, ultimately increases site fertility, but increased soil N availability decreases C allocation to 

roots, further exacerbating the effects of O3 exposure on roots (Grulke et al., 1998a). Increased 

N availability also increases foliage turnover: fewer needle age classes are retained (Gower 

et al., 1993). Therefore, the combination of both increased N and O3 exposure increases foliar 

turnover. Finally, N deposition and increased plant N nutrition can increase stomatal 

conductance, leading to increased O3 uptake. Alternatively, increased N counteracts the effect 

of O3 on photosynthesis by increasing photosynthetic pigments and enzymes. Nitrogen 

deposition may mitigate the degree of foliar injury from oxidant pollution via higher available 

N for reparation of photosynthetic pigments. Nitrogen amendments also modify the antioxidant 

defense system in complex ways (Polle, 1998). 

Attributes of O3 injury to trees (foliar injury, needle retention, and canopy transparency), 

as well as presence of pathogens and insect infestation, are routinely inventoried in established 

sample plots distributed on Federal lands across the United States (Forest Health Protection, 

USDA Forest Service). Foliar injury to several widespread, herbaceous species nationally 

AX9-297
recognized as sensitive (bioindicators) is also assessed (U.S. National Park Service, 2003). This 

assessment is part of a larger assessment of forest tree growth and dynamics (the Forest 

Inventory and Analysis Program) (Smith et al., 2003; Smith, 2002)). Risk of O3 injury is then 

estimated for the dominant forest tree species in the sample plots. For example, 12% of sampled 

black cherry, 15% of loblolly pine, and 24% of sweetgum (Liquidambar styraciflua) were found 

to be in the highest risk category in the northeast and mid-Atlantic states (Coulston et al., 2003). 

In the Carpathian Mountains, 12 to 13% of all trees (broadleaf and coniferous) have greater than 

26% crown defoliation (Bytnerowicz et al., 2002b). In general, broadleaves (primarily beech) 

trees were less affected (8 to 45%) than spruce (up to 37%) and fir (up to 50%) (Bytnerowicz 

et al., 2002b). Ozone injury was directly correlated with cumulative O3 exposure in the Sierra 

Nevada Mountains (Arbaugh et al., 1998); with the best correlation being found across sites 

where >90% of the trees had O3 injury. Although direct links of visible foliar symptoms induced 

by O3 to adverse effects on biomass are not always found, visible foliar symptoms have been 

linked to decreased vegetative growth (Karnosky et al., 1996; Peterson et al., 1987; Somers 

et al., 1998), as well as reproductive function (Black et al., 2000; Chappelka, 2002). 

Foliar O3 injury has also been associated with adverse effects on competitive ability and 

survival in forest communities (Karnosky, 1981; Karnosky et al., 2003a; McDonald et al., 2002). 

Competition can alter organism condition and affect susceptibility to O3. Ponderosa pine 

seedlings were more susceptible to O3, as determined by decreased plant biomass, when grown 

in competition with blue wild-rye grass (Andersen et al., 2001). Similarly, the magnitude of O3 

effects on height and diameter growth depended on the competitive status of Populus 

tremuloides trees (McDonald et al., 2002). These studies show the importance of including 

competition as a concurrent stressor in assessing whole plant responses to O3. The age of the 

community (“time since disturbance”) may also affect the ability of individuals to effectively 

respond to O3. Unfortunately, the vast majority of O3 studies have been conducted on 

open-grown plants, often grown in pots where competition is absent both above and below 

ground. 

Clearly, age-dependent O3 responsiveness and juvenile-mature correlations remain 

important research questions in attempting to scale up to ecosystem level responses. Patterns of 

allocation between root, stem, and leaf differ between immature and mature trees. Tree 

architecture varies with tree age, and leaf area distribution in space and time may change in 

AX9-298
response to elevated O3. All of these factors influence gas exchange in the canopy. 

Furthermore, there may be few generalities that can be made from seedling to mature tree 

response to O3 within a plant functional group (Karnosky, 2003a; Norby et al., 1999). 

Consequently, modeling ecosystem response is limited to either dealing with monospecific 

plantations or assigning average responses to a mix of species. 

Below-Ground Responses 

The effect of O3 on the soil ecosystem is thought to occur through physiological changes in 

the root and interactions with soil organisms (Andersen, 2003). Comparatively little is known 

about how changes in root growth and metabolism are translated through the soil food web, 

resulting in changes in soil and, hence, ecosystem processes. An overview of physiological 

changes likely to lead to changes at the ecosystem level is provided below. 

Ozone stress decreases C allocation to roots (Cooley and Manning, 1987; Gorissen et al., 

1994; Gorissen and Van Veen, 1988; Manning et al., 1971; McCool and Menge, 1983; 

McLaughlin and McConathy, 1983; Rennenberg et al., 1996; Spence et al., 1990; U.S. 

Environmental Protection Agency, 1996). Since roots are often dependent on current 

photosynthate for their structural development (Marshall and Waring, 1985; Ritchie and Dunlap, 

1980; Van Den Driessche, 1978, 1991), C-limiting stressor such as O3 can have rapid and 

significant effects on root growth. In many cases, decreased allocation to roots in response to O3 

occurs quickly, with reductions in root growth occurring within one growing season (Andersen 

and Rygiewicz, 1995a; Andersen and Rygiewicz, 1991; Gorissen et al., 1991b; Gorissen and 

Van Veen, 1988; Spence et al., 1990; U.S. Environmental Protection Agency, 1996). Decreased 

C allocation below ground is often associated with decreased root-shoot ratio, but observed 

responses in root-shoot ratio are highly variable owing to several factors including intra- and 

interspecies variation, culture conditions, and ontogenetic drift (Reich, 2002). Root-shoot ratio 

is a point-in-time measurement that does not include C lost to exudation, respiration, or turnover. 

Therefore, biomass and ratios of biomass (such as root-shoot ratio) do not necessarily reveal 

physiological changes in response to O3 stress. 

Decreased C acquisition leads to reduced carbohydrate levels and storage pools 

in O3-exposed plants (Andersen and Scagel, 1997; Cooley and Manning, 1987; Gorissen et al., 

1994; Ito et al., 1985; McLaughlin et al., 1982; Rebbeck et al., 1988; Tingey et al., 1976b). 

AX9-299
Although it is difficult to quantify changes in the field, Grulke et al. (1998a) found decreased 

medium and fine root biomass with increased pollutant load across an O3 gradient in southern 

California. Coarse and fine root starch concentrations also were lowest in mature trees at the 

most polluted site (Grulke et al., 2001). The effects of O3 could not be completely separated 

from other known stresses across the pollutant gradient, but it appeared that O3 was an important 

factor in the patterns observed. 

Decreased storage pools can lead to carry-over effects on root growth that are compounded 

over time. Decreased carbohydrate storage pools were associated with decreased root growth 

during the spring following exposure to O3, even in the absence of additional O3 exposure 

(Andersen et al., 1991, 1997). Decreased spring root growth was attributed to decreased stored 

C reserves as well as to premature loss of older foliage age classes during the previous fall. 

Aside from the loss of photosynthetic surface area associated with premature senescence, early 

loss of foliage in the fall occurs when allocation to roots is at a maximum in many species 

(Kozlowski and Pallardy, 1997). Older needle age classes preferentially allocate photosynthate 

basipetally to stems and roots (Gordon and Larson, 1970; Rangnekar et al., 1969); and the loss of 

older needles in the fall during allocation to root growth and storage; and in the spring during 

periods of root growth, preferentially impacts roots and root processes. 

Ozone has also been shown to affect root metabolism as evidenced by changes in root 

respiration. Edwards (1991) found decreased root and soil CO2 efflux during a 2-year exposure 

of loblolly pine to O3. Fine root respiration increased in mature red oak exposed to O3, while 

total soil CO2 efflux increased in the spring and decreased in the summer and fall (Kelting et al., 

1995). The authors attributed increased root respiration to increased nutrient uptake in support 

of increased demands in the shoot. Ozone decreased root system respiration in aspen after 

12 weeks of exposure, but the decrease was closely associated with decreased root biomass and 

probably not metabolic processes (Coleman et al., 1996). Whether other metabolic shifts occur 

in the roots of plants exposed to O3 needs to be examined. 

Measurable effects on roots may occur before effects on shoots are observed because 

shoots have immediate access to C for repair and compensation, whereas roots must compete 

with shoots for C. Mortensen (1998) found decreased root, but not shoot, growth in Betula 

pubescens at O3 exposures of 42 nMol mol-L (applied 12 h day!1

), whereas both root and shoot 

growth were reduced at higher exposures. Chromosomal aberrations were found in root tips of 

AX9-300
Norway spruce exposed to O3, even in the absence of biochemical changes in needles (Wonisch 

et al., 1998, 1999). Using relatively high O3 concentrations (0.15 ppm O3 6 h day!1

), Hofstra 

et al. (1981) found metabolic changes in Phaseolus vulgaris root tips prior to the development of 

leaf injury. Morphological changes in root tips occurred within 2 to 3 days, and metabolism 

declined within 4 to 5 days of initiation of O3 exposure. 

Feedback signals from roots can influence the degree of O3 response. Stolzy et al. (1964) 

exposed tomato roots to periods of anaerobic conditions and tracked a change in leaf 

susceptibility to O3. An exposure of roots to low oxygen conditions for 3 h did not alter 

photosynthesis, but foliar damage was decreased when the roots were subsequently exposed 

to O3. In this case, a signal originating in the root appeared to alter leaf sensitivity to O3, the 

signal possibly being hydraulic in nature and leading to decreased O3 uptake. 

SYMPTOMS OF DISEASE OR TRAUMA, SIGNS OF DISEASE 

Although insects and diseases are dynamic components of forest ecosystems, trees can be 

especially susceptible to outbreaks due to the presence of multiple stressors such as drought and 

pollutant exposure. Ozone can have direct effects on insect or disease organisms, indirect effects 

on the insect or pathogen through changes to the host, and direct or indirect effects on natural 

enemies of the insect or pathogen (Pronos et al., 1999). A full discussion of O3 effects on insect 

and pathogen interactions can be found in Section AX9.4. 

Although the multitude of interacting factors makes it difficult to identify causative factors 

in the field, some recent examples suggest a role for O3 in the timing or magnitude of disease 

attacks in the field. After periods of drought stress (such as 1995 in central Europe), the 

incidence of bark beetle (Ips spp.) appears to increase. In 1998 and 1999, the mean daily capture 

of Ips was lowest in plots with low O3 exposure; the converse was also true (Grodzki et al., 

2004). Elevation confounded the relationships, but the differences in Ips frequency in relation to 

O3 concentrations were highly significant at lower elevations. In the Valley of Mexico, a 1982 

to 1983 drought was documented, but not described as precipitating a bark beetle attack in the 

early 1980s. However, a link between air pollutant exposure and bark beetle attacks were 

implicated, because attacked trees were already O3-stressed at the time of the bark beetle attack 

(Alvarado et al., 1993). 

AX9-301
An early study showed that oxidant exposure predisposed Ponderosa pine to the root 

pathogen Fomes annosus (James et al., 1980). Both root diseases (Pronos et al., 1999) and O3 

exposure (Grulke and Balduman, 1999) can each reduce root biomass, leading to increased 

drought stress, insect attack, and subsequent windthrow or death. Trees may take several years 

to die, and the patterns of precipitation and annual total rainfall interact to drive the level of 

drought stress experienced by the tree (Pronos et al., 1999). Additional research is necessary to 

fully understand the complex interactions occurring between O3 stress and other biotic stresses. 

AX9.6.6 Ecosystem, Chemical, and Physical Characteristics (water, soil) 

AX9.6.6.1 Nutrient Concentrations, Trace Inorganic, and Organic Chemicals 

Ozone exposure reduces the nutritional content of tissues, as well as causing elemental 

imbalances. Foliar nutrient content may be too high (toxic) or too low (deficient), but the 

relative amounts and ratios among all nutrients can also result in imbalances. 

Although N deposition and foliar N content increased with O3 exposure in the San 

Bernardino Mountains, K, Mg, Fe, and Al were all also higher in Ponderosa pine at sites more 

exposed to air pollution (Poth and Fenn, 1998). Trees with greater foliar injury (due to O3 

exposure) had higher current-year needle concentrations of P, K, Zn, and Fe than trees at the 

same site that were less injured. In drought-stressed Ponderosa pine with O3 exposure, foliar N 

was also elevated and retained in the remaining needles (Temple et al., 1992). At a relatively 

clean site in the eastern San Bernardino Mountains, N, P, and K were efficiently readsorbed, but 

the amount of P remaining in the foliage was relatively high compared to defined thresholds. 

The fact that other elements were modified besides the N being deposited emphasizes the degree 

of chemical imbalance in the tissue. Foliar micronutrients were within the normal ranges 

reported for Ponderosa pine (Poth and Fenn, 1998; Powers, 1981). Because both N deposition 

and O3 exposure reduce root biomass, it was unlikely that the foliar nutrient content was higher 

due to greater uptake. Instead, it appears that retranslocation from senescing tissue was 

responsible. Across a pollution gradient in the Carpathian Mountains in eastern Europe, only 

S/N (expected due to high S deposition) and Fe/Mn ratios were out of balance relative to 

established norms. The S was relatively high due to SO2 deposition, and the Fe was relatively 

high due to smelter plumes. No imbalances could be directly attributed to O3 exposure 

(Mankovska et al., 2004). 

AX9-302
AX9.6.7 Ecological Processes 

AX9.6.7.1 Energy Flow 

All green plants generate and use energy-containing C compounds through the processes of 

photosynthesis and respiration. Whole-plant C uptake is dependent on photosynthesis rates, leaf 

area, and leaf phenology. The effects of O3 at the site of action in the leaf are discussed in 

Section AX9.2. Here, the main focus is on whole-plant carbon dynamics resulting from changes 

in C acquisition or use under O3 stress. 

In natural ecosystems, O3 has been shown to depress photosynthesis in sensitive tree 

species including Ponderosa pine (Grulke et al., 2002b; Miller et al., 1969; Takemoto et al., 

1997; Weber et al., 1993) and aspen (Coleman et al., 1995a; Noormets et al., 2001a,b; Sharma 

et al., 2003; Yun and Laurence, 1999a). In a study of mature Jeffrey pine, trees in mesic 

microsites had greater O3 uptake over the growing season in comparison to trees in xeric 

microsites (Grulke et al., 2003a) and greater O3 uptake was correlated with lower mid-canopy 

needle retention, lower branch diameters, and lower foliar N content (Grulke et al., 2003a). 

Chamber studies have also shown negative effects of O3 on tree seedling canopy structure 

(Dickson et al., 2001) and leaf area (Neufeld et al., 1995; Wiltshire et al., 1994). It is well 

known from chamber and field studies that O3 exposure is correlated with lower foliar retention 

(Grulke and Lee, 1997; Karnosky et al., 1996; Miller et al., 1963, 1972; Pell et al., 1999; Topa 

et al., 2001). 

In contrast to the relatively consistent findings for photosynthesis, O3 effects on respiration 

have been more variable. Stem respiration was unaffected by O3 exposure (Matyssek et al., 

2002), suggesting that construction costs of new stems are not affected by O3. However, the bole 

represents a relatively large storage pool of carbohydrates in mature trees, and the timing of 

phenological events among individual trees may help to confound the ability to statistically 

detect differences in stem respiration across pollutant gradients (Grulke et al., 2001). Belowground respiration has been found to both increase and decrease in response to O3, depending on 

the approach and timing of CO2 measures (Andersen and Scagel, 1997; Coleman et al., 1996; 

King et al., 2001; Scagel and Andersen, 1997). The decreased soil respiration is thought to be 

due to reduced root growth under O3 exposure, but could also be partially explained by 

decreased microbial respiration in response to O3. Additional research is necessary to identify 

AX9-303
the role of O3 in affecting root versus heterotrophic respiration, particularly over long time 

intervals. 

Carbohydrate availability and use influence the degree to which plants respond to O3 

exposure. A model simulation of the effect of O3 exposure on bole growth of Ponderosa pine 

showed a 15% reduction in mass (Weber and Grulke, 1995), largely influenced by differences in 

carbohydrates allocated and partitioned in repair processes elsewhere in the tree. Foliar 

respiration is thought to increase under elevated O3 because maintenance costs (energy needs) of 

leaves damaged by O3 are higher than normal (Grulke and Balduman, 1999; Noormets et al., 

2001b). However, differences in foliar respiration are subtle and difficult to detect statistically. 

Foliar carbohydrate studies also suggest that more C is used under O3 stress for repair processes 

(Grulke et al., 2001; Topa et al., 2001) which would result in increased respiration. Ozone 

exposure also reduced enzymatic activities of carbohydrate metabolism related to the breakdown 

of sucrose (Einig et al., 1997). Changes in soil respiration in response to O3, even though O3 

does not penetrate into the soil, illustrates the tight coupling of plant C balance and soil biota and 

illustrates the potential role O3 plays in altering ecosystem C balances (Andersen, 2000). 

Ozone can affect plant allometry through changes in energy use, potentially affecting net 

primary production (NPP) at larger scales. The net effect of O3 impacts on photosynthesis and 

respiration for sensitive components of natural ecosystems is that height growth (Isebrands et al., 

2001; Oksanen, 2003a) and radial growth (Isebrands et al., 2001; Oksanen, 2003b; Peterson 

et al., 1987, 1991) can be negatively affected by O3. This has been extrapolated to decreased 

NPP (Hogsett et al., 1997; Laurence et al., 2000). 

Energy flow in plant communities can be altered by O3 through changes in C allocation. 

It is well known that elevated O3 affects C allocation to roots (Andersen et al., 1997; Coleman 

et al., 1995b; Grulke et al., 1998a, 2001; Grulke and Balduman, 1999) by decreasing or 

inhibiting phloem loading of carbohydrates (Grantz and Farrar, 1999; Landolt et al., 1997), or of 

carbohydrate metabolism (Einig et al., 1997). This leads to depressed root growth (Andersen 

et al., 1991; Coleman et al., 1996; Grulke et al., 1998a) and the potential for plant communities 

to have an increased susceptibility to drought through altered root-shoot balance. Furthermore, it 

can negatively affect below-ground food webs (Phillips et al., 2002; Scagel and Andersen, 

1997). 

AX9-304
Another energetically costly response to O3 exposure is that the production of defense 

compounds, such as antioxidants, tend to increase under elevated O3 conditions (Sheng et al., 

1997; Tausz et al., 1999c, 2002). Antioxidants help the plant scavenge free radicals before they 

can cause damage to membranes or cell walls, but they demand C for production such that 

growth can be adversely affected. In mature Jeffrey pine, stomatal uptake of O3 elicited one 

complex of antioxidant defenses in mesic microsites, while endogenously generated free radicals 

in the chloroplast elicited a second complex of antioxidant defenses in xeric microsites (Grulke 

et al., 2003b). 

AX9.6.7.2 Material Flow 

Plants as producers are responsible for using inorganic atmospheric C and reducing it into 

organic forms used by consumers, thus driving nutrient processes in ecosystems. Ozone has the 

potential to disrupt material flow through organic C cycling and changes in nutrient cycling, 

particularly N and P cycling. Although there is indirect evidence that O3 is disrupting C and 

nutrient cycling at the ecosystem level, there is little direct evidence that O3 alters nutrient 

processing at ecosystem scales. 

The greatest annual nutrient and C input to ecosystems is from foliar and root turnover. 

Excision of plant parts and whole plant mortality are potentially much larger, but syncopated, 

ecosystem inputs. Ozone exposure alters C cycling in the ecosystem by affecting the 

within-plant C allocation and partitioning in dominant, O3-sensitive plants, and through chemical 

composition and rate of decomposition of sloughed plant parts (roots, branches, leaves) 

(Figure AX9-19). 

In addition to O3-induced changes in the quantity of C and nutrient inputs into ecosystems, 

O3 also can alter the nutrient quality of inputs. Ozone exposure alters nutrient levels in the 

foliage (Boerner and Rebbeck, 1995; Fenn and Poth, 1998; Lindroth et al., 2001; Momen et al., 

2002) and affects the C:N ratio (Andersen et al., 2001; Grulke et al., 2003b; Grulke and Lee, 

1997; Lindroth et al., 2001). Concentrations of compounds such as tannins, lignin, and 

phenolics (Baumgarten et al., 2000; Findlay et al., 1996; Kim et al., 1998; Saleem et al., 2001) 

are also affected by O3 exposure, which in turn alters decomposability (Fenn and Dunn, 1989) 

and litter buildup in the ecosystem. 

AX9-305
There are several possible pathways by which O3 may affect litter quality and, hence, litter 

decomposition, thus altering nutrient flow in ecosystems. These include altered C quality, 

altered nutrient quality, and alteration of leaf surface organisms important in decomposition 

pathways. For example, yellow poplar and black cherry litter exposed to O3 showed greater N 

loss during decomposition than charcoal-filtered controls, although mass loss did not vary 

among O3 treatments (Boerner and Rebbeck, 1995). Subsequent studies showed that although 

foliar N was not affected by O3 exposure in yellow poplar leaves, foliage decomposed more 

slowly (Scherzer et al., 1998). Other studies have also shown a change in foliar N concentration 

in response to O3 treatment, affecting the C:N ratio and possible litter quality (Andersen et al., 

2001). 

In some cases, it appears that N remobilization from foliage into the plant is not complete 

at the time of foliage abscission in O3-exposed plants (Findlay and Jones, 1990; Matyssek et al., 

1993; Patterson and Rundel, 1995; Stow et al., 1992). Greater N content of senesced litter could 

increase rates of decomposition. When O3-exposed cottonwood (Populus deltoides) leaves 

abscissed at the same time as control leaves, they decomposed at similar rates; however, 

prematurely senesced foliage from O3-exposed cottonwood decomposed more slowly than 

controls despite their higher N content (Findlay et al., 1991; Findlay and Jones, 1990). Higher N 

in senesced leaves appeared to be related to organic complexes formed by bound phenolics 

in O3-exposed leaves, making the litter less palatable to decomposers, and slowing 

decomposition rates (Findlay et al., 1996; Jones et al., 1994). Increased phenolics also have 

been found in European silver birch (Betula pendula) exposed to O3 (Saleem et al., 2001). 

Carbon quality in leaf litter also changes in O3-exposed foliage. Compositional changes in 

leaf structural characteristics, such as lignin content, would be expected to alter rates of litter 

decomposition (Fogel and Cromack, 1977; Kim et al., 1998; Meentemeyer, 1978). Blackberry 

litter exposed to elevated O3 had greater permanganate lignin than the control treatment, a 

difference that was inversely related to mass-loss rates in decomposition studies (Kim et al., 

1998). 

Ozone may affect early stages of decomposition by altering populations of leaf surface 

organisms before or after senescence. Magan et al. (1995) found a shift in phyllosphere fungi on 

Scots pine, Sitka spruce (Picea sitchensis), and Norway spruce exposed to O3, but the potential 

effect of these changes on subsequent litter decomposition was uncertain. The slowest 

AX9-306
decomposition rates of preexposed blackberry leaves were found when senesced foliage was 

exposed to O3 during decomposition, suggesting a possible direct effect of O3 on microorganisms 

in decomposing litter (Kim et al., 1998). Whether O3 concentrations at the soil surface influence 

initial stages of litter decomposition remains to be addressed. 

Ozone exposure also reduces nutritional content of foliage because of the degradation of 

chlorophyll. Resynthesis of chlorophyll may be limited by nutritionally poor soils or low soil 

moisture, as well as alteration of root uptake by O3 exposure and other stressors. Foliar exposure 

to O3 may also increase leaching of nutrients (Kerstiens and Lendzian, 1989). Ozone exposure 

promotes early senescence of foliage (Heath and Taylor, 1997; Miller and Elderman, 1977), with 

higher nutrient content than if excised later in the growing season (Poth and Fenn, 1998). 

Since O3 can slow decomposition through changes in leaf quality and quantity, leaf litter 

can accumulate (Fenn and Dunn, 1989). The accumulation of soil organic matter from increased 

leaf litter, even in the absence of acidic deposition, can lower soil pH (Binkley, 1992). Lower 

soil pH can promote loss of nutrients, particularly cations, from the system, further reducing 

nutrient availability to the plant. Complex organic compounds in decomposing litter may also 

tie up nutrients, rendering them less available to plants. 

Other factors such as nutrient deposition also affect the degree to which O3 influences C 

and nutrient flow through ecosystems. In high-pollution sites, the effect of N deposition is 

difficult to separate from O3 exposure, and reviews of the effect of acidic (and N) deposition on 

ecosystem nutrient dynamics are important to consider (see Binkley [1992]; Fenn et al. [1998, 

2003b]). For example, at moderately high pollution sites, foliar content of N is higher than that 

at lower pollution sites, but so are P, Mg, and Fe contents (Poth and Fenn, 1998). Although 

significant changes in foliar tissue chemistry have occurred in response to long-term pollutant 

deposition in the Carpathians (Fenn et al., 2002; Mankovska et al., 2004), much of this response 

is correlated to heavy metal and N and S deposition. At this point, the contribution of O3 

exposure alone cannot be isolated without careful between-site comparisons of plant response 

and understanding the depositional velocities of constituent atmospheric species. The significant 

effects of plant response along known pollution gradients are important to consider, because 

heavy metal and N and S deposition has significantly declined over the last decade, while O3 

exposures remain high and declines in forest health have been sustained. 

AX9-307
Models provide a means to track material flows through ecosystems. Two biogeochemical 

models were parameterized to capture long-term effects of O3 exposure, N deposition, and 

climate on a Ponderosa pine-dominated site in the eastern San Bernardino Mountains. 

Simulated O3 exposure resulted in faster production and turnover of foliage and a shift in C from 

the canopy (15% reduction) to the forest floor (increase of 50 to 60%) (Arbaugh et al., 1999). 

When O3 exposure was combined with that of N deposition, litter mass increased exponentially. 

The direct effect of O3 exposure on below-ground nutrient dynamics and ecosystem 

material flow is poorly understood. Additional research will be necessary to understand spatial 

and temporal dynamics of nutrient and C flow in ecosystems and to separate the effects of O3 

from those attributable to N deposition. 

AX9.6.8 Hydrological and Geomorphological 

At present, there are no publications on the effects of O3 exposure that are carried through 

at the ecosystem level to changes in mass water flow, channel morphology, riparian habitat 

complexity, or sediment movement. It is possible that processes occurring at smaller scales are 

affecting geomorphological processes in ecosystems; however, difficulty in scaling these 

responses spatially and temporally have made it difficult to show experimentally. It is possible 

that O3 exposure affects water quality through changes in energy and material flows, as 

discussed previously. 

AX9.6.9 Natural Disturbance Regimes 

There has been little research on how natural disturbances interact with O3 to affect 

performance of plants, communities, and ecosystems. The frequency, intensity, extent, and 

duration of natural disturbances are variable and unpredictable. However, there have been 

enough ecophysiological studies to suggest that O3 could predispose plant communities to 

certain natural stresses, e.g., drought stress or extreme low-temperature stress during the winter. 

While several studies have shown that drought stress reduces O3 uptake through stomatal 

closure, evidence also suggests that O3 can alter plant water use and susceptibility to drought. In 

controlled studies, Reich and Lassoie (1984) showed that relatively low O3 concentrations could 

diminish stomatal control and alter water use efficiency. Ash trees (Fraxinus excelsior) exposed 

to elevated O3 had greater water use early in the growing season, but less water use late in the 

AX9-308
growing season when exposed to elevated O3 (Wiltshire et al., 1994). Under moderate drought 

stress, Norway spruce trees grown under elevated O3 consumed water faster and showed higher 

stomatal conductances than controls (Karlsson et al., 1995). Pearson and Mansfield (1993) 

showed that successive O3 episodes disrupted stomatal function, making beech seedlings more 

susceptible to drought. Previous year O3 exposure was shown to have a carry-over effect in the 

following growing season for beech (Pearson and Mansfield, 1994). 

Few studies showing the effects of O3 on water relations of field-grown trees are found in 

the literature. However, Grulke et al. (2003a) examined the effects of O3 on canopy transpiration 

of Jeffrey pine from mesic and xeric microsites and found that trees from mesic sites had 20% 

more O3 uptake than those in the xeric sites. The authors also concluded that the mesic trees had 

greater O3 injury as evidenced by lower needle retention, whereas trees in xeric microsites had 

greater chlorotic mottle. Chlorotic mottle induced by stomatal uptake of O3 is indistinguishable 

from that of endogenously produced oxidants resulting from partially closed stomata, a reduction 

of CO2 inside the leaf, and production of strong oxidizers within the chloroplast when excited 

electrons are passed to O2 instead of CO2 under high light levels. 

Trees living near the limits of their freezing-tolerance range may be especially susceptible 

to predisposition of freezing injury by O3 (Sheppard et al., 1989). However, Aleppo pine 

exposed to elevated O3 had enhanced winter hardiness (Wellburn and Wellburn, 1994). As with 

the seasonal carryover of drought susceptibility, the influence of elevated O3 on freezing 

tolerance is carried over from summer to winter. Such effects have been demonstrated for Sitka 

spruce (Lucas et al., 1988) and for red spruce (Waite et al., 1994). Sorting out the role of 

elevated O3 in contributing to frost or low-temperature damage in forests remains difficult due to 

the presence of other factors that may affect senescence. 

AX9.6.10 Scaling to Ecosystem Levels 

The vast majority of literature describing O3 effects comes from short-duration herbaceous 

plant or tree seedling studies under controlled conditions. Scaling results from these studies 

requires extrapolation over both space and time in order to understand the full extent of changes 

in ecosystems. In addition to spatial, temporal, and age-related complexities, ecosystems are 

composed of organisms whose lifetimes range from hours to centuries (Laurence and Andersen, 

2003). Forested ecosystems are affected by environmental conditions such as water and nutrient 

AX9-309
availability, as well as by intra- and interspecies competition. Therefore, direct experimentation 

to determine the response of forested ecosystems is not simply a matter of determining the effect 

of O3 on individual mature trees. In addition, even if an experiment can be conducted, 

extrapolation of the results across landscapes and regions remains challenging. Nonetheless, 

models provide a means to explore possible long-term changes and to identify important 

research uncertainties. 

Approaches to scaling fall roughly into two categories: (1) process-based modeling to 

extrapolate physiological responses to O3 based on seedling studies and (2) field assessments 

using surveys and growth correlations, often in association with stand-level models to address 

ecosystem complexity. Comparatively good information is available on process level effects 

of O3 in seedlings and, therefore, some models offer the opportunity to use this information to 

scale O3 effects at the stand and regional scales (Chappelka and Samuelson, 1998; Fuhrer et al., 

1997; Hogsett et al., 1997; Laurence et al., 2000). Field assessments offer the opportunity to 

examine larger plot sizes, older trees, and trees growing under realistic competition. These two 

approaches are discussed in more detail below. 

AX9.6.10.1 Scaling from Seedlings to Mature Trees 

A number of investigators have used simulation models based on physiological processes 

to integrate available data and predict the effects of O3 on mature trees. Such models predict tree 

growth by simulating fundamental mechanisms, rather than through a statistical analysis of 

empirical data. For instance, the process of photosynthesis is simulated based on environmental 

conditions and physiological characteristics, and then the fixed C is allocated to plant growth 

using principles of plant physiology. Models based on mechanisms should be applicable across 

wide areas if the important functional relationships are represented accurately in the models and 

if the environmental conditions are accurately identified. The ability of six models (TREGRO, 

CARBON, ECOPHYS, PGSM, TREE-BGC, and W91) to simulate the effects of climatic 

change and O3 have been reviewed by Constable and Friend (2000). Of these models, only 

PGSM and TREGRO explicitly simulated the effects of O3 on foliar processes. 

The TREGRO model was used to simulate C allocation and tissue growth in seedlings and 

mature red oak trees based on the experimental data discussed above (Weinstein et al., 1998). 

For seedlings at 2×-ambient O3, only the total nonstructural carbohydrate (TNC) storage pool 

AX9-310
was predicted to be affected. For mature trees, large decreases were predicted for TNC, leaves, 

stem, branch, and both fine and coarse roots. Most predicted effects in mature trees were 

consistent with observations in the field, but the simulations overestimated the effect of 

2×-ambient O3 on root TNC and growth. The authors suggested that this discrepancy may have 

been due to trees reducing respiration in response to O3 stress, a response not simulated in the 

model. 

For Abies concolor, TREGRO was parameterized and simulated growth of a mature tree 

for 3 years to test for effects of O3 exposure and drought stress (Retzlaff et al., 2000). 

Reductions in O3 exposure-mediated carbon assimilation were translated to losses in whole tree 

biomass that probably would not be detectable in the field. However, TNC levels in branch 

tissue were simulated to be lowered by over 50%, and branch growth was reduced in a 

moderately polluted site relative to a clean site. Low O3 exposure (sufficient to decrease C 

assimilation by 2.5%) and drought stress (25% reduction in annual precipitation, which is 

common on a decadal scale) acted synergistically to reduce C gain of whole tree biomass of 

A. concolor. Simulated results of the tests were comparable to effects found in OTCs for 

seedlings and pole-sized trees in clean and moderately polluted sites. 

Models such as TREGRO are usually parameterized from many different sources of data, 

including chamber experiments and plantations, from seedlings to mature trees, making it 

difficult to validate that they reproduce changes that occur as trees develop from seedlings to 

maturity. To address this issue, physiological and growth data were collected from a natural 

stand of P. ponderosa and used to parameterize the TREGRO model (Grulke and Retzlaff, 

2001). Representative trees of each of five tree age classes were selected based on population 

means of morphological, physiological, and nearest neighbor attributes. Seedlings were 

observed to differ significantly from pole-sized and older trees in most physiological traits. The 

changes in biomass with tree age predicted from the model closely matched those of trees in the 

natural stand. 

The PGSM model was used to simulate Pinus ponderosa seedling growth responses to O3 

exposure and drought stress (Chen et al., 1994). Drought stress was predicted to reduce the 

effect of O3 on growth, as was observed in the experimental data. The TREGRO model was 

used to simulate responses of Pacific Coast and interior varieties of P. ponderosa to five 

simulated O3 exposures between subambient and 3×-ambient (Constable and Taylor, 1997). 

AX9-311
Simulated growth of var. ponderosa was reduced more than that in var. scopulorum with all O3 

exposures. Drought was protective of O3 exposure. Similar results were also found in a 

relatively moist Ponderosa pine plantation (Panek and Goldstein, 2001), whereas drought was 

synergistically deleterious with cumulative O3 exposure in a natural stand (Grulke et al., 2002b). 

For Pinus taeda, the Plant Growth Stress Model (PGSM) was calibrated with seedling data 

and then used to simulate the growth of mature trees over a 55 year period in the Duke Forest, 

NC, using estimates of historical O3 concentrations (Chen et al., 1998). Simulated stem diameter 

and tree height were comparable to observed values. In another simulation using TREGRO, 

loblolly pine was more sensitive (greatest reduction in C gain) to a peak O3 episode in July 

(Constable and Retzlaff, 1997), whereas mature yellow poplar (Liriodendron tulipifera) was 

more sensitive to a peak O3 episode in August. 

For Populus tremuloides, the ECOPHYS model was used to simulate the relative aboveground growth response of an O3-sensitive clone (259) exposed to square-wave variation in O3 

concentration (Martin et al., 2001). The model adequately simulated several effects of O3, 

including a greater effect on stem diameter than on stem height, earlier leaf abscission, and 

reduced stem and leaf dry matter production at the end of the growing season. For Acer 

saccharum, the TREGRO model was use to predict effects of a 10-year O3 exposure on root and 

stem growth of a simulated 160-year-old tree (Retzlaff et al., 1996). Twice-ambient O3 exposure 

(for Ithaca, NY) was predicted to deplete the TNC pools and reduce fine root production. 

AX9.6.10.2 Surveys, Growth Correlations, and Stand-Level Modeling 

Stand-level studies have included surveys of O3 symptoms, correlations of radial growth 

with O3 and other environmental factors, and regional-scale modeling. In addition, open air O3 

exposure systems, such as those being used on Populus tremuloides-mixed stands in northern 

Wisconsin (Karnosky et al., 2003b) and on Fagus sylvatica and Picea abies in Germany (Nunn 

et al., 2002) offer an opportunity to examine larger plot sizes, older trees, and trees growing 

under realistic competition. Plots along natural O3 gradients, as have been used very effectively 

in southern California forest studies (Miller and McBride, 1999b) and with P. tremuloides stands 

in the Great Lakes region (Karnosky et al., 1999), offer additional insights into ecosystem level 

responses. Undoubtedly, however, simulation modeling will have to become an integral 

component of research in order to predict adequately ecosystem responses to O3 (Laurence and 

AX9-312
Andersen, 2003). Results of these approaches are discussed below, organized into three U.S. 

regions: (1) northern states (including the upper Midwest and the Northeast), (2) southeastern 

states, and (3) western states (primarily California). A fourth section contains selected 

information from Europe. 

Northern and Midwestern United States. In recent years, the USDA Forest Service has 

conducted systematic O3 biomonitoring surveys in most north-central and northeastern states 

(Coulston et al., 2003; Smith et al., 2003). Plots are located on a systematic grid, and trained 

field crews evaluate up to 30 plants of up to six species that have foliar injury symptoms 

diagnostic of O3 damage. For the United States as a whole, injury has been found more often in 

eastern than in interior or west-coast states. As expected, O3 injury is more common and more 

severe in areas with higher O3 concentrations. Of sampled Prunus serotina plots, ~12% were 

estimated to be at high risk for injury based on a injury index derived from the survey data 

(Coulston et al., 2003). P. serotina was estimated to be at risk for injury on the Allegheny 

Plateau and the Allegheny Mountains (in Pennsylvania, West Virginia, and Maryland), as well as 

in the coastal plain of Maryland and Virginia. 

Ozone concentration, foliar injury, and physiological traits were measured on P. serotina 

trees of different sizes in Pennsylvania (Fredericksen et al., 1995). The proportion of foliage 

injured was 46% for seedlings, 15% for saplings, and 20% for canopy trees. Cumulative O3 flux 

was the most useful O3 metric for predicting injury. Injury was negatively correlated (r2

 = 0.82) 

with net photosynthetic rates, but was not related to stomatal conductance. P. serotina is 

discussed further below, because detailed surveys have been conducted in the Shenandoah and 

Great Smoky Mountains NPs in the southeastern United States. 

Over the past several decades, some surveys of white pine (Pinus strobus) have reported 

significant associations between foliar injury and reduced growth (Anderson et al., 1988; Benoit 

et al., 1982). However, a review of 93 surveys conducted from 1900 through the late 1980s 

concluded that methodological problems were pervasive, including such issues as proximity to 

roads, lack of peer review, lack of random sampling, small sample sizes, and lack of quantitative 

methods to estimate severity (Bennett et al., 1994). Because of these problems, along with 

evidence of adequate growth rates for P. strobus regionally and contradictory evidence from 

numerous studies of symptom production in response to controlled O3 exposure, these authors 

AX9-313
concluded that there was no clear evidence of decline in P. strobus (Bennett et al., 1994). 

A more recent study in Acadia National Park (ANP) in Maine found no association between O3 

exposure in OTCs and symptom development in P. strobus, calling into question whether 

symptoms previously ascribed to O3 may be caused by some other stress (Kohut et al., 2000). 

However, another ANP study found significant correlations between O3 exposure and the radial 

growth of trees during 10 years in 7 of 8 stands examined (102 trees total; (Bartholomay et al., 

1997). Taken together, these results suggest that there may not be an association between 

growth of P. strobus trees and putative O3 symptoms, but there may be an association between 

O3 exposure and radial growth of mature trees in the field. 

The study of O3 effects was undertaken from 1990 to 1993 in the ANP in Maine, because 

this location experiences elevated O3 exposures due to transport from urban areas located upwind 

(Kohut et al., 2000). Thirty-two species of plants found in the park were propagated and 

exposed to O3 in OTCs. In addition, ambient O3 concentrations were monitored at the study site 

at 15 m above sea level and near the top of Cadillac Mountain at 470 m above sea level. At the 

study site, the maximum 1-h O3 concentration was 140 ppb, which occurred in both 1990 and 

1991. Daytime 12-h O3 concentrations were 35, 41, 36, and 37 ppb during the four years; and O3 

concentrations were consistently higher at the high-elevation site. Species showing foliar injury 

at ambient O3 concentrations included black cherry (Prunus serotina), quaking aspen (Populus 

tremuloides), white ash (Fraxinus americana), jack pine (Pinus banksiana), big-leaf aster (Aster 

macrophyllus), and spreading dogbane (Apocynum androsaemifolum). Species showing foliar 

injury at 1.5 × ambient O3 concentrations included grey birch (Betula populifolia), small 

sundrops (Oenothera perennis), and bunchberry (Cornus canadensis). Species remaining 

uninjured at 2 × ambient O3 concentrations included paper birch (Betula papyrifera), white pine 

(Pinus strobus), pitch pine (Pinus rigida), red spruce (Picea rubens), Eastern white cedar (Thuja 

occidentalis), English oak (Quercus robur), Canada bluejoint grass (Calamagrostis canadensis), 

wild radish (Raphanus raphanistrum), and Canada mayflower (Maianthemum canadense). 

Because of their O3 sensitivity and diagnostic symptoms, big-leaf aster, spreading dogbane, 

quaking aspen, white ash, and black cherry were recommended as bioindicators for the ANP. 

The PnET-II model was applied to 64 locations across the northeastern United States to 

simulate the effects of ambient O3 on mature hardwood forests (Ollinger et al., 1997). In this 

model, O3 effects on each of several layers of the forest canopy were represented by a single 

AX9-314
linear equation relating predicted O3 uptake to decreased net photosynthetic rate. Wood growth 

was predicted to decrease between 3 to 22%, with greatest reductions in southern portions of the 

region where O3 levels were highest and on soils with high water-holding capacity where 

drought stress was absent. Little variation was predicted among years, because high O3 often 

coincided with hot, dry weather conditions that reduced predicted stomatal conductance and O3 

uptake. 

In order to estimate the impact of O3 on forests, effects must be evaluated not only on 

individuals, but also on mixtures of species and the composition of forest stands. The PnET 

model described above evaluated the effects of O3 on broad forest types (an evergreen/deciduous 

mix), but did not address specific forest species composition. In order to address competition 

among species, the TREGRO model was linked to the ZELIG forest stand growth model and a 

geographic information system was created to predict the effects of O3 across the north-central 

and northeastern United States (Laurence et al., 2000). ZELIG is a gap-succession model used 

to simulate succession in mixed stands typical of eastern and northern forests. Ambient O3 

generally caused a reduction of 2 to 4% in the growth of Quercus rubra across the region during 

the 100-year simulation. The response followed the pattern of O3 exposure, with little effect in 

the northwest part of the region, but with greater effect in southern locations. The O3 response of 

Acer saccharum to O3 varied widely, but the overall growth response was always positive, 

indicating that the evergreen/deciduous mix was able to take advantage of the decrease in the 

growth of Q. robur and other species caused by O3. In the northernmost part of the region, A. 

saccharum growth increased by up to 3%, but in the southern part of the region, its growth 

increased by up to 12%. The authors ascribed this enhanced growth to a combination of warmer 

temperatures and reduction in the growth of Prunus serotina, a minor component of the 

simulated stand that was very sensitive to O3. 

Southeast United States. In a survey of the Great Smoky Mountains NP, foliar injury 

attributed to O3 was found on 47% of the more than 1,600 plants examined (Chappelka et al., 

1997). In subsequent surveys of injury in the park, injury was found on mature trees of the 

following species: Sassafras albidum, Prunus serotina, and Liriodendron tulipfera (Chappelka 

et al., 1999a,b). In a similar study in Shenandoah National Park, injury was found on Fraxinus 

americana, P. serotina, and L. tulipfera (Hildebrand et al., 1996). 

AX9-315
For Prunus serotina seedlings grown in soil in OTCs and exposed to relatively low 

ambient levels of O3 in Pennsylvania, there was no correspondence between visible foliar stipple, 

leaf gas exchange, and seedling growth between two families previously shown to differ in O3 

symptoms (Kouterick et al., 2000). However, significant exposure-response relationships were 

found for foliar injury in the Great Smoky Mountains and Shenandoah National Parks. In each 

park, foliar injury was evaluated on mature P. serotina trees on three plots at different elevations 

near O3 monitors during 1991 to 1993 (Chappelka et al., 1999a,b; Hildebrand et al., 1996). In 

1991, incidence was 60% and 45% for the two parks and 33% in both parks during 1992 and 

1993. Symptoms were greater at the highest elevations where O3 concentrations were highest. 

In another study, radial growth rates were measured in 44 P. serotina trees ranging in age from 

19 to 56 years old with and without O3 symptoms, at three sites in the Great Smoky Mountains 

National Park. Trees with O3 symptoms were compared to similar-sized trees with few 

symptoms. There was no evidence that trees with O3 symptoms had lower growth rates (p = 0.6) 

(Somers et al., 1998). 

In the Great Smoky Mountains National Park, radial growth rates were measured for 44 L. 

tulipfera trees ranging in age from 30 to 58 years old, with and without O3 symptoms, at three 

sites at different elevations (Somers et al., 1998). Trees with O3 symptoms averaged 30% lower 

growth rates over ten years (p = 0.0005). Seedlings of Liriodendron tulipfera were exposed for 

two seasons to 2×-ambient O3 exposures in OTCs in Delaware, OH (seasonal SUM00 exposures 

of 107 and 197 ppmh) (Rebbeck, 1996). Foliar O3 symptoms were observed, but growth was not 

reduced. 

In order to evaluate the influence of interspecies competition on O3 effects, the linked 

TREGRO and ZELIG modeling system was used to predict the effects of O3 over 100 years on 

the basal area of species in a Liriodendron tulipfera-dominated forest in the Great Smoky 

Mountains NP (Weinstein et al., 2001). Ambient O3 was predicted to reduce the basal area of L. 

tulipfera by 10%, whereas a 1.5×-ambient exposure was predicted to cause a 30% reduction. 

Basal area of Acer rubrum and Prunus serotina was predicted to increase for some years, but 

then decrease by up to 30%, with few changes in the total basal area of all species by the end of 

the simulation. 

In order to evaluate the influence of interspecies competition on O3 effects, the linked 

TREGRO and ZELIG modeling system was used to predict the effects of O3 on the basal area of 

AX9-316
Pinus taeda and Liriodendron tulipfera growth throughout their ranges (Laurence et al., 2003). 

The models were parameterized using biological and meteorological data from three sites in the 

southeastern United States (in Alabama, Louisiana, and North Carolina). Forest stand response 

to five O3-exposure regimes with annual SUM06 values ranging from 0 to 100 ppmh per year 

was simulated for 100 years. The simulated basal area of the two species was generated within 

the context of four other tree species common in southeastern forests. Basal area of P. taeda was 

highly responsive to precipitation and O3 exposure, with the greatest increases under highprecipitation, low O3-exposure scenarios and the greatest decreases under low-precipitation, high 

O3-exposure scenarios. The basal area of L. tulipifera did not significantly differ (+10%) from 

simulations using a “base case” (ambient O3, average precipitation). 

Systematic biomonitoring surveys found that approximately 24% of sampled sweetgum 

and 15% of sampled Pinus taeda plots were estimated to be at high risk for foliar injury on the 

coastal plain of Maryland and Virginia (Coulston et al., 2003; Smith et al., 2003). In a study in 

Tennessee, the effect of ambient (uncontrolled) O3 on 28 mature canopy-dominant 50 to 

90-year-old P. taeda trees in five stands was measured over a 5-year period (McLaughlin and 

Downing, 1995, 1996). Of many O3 metrics, a 3-day average of hourly O3 values $ 40 ppb 

(AOT40) was found to best explain short-term variation in stem expansion as measured with 

dendrometer bands. Interactions between O3, temperature, and drought stress (as indicated by 

the weekly moisture stress index) accounted for 63% of the short-term variation in stem growth 

rates. Because there are interactions among O3, drought stress, and temperature that may differ 

with the averaging time (days to years), this type of study cannot provide conclusive proof of 

cause and effect (Reams et al., 1995). However, the results do suggest that the effects of O3 

measured on loblolly seedlings may also be occurring in mature trees in both wet and dry sites. 

The magnitude of effects of O3 on growth, including interactions with other variables in this 

study, ranged from 0 to 15% over 5 years, with an average of 5.5%. 

Western United States. The USDA Forest Service conducted O3 biomonitoring surveys in 

Washington, Oregon, and California during one year (1998), and in the Sierra Nevada and 

Sequoia NFs every other year for several decades (Campbell et al., 2000). Overall, only one plot 

showed any symptoms of O3 injury outside of the Sierra Nevada and Sequoia National Forests. 

In the Sierra Nevada National Forest, between 30 and 40% of trees showed injury from 1989 

AX9-317
through 1997. In the Sequoia National Forest, between 40 and 50% of the trees surveyed 

showed injury from 1990 through 1998. 

For Pinus ponderosa along a well-studied gradient of O3 exposure in the San Bernardino 

mountains, chlorotic mottle was highest on foliage at the most polluted site, as has been found 

previously (Grulke and Balduman, 1999). Based on whole-tree harvests, root biomass was 

lowest at the most polluted sites, confirming previous studies with seedlings under controlled 

conditions, as discussed above. Ozone responses in highly polluted environments such as 

Southern California may not be predicted adequately by extrapolating effects from single-factor 

experiments. Instead, combined approaches utilizing field experiments and modeling efforts 

may be required to properly account for a combination of stressors including O3, N deposition, 

and drought. Furthermore, the available studies underscore the lack of correlation between O3 

symptoms and mature tree effects. If a field survey fails to find a correlation between mature 

tree growth and O3, this result may be due to the dominant effect of another factor such as N 

deposition and may not be evidence that O3 does not reduce the growth of mature trees. 

The TREGRO model was used to evaluate how projected future temperature and CO2 

concentrations might affect the response of individual Ponderosa pine to O3 at seven sites in 

California, Oregon, and Washington (Tingey et al., 2001). As expected, growth decreased with 

increasing O3 exposure. Differences in O3 response among sites appeared to be due primarily to 

differences in precipitation. 

Often air quality standards do not translate directly into measurable improvements in tree 

growth or productivity. To evaluate whether past improvements in air quality have improved 

Ponderosa pine growth, TREGRO was used to simulate growth at sites in the San Bernardino 

Mountains in California (Tingey et al., 2004). Ozone and meteorological data from the past 

37 years was used to run the simulations. Despite variation in precipitation and temperature, O3 

was found to reduce simulated tree growth. The authors were able to simulate growth 

improvements as air quality improved during the 1980s and 1990s, suggesting that 

improvements in emission control strategies benefitted Ponderosa pine. The model simulations 

were qualitatively consistent with improvements in canopy condition that were observed at sites 

where O3 reductions were the greatest. 

AX9-318
Studies in Europe. In a 4-year study of Fagus sylvatica in Switzerland at 57 forest sites 

ranging in age from 65 to 173 years, stem increment was found to decrease with increasing 

maximum O3 exposure (Braun et al., 1999). In this study, O3 concentration was estimated by 

interpolation among monitoring stations, and other site conditions such as soil water status and 

temperature were interpolated from weather stations. Other factors such as N deposition, tree 

diameter, and canopy dominance were also found to be significantly associated with stem 

increment. The maximum annual O3 dose (expressed as AOT40) was found to be more strongly 

associated with decreased stem increment than was the average O3 dose over the 4 years. 

A growth reduction of 22.5% (confidence interval 14.3 to 28.6%) was associated with each 

10 ppmh increment of O3 (expressed as AOT40). This decrease was steeper than the 6.1% 

growth reduction summarized previously from several OTC studies with F. sylvatica seedlings 

(Fuhrer et al., 1997). However, the authors suggest that this difference may be explained largely 

by the 4 years of exposure in their forest survey study as compared to the 1-year exposures for 

seedlings. As with any forest survey, these results must be interpreted with caution because O3 

exposure was correlated with other variables, such as tree age and the deposition of NO2 

and SO2. 

AX9.6.11 Summary of Ecological Effects of Ozone Exposure on 

Natural Ecosystems 

In this chapter, an effort has been made to discuss the adverse effects of O3 on natural 

ecosystems within the context of the SAB framework for assessing and reporting ecological 

conditions (Young and Sanzone, 2002) (Figure AX9-21). Using this framework, there is 

evidence that tropospheric O3 is an important stressor of natural ecosystems, with 

well-documented impacts on the biotic condition, ecological processes, and chemical/physical 

nature of natural ecosystems. In turn, the effects of O3 on individual plants and processes are 

scaled up through the ecosystem affecting processes such as energy and material flow, intra- and 

interspecies competition, and NPP. Thus, effects on individual keystone species and their 

associated microflora and fauna may cascade through the ecosystem to the landscape level. This 

suggests that by affecting water balance, cold hardiness, tolerance to wind, and by predisposing 

plants to insect and disease pests, O3 may even influence the occurrence and impact of natural 

disturbance. Despite the probable occurrence of such effects, however, there are essentially no 

AX9-319
Figure AX9-21. Common anthropogenic stressors and the essential ecological attributes 

they affect. 

Source: Modified from Young and Sanzone (2002). 

instances where highly integrated ecosystem-level studies have conclusively shown that O3 is 

indeed altering ecosystem structure and/or function. 

Systematic injury surveys demonstrate that foliar injury occurs on O3-sensitive species in 

many regions of the United States. However, the frequent lack of correspondence between foliar 

symptoms and growth effects means that other methods must be used to estimate the regional 

AX9-320
effects of O3 on tree growth rates. Investigations of the radial growth of mature trees, in 

combination with data from many controlled studies with seedlings, as well as a few studies with 

mature trees suggest that ambient O3 is reducing the growth of mature trees in some locations. 

Studies using models based on tree physiology and forest stand dynamics suggest that modest 

effects of O3 on growth may accumulate over time and may interact with other stresses. For 

mixed-species stands, such models predict that overall stand growth rate is generally not likely to 

be affected. However, competitive interactions among species may change as a result of growth 

reductions of sensitive species. These results suggest that O3 exposure over decades may be 

altering the species composition of forests in some regions. 

RESEARCH NEEDS 

The knowledge base for examining the range of ecological effects of O3 on natural 

ecosystems is growing, but significant uncertainties remain regarding O3 effects at the ecosystem 

level. For example, there is a need for information on the following ecosystem-level responses: 

• Ecosystem Processes. Little is known about the effects of O3 on water, C, and nutrient 

cycling, particularly at the stand and community levels. Effects on belowground 

ecosystem processes in response to O3 exposure alone and in combination with other 

stressors are critical to projections at the watershed and landscape scales. Little is yet 

known about the effects of O3 on structural or functional components of soil food webs, or 

how these impacts could affect plant species diversity (Andersen, 2003). 

• Biodiversity and Genetic Diversity. The study of genetic aspects of O3 impacts on natural 

ecosystems has been largely correlational in nature and it remains to be shown 

conclusively whether O3 affects biodiversity or genetic diversity (Davison and Barnes, 

1998; Pitelka, 1988; Winner et al., 1991). Studies of competitive interactions under 

elevated O3 levels are needed (Laurence and Andersen, 2003), and reexamination via new 

sampling of population studies to bring in a time component to previous studies showing 

spatial variability in population responses to O3 are needed. These studies could be 

strengthened by modern molecular methods to quantify impacts on diversity. 

• Natural Ecosystem Interactions with the Atmosphere. Little is known about feedbacks 

between O3 and climate change on volatile organic compound (VOC) production, which in 

turn, could affect O3 production (Fuentes et al., 2001). At moderate-to-high O3 exposure 

sites, aberrations in stomatal behavior could significantly affect individual tree water 

balance of sensitive trees, and if the sensitive tree species is dominant, hydrologic balance 

at the watershed and landscape level could be affected. This has not been addressed in any 

model because O3 exposure effects, if included in the modeling effort have assumed a 

linear relationship between assimilation and stomatal conductance. 

AX9-321
• Below-Ground Interactions. While the negative effects of O3 on below ground growth are 

well characterized, interactions of roots with the soil or microorganisms are not. 

• Other Interactive Effects. Interaction studies with other components of global change 

(e.g.,warming, increasing atmospheric CO2, N deposition, etc.) or with various biotic 

stressors are needed to better predict complex interactions likely in the future (Laurence 

and Andersen, 2003). Whether O3 will negate the positive effects of an elevated CO2 

environment on plant carbon and water balances is not yet known; nor is it known if these 

effects will scale up through the ecosystem. How might O3 affect the progress of pest 

epidemics and insect outbreaks as concentrations increase is unclear (Ball et al., 1998). 

• Reproduction Effects. Information concerning the impact of O3 on reproductive processes 

and reproductive development under realistic field or forest conditions are needed as well 

as examination of reproductive effects under interacting pollutants (Black et al., 2000). 

• Comparative Extrapolation. The vast majority of O3 studies of trees have been conducted 

with young, immature trees and in trees that have not yet formed a closed canopy. 

Questions remain as to the comparability of O3 effects on juvenile and mature trees and on 

trees grown in the open versus those in a closed forest canopy in a competitive 

environment (Chappelka and Samuelson, 1998; Kolb and Matyssek, 2001; Samuelson and 

Kelly, 2001). 

• Scaling-Up Issues. Scaling the effects of O3 from the responses of single or a few plants 

to effects on communities and ecosystems is a complicated matter that will require a 

combination of manipulative experiments with model ecosystems, community and 

ecosystem studies along natural O3 gradients, and extensive modeling efforts to project 

landscape-level, regional, national and international impacts of O3. Linking these various 

studies via impacts on common research quantification across various scales using 

measures of such factors as leaf area index or spectral reflective data, which could 

eventually be remotely sensed (Kraft et al., 1996; Panek et al., 2003), would provide 

powerful new tools for ecologists. 

• Comparative Risk Assessment Methodologies. Methodologies to determine the important 

values of services and benefits derived from natural ecosystems such that these could be 

used in comprehensive risk assessment for O3 effects on natural ecosystems (Heck et al., 

1998). 

AX9-322
AX9.7 ECONOMIC EVALUATION OF OZONE EFFECTS ON 

AGRICULTURE, FORESTRY, AND NATURAL ECOSYSTEMS 

AX9.7.1 Introduction 

The adverse consequences of ambient air pollutant exposures on vegetation, ecosystems, 

and components of the material environment have been documented since the beginning of the 

industrial revolution. Attempts to quantify the monetary damage and injury resulting from 

tropospheric O3 exposures to managed agriculture, forests, and natural ecosystems date back at 

least to the 1950s. 

Both methodological and data problems plagued early efforts to assess the monetary 

damages of air pollution to crops and natural vegetation. Adams and Crocker (1989) discussed 

the methodological issues, e.g., a lack of reliable data on effects from air pollutants on crop 

yields or the failure to develop and apply appropriate economic models. Some of these problems 

were remedied by the EPA’s National Crop Loss Assessment Network (NCLAN) in the 1980s. 

The EPA’s NCLAN facilitated the performance of economic assessments by providing O3-crop 

yield data with which to estimate O3-crop yield response functions (see Heagle [1988] for a 

review of NCLAN procedures and findings). NCLAN also funded a series of economic 

assessments that, along with subsequent economic assessments, documented substantial 

economic damages to agriculture. (See Spash [1997] for a detailed review of economic 

assessments, many of which used NCLAN data.) 

Since the completion of the NCLAN program in the late 1980s, the number of economic 

assessments of air pollution studies focusing on terrestrial ecosystems in general, and agriculture 

in particular, has declined. For example, for the period of 1980 to 1990, 33 economic studies 

of O3 and other air pollutant effects on U.S. crops were published in peer-reviewed journal 

outlets (Spash, 1997). However, in preparing this section of the current O3 AQCD, only four 

peer-reviewed economic assessments were found for the decade of 1991 to 2000 that addressed 

vegetation in the United States. In addition, one peer-reviewed article (Kuik et al., 2000) was 

found dealing with agriculture in the Netherlands. Recent interest in global climate change, and 

the potential effects of global warming on O3 and other photochemical oxidants, has renewed 

interest in the effects of air pollution on both managed and unmanaged terrestrial ecosystems 

(Adams et al., 1998). In addition, concern is growing regarding the effects of air pollutants on 

natural ecosystems and on the services they provide (Daily, 1997). Unfortunately, this interest 

AX9-323
has not yet translated into additional peer-reviewed publications addressing O3 or other air 

pollutant effects on ecosystems. 

This section of the current O3 criteria document first discusses the availability of economic 

information and its usefulness in forming environmental policy. Next, economic assessments of 

air pollution effects and findings from the 1996 AQCD (U.S. Environmental Protection Agency, 

1996) are discussed, followed by a synthesis of the limited literature available since the 1996 

AQCD with respect to O3 effects on agriculture, forestry, and ecosystems. Finally, limitations 

and continuing uncertainties are reviewed. The most fundamental of these is the lack of 

measurements of the economic effects of air pollution on natural ecosystems. Other issues 

include the variability of performance in both managed and natural ecosystems under increased 

climatic and air pollution variability as well as the challenges related to spatial and temporal 

scales used in performing economic assessments. To date, this set of effects has been sparsely 

addressed. 

AX9.7.2 The Measurement of Economic Information 

Economic science is an exercise in deductive logic in which testable hypotheses about the 

behavior of economic agents (i.e., farmers, consumers, resource owners) and markets are 

deduced from a body of theory. That body of theory is based on a series of premises proposed 

by economists and philosophers dating back over two centuries to Adam Smith. These premises 

gradually evolved into a theoretical foundation primarily dealing with microeconomics and 

culminating in structural relationships that define the operation of markets. This foundation was 

first laid out in a comprehensive and rigorous fashion by Alfred Marshall in 1920. Samuelson 

(1948) formalized these theoretical relationships, resulting in what is sometimes referred to as 

modern, or neoclassical, economics. 

The insights gained from the theoretical foundation of economics helped shape the nature 

of applied economic, or policy-relevant research. An example of such applied research is when 

economists seek to measure the economic consequences of air pollution on agriculture. Such an 

application is described in Adams and Horst (2003), who provided a graphical representation of 

the measurement of the effects of air pollution on the well-being of producers and consumers. 

Economic theory is applied to real-world problems when the methods of economics and the need 

AX9-324
for data from other disciplines come into play. When estimating the economic effects of an 

environmental change, economists need an economic model or method that is theoretically 

consistent, i.e, defensible, as well as data with which to estimate both environmental science and 

economic relationships for use in the model. Among accepted economic assessment methods, 

the actual choice is frequently determined by the nature of the problem to be addressed. 

It should be noted that the choice of assessment method can affect the type of economic 

information that is obtained. Even with a given assessment method, results are sensitive to 

specific data treatment or assumptions (Adams, 1999). For example, some methods only 

measure effects on a particular group e.g., farmers. Other methods may measure effects across 

several groups. Thus, one should not expect the magnitudes of damage or benefits to be 

identical across economic assessments. One should, however, expect that the direction of the 

effects will be similar. 

Once it is established that an assessment meets basic economic criteria, e.g., including 

human behavioral responses, the selection of the specific economic assessment method is often a 

relatively minor issue in terms of estimating benefits of air pollution control (or disadvantages of 

increases in air pollution). Although results differ across approaches, the differences are largely 

attributable to specific features of the assessment (e.g., whether the natural science data include a 

particular effect or relationship, whether effects on consumers are measured, and so forth). The 

nature and quality of the air quality forecasts used in the assessments can greatly influence the 

sensitivity of the assessments (Adams and Crocker, 1989). This is particularly noticeable when 

dealing with forecasts of seasonal air pollution changes (Adams et al., 1988). From the 

standpoint of providing policy guidance, the differences in economic estimates attributable to the 

assessment methods are often swamped by uncertainty in the natural and physical science 

forecasts. This has also been noted in recent economic assessments of climate change (Adams 

et al., 1998). In many settings, the quality of economic assessments of air pollution is likely to 

be improved more by refining the physical and natural science data used in the assessments than 

by intensive efforts to fine-tune the assessment techniques (Adams, 1999). 

AX9-325
AX9.7.3 Understanding of Air Pollutants Effects on the Economic Valuation 

of Agriculture and Other Vegetation in the 1996 Criteria Document 

Evidence from the plant science literature cited in the 1996 O3 AQCD (U.S. Environmental 

Protection Agency, 1996) is unambiguous with respect to the adverse effects of tropospheric O3 

on some types of vegetation. For example, the 1996 AQCD noted that findings from the EPA 

multiyear NCLAN program in the 1980s provided rigorous corroboration of at least two decades 

of previous research and a century of anecdotal observations which showed that O3 at ambient 

levels caused physical damage to plants in general and to important crop species in particular. 

Specifically, NCLAN established that ambient O3 levels resulted in statistically significant 

reductions in yields for some crop species (Heagle et al., 1988). The 1996 AQCD also assessed 

the results of studies regarding O3 effects on crops, forests, and natural vegetation in more detail. 

More recent reviews, such as the comprehensive survey of the economic literature on 

agricultural effects by Spash (1997), corroborate the synthesis of results reported in the 1996 

AQCD. 

The number and quality of assessments of the economic consequences of O3 exposures on 

vegetation reported in the 1996 AQCD are primarily a function of the state of evidence obtained 

from scientific studies in each vegetation category. For example, the plant science evidence 

reviewed in the 1996 AQCD concerning effects of O3 exposures on agricultural crops was 

reported to be more valid than for individual tree species or plant communities (ecosystems). 

As a result, most economic assessments discussed in the 1996 AQCD focused on the data 

obtained from studies of agricultural crops. The economic literature dealing with O3 effects on 

forest productivity in the 1996 AQCD was sparse. The few economic assessments of tree or 

forest effects reported in the 1996 AQCD were confined to evaluations of assumed or 

hypothetical changes in output, such as board feet of lumber (e.g., Haynes and Adams, 1992). 

As noted in the 1996 AQCD, O3 effects on ecosystems and their services had not been measured 

in any systematic fashion and no peer-reviewed economic assessments were yet reported. 

This section first briefly reviews economic assessments drawn from the review in the 1996 

criteria document. This review is the benchmark against which recent articles are then discussed 

in the subsequent section. As was the case in 1996, the discussion of economic valuation of 

ecosystem effects is generally limited to conceptual and methodological issues, given the 

continued lack of empirical analyses in this category. 

AX9-326
AX9.7.3.1 Agriculture 

In view of the importance of U.S. agriculture for both domestic and world consumption of 

food and fiber, reductions in U.S. 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 O3 exposures, or the benefits of O3 control, to agriculture. Twenty-three 

assessments of the economic effects of O3 exposures on agriculture were reviewed in the 1996 

AQCD, highlighting key issues in the validity of these assessments (U.S. Environmental 

Protection Agency, 1996). First, the evidence should reflect how crop yields will respond under 

actual field conditions to O3 exposures. Second, the air quality data used to frame current or 

hypothetical effects of O3 on crops should represent actual exposures sustained by crops at 

individual sites or production areas. 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 O3 air quality; 

(2) accurately reflect institutional considerations, such as regulatory programs and income 

support policies (e.g., provisions of federal Farm Bill legislation), that may result in market 

distortions; and (3) use measures of well-being that are consistent with economic principles. 

Assessments of O3 damage to agricultural crops reported in the 1996 AQCD employed 

procedures for calculating economic losses that met the conditions described above. More 

specifically, the assessments provided 23 quantitative estimates of the economic consequences 

of exposures to O3 and other air pollutants to agriculture that reflect producer-consumer 

decision-making processes, associated market adjustments, and some measure of distributional 

consequences between affected parties. Many of the economic assessments reviewed in 

previous O3 documents also focused on O3 effects in specific regions, primarily California and 

the Corn Belt (e.g., Garcia et al., 1986). This regional emphasis in the earlier literature may be 

attributed to the relative abundance of data on crop response and air quality for selected U.S. 

regions, as well as the importance of some agricultural regions (such as California) for the U.S. 

agricultural economy. 

Two U.S. national studies described in previous O3 criteria documents that are worthy of 

additional comment are Kopp et al. (1985) and Adams et al. (1986). These were judged to be 

“adequate” in terms of the three critical areas of data inputs in the 1996 AQCD. Together, it was 

argued, they provide a reasonably comprehensive estimate of the economic consequences of 

AX9-327
changes in ambient air O3 levels on agriculture. Because of their central role in the 1996 criteria 

document, these two studies are reviewed briefly below. 

The Kopp et al. (1985) and Adams et al. (1986) studies indicated that ambient levels of O3 

were imposing substantial economic costs of ~$3.4 billion (in 2000 U.S. dollars) on agriculture. 

Both were judged to suffer from several sources of uncertainty, but the document concluded that 

these possible improvements in future assessments were not likely to greatly alter the range of 

agricultural benefit estimates arising from O3 reductions for several reasons. First, the studies 

covered about 75 to 80 % of U.S. agricultural crops (by value). For inclusion of the other 20% 

to significantly change the estimates would require that their sensitivities to O3 be much greater 

than for the crops that have been included to date. Second, model-sensitivity analyses reported 

in past studies indicate that changes in plant exposure-response relationships must be substantial 

to translate into major changes in economic benefits estimates. For example, it was assumed 

unlikely that use of different exposure measures, or inclusion of interaction effects, would 

greatly alter the magnitude of the economic estimates. Third, it was believed that countervailing 

effects would mitigate against large swings in the estimates, e.g., longer O3-exposure periods 

may predict greater yield losses, but O3-water stress interactions tend to reduce the 

yield estimates. 

Other national assessments reported in the 1996 AQCD provided general corroboration of 

the results of Kopp et al. (1985) and Adams et al. (1986). An evaluation of these studies, in 

terms of the adequacy of information from plant exposure studies and aerometric and economic 

data, was presented in the 1996 AQCD, along with estimates of benefits or damages associated 

with changes in O3. Most of the studies added onto either Kopp et al. (1985) or Adams et al. 

(1986). A relevant question was whether subsequent studies provided any “surprises” in terms 

of magnitude of economic effects. 

Common themes or findings from these and earlier O3 and other air pollution studies were 

summarized in two synthesis papers, those of Adams and Crocker (1989) and Segerson (1991). 

The major conclusion is that the agricultural effects of tropospheric O3 at ambient levels impose 

economic costs to society or, conversely, that reductions in ambient O3 should result in societal 

benefits. 

AX9-328
Several studies contained in the 1996 AQCD still are of interest. For example, one finding 

pertains to the relationship between federal farm programs and air pollution regulations 

(McGartland, 1987). In each case, the inclusion of farm programs in the economic models 

resulted in modest reductions in the economic benefits of O3 control due to increased farm 

program costs. As Segerson (1987) noted, however, it is not clear that these increased costs 

should be charged against the potential benefits of an O3-regulatory standard but, rather, 

considered as an additional cost associated with the inefficiencies of the farm program. It should 

also be noted that the nature of federal farm programs was changed dramatically by Congress in 

1996 in an attempt to reduce the federal government’s role in agriculture. Although more recent 

federal legislation, such as the 2002 Farm Bill (U.S. Congress, 2002), appears to be restoring the 

federal government’s role in the farming sector, this issue currently is not as important as 

suggested by earlier studies, due to the declining reliance on deficiency payments to farmers, 

which tend to distort resource allocation. 

Another national study (Adams et al., 1988) analyzed economic benefits under a regulatory 

alternative involving a seasonal (crop growing season) O3-exposure index measured as a 12-h 

mean, instead of hourly levels or percent changes from ambient as reported in earlier studies. 

Specifically, a seasonal average of 50 ppb O3 (measured as a 12-h seasonal mean), with a 

95% compliance level, was reported in Adams et al. (1988). The result (a $2.9 billion benefit in 

2000 dollars) is similar to the assumed 25% reduction across all regions reported by Adams et al. 

(1986). At least one study also combined environmental stressors (e.g., O3, UV-B radiation) in 

performing economic assessments. Adams and Rowe (1990), using the same model as Adams 

et al. (1986), reported that a 15% depletion of stratospheric O3 (resulting in a 13% increase in 

tropospheric O3) would cause an economic loss of ~$1.4 billion in 2000 dollars attributed to the 

tropospheric O3 increase. Reducing VOCs/NOx motor vehicle emissions by 10% would result in 

a benefit of ~$0.3 billion, while a complete elimination of motor vehicle emissions would yield a 

benefit of ~$3.4 billion (1990 dollars). The range of these numbers is consistent with values 

reported in Adams et al. (1986), Kopp et al. (1985), and other national-level analyses, i.e., 

estimates of from $1.0 to 2.0 billion for reductions in ambient O3 of 25 to 40%. 

AX9-329
AX9.7.3.2 Forests (Tree Species) and Natural Ecosystems 

The long-term nature of air pollution effects on perennial species such as trees creates 

challenges to plant scientists in attempts to sort out the effects of specific individual stressors 

such as O3 from among the many other potential causal factors (Skelly, 1988). It also creates 

problems in terms of measuring the impacts on direct economic value of goods, such as 

reductions in board-feet of lumber produced per unit of time. 

Most of the literature in the 1996 AQCD dealing with forest species reported the effects 

of O3 exposures in terms of foliar injury (Davis and Skelly, 1992; Freer-Smith and Taylor, 1992; 

Simini et al., 1992; Taylor and Hanson, 1992). This emphasis on foliar effects in the forest 

effects literature (rather than marketable yield) is similar to the state of science for agricultural 

crops prior to 1980. More recent studies address the effects of air pollutants on forest tree 

species diversity (Bringmark and Bringmark, 1995; Vacek et al., 1999; Weiner et al., 1997). 

However, such information is of limited use in economic assessments. The exception is in 

measuring the economic value of aesthetic changes in a forest stock, where changes in species 

composition may affect recreational values (Crocker, 1985). 

The data concerning changes in marketed output, such as board-feet of lumber or changes 

in growth rates in managed forests or effects on the growth of almond, peach, apricot, pear and 

plum trees in orchards, cited in the 1996 document (U.S. Environmental Protection Agency, 

1996) have not been quantified. In addition, the economic impact of reductions in growth of 

seedlings of evergreen trees, e.g., slash pine, presented in the same document have not been 

valued. The few studies which attempted to measure economic losses arising from exposures 

to O3 or other pollutants circumvented the lack of plant science data by assuming (often 

arbitrary) reductions in forest species growth or harvest rates (Adams, 1986; Callaway et al., 

1986; Crocker and Forster, 1986; Haynes and Adams, 1992). Although the economic estimates 

reported in the 1996 AQCD are comparable to those reported for agriculture (e.g., $2.6 billion 

for eastern Canada forests, $2.9 billion for eastern U.S. forests in 2000 dollars), the lack of yield 

and/or growth effects data makes these studies only suggestive at best, of the economic 

consequences of forest effects directly attributable to O3 exposures questionable. Recent 

developments in forestry economic modeling capabilities, in support of climate change research, 

have enhanced the ability to measure the effects of environmental stressors on this sector 

(Adams et al., 1996; McCarl et al., 1998). However, these models need data on changes in either 

AX9-330
timber production or growth rates, both of which are lacking for forest species under alternative 

O3 levels. 

AX9.7.4 Studies Since 1996 of Ozone Exposure Effects on the Economic 

Value of Agriculture, Forests, and Ecosystems 

Of the few current (post-1996) economic studies addressing agricultural effects, none offer 

new insights of value in determining the economic cost of O3 exposures. These post-1996 

studies used variants of the economic methods from earlier assessments and measure yield 

changes from response functions arising from the NCLAN or similar data. For example, Kim 

et al. (1998) used a mathematical programming model of the San Joaquin Valley agricultural 

sector in California, combined with crop yield response functions, to assess the economic effects 

of O3 on California crops. Their results showed net benefits from reductions in ambient O3 

levels, a finding consistent with all previous economic assessments. In another study, 

Westenbarger and Frisvold (1994) measured agricultural exposures to O3 (and acid precipitation) 

in the United States. Though not an economic analyses of the costs of ambient exposures, they 

identified areas of the United States of greatest potential economic damage based on the 

interface between regional pollution levels and the value of crop production in each region. 

A study by Murphy et al. (1999) of the economic effects of tropospheric O3 on U.S. 

agriculture is of note here, because it confirms the general magnitude of economic effects 

reported by the two key studies performed a decade earlier (Adams, 1986; Adams et al., 1985). 

Specifically, Murphy et al. (1999) evaluated benefits to eight major crops associated with several 

scenarios concerning the reduction or elimination of O3 precursor emissions from motor vehicles 

in the United States. Their analysis reported a $2.8 to 5.8 billion (1990 dollars) benefit from 

complete elimination of O3 exposures from all sources, i.e., ambient O3 reduced to a background 

level assumed to be 0.025 to 0.027 ppm. While the analytical framework is similar to Adams 

et al. (1986) in the use of NCLAN-based yield response functions and a mathematical 

programming-based economic optimization model, the study is novel in its focus on the role of 

motor vehicle emissions of VOCs/NOx in total anthropogenic O3 levels. The study is also 

notable in its careful attention to federal farm program effects, particularly the deficiency 

payment component. 

AX9-331
In addition to these studies in peer-reviewed journals, a number of site-specific effects 

studies have been performed, primarily by consulting companies for state public utility 

commissions. Although perhaps of use to public utility commissioners concerned with effects 

from single power plants or other localized sources, these regional studies generally contribute 

little to the assessment of air pollution effects at the national level. Also, such reports are not 

peer reviewed. Therefore, they are not discussed here. 

There have been a number of recent studies of air pollutant effects on tree species in the 

literature. Some have reported changes in total biomass and focused on European species 

(Kurczynska et al., 1997). Other studies have assessed changes in composition of forest species 

(biodiversity) or forest health due to exposure to air pollutants (Bringmark and Bringmark, 1995; 

McLaughlin and Percy, 1999; Vacek et al., 1999). As noted previously, changes in forest 

biomass and composition are more difficult to value than marketable products. However, 

measures of forest composition or health have implications for an area of increasing policy 

concern, that being the effect of air pollutants and other environmental stressors on unmanaged 

(natural) ecosystems and the services they provide (Goulder and Kennedy, 1997; Pimentel et al., 

1997). Considerable discussion has occurred among ecologists and economists as to the 

appropriate means for valuing these services (Anderson, 1990; Carpenter and Dixon, 1985; 

Common and Perrings, 1992). A number of conceptual articles have been published on this 

issue in both economic and ecological journals (Bergstrom, 1990; Castle, 1993; Pearce, 1993; 

Suter, 1990). 

A continuing empirical challenge concerns the lack of information on how changes in 

biodiversity affect ecosystem performance resulting from O3 stresses and the problem of 

establishing economic values for such changes (Cairnes and Lackey, 1992; Norton, 1987; Pimm, 

1997; Polasky, 2001; Randall, 1991). As noted in the 1996 AQCD, and more recently by Daily 

(1999, 2000) and Polasky (2001), there continues to be a lack of empirical studies that actually 

assess the economic value of changes in biodiversity or in service flows due to any 

environmental stressors. While some studies report monetary estimates, the estimates are 

generally for expository purposes and those would not be as defensible as the agricultural studies 

described earlier. For example, Costanza et al. (1997) assigned a value to the world’s 

ecosystems, but the procedures used render this an exploratory study at best. As assessed by 

Polasky (2001), “In general, the field of valuation of ecosystem services is in its infancy.” 

AX9-332
He attributed the lack of empirical studies due to both a lack of the understanding of ecology of 

ecosystem services and to the absence of reliable methods to estimate the value of these services. 

In summary, the studies of crop and forest responses in the economic literature indicate 

that O3 reduces crop yields and imposes economic costs. The economic literature also indicates 

that O3 adversely influences the physiological performance of tree species and demonstrates, as 

expected, that changes in growth have economic consequences. However, the economic data 

and literature available on ecosystem effects of O3 exposures are not sufficient to determine the 

economic costs. 

AX9.7.5 Limitations of Scientific Studies and Economic Information 

The 1996 O3 AQCD discussed the need for additional research on both ecological 

functions and economic methodology in order to better understand the economic implications of 

air pollutants on ecosystem services. As noted by Daily (1999, 2000), Polasky (2001), and 

others, this research agenda continues to be important. Despite the large number of discussion 

and survey articles published since 1996, there still are not sufficient data by which to estimate 

confidently the magnitude of economic effects of O3 to forests and natural ecosystems. Nor is it 

apparent that ongoing research is adequate to answer this question in the next criteria document 

cycle. Specifically, there do not appear to be any comprehensive, ecological studies underway 

that attempt to measure changes in ecosystem outputs under alternative O3 or other air pollutant 

levels. Thus, in the near term, ecosystem services can only be discussed in qualitative terms. 

However plausible the likelihood of economic damages to ecosystems, the available scientific 

and economic information does not provide specific guidance on the magnitudes of these effects. 

Beside the need to improve our understanding of the effects of O3 exposures on natural 

ecosystems, a number of areas of research could help assess the full economic consequences of 

such pollutants on managed ecosystems. The first of these is the relationship between O3 

exposure levels and the variability of crop yields or changes in forest biodiversity. Most 

assessments are based on the average or expected yield response of a crop to air pollution 

exposure. However, the variability in yields (the spread or dispersion around the mean) appears 

to be affected by the nature of plant exposure to pollutants (Hogsett et al., 1985a; Musselman 

et al., 1983). Plants exposed to the same mean dose but with different second moments of the 

distribution of exposure may have different mean yields (Altshuller and Lefohn, 1996; Lefohn 

AX9-333
and Benedict, 1982). In addition, the variability of the yield of the plant may also be increased 

by greater variability in exposures (e.g., a higher frequency of extreme events). The economic 

significance of higher yield varieties is such that variability may impose additional economic 

costs, because most farmers have been reported to be averse to risk and prefer less variability for 

a given yield (or profit). To date, no economic assessments of O3 damages to agriculture or 

vegetation in general include risk-averse behavior (studies cited here assumed risk neutrality). 

To assess the economic consequences of a relationship between farmers’ risk preferences 

and O3-induced changes in yield variability will require more information on the potential 

effects of changes in O3 on crop yield variability. While no economic studies were found on the 

effects of O3-induced changes on yield variability, a number of recent studies of climate change 

effects on crop yields have indicated increased economic costs in the presence of increased 

climatic variability (e.g., Mearns et al., 1997; Dixon and Segerson, 1999). Analogous economic 

costs would be expected for changes in air pollution distributions; and these effects need to be 

examined and quantified. 

Another research area concerns the need improvement in our understanding of temporal 

(dynamic) and spatial characteristics of O3 exposures and their implications for crop yields, 

production and producer profits. Most economic studies are static, in that they compare two 

states of the world (e.g., economic activity at one O3 level versus at an alternative level). 

In addition, most national-level studies (the type needed to evaluate Secondary National 

Ambient Air Quality Standards [SNAAQS]) display coarse regional-level resolution in terms of 

crop response, O3 exposure, and economic behavior. The responses of producers to changes in 

yields due to changes in O3 levels are generally assumed to be similar over geographical areas up 

to several states in size. However, the changes between air quality scenarios are more likely to 

be characterized by transient changes in exposure levels, which means the producer responses 

are also likely to be gradual, rather than abrupt. Similarly, the lack of finer scale (regional-level) 

data and modeling capabilities suggests that important micro-level physical and economic effects 

are ignored. To what extent these abstractions influence net economic effects is an empirical 

question. Research on these types of abstractions and assumptions within other economic 

settings, such as climatic change, have shown that they have implications for economic 

measurements (Adams, 2002). 

AX9-334
Another issue is the natural or background level of O3 (or other pollutants of interest) 

assumed in economic studies. While many economic studies focus on changes in pollution 

levels from current conditions, some studies have measured the economic damages between an 

assumed, or pristine, level and current levels in agricultural areas. Such an analysis, it is 

reasoned, will provide a measure of the net damages due to anthropogenic sources. The 

challenge here is to have a correct measure of the background level of the pollutant. Recent 

research by Lefohn et al. (2001) has suggested that background levels may be considerably 

higher than assumed in some of the previous economic assessments reported in the 1996 AQCD 

(25 to 30 ppb in most studies). For example, Lefohn et al. (Lefohn et al., 2001) detected hourly 

readings of from 40 to 80 ppb during winter and spring months in remote areas of the United 

States. If background O3 levels are in the range, then the economic damages estimated in studies 

with lower background levels will be overstated. The issue of the background O3 level is 

important to all assessments of vegetative damages due to O3. 

In terms of expanding economic methods for future assessment, analysts should consider 

using more “reduced form” estimation methods, particularly in situations where the availability 

of dose-response functions is limited. This estimation approach is exemplified by Garcia et al. 

(1986). Specifically, their econometric study of the impact of O3 on producer profits used 

such a reduced-form approach. In this approach, farmer actions are modeled as a function of 

ambient O3 levels (along with other explanatory variables) without the direct use of dose 

response functions. The advantage of this procedure is that one source of modeling uncertainty, 

the need for dose-response functions, including time-consuming crop experiments to generate 

data, is reduced. Actual responses of farmers’ profits across air pollution gradients of ambient 

pollution are observed instead of hypothesized. Although this procedure has not been widely 

used in air pollution economic assessments, it has been used in a number of relatively recent 

climate change studies (e.g., Mendelsohn et al., 1994). The reduced-form method suffers from 

the fact that if proposed O3 levels are lower than those observed in the estimation sample, then 

the prediction accuracy of the method deteriorates. Also, some dose-response information is 

needed, if only to establish the plausibility of the economic estimates. 

Another area that may improve economic assessments is incorporating consideration of 

livestock issues. To date, most agricultural economic assessments of O3 impacts have ignored 

the livestock sector. Presumably this is because ambient O3 levels do not noticeably affect meat 

AX9-335
yields. However, if feed prices and pasture conditions are affected by ambient O3 levels, then a 

more accurate estimate of economic impacts would be forthcoming by including this link to 

livestock in the assessment. These types of feed production and feed price effects are included 

in the mathematical programming model used in Adams et al. (1986), but not in most other O3 

effects assessments. The significance of livestock-feed linkages are demonstrated in a recent 

study in the Netherlands by Kuik et al (2000). Using a mathematical programming model 

similar to that in Adams et al. (1986), Kuik et al. (2000) found that livestock effects were 

prominent, due mainly to improved pasture yields under reduced ambient O3 levels. 

AX9.7.6 Conclusions 

Substantial progress has been made over the past two decades in our understanding of the 

effects of O3 and other oxidants on vegetation, particularly for agriculturally important plant 

species. The physical and economic effects on agriculture are well documented and provide 

useful information for the setting of SNAAQS. Effects on forests and natural ecosystems remain 

problematic, due to limitations in natural science data and economic methods. The problem is 

most acute for valuing natural ecosystem service flows. 

The current limitations surrounding forests and natural ecosystems present a rich research 

agenda. However, not all research needs are likely to lead to better policies. Thus, areas of 

greatest potential value in terms of regional policymaking need to be prioritized. Such priority 

setting can be assisted by sensitivity analyses with existing economic models. By measuring the 

changes in economic effects arising from changes in key parameters, research data gaps most 

likely to affect economic values can be identified. 

AX9-336
REFERENCES 

Abeles, F. B.; Biles, C. L.; Dunn, L. J. (1989a) Hormonal regulation and distribution of peroxidase isoenzymes in 

the cucurbitaceae. Plant Physiol. 91: 1609-1612. 

Abeles, F. B.; Hershberger, W. L.; Dunn, L. J. (1989b) Hormonal regulation and intracellular localization of a 

33 kda cationic peroxidase in excised cucumber cotyledons. Plant Physiol. 89: 664-668. 

Adams, R. M. (1986) Agriculture, forestry, and related benefits of air pollution control: a review and some 

observations. Am. J. Agric. Econ. 68: 464-472. 

Adams, R. M. (1999) On the search for the correct economic assessment method: an editorial comment. Clim. 

Change 41: 363-370. 

Adams, W. C. (2002) Comparison of chamber and face-mask 66-hour exposures to ozone on pulmonary function 

and symptoms responses. Inhalation Toxicol. 14: 745-764. 

Adams, R. M.; Crocker, T. D. (1989) The agricultural economics of environmental change: some lessons from air 

pollution. J. Environ. Manage. 28: 295-307. 

Adams, R. M.; Horst, R. L., Jr. (2003) Future directions in air quality research economic issues. Environ. Int. 

29: 289-302.

Adams, M. B.; O'Neill, E. G. (1991) Effects of ozone and acidic deposition on carbon allocation and mycorrhizal 

colonization of Pinus taeda L seedlings. For. Sci. 37: 5-16. 

Adams, R. M.; Rowe, R. D. (1990) The economic effects of stratospheric ozone depletion on US agriculture: 

a preliminary assessment. J. Environ. Manage. 30: 321-335. 

Adams, R. M.; Hamilton, S. A.; McCarl, B. A. (1985) An assessment of the economic effects of ozone on U.S. 

agriculture. J. Air Pollut. Control Assoc. 35: 938-943. 

Adams, R. M.; Hamilton, S. A.; McCarl, B. A. (1986) The benefits of pollution control: the case of ozone and U.S. 

agriculture. Am. J. Agric. Econ. 68: 886-893. 

Adams, R. M.; Glyer, J. D.; McCarl, B. A. (1988) The NCLAN economic assessment: approach, findings and 

implications. In: Heck, W. W.; Taylor, O. C.; Tingey, D. T., eds. Assessment of crop loss from air pollutants: 

proceedings of an international conference; October 1987; Raleigh, NC. New York, NY: Elsevier Applied 

Science; pp. 473-504. 

Adams, D. M.; Alig, R. J.; Callaway, J. M.; McCarl, B. A.; Winnett, S. M. (1996) The forest and agricultural sector 

optimization model (FASOM): model structure and policy application. Portland, OR: U.S. Department of 

Agriculture, Pacific Northwest Research Station; PNW-RP-495. 

Adams, R. M.; Hurd, B. H.; Lenhart, S.; Leary, N. (1998) Effects of global climate change on agriculture: 

an interpretative review. Clim. Res. 11: 19-30. 

Agrawal, M.; Agrawal, S. B. (2000) Effects of air pollution on plant diversity. In: Agrawal, S. B.; Agrawal, M., eds. 

Environmental Pollution and Plant Responses. Lewis Publishers; pp. 137-152. 

Ainsworth, E. A.; Long, S. P. (2005) What have we learned from 15 years of free-air CO2 enrichment (FACE)? 

A meta-analytic review of the responses of photosynthesis, canopy properties and plant production to rising 

CO2. New Phytol. 165: 351-371. 

Ainsworth, N.; Fumagalli, I.; Giorcelli, A.; Mignanego, L.; Schenoe, G.; Vietto, L. (1996) Assessment of EDU stem 

injections as a technique to investigate the response of trees to ambient ozone in field conditions. Agric. 

Ecosyst. Environ. 59: 33-42. 

Ainsworth, E. A.; Davey, P. A.; Bernacchi, C. J.; Dermody, O. C.; Heaton, E. A.; Moore, D. J.; Morgan, P. B.; 

Naidu, S. L.; Yoora, H. S.; Zhu, X. G.; Curtis, P. S.; Long, S. P. (2002) A meta-analysis of elevated [CO2] 

effects on soybean (Glycine max) physiology, growth and yield. Global Change Biol. 8: 695-709. 

Alexis, A.; Garcia, A.; Nystrom, M.; Rosenkranz, K. (2001) The 2001 California almanac of emissions and air 

quality. Sacramento, CA: California Air Resources Board, California Planning and Technical Support 

Division. Available: http://www. arb. ca. gov/aqd/almanac/almanac01/almanac01. htm [10 August, 2005]. 

Alscher, R. G.; Wellburn, A. R. (1994) Plant responses to the gaseous environment. Molecular, metabolic and 

physiological aspects. London, UK: Chapman and Hall. 

Alscher, R. G.; Donahue, J. L.; Cramer, C. L. (1997) Reactive oxygen species and antioxidants: relationships in 

green cells. Physiol. Plant. 100: 224-233. 

Altimir, N.; Vesala, T.; Keronen, P.; Kulmala, M.; Hari, P. (2002) Methodology for direct field measurements of 

ozone flux to foliage with shoot chambers. Atmos. Environ. 36: 19-29. 

Altimir, N.; Tuovinen, J. P.; Vesala, T.; Kulmala, M.; Hari, P. (2004) Measurements of ozone removal by Scots pine 

shoots: calibration of a stomatal uptake model including the non-stomatal component. Atmos. Environ. 

38: 2387-2398.

AX9-337 
Altshuller, A. P.; Lefohn, A. S. (1996) Background ozone in the planetary boundary layer over the United States. 

J. Air Waste Manage. Assoc. 46: 134-141. 

Alvarado, D.; De Bauer, L. I.; Galindo, J. (1993) Decline of sacred fir (Abies religiosa) in a forest park south of 

Mexico City. Environ. Pollut. 80: 115-121. 

Amiro, B. D.; Gillespie, T. J. (1985) Leaf conductance response of Phaseolus vulgaris to ozone flux density. 

Atmos. Environ. 19: 807-810. 

Amiro, B. D.; Gillespie, T. J.; Thurtell, G. W. (1984) Injury response of Phaseolus vulgaris to ozone flux density. 

Atmos. Environ. 18: 1207-1215. 

Amthor, J. S.; Goulden, M. L.; Munger, J. W.; Wofsy, S. C. (1994) Testing a mechanistic model of forest-canopy 

mass and energy exchange using eddy correlation: carbon dioxide uptake and ozone uptake by a mixed 

oak-maple stand. Aust. J. Plant Physiol. 211: 623-651. 

Anderson, E. (1990) The ethical limitations of the market. Econ. Philos. 6: 179-205. 

Andersen, C. P. (2000) Ozone stress and changes below-ground: linking root and soil processes. Phyton 

40(Special issue 4): 7-12. 

Andersen, C. P. (2003) Source-sink balance and carbon allocation below ground in plants exposed to ozone. 

New Phytol. 157: 213-228. 

Andersen, C. P.; Rygiewicz, P. T. (1991) Stress interactions and mycorrhizal plant response: understanding carbon 

allocation priorities. Environ. Pollut. 73: 217-244. 

Andersen, C. P.; Rygiewicz, P. T. (1995a) Allocation of carbon in mycorrhizal Pinus ponderosa seedlings exposed 

to ozone. New Phytol. 131: 471-480. 

Andersen, C. P.; Rygiewicz, P. T. (1995b) Effect of ozone on temporal allocation of carbon in mycorrhizal of 

Pinus ponderosa seedlings. New Phytol. 131: 471-480. 

Andersen, C. P.; Scagel, C. F. (1997) Nutrient availability alters belowground respiration of ozone-exposed 

ponderosa pine. Tree Physiol. 17: 377-387. 

Anderson, R. L.; Brown, H. D.; Chevone, B. I.; McCartney, T. C. (1988) Occurence of air pollution symptoms 

(needle tip necrosis and chlorotic mottling) on eastern white pine in the southern Appalachian Mountains. 

Plant Dis. 72: 130-132. 

Andersen, C. P.; Hogsett, W. E.; Wessling, R.; Plocher, M. (1991) Ozone decreases spring root growth and root 

carbohydrate content in ponderosa pine the year following exposure. Can. J. For. Res. 21: 1288-1291. 

Andersen, C. P.; Wilson, R.; Plocher, M.; Hogsett, W. E. (1997) Carry-over effects of ozone on root growth and 

carbohydrate concentrations of ponderosa pine seedlings. Tree Physiol. 17: 805-811. 

Andersen, C. P.; Hogsett, W. E.; Plocher, M.; Rodecap, K.; Lee, E. H. (2001) Blue wild-rye grass competition 

increases the effect of ozone on ponderosa pine seedlings. Tree Physiol. 21: 319-327. 

Antonielli, M.; Pasqualini, S.; Ederli, L.; Batini, P.; Moscatello, S.; Loreto, F. (1997) Physiological characteristics of 

tobacco cultivars with contrasting sensitivity to ozone. Environ. Exp. Bot. 38: 271-277. 

Anttonen, S.; Sutinen, M. L.; Heagle, A. S. (1996) Ultrastructure and some plasma membrane characteristics of 

ozone-exposed loblolly pine needles. Physiol. Plant. 98: 309-319. 

Aono, M.; Saji, H.; Fujiyama, K.; Sugita, M.; Kondo, N.; Tanaka, K. (1995) Decrease in activity of glutathione 

reductase enhances paraquat sensitivity in transgenic nicotiana tabacum. Plant Physiol. 107: 645-648. 

Arbaugh, M. J.; Miller, P. R.; Carroll, J. J.; Takemoto, B. L.; Proctor, T. (1998) Relationships of ozone exposure to 

pine injury in the Sierra Nevada and San Bernardino Mountains of California, USA. Environ. Pollut. 

101: 291-301. 

Arbaugh, M. J.; Johnson, D. W.; Pulliam, W. M. (1999) Simulated effects of N deposition, ozone injury, and climate 

change on a forest stand in the San Bernardino Mountains. In: Miller, P. R.; McBride, J. R., eds. Oxidant air 

pollution impacts in the montane forests of southern California. New York, NY: Springer (Ecological Studies, 

Vol. 134); pp. 179-207. 

Arbaugh, M.; Bytnerowicz, A.; Grulke, N.; Fenn, M.; Poth, M.; Temple, P.; Miller, P. (2003) Photochemical smog 

effects in mixed conifer forests along a natural gradient of ozone and nitrogen deposition in the 

San Bernardino Mountains. Environ. Int. 29: 401-406. 

Arndt, U. (1995) Air pollutants and pheromones - a problem? Chemosphere 30: 1023-1031. 

Arp, W. J.; Drake, B. G. (1991) Increased photosynthetic capacity of Scirpus olneyi after 4 years of exposure to 

elevated CO2. Plant Cell Environ. 14: 1003-1006. 

Asada, K. (1999) The water-water cycle in chloroplasts: scavenging of active oxygens and dissipation of excess 

photons. Ann. Rev. Plant Physiol. Mol. Biol. 50: 601-639. 

Ashenden, T. W.; Bell, S. A.; Rafarel, C. R. (1995) Responses of white clover to gaseous pollutants and acid mist: 

implications for setting critical levels and loads. New Phytol. 130: 89-96. 

AX9-338
Ashenden, T. W.; Bell, S. A.; Rafarel, C. R. (1996) Interactive effects of gaseous air pollutants and acid mist on two 

major pasture grasses. Agric. Ecosyst. Environ. 57: 1-8. 

Ashmore, M. R.; Ainsworth, N. (1995) The effects of ozone and cutting on the species composition of artificial 

grassland communities. Funct. Ecol. 9: 708-712. 

Ashmore, M. R.; Bell, J. N. B.; Mimmack, A. (1988) Crop growth along a gradient of ambient air pollution. 

Environ. Pollut. 53: 99-121. 

Ashmore, M. R.; Thwaites, R. H.; Ainsworth, N.; Cousins, D. A.; Power, S. A.; Morton, A. J. (1995) Effects of 

ozone on calcareous grassland communities. Water Air Soil Pollut. 85: 1527-1532. 

Ashmore, M.; Emberson, L.; Karlsson, P. E.; Pleijel, H. (2004a) New directions: a new generation of ozone critical 

levels for the protection of vegetation in Europe (correspondence). Atmos. Environ. 38: 2213-2214. 

Ashmore, M. E.; Karlsson, P. E.; Pleijel, H. (2004b) Introduction for ozone deposition special issue. Atmos. 

Environ. 38: 2211-2212. 

Assmann, S. M. (2003) Open stomata 1 opens the door to ABA signaling in Arabidopsis guard cells. Trends Plant 

Sci. 8: 151-153. 

Assmann, S. M.; Wang, X. Q. (2001) From milliseconds to millions of years: guard cells and environmental 

responses. Curr. Opin. Plant Biol. 4: 421-428. 

Atkinson, R. (1990) Gas-phase tropospheric chemistry of organic compounds: a review. Atmos. Environ. Part A 

24: 1-41. - Auh, C. K.; Murphy, T. M. (1995) Plasma membrane redox enzyme is involved in the synthesis of O2 and H2O2 by 

phytophthora elicitor-stimulated rose cells. Plant Physiol. 107: 1241-1247. 

Awmack, C. S.; Harrington, R.; Lindroth, R. L. (2004) Aphid individual performance may not predict population 

responses to elevated CO2 or O3. Global Change Biol. 10: 1414-1423. 

Ayres, P. G. (1991) Growth responses induced by pathogens and other stresses. In: Mooney, H. A.; Winner, W. E.; 

Pell, E. J., eds. Responses of plants to multiple stresses. San Diego, CA: Academic Press; pp. 227-248. 

Azevedo, R. A.; Alas, R. M.; Smith, R. J.; Lea, P. J. (1998) Response of antioxidant enzymes to transfer from 

elevated carbon dioxide to air and ozone fumigation, in the leaves and roots of wild-type and a 

catalase-deficient mutant of barley. Physiol. Plant. 104: 280-292. 

Bäck, J.; Vanderklein, D. W.; Topa, M. A. (1999) Effects of elevated ozone on CO2 uptake and leaf structure in 

sugar maple under two light environments. Plant Cell Environ. 22: 137-147. 

Bahl, A.; Kahl, G. (1995) Air pollutant stress changes the steady-state transcript levels of three photosynthesis 

genes. Environ. Pollut. 88: 57-65. 

Bahl, A.; Loitsch, S. M.; Kahl, G. (1995) Transcriptional activation of plant defence genes by short-term air 

pollutant stress. Environ. Pollut. 89: 221-227. 

Baker, T. R.; Allen, H. L.; Schoeneberger, M. M.; Kress, L. W. (1994) Nutritional response of loblolly pine exposed 

to ozone and simulated acid rain. Can. J. For. Res. 24: 453-461. 

Baker, C. J.; Harmon, G. L.; Glazener, J. A.; Orlandi, E. W. (1995) A noninvasive technique for monitoring 

peroxidative and H2O2-scavenging activities during interactions between bacterial plant pathogens and 

suspension cells. Plant Physiol. 108: 353-359. 

Balaguer, L.; Barnes, J. D.; Panicucci, A.; Borland, A. M. (1995) Production and utilization of assimilates in wheat 

(Triticum aestivum L) leaves exposed to elevated O3 and/or CO2. New Phytol. 129: 557-568. 

Ball, J. T. (1987) Calculations related to gas exchange. In: Zeiger, E.; Farquhar, G. D.; Cowan, I. R. S.; pp. 445-476, 

eds. ; Farquhar, G. Stanford Press; pp. 445-476.; pp. G-47. 

Ball, G. R.; Benton, J.; Palmer-Brown, D.; Fuhrer, J.; Skärby, L.; Gimeno, B. S.; Mills, G. (1998) Identifying factors 

which modify the effects of ambient ozone on white clover (Trifolium repens) in Europe. Environ. Pollut. 

103: 7-16. 

Ball, G. R.; Palmer-Brown, D.; Fuhrer, J.; Skärby, L.; Gimeno, B. S.; Mills, G. (2000) Identification of non-linear 

influences on the seasonal ozone dose-response of sensitive and resistant clover clones using artificial neural 

networks. Ecol. Model. 129: 153-168. 

Balls, G. R.; Palmer-Brown, D.; Cobb, A. H.; Sanders, G. E. (1995) Towards unravelling the complex interactions 

between microclimate, ozone dose, and ozone injury in clover. Water Air Soil. Pollut. 85: 1467-1472. 

Balls, G. R.; Palmer-Brown, D.; Sanders, G. E. (1996) Investigating microclimatic influences on ozone injury in 

clover (Trifolium subterraneum) using artificial neural networks. New Phytol. 132: 271-280. 

Banerjee, S. K.; Mudd, J. B. (1992) Reaction of ozone with glycophorin in solution and in lipid vesicles. 

Arch. Biochem. Biophys. 295: 84-89. 

Barbo, D. N.; Chappelka, A. H.; Somers, G. L.; Miller-Goodman, M. S.; Stolte, K. (1998) Diversity of an early 

successional plant community as influenced by ozone. New Phytol. 138: 653-662. 

AX9-339
Barbo, D. N.; Chappelka, A. H.; Somers, G. L.; Miller-Goodman, M. S.; Stolte, K. (2002) Ozone impacts on loblolly 

pine (Pinus taeda L) grown in a competitive environment. Environ. Pollut. 116: 27-36. 

Barnes, J. D.; Pfirrmann, T. (1992) The influence of CO2 and O3, singly and in combination, on gas exchange, 

growth and nutrient status of radish (Raphanus sativus L). New Phytol. 121: 403-412. 

Barnes, J. D.; Wellburn, A. R. (1998) Air pollutant combinations. In: De Kok, L. J.; Stulen, I., eds. Responses of 

plant metabolism to air pollution and global change. Leiden, The Netherlands: Backhuys Publishers; 

pp. 147-164. 

Barnes, J. D.; Ollerenshaw, J. H.; Whitfield, C. P. (1995) Effects of elevated CO2 and/or O3 on growth, development 

and physiology of wheat (Triticum aestivum L). Global Change Biol. 1: 129-142. 

Barnes, J.; Zheng, Y.; Lyons, T. (2002) Plant resistance to ozone: the role of ascorbate. In: Omasa, K.; Saji, H.; 

Youssefian, S.; Kondo, N., eds. Air pollution and plant biotechnology - prospects for phytomonitoring and 

phytoremediation. Air pollution and plant biotechnology - prospects for phytomonitoring and 

phytoremediation. Springer-Verlag; pp. 235-252. 

Barrett, S. C. H.; Bush, E. J. (1991) Population processes in plants and the evolution of resistance to gaseous air 

pollutants. In: Taylor, G. E., Jr.; Pitelka, L. F.; Clegg, M. T., eds. Ecological genetics and air pollution. 

Springer-Verlag; pp. 137-165. 

Bartholomay, G. A.; Eckert, R. T.; Smith, K. T. (1997) Reductions in tree-ring widths of white pine following ozone 

exposure at Acadia National Park, Maine, USA. Can. J. For. Res. 27: 361-368. 

Bassin, S.; Calanca, P.; Weidinger, T.; Gerosa, G.; Fuhrer, J. (2004) Modeling seasonal ozone fluxes to grassland 

and wheat: model improvement, testing, and application. Atmos. Environ. 38: 2349-2359. 

Batts, G. R.; Morison, J. I. L.; Ellis, R. H.; Hadley, P.; Wheeler, T. R. (1997) Effects of CO2 and temperature on 

growth and yield of crops of winter wheat over four seasons. Eur. J. Agron. 7: 43-52. 

Bauer, M. R.; Hultman, N. E.; Panek, J. A.; Goldstein, A. H. (2000) Ozone deposition to a ponderosa pine plantation 

in the Sierra Nevada Mountains (CA): a comparison of two different climatic years. J. Geophys. Res. 

[Atmos.] 105: 22,123-22,136. 

Baumbusch, L. O.; Eiblmeier, M.; Schnitzler, J. P.; Heller, W.; Sandermann, H., Jr.; Polle, A. (1998) Interactive 

effects of ozone and low UV-B radiation on antioxidants in spruce (Picea abies) and pine (Pinus sylvestris). 

Physiol. Plant. 104: 248-254. 

Baumgardner, R. E.; Edgerton, E. S. (1998) Rural ozone across the eastern United States: analysis of CASTNet data, 

1988-1995. J. Air Waste Manage. Assoc. 48: 674-688. 

Baumgarten, M.; Werner, H.; Häberle, K. H.; Emberson, L. D.; Fabian, P.; Matyssek, R. (2000) Seasonal ozone 

response of mature beech trees (Fagus sylvatica) at high altitude in the Bavarian Forest (Germany) 

in comparison with young beech grown in the field and in phytotrons. Environ. Pollut. 109: 431-442. 

Beare, J. A.; Archer, S. A.; Bell, J. N. B. (1999a) Effects of Melampsora leaf rust disease and chronic ozone 

exposure on poplar. Environ. Pollut. 105: 419-426. 

Beare, J. A.; Archer, S. A.; Bell, J. N. B. (1999b) Marssonina leafspot disease of poplar under elevated ozone: 

pre-fumigated host and in vitro studies. Environ. Pollut. 105: 409-417. 

Beligni, M. V.; Lamattina, L. (2001) Nitric oxide in plants: the history is just beginning. Plant Cell Environ. 

24: 267-278.

Beligni, M.; Lamattina, L. (2002) Nitric oxide interferes with plant photo-oxidative stress by detoxifying reactive 

oxygen species. Plant Cell Environ. 25: 737-748. 

Bell, J. N. B.; Treshow, M. (2002) Air pollution and plant life. Chichester, UK: John Wiley & Sons Ltd., 465. 

Bell, J. N. B.; Ashmore, M. R.; Wilson, G. B. (1991) Ecological genetics and chemical modifications of the 

atmosphere. In: Taylor, G. E., Jr.; Pitelka, L. F.; Clegg, M. T., eds. Ecological genetics and air pollution. 

Springer-Verlag; pp. 33-59. 

Bender, J.; Weigel, H. J. (1992) Crop responses to mixtures of air pollutants. In: Jäger, H. J.; Unsworth, M.; 

De Temmerman, L.; Mahy, P., eds. Effects of air pollution on agricultural crops in Europe. Commission of 

the European Communities; pp. 445-454. 

Bender, J.; Weigel, H. J.; Jager, H. J. (1991) Response of nitrogen metabolism in beans (Phaseolus vulgaris L) 

after exposure to ozone and nitrogen dioxide, alone and in sequence. New Phytol. 119: 261-267. 

Bender, J.; Weigel, H. J.; Wegner, U.; Jäger, H. J. (1994) Response of cellular antioxidants to ozone in wheat flag 

leaves at different stages of plant development. Environ. Pollut. 84: 15-21. 

Bender, J.; Herstein, U.; Black, C. R. (1999) Growth and yield responses of spring wheat to increasing carbon 

dioxide, ozone and physiological stresses: a statistical analysis of 'ESPACE-wheat' results. Eur. J. Agron. 10: 

185-195. 

AX9-340
Benes, S. E.; Murphy, T. M.; Anderson, P. D.; Houpis, J. L. J. (1995) Relationship of antioxidant enzymes to ozone 

tolerance in branches of mature ponderosa pine (Pinus ponderosa) trees exposed to long term, low 

concentration, ozone fumigation and acid precipitation. Physiol. Plant. 94: 124-134. 

Bennett, J. H. (1979) Foliar exchange of gases. In: Heck, W. W.; Krupa, S. V.; Linzon, S. N. eds. Methodology for 

the assessment of air pollution effects on vegetation: a handbook from a specialty conference; April 1978; 

Minneapolis, MN; Pittsburgh, PA: Air Pollution Control Association; pp. 10-1-10-29. 

Bennett, J. P.; Anderson, R. L.; Mielke, M. L.; Ebersole, J. J. (1994) Foliar injury air pollution surveys of eastern 

white-pine (Pinus strobus L) - a review. Environ. Monit. Assess. 30: 247-274. 

Benoit, L. F.; Skelly, J. M.; Moore, L. D.; Dochinger, L. S. (1982) Radial growth reductions of Pinus strobus L 

correlated with foliar ozone sensitivity as an indicator of ozone-induced losses in eastern forests. Can. J. For. 

Res. 12: 673-678. 

Benoit, L. F.; Skelly, J. M.; Moore, L. D.; Dochinger, L. S. (1983) The influence of ozone on Pinus strobus L pollen 

germination. Can. J. For. Res. 13: 184-187. 

Benton, J.; Fuhrer, J.; Skärby, L.; Sanders, G. (1995) Results from the UN/ECE ICP-crops indicate the extent of 

exceedance of the critical levels of ozone in Europe. Water Air Soil Pollut. 85: 1473-1478. 

Benton, J.; Fuhrer, J.; Gimeno, B. S.; Skarby, L.; Palmer-Brown, D.; Ball, G.; Roadknight, C.; Sanders-Mills, G. 

(1996) The critical level of ozone for visible injury on crops and natural vegetation (ICP-Crops). In: 

Kärenlampi, L.; Skärby, L. eds. Critical levels for ozone in Europe: testing and finalizing the concepts: 

UN-ECE workshop report. UN-ECE Convention on Long-Range Transboundary Air Pollution; April; 

Kuopio, Finland; Kuopio, Finland: University of Kuopio, Department of Ecology and Environmental Science; 

pp. 44-57. 

Benton, J.; Fuhrer, J.; Gimeno, B. S.; Skärby, L.; Palmer-Brown, D.; Ball, G.; Roadknight, C.; Mills, G. (2000) 

An international cooperative programme indicates the widespread occurrence of ozone injury on crops. Agric. 

Ecosyst. Environ. 78: 19-30. 

Bergmann, E.; Bender, J.; Weigel, H.-J. (1999) Ozone threshold doses and exposure-response relationships for the 

development of ozone injury symptoms in wild plant species. New Phytol. 144: 423-435. 

Bergstrom, J. C. (1990) Concepts and measures of the economic value of environmental quality: a review. 

J. Environ. Manage. 31: 215-228.

Bergweiler, C. J.; Manning, W. J. (1999) Inhibition of flowering and reproductive success in spreading dogbane 

(Apocynum androsaemifolium) by exposure to ambient ozone. Environ. Pollut. 105: 333-339. 

Bermejo, V.; Gimeno, B. S.; Granados, I.; Santamaria, J.; Irigoyen, J. J.; Bermajo, R.; Porcuna, J. L.; Mills, G. 

(2002) Investigating indices to explain the impacts of ozone on the biomass of white clover (Trifolium repens 

L. cv. Regal) at island and coastal sites in Spain. New Phytol. 156: 43-55.

Berrang, P.; Karnosky, D. F.; Mickler, R. A.; Bennett, J. P. (1986) Natural selection for ozone tolerance in 

Populus tremuloides. Can. J. For. Res. 16: 1214-1216. 

Berrang, P.; Karnosky, D. F.; Bennett, J. P. (1989) Natural selection for ozone tolerance in Populus tremuloides: 

field verification. Can. J. For. Res. 19: 519-522. 

Berrang, P.; Karnosky, D. F.; Bennett, J. P. (1991) Natural selection for ozone tolerance in Populus tremuloides: 

an evaluation of nationwide trends. Can. J. For. Res. 21: 1091-1097. 

Bertrand, A.; Robitaille, G.; Nadeau, P.; Castonguay, Y. (1999) Influence of ozone on cold acclimation in sugar 

maple seedlings. Tree Physiol. 19: 527-534. 

Beyers, J. L.; Riechers, G. H.; Temple, P. J. (1992) Effects of long-term ozone exposure and drought on the 

photosynthetic capacity of ponderosa pine (Pinus ponderosa Laws). New Phytol. 122: 81-90. 

Bichele, I.; Moldau, H.; Padu, E. (2000) Estimation of plasmalemma conductivity to ascorbic acid in intact leaves 

exposed to ozone. Physiol. Plant. 108: 405-412. 

Bidartondo, M. I.; Ek, H.; Wallander, H.; Soderstrom, B. (2001) Do nutrient additions alter carbon sink strength of 

ectomycorrhizal fungi? New Phytol. 151: 543-550. 

Bielenberg, D. G.; Lynch, J. P.; Pell, E. J. (2001) A decline in nitrogen availability affects plant responses to ozone. 

New Phytol. 151: 413-425. 

Binkley, D. (1992) Sensitivity of forest soils in the western US to acidic deposition. In: Olson, R. K.; Binkley, D.; 

Böhm, M., eds. The Response of Western Forests to Air Pollution. Springer-Verlag; pp. 153-182. 

Birecka, H.; Catalfamo, J.; Urban, P. (1976) Cell isoperoxidases in sweet potato plants in relation to mechanical 

injury and ethylene. Plant Physiol. 57: 74-79. 

Björkman, O.; Demmig-Adams, B. (1995) Regulation of photosynthetic light energy capture, conversion, and 

dissipation in leaves of higher plants. In: Schulze, E.-D.; Caldwell, M. M., eds. Ecophysiology of 

photosynthesis. New York, NY: Springer-Verlag; pp. 17-47. 

AX9-341
Black, V. J.; Black, C. R.; Roberts, J. A.; Stewart, C. A. (2000) Impact of ozone on the reproductive development of 

plants. New Phytol. 147: 421-447. 

Blankenship, E. E.; Stefanski, L. A. (2001) Statistical estimation of ozone exposure metrics. Atmos. Environ. 

35: 4499-4510.

Boerner, R. E. J.; Rebbeck, J. (1995) Decomposition and nitrogen release from leaves of three hardwood species 

grown under elevated O3 and/or CO2. Plant Soil 170: 149-157. 

Böhm, M.; McCune, B.; Vandetta, T. (1995) Ozone regimes in or near forests of the western United States: 

I regional patterns. J. Air Waste Manage. Assoc. 45: 235-246. 

Bonello, P.; Heller, W.; Sandermann, H., Jr. (1993) Ozone effects on root-disease susceptibility and defence 

responses in mycorrhizal and non-mycorrhizal seedlings of Scots pine (Pinus sylvestris L). New Phytol. 

124: 653-663. 

Booker, F. L. (2000) Influence of carbon dioxide enrichment, ozone and nitrogen fertilization on cotton (Gossypium 

hirsutum L) leaf and root composition. Plant Cell Environ. 23: 573-583. 

Booker, F. L.; Miller, J. E. (1998) Phenylpropanoid metabolism and phenolic composition of soybean [Glycine max 

(L) Merr] leaves following exposure to ozone. J. Exp. Bot. 49: 1191-1202.

Booker, F. L.; Fiscus, E. L.; Miller, J. E. (1991) Ozone-induced changes in soybean cell wall physiology. In: 

Pell, E.; Steffen, K. , eds. Active Oxygen/Oxidative Stress and Plant Metabolism. Rockville, MD: American 

Society of Plant Physiologists, pp. 229-232. (Current topics in plant physiology: an American Society of Plant 

Physiologists series, v. 6). 

Booker, F. L.; Anttonen, S.; Heagle, A. S. (1996) Catechin, proanthocyanidin and lignin contents of loblolly pint 

(Pinus taeda) needles after chronic exposure to ozone. New Phytol. 132: 483-492. 

Booker, F. L.; Reid, C. D.; Brunschön-Harti, S.; Fiscus, E. L.; Miller, J. E. (1997) Photosynthesis and 

photorespiration in soybean [Glycine max (L) Merr] chronically exposed to elevated carbon dioxide and 

ozone. J. Exp. Bot. 48: 1843-1852. 

Booker, F. L.; Burkey, K. O.; Overmyer, K.; Jones, A. M. (2004) Differential responses of G-protein Arabidopsis 

thaliana mutants to ozone. New Phytol. 162: 633-641. 

Booker, F. L.; Fiscus, E. L.; Miller, J. E. (2004) Combined effects of elevated atmospheric carbon dioxide and ozone 

on soybean whole-plant water use. Environ. Manage. 33: S355-S362. 

Booker, F. L.; Miller, J. E.; Fiscus, E. L.; Pursley, W. A.; Stefanski, L. A. (2005) Comparative responses of 

container- versus ground-grown soybean to elevated CO2 and O3. Crop Sci. 45: 883-895. 

Bortier, K.; Ceulemans, R.; De Temmerman, L. (2000a) Effects of tropospheric ozone on woody plants. In: 

Agrawal, S. B.; Agrawal, M., eds. Environmental pollution and plant responses; pp. Lewis Publishers.; 

pp. 153-182. 

Bortier, K.; DeTemmerman, L.; Ceulemans, R. (2000b) Effects of ozone exposure in open-top chambers on poplar 

(Populus nigra) and beech (Fagus sylvatica): a comparison. Environ. Pollut. 109: 509-516. 

Bortier, K.; Ceulemans, R.; De Temmerman, L. (2000c) Effects of ozone exposure on growth and photosynthesis of 

beech seedlings (Fagus sylvatica). New Phytol. 146: 271-280. 

Bortier, K.; Dekelver, G.; De Temmerman, L.; Ceulemans, R. (2001a) Stem injection of Populus nigra with EDU to 

study ozone effects under field conditions. Environ. Pollut. 111: 199-208. 

Bortier, K.; Vandermeiren, K.; De Temmerman, L.; Ceulemans, R. (2001b) Growth, photosynthesis and ozone 

uptake of yound beech (Fagus sylvatica L) in response to different ozone exposures. Trees-Struct. Funct. 15: 

75-82. 

Bosac, C.; Black, V. J.; Roberts, J. A.; Black, C. R. (1998) The impact of ozone on seed yield and quality and 

seedling vigour in oilseed rape (Brassica napus L). J. Plant Physiol. 153: 127-134. 

Bosley, A.; Peterson, R.; Rebbeck, J. (1998) The resistance of the moss Polytrichum commune to acute exposure of 

simulated acid rain or ozone compared to two fern species: spore germination. Bryologist 101: 512-518. 

Bowler, C.; Van Montagu, M.; Inzé, D. (1992) Superoxide dismutase and stress tolerance. Ann. Rev. Plant Physiol. 

Mol. Biol. 43: 83-116. 

Bradshaw, A. D.; McNeilly, T. (1991) Evolution in relation to environmental stress. In: Taylor, G. E., Jr.; Pitelka, 

L. F.; Clegg, M. T., eds. Ecological genetics and air pollution. Springer-Verlag; pp. 11-31. 

Braun, S.; Fluckiger, W. (1995) Effects of ambient ozone on seedlings of Fagus sylvatica L and Picea abies (L) 

Karst. New Phytol. 129: 33-44. 

Braun, S.; Rihm, B.; Schindler, C.; Flückiger, W. (1999) Growth of mature beech in relation to ozone and nitrogen 

deposition: an epidemiological approach. Water Air Soil Pollut. 116: 357-364. 

Brennan, E. G.; Leone, I. A.; Clarke, B. B. (1987) EDU: a chemical for evaluating ozone foliar injury and yield 

reduction in field-grown crops. Int. J. Trop. Plant Disease 5: 35-42. 

AX9-342
Bringmark, E.; Bringmark, L. (1995) Disappearance of spatial variability and structure in forest floors - a distinct 

effect of air pollution? Water Air Soil Pollut. 85: 761-766. 

Broadbent, P.; Creisson, G. P.; Kular, B.; Wellburn, A. R.; Mullineaux, P. M. (1995) Oxidative stress responses in 

transgenic tobacco containing altered levels of glutathione reductase activity. Plant J. 8: 247-255. 

Broadmeadow, M. (1998) Ozone and forest trees. New Phytol. 139: 123-125. 

Broadmeadow, M. S. J.; Heath, J.; Randle, T. J. (1999) Environmental limitations to O3 uptake - some key results 

from young trees growing at elevated CO2 concentrations. Water Air Soil Pollut. 116: 299-310. 

Broadmeadow, M. S. J.; Jackson, S. B. (2000) Growth responses of Quercus petraea, Fraxinus excelsior and Pinus 

sylvestris to elevated carbon dioxide, ozone and water supply. New Phytol. 146: 437-451. 

Brook, J. R.; DiGiovanni, F.; Cakmak, S.; Meyers, T. P. (1997) Estimation of dry deposition velocity using 

inferential models and site-specific meteorology–uncertainty due to siting of meteorological towers. Atmos. 

Environ. 31: 3911-3919. 

Brown, V. C. (1995) Insect herbivores and gaseous air pollutants - current knowledge and predictions. In: 

Harrington, R.; Stork, N. E., eds. Insects in a changing environment: 17th symposium of the Royal 

Entomological Society. 17th symposium of the Royal Entomological Society. New York, NY: Academic 

Press; pp. 219-249. 

Brown, K. A.; Roberts, T. M. (1988) Effects of ozone on foliar leaching in Norway spruce (Picea abies L Karst): 

confounding factors due to NOx production during ozone generation. Environ. Pollut. 55: 55-73. 

Brown, V. C.; Ashmore, M. R.; McNeill, S. (1993) Experimental investigations of the effects of air pollutants on 

aphids on coniferous trees. Forstwiss. Centralbl. 112: 129-132. 

Brühl, C.; Crutzen, P. J. (1989) On the disproportionate role of tropospheric ozone as a filter against solar UV-B 

radiation. Geophys. Res. Lett. 16: 703-706. 

Bruhn, J. N. (1980) Effects of oxidant air pollution on Ponderosa and Jeffrey pine foliage decomposition 

[dissertation]. Berkeley, CA: University of California; Dissertation Abstracts Iinternational v. 42-07B. 

Brunschön-Harti, S.; Fangmeier, A.; Jäger, H. J. (1995) Influence of ozone and ethylenediurea (EDU) on growth and 

yield of bean (Phaseolus vulgaris L) in open-top field chambers. Environ. Pollut. 90: 89-94. 

Buchanan, B. B.; Gruissem, W.; Jones, R. L. (2000) Biochemistry and molecular biology of plants. Rockville, MD: 

American Society of Plant Physiologists. 

Bungener, P.; Balls, G. R.; Nussbaum, S.; Geissmann, M.; Grub, A.; Fuhrer, J. (1999a) Leaf injury characteristics of 

grassland species exposed to ozone in relation to soil moisture condition and vapour pressure deficit. 

New Phytol. 142: 271-282. 

Bungener, P.; Nussbaum, S.; Grub, A.; Fuhrer, J. (1999b) Growth response of grassland species to ozone in relation 

to soil moisture and plant strategy. New Phytol. 142: 283-293. 

Burkey, K. O. (1999) Effects of ozone on apoplast/cytoplasm partitioning of ascorbic acid in snap bean. Physiol. 

Plant. 107: 188-193. 

Bussotti, F.; Ferretti, M. (1998) Air pollution, forest condition and forest decline in Southern Europe. Environ. 

Pollut. 101: 49-65. 

Bytnerowicz, A. (2002) Physiological/ecological interactions between ozone and nitrogen deposition in forest 

ecosystems. Phyton 42: 13-28. 

Bytnerowicz, A.; Turunen, M. (1994) Effects of ozone exposures on epicuticular wax of ponderosa pine needles. In: 

Percy, K. E.; Cape, J. N.; Jagels, R.; Simpson, C. J., eds. Air pollutants and the leaf cuticle. Springer-Verlag; 

pp. 305-313. 

Bytnerowicz, A.; Olszyk, D. M.; Fox, C. A.; Dawson, P. J.; Kats, G.; Morrison, C. L.; Wolf, J. (1988) Responses of 

desert annual plants to ozone and water stress in an in situ experiment. JAPCA 38: 1145-1151. 

Bytnerowicz, A.; Poth, M.; Takemoto, B. K. (1990) Effects of photochemical smog and mineral nutrition on 

ponderosa pine seedlings. Environ. Pollut. 67: 233-248. 

Bytnerowicz, A.; Fenn, M. E.; Miller, P. R.; Arbaugh, M. J. (1999) Wet and dry pollutant deposition to the mixed 

conifer forest. In: Miller, P. R.; McBride, J. R., eds. Oxidant air pollution impacts in the montane forests of 

southern California: a case study of the San Bernardino Mountains. New York, NY: Springer; pp. 235-269. 

Bytnerowicz, A.; Frczek, W.; Grodzi1⁄2ska, K.; Godzik, B.; Fleischer, P.; Krzan, Z.; Skawinski, P. (2002a) 

Distribution of ambient ozone concentrations in the forested areas of the Tatra Mountains. In: Widacki, W.; 

Bytnerowicz, A.; Riebau, A., eds. A Message from the Tatra–geographical information systems and remote 

sensing in mountain environmental research. Krakow, Poland: Jagiellonian University Press, 233 pp. 

AX9-343
Bytnerowicz, A.; Godzik, B.; Frczek, W.; Grodzi1⁄2ska, K.; Krywult, M.; Badea, O.; Baran...ok, P.; Blum, O.; „erny, 

M.; Godzik, S.; Mankovska, B.; Manning, W.; Morav...ik, P.; Musselman, R.; Oszlanyi, J.; Postelnicu, D.; 

Szdïuj, J.; Varšavova, M.; Zota, M. (2002b) Distribution of ozone and other air pollutants in forests of the 

Carpathian Mountains in central Europe. Environ. Pollut. 116: 3-25. 

Bytnerowicz, A.; Tausz, M.; Alonso, R.; Jones, D.; Johnson, R.; Grulke, N. (2002c) Summer-time distribution of air 

pollutants in Sequoia National Park, California. Environ. Pollut. 118: 187-203. 

Bytnerowicz, A.; Godzik, B.; Grodzi1⁄2ska, K.; Frczek, W.; Musselman, R. C.; Manning, W. J.; Badea, O.; Popescu, 

F.; Fleischer, P. (2004) Ambient ozone in forests of the central and eastern European mountains. Environ. 

Pollut. 130: 5-16. 

Cairnes, M. S.; Lackey, R. T. (1992) Biodiversity and management of natural resources: the issues. Fisheries 

17: 6-10.

Callaway, J. M.; Darwin, R. F.; Nesse, R. J. (1986) Economic valuation of acidic deposition damages: preliminary 

results from the 1985 NAPAP assessment. In: Proceedings of International Symposium on Acidic 

Precipitation; September; Muskoka, Ontario, Canada; Richland, WA: U. S. Environmental Protection 

Agency; Pacific Northwest Laboratory; Water Air Soil Pollut. 31: 1019-1034. 

Campbell, S.; Temple, P.; Pronos, J.; Rochefort, R.; Andersen, C. (2000) Monitoring for ozone injury in west coast 

(Oregon, Washington, California) forests in 1998. Portland, OR: U.S. Department of Agriculture, Forest 

Service, Pacific Northwest Research Station; general technical report no. PNW-GTR-495. Available: 

http://www.fs.fed.us/pnw/gtrs.htm [11 April, 2003]. 

Cao, W.; Tibbitts, T. W.; Wheeler, R. M. (1994) Carbon dioxide interactions with irradiance and temperature in 

potatoes. In: MacElroy, R. D. Life sciences and space research XXV(3): natural and artificial ecosystems: 

proceedings of the topical meeting of the COSPAR Interdisciplinary Scientific Commission F (Meeting F4) 

of the COSPAR twenty-ninth plenary meeting; August-September, 1992; Washington, DC. New York, NY: 

Committee on Space Research; pp. 243-250. (Advances in space research: v. 14, no. 11). 

Cape, J. N. (1996) Surface wetness and pollutant deposition. In: Kerstiens, G., ed. Plant cuticles: an integrated 

functional approach. an integrated functional approach. Oxford, United Kingdom: BIOS Scientific Publishers; 

pp. 283-300. 

Cardoso-Vilhena, J.; Balaguer, L.; Daymond, J.; Barnes, J. D. (1998) Amelioration of ozone damage by elevated 

CO2. In: DeKok, L. J.; Stulen, I., eds. Responses of plant metabolism to air pollution and global change. 

Leiden, The Netherlands: Backhuys Publishers; pp. 281-282. 

Cardoso-Vilhena, J.; Barnes, J. (2001) Does nitrogen supply affect the response of wheat (Triticum aestivum cv 

Hanno) to the combination of elevated CO2 and O3? J. Exp. Bot. 52: 1901-1911. 

Carey, W. A.; Kelley, W. D. (1994) Interaction of ozone exposure and Fusarium subglutinans inocculation on 

growth and disease development of loblolly pine seedlings. Environ. Pollut. 84: 35-43. 

Carnahan, J. E.; Jenner, E. L.; Wat, E. K. W. (1978) Prevention of ozone injury to plants by a new protectant 

chemical. Phytopathology 68: 1225-1229. 

Carpenter, R. A.; Dixon, J. A. (1985) Ecology meets economics: a guide to sustainable development. Environment 

27: 6-32.

Carroll, J. J.; Miller, P. R.; Pronos, J. (2003) Historical perspectives on ambient ozone and its effects on the Sierra 

Nevada. In: Bytnerowicz, A.; Arbaugh, M. J.; Alonso, R., eds. Ozone air pollution in the Sierra Nevada: 

distribution and effects on forests; pp. distribution and effects on forests; pp. 33-54. 

Castillo, F. J.; Greppin, H. (1986) Balance between anionic and cationic extracellular peroxidase activities in sedum 

album leaves after ozone exposure Analysis by high-performance liquid chromatography. Physiol. Plant. 68: 

201-208. 

Castillo, F. J.; Heath, R. L. (1990) CA2+ transport in membrane vesicles from pinto bean leaves and its alteration 

after ozone exposure. Plant Physiol. 94: 788-795. 

Castillo, F. J.; Penel, C.; Greppin, H. (1984) Peroxidase release induced by ozone in sedum album leaves: 

involvement of Ca2+. Plant Physiol. 74: 846-851. 

Castillo, F. J.; Miller, P. R.; Greppin, H. (1987) Extracellular biochemical markers of photochemical oxidant air 

pollution damage to Norway spruce. Experientia (Basel) 43: 111-115. 

Castle, E. N. (1993) A pluralistic, pragmatic and evolutionary approach to natural resource management. For. Ecol. 

Manage. 56: 279-295. 

Cazalé, A. C.; Rouet-Mayer, M. A.; Barbier-Brygoo, H.; Mathieu, Y.; Laurière, C. (1998) Oxidatiave burst and 

hypoosmotic stress in tobacco cell suspensions. Plant Physiol. 116: 659-669. 

Chappelka, A. H. (2002) Reproductive development of blackberry (Rubus cuneifolius) as influenced by ozone. 

New Phytol. 155: 249-255. 

AX9-344
Chappelka, A. H., III; Chevone, B. I. (1986) White ash seedling growth response to ozone and simulated acid rain. 

Can. J. For. Res. 16: 786-790. 

Chappelka, A. H.; Freer-Smith, P. H. (1995) Predisposition of trees by air pollutants to low temperatures and 

moisture stress. Environ. Pollut. 87: 105-117. 

Chappelka, A. H.; Samuelson, L. J. (1998) Ambient ozone effects on forest trees of the eastern United States: 

a review. New Phytol. 139: 91-108. 

Chappelka, A. H., III; Chevone, B. I.; Burk, T. E. (1985) Growth response of yellow-poplar (Liriodendron tulipifera 

L) seedlings to ozone, sulfur dioxide, and simulated acidic precipitation, alone and in combination. Environ. 

Exp. Bot. 25: 233-244. 

Chappelka, A.; Renfro, J.; Somers, G.; Nash, B. (1997) Evaluation of ozone injury on foliage of black cherry 

(Prunus serotina) and tall milkweed (Asclepias exaltata) in Great Smoky Mountains National Park. Environ. 

Pollut. 95: 13-18. 

Chappelka, A.; Skelly, J.; Somers, G.; Renfro, J.; Hildebrand, E. (1999a) Mature black cherry used as a bioindicator 

of ozone injury. Water Air Soil Pollut. 116: 261-266. 

Chappelka, A.; Somers, G.; Renfro, J. (1999b) Visible ozone injury on forest trees in Great Smoky Mountains 

National Park, USA. Water Air Soil Pollut. 116: 255-260. 

Chappelka, A. H.; Neufeld, H. S.; Davison, A. W.; Somers, G. L.; Renfro, J. R. (2003) Ozone injury on cutleaf 

coneflower (Rudbeckia laciniata) and crown-beard (Verbesina occidentalis) in Great Smoky Mountains 

National Park. Environ. Pollut. 125: 53-60. 

Chen, C. W.; Tsai, W. T.; Gomez, L. E. (1994) Modeling responses of Ponderosa pine to interacting stresses of 

ozone and drought. For. Sci. 40: 267-288. 

Chen, C. W.; Tsai, W. T.; Lucier, A. A. (1998) A model of air-tree-soil system for ozone impact analysis. Ecol. 

Modell. 111: 207-222. 

Cheong, Y. H.; Chang, H. S.; Gupta, R.; Wang, X.; Zhu, T.; Luan, S. (2002) Transcriptional profiling reveals novel 

interactions between wounding, pathogen, abiotic stress, and hormonal responses in arabidopsis. Plant 

Physiol. 129: 661-677. 

Chernikova, T.; Robinson, J. M.; Lee, E. H.; Mulchi, C. L. (2000) Ozone tolerance and antioxidant enzyme activity 

in soybean cultivars. Photosynth. Res. 64: 15-26. 

Chevrier, N.; Chung, Y. S.; Sarhan, F. (1990) Oxidative damages and repair in Euglena gracilis exposed to ozone: II 

membrane permeability and uptake of metabolites. Plant Cell Physiol. 31: 987-992. 

Clark, C. S.; Weber, J. A.; Lee, E. H.; Hogsett, W. E. (1995) Accentuation of gas exchange gradients in flushes of 

ponderosa pine exposed to ozone. Tree Physiol. 15: 181-189. 

Clarke, B. B.; Henninger, M. R.; Brennan, E. (1983) An assessment of potato losses caused by oxidant air pollution 

in New Jersey. Phytopathology 73: 104-108. 

Cobb, F. W., Jr.; Wood, D. L.; Stark, R. W.; Parmeter, J. R., Jr. (1968) Photochemical oxidant injury and bark beetle 

(Coleoptera: Scolytidae) infestation of ponderosa pine IV Theory on the relationships between oxidant injury 

and bark beetle infestation. Hilgardia 39: 141-152. 

Coe, H.; Gallagher, M. W.; Choularton, T. W.; Dore, C. (1995) Canopy scale measurements of stomatal and 

cuticular O3 uptake by Sitka spruce. Atmos. Environ. 29: 1413-1423. 

Coleman, J. S.; Jones, C. G.; Smith, W. H. (1987) The effect of ozone on cottonwood - leaf rust interactions: 

independence of abiotic stress, genotype, and leaf ontogeny. Can. J. Bot. 65: 949-953. 

Coleman, M. D.; Dickson, R. E.; Isebrands, J. G.; Karnosky, D. F. (1995a) Carbon allocation and partitioning in 

aspen clones varying in sensitivity to tropospheric ozone. Tree Physiol. 15: 593-604. 

Coleman, M. D.; Isebrands, J. G.; Dickson, R. E.; Karnosky, D. F. (1995b) Photosynthetic productivity of aspen 

clones varying in sensitivity to tropospheric ozone. Tree Physiol. 15: 585-592. 

Coleman, M. D.; Dickson, R. E.; Isebrands, J. G.; Karnosky, D. F. (1996) Root growth physiology of potted and 

field-grown trembling aspen exposed to tropospheric ozone. Tree Physiol. 16: 145-152. 

Colls, J. J.; Unsworth, M. H. (1992) Air pollution interactions with natural stressors. In: Barker, J. R.; Tingey, D. T., 

eds. Air pollution effects on biodiversity. New York, NY: Van Nostrand Reinhold; pp. 93-108. 

Common, M.; Perrings, C. (1992) Towards an ecological economics of sustainability. Ecol. Econ. 6: 7-34. 

Conklin, P. L.; Barth, C. (2004) Ascorbic acid, a familiar small molecule intertwined in the response of plants to 

ozone, pathogens, and the onset of senescence. Plant Cell Environ. 27: 959-970. 

Conklin, P. L.; Last, R. L. (1995) Differential accumulation of antioxidant mRNAs in Arabidopsis thaliana exposed 

to ozone. Plant Physiol. 109: 203-212. 

Conklin, P. L.; Williams, E. H.; Last, R. L. (1996) Environmental stress sensitivity of an ascorbic acid-deficient 

Arabidopsis mutant. Proc. Natl. Acad. Sci. U. S. A. 93: 9970-9974. 

AX9-345
Constable, J. V. H.; Friend, A. L. (2000) Suitability of process-based tree growth models for addressing tree 

response to climate change. Environ. Pollut. 110: 47-59. 

Constable, J. V. H.; Retzlaff, W. A. (1997) Simulating the response of mature yellow poplar and loblolly pine trees 

to shifts in peak ozone periods during the growing season using the TREGRO model. Tree Physiol. 

17: 627-635.

Constable, J. V. H.; Taylor, G. E., Jr. (1997) Modeling the effects of elevated tropospheric O3 on two varieties of 

Pinus ponderosa. Can. J. For. Res. 27: 527-537. 

Constable, J. V. H.; Taylor, G. E., Jr.; Laurence, J. A.; Weber, J. A. (1996) Climatic change effects on the 

physiology and growth of Pinus ponderosa: expectations from simulation modeling. Can. J. For. Res. 26: 

1315-1325. 

Cook, E. (1990) A conceptual linear aggregate model for tree rings. In: Cook, E.; Kairiukstis, A., eds. Methods of 

dendrochronology: applications in the environmental sciences. Dordrecht, The Netherlands: Kluwer 

Academic Publ.; pp. 98-104. 

Cooley, D. R.; Manning, W. J. (1987) The impact of ozone on assimilate partitioning in plants: a review. Environ. 

Pollut. 47: 95-113. 

Coruzzi, G. M.; Zhou, L. (2001) Carbon and nitrogen sensing and signaling in plants: emerging 'matrix effects'. 

Curr. Opin. Plant Biol. 4: 247-253. 

Costa, S. D.; Kennedy, G. G.; Heagle, A. S. (2001) Effect of host plant ozone stress on Colorado potato beetles. 

Environ. Entomol. 30: 824-831. 

Costanza, R.; d'Arge, R.; De Groot, R.; Farber, S.; Grasso, M.; Hannon, B.; Limburg, K.; Naeem, S.; O'Neill, R. V.; 

Paruelo, J.; Raskin, R. G.; Sutton, P.; Van Den Belt, M. (1997) The value of the world's ecosystem services 

and natural capital. Nature (London) 387: 253-259. 

Coulston, J. W.; Smith, G. C.; Smith, W. D. (2003) Regional assessment of ozone sensitive tree species using 

bioindicator plants. Environ. Monit. Assess. 83: 113-127. 

Cox, R. M.; Malcolm, J. W. (1999) Passive ozone monitoring for forest health assessment. Water Air Soil Pollut. 

116: 339-344. 

Craigon, J.; Fangmeier, A.; Jones, M.; Donnelly, A.; Bindi, M.; De Temmerman, L.; Persson, K.; Ojanpera, K. 

(2002) Growth and marketable-yield responses of potato to increased CO2 and ozone. Eur. J. Agron. 

17: 273-289.

Crocker, T. D. (1985) On the value of the condition of a forest stock. Land Econ. 61: 244-254. 

Crocker, T. D.; Forster, B. A. (1986) Atmospheric deposition and forest decline. Water Air Soil Pollut. 31: 

1007-1017. 

Croft, P. C.; Jüttner, F.; Slusarenko, A. J. (1993) Volatile products of the lipoxygenase pathway evolved from 

Phaseolus vulgaris (L) leaves inoculated with pseudomonas syringae pv phaseolicola. Plant Physiol. 

101: 13-24. 

Curtis, P. S. (1996) A meta-analysis of leaf gas exchange and nitrogen in trees grown under elevated carbon dioxide. 

Plant Cell Environ. 19: 127-137. 

Dahlsten, D. L.; Rowney, D. L.; Kickert, R. N. (1997) Effects of oxidant air pollutants on western pine beetle 

(Coleoptera: Scolytidae) populations in southern California. Environ. Pollut. 96: 415-423. 

Daily, G. C. (1997) Nature's services: societal dependence on natural ecosystems. 

Daily, G. C. (1999) Developing a scientific basis for managing Earth's life support systems. Conserv. Ecol. 3: 14. 

Daily, G. C. (2000) Ecological assessment then and now. Sci. Technol. 16: 6-15. 

Dämmgen, U.; Grünhage, L.; Haenel, H. D.; Jäger, H. J. (1993) Climate and stress in ecotoxicology: a coherent 

system of definitions and terms. Angew Bot. 67: 157-162. 

Dämmgen, U.; Grünhage, L.; Kusters, A.; Jäger, H. J. (1994) Response of a grassland ecosystem to air pollutants:--II 

the chemical climate: fluxes of sedimenting airborne matter. Environ. Pollut. 85: 35-42. 

Danielsson, H.; Gelang, J.; Pleijel, H. (1999) Ozone sensitivity, growth and flower development in Phleum 

genotypes of different geographic origin in the Nordic countries. Environ. Exp. Bot. 42: 41-49. 

Danielsson, H.; Karlsson, G. P.; Karlsson, P. E.; Pleijel, J. (2003) Ozone uptake modelling and flux-response 

relationships–an assessment of ozone-induced yield loss in spring wheat. Atmos. Environ. 37: 475-485. 

Darrall, N. M. (1989) The effect of air pollutants on physiological processes in plants. Plant Cell Environ. 12: 1-30. 

Dass, H. C.; Weaver, G. M. (1968) Modification of ozone damage to Phaseolus vulgaris by antioxidants, thiols and 

sulfhydryl reagents. Can. J. Plant Sci. 48: 569-574. 

Dass, H. C.; Weaver, G. M. (1972) Enzymatic changes in intact leaves of Phaseolus vulgaris following ozone 

fumigation. Atmos. Environ. 6: 759-763. 

Davidson, A. (1993) Update of ozone trends in California's South Coast Air Basin. Air Waste 43: 226-227. 

AX9-346
Davis, D. D.; Skelly, J. M. (1992) Foliar sensitivity of eight eastern hardwood tree species to ozone. Water Air Soil 

Pollut. 62: 269-277. 

Davison, A. W.; Barnes, J. D. (1998) Effects of ozone on wild plants. New Phytol. 139: 135-151. 

Davison, A. W.; Reiling, K. (1995) A rapid change in ozone resistance of Plantago major after summers with high 

ozone concentrations. New Phytol. 131: 337-344. 

Davison, A. W.; Neufeld, H. S.; Chappelka, A. H.; Wolff, K. W.; Finkelstein, P. L. (2003) Interpreting spatial 

variation in ozone symptoms shown by cutleaf cone flower, Rudbeckia laciniata L. Environ. Pollut. 

125: 61-70. 

De Bauer, L. I. (1972) Uso de plantas indicadoras en la determinacion de aeropolutos en la Ciudad de México 

[Use of indicative plants to determine air pollution in Mexico city]. Agrociencia 9: 139-141. 

De Bauer, L. I.; Hernández-Tejeda, T. (1986) Contaminación: una amenaza para la vegetación en México. 

Chapingo, Mexico: Colegio de Postgraduados. 

De Bauer, L. I.; Krupa, S. V. (1990) The Valley of Mexico: summary of observational studies of air quality and 

effects on vegetation. Environ. Pollut. 65: 109-118. 

De Kok, L. J.; Stulen, I. (1998) Responses of plant metabolism to air pollution and global change. Leiden, 

The Netherlands: Backhuys Publishers. 

De Kok, L. J.; Tausz, M. (2001) The role of glutathione in plant reaction and adaptation to air pollutants. In: Grill, 

D.; Tausz, M.; De Kok, L. J., eds. Significance of glutathione in plant adaptation to the environment. 

Amsterdam, The Netherlands: Kluwer Publishers; pp. 185-208. 

De Ruiter, P. C.; Neutel, A. M.; Moore, J. C. (1998) Biodiversity in soil ecosystems: the role of energy flow and 

community stability. Appl. Soil Ecol. 10: 217-228. 

De Santis, F. (1999) New directions: will a new European vegetation ozone standard be fair to all European 

countries? Atmos. Environ. 33: 3873-3874. 

De Santis, F. (2000) Reply to 'Use and abuse of the AOT40 concept' by M Ashmore and J Fuhrer [letter]. Atmos. 

Environ. 34: 1158-1159. 

De Temmerman, L.; Hacour, A.; Guns, M. (2002a) Changing climate and potential impacts on potato yield and 

quality 'CHIP': introduction, aims and methodology. Eur. J. Agron. 17: 233-242. 

De Temmerman, L.; Wolf, J.; Colls, J.; Bindi, M.; Fangmeier, A.; Finnan, J.; Ojanperä, K.; Pleijel, H. (2002b) Effect 

of climatic conditions on tuber yield (Solanum tuberosum L) in the European 'CHIP' experiments. Eur. J. 

Agron. 17: 243-255. 

DeCaria, A. J.; Pickering, K. E.; Stenchikov, G. L.; Scala, J. R.; Stith, J. L.; Dye, J. E.; Ridley, B. A.; Laroche, P. 

(2000) A cloud-scale model study of lightning-generate NOx in an individual thunderstorm during 

STERAO-A. J. Geophys. Res. [Atmos. ] 105: 11601-11616. 

Degousée, N.; Triantaphylidès, C.; Montillet, J. L. (1994) Involvement of oxidative processes in the signaling 

mechanisms leading to the activation of glyceollin synthesis in soybean (Glycine max). Plant Physiol. 

104: 945-952. 

DeLucia, E. H.; Hamilton, J. G.; Naidu, S. L.; Thomas, R. B.; Andrews, J. A.; Finzi, A.; Lavine, M.; Matamala, R.; 

Mohan, J. E.; Hendrey, G. R.; Schlesinger, W. H. (1999) Net primary production of a forest ecosystem with 

experimental CO2 enrichment. Science (Washington, DC) 284: 1177-1179. 

Desikan, R.; Mackerness, S.; Hancock, J. T.; Neill, S. J. (2001) Regulation of the Arabidopsis transcriptome by 

oxidative stress. Plant Physiol. 127: 159-172. 

Diaz, G.; Barrantes, O.; Honrubia, M.; Gracia, C. (1996) Effect of ozone and sulphur dioxide on mycorrhizae of 

Pinus halepensis Miller. Ann. Sci. For. 53: 849-856. 

Dickson, R. E.; Isebrands, J. G. (1991) Leaves as regulators of stress response. In: Mooney, H. A.; Winner, W. E.; 

Pell, E. J., eds. Response of plants to multiple stresses. Academic Press, Inc.; pp. 3-34. 

Dickson, R. E.; Coleman, M. D.; Reimenschneider, D. E.; Isebrands, J. G.; Hogan, G. D.; Karnosky, D. F. (1998) 

Growth of five hybrid poplar genotypes exposed to interacting elevated CO2 and O3. Can. J. For. Res. 

28: 1706-1716.

Dickson, R. E.; Coleman, M. D.; Pechter, P.; Karnosky, D. (2001) Growth and crown architecture of two aspen 

genotypes exposed to interacting ozone and carbon dioxide. Environ. Pollut. 115: 319-334. 

Dietz, K. J. (1997) Functions and responses of the leaf apoplast under stress. Prog. Bot. 58: 221-254. 

Dighton, J.; Jansen, A. E. (1991) Atmospheric pollutants and ectomycorrhizae: more questions than answers? 

Environ. Pollut. 73: 179-204. 

Dijak, M.; Ormrod, D. P. (1982) Some physiological and anatomical characteristics associated with differential 

ozone sensitivity among pea cultivars. Environ. Exp. Bot. 22: 395-402. 

AX9-347
Dixon, B. L.; Segerson, K. (1999) Impacts of increased climate variability on the profitability of Midwest 

agriculture. J. Agric. Appl. Econ. 31: 537-550. 

Dixon, M.; Le Thiec, D.; Garrec, J. P. (1998) Reactions of Norway spruce and beech trees to 2 years of ozone 

exposure and episodic drought. Environ. Exp. Bot. 40: 77-91. 

Docherty, M.; Salt, D. T.; Holopainen, J. K. (1997) The impacts of climate change and pollution on forest pests. In: 

Watt, A. D.; Stork, N. E.; Hunter, M. D., eds. Forests and insects. Chapman and Hall; pp. 229-247. 

Dominy, P. J.; Heath, R. L. (1985) Inhibition of the K+

-stimulated ATpase of the plasmalemma of pinto bean leaves 

by ozone. Plant Physiol. 77: 43-45. 

Donnelly, A.; Jones, M. B.; Schneiders, B. J.; Burke, J. I. (1998) The interactive effects of CO2, O3 and nitrogen on 

the photosynthetic response of juvenile wheat plants. In: De Kok, L. J.; Stulen, I., eds. Responses of plant 

metabolism to air pollution and global change. Backhuys Publishers; pp. 291-293. 

Donnelly, A.; Jones, M. B.; Burke, J. I.; Schnieders, B. (1999) Does elevated CO2 protect grain yield of wheat from 

the effects of ozone stress? Z. Naturforsch. C: J. Biosci. 54: 802-811.: J-811. 

Donnelly, A.; Jones, M. B.; Burke, J. I.; Schneiders, B. (2000) Elevated CO2 provides protection from O3 induced 

photosynthetic damage and chlorophyll loss in flag leaves of spring wheat (Triticum aestivum L, cv'Minaret'). 

Agric. Ecosyst. Environ. 80: 159-168. 

Donnelly, A.; Craigon, J.; Black, C. R.; Colls, J. J.; Landon, G. (2001a) Does elevated CO2 ameliorate the impact of 

O3 on the chlorophyll content and photosynthesis in potato (Solanum tuberosum)? Physiol. Plant. 

111: 501-511. 

Donnelly, A.; Craigon, J.; Black, C. R.; Colls, J. J.; Landon, G. (2001b) Elevated CO2 increases biomass and tuber 

yield in potato even at high ozone concentrations. New Phytol. 149: 265-274. 

Donnelly, A.; Lawson, T.; Craigon, J.; Black, C. R.; Coils, J. J.; Landon, G. (2001b) Effects of elevated CO2 and O3 

on tuber quality in potato (Solanum tuberosum L). Agric. Ecosyst. Environ. 87: 273-285. 

Dowding, P. (1988) Air pollutant effects on plant pathogens. In: Schulte-Hostede, S.; Darrall, N. M.; Blank, L. W.; 

Wellburn, A. R., eds. Air pollution and plant metabolism. London, United Kingdom: Elsevier Applied 

Scientific; pp. 329-355. 

Drogoudi, P. D.; Ashmore, M. R. (2000) Does elevated ozone have differing effects in flowering and deblossomed 

strawberry? New Phytol. 147: 561-569. 

Drogoudi, P. D.; Ashmore, M. (2001) 14C-allocation of flowering and deblossomed strawberry in response to 

elevated ozone. New Phytol. 152: 455-461. 

Duchelle, S. F.; Skelly, J. M. (1981) Response of common milkweed to oxidant air pollution in the Shenandoah 

National Park in Virginia. Plant Dis. 65: 661-663. 

Duchelle, S. F.; Skelly, J. M.; Chevone, B. I. (1982) Oxidant effects on forest tree seedling growth in the 

Appalachian mountains. Water Air Soil Pollut. 18: 363-373. 

Duchelle, S. F.; Skelly, J. M.; Sharick, T. L.; Chevone, B. I.; Yang, Y. S.; Nellessen, J. E. (1983) Effects of ozone on 

the productivity of natural vegetation in a high meadow of the Shenandoah National Park of Virginia. J. 

Environ. Manage. 17: 299-308. 

Dueck, T.; Zuin, A.; Elderson, J. (1998) Influence of ammonia and ozone on growth and drought sensitivity of 

Pinus sylvestris. Atmos. Environ. 32: 545-550. 

Dugger, W. M.; Ting, I. P. (1970) Air pollution oxidants–their effects on metabolic processes in plants. Ann. Rev. 

Plant Physiol. 21: 215-234. 

Dunajski, A. (2002) Mountain forest vegetation dynamics after huge forest dieback: landscape level approach in the 

Polish part of Karkonosze Mountains. In: Widacki, W.; Bytnerowicz, A.; Riebau, A., eds. A message from 

the Tatra–geographical information systems and remote sensing in mountain environmental research. 

Krakow, Poland: Jagiellonian University Press, 233 pp. 

Dunn, D. B. (1959) Some effects of air pollution on Lupinus in the Los Angeles area. Ecology 40: 621-625. 

Eamus, D. (1996) Responses of field grown trees to CO2 enrichment. Commonw. For. Rev. 75: 39-47. 

Eamus, D.; Murray, M. (1991) Photosynthetic and stomatal conductance responses of Norway spruce and beech to 

ozone, acid mist and frost - a conceptual model. Environ. Pollut. 72: 23-44. 

Eckey-Kaltenbach, H.; Ernst, D.; Heller, W.; Sandermann, H., Jr. (1994a) Biochemical responses to ozone IV 

Cross-induction of defensive pathways in parsley (Petroselinum crispum L) plants. Plant Physiol. 104: 67-74. 

Eckey-Kaltenbach, H.; Grosskopf, E.; Sandermann, H., Jr.; Ernst, D. (1994b) Induction of pathogen defence genes 

in parsley (Petroselinum crispum L) plants by ozone. Proc. R. Soc. Edinburgh Sect. B: Biol. Sci. 102: 63-74.: 

Biol-74. 

AX9-348
Edinger, J. G.; McCutchan, M. H.; Miller, P. R.; Ryan, B. C.; Schroeder, M. J.; Behar, J. V. (1972) Penetration and 

duration of oxidant air pollution in the South Coast Air Basin of California. J. Air Pollut. Control Assoc. 22: 

882-886. 

Edwards, N. T. (1991) Root and soil respiration responses to ozone in Pinus taeda L seedlings. New Phytol. 

118: 315-321. 

Edwards, G. S.; Kelly, J. M. (1992) Ectomycorrhizal colonization of loblolly pine seedlings during three growing 

seasons in response to ozone, acidic precipitation, and soil Mg status. Environ. Pollut. 76: 71-77. 

Edwards, G. S.; Wullschleger, S. D.; Kelly, J. M. (1994) Growth and physiology of northern red oak: preliminary 

comparisons of mature tree and seedling responses to ozone. Environ. Pollut. 83: 215-221. 

Einig, W.; Lauxmann, U.; Hauch, B.; Hampp, R.; Landolt, W.; Maurer, S.; Matyssek, R. (1997) Ozone-induced 

accumulation of carbohydrates changes enzyme activities of carbohydrate metabolism in birch leaves. 

New Phytol. 137: 673-680. 

Elagöz, V.; Manning, W. J. (2002) Bean plants and ozone: morphology matters. Environ. Pollut. 120: 521-524. 

Elagöz, V.; Manning, W. J. (2005a) Factors affecting the effects of EDU on growth and yield of field-grown bush 

beans (Phaseolus vulgaris L), with varying degrees of sensitivity to ozone. Environ. Pollut. 136: 385-395. 

Elagöz, V.; Manning, W. J. (2005b) Responses of sensitive and tolerant bush beans (Phaseolus vulgaris L.) to ozone 

in open-top chambers are influenced by phenotypic differences, morphological characteristics, and the 

chamber environment. Environ. Pollut. 136: 371-383. 

Elkiey, T.; Ormrod, D. P. (1979) Leaf diffusion resistance responses of three petunia cultivars to ozone and/or sulfur 

dioxide. J. Air Pollut. Control Assoc. 29: 622-625. 

Elkiey, T.; Ormrod, D. P. (1981) Sorption of O3, SO2, NO2 or their mixture by nine Poa pratensis cultivars of 

differing pollutant sensitivity. Atmos. Environ. 15: 1739-1743. 

Ellsworth, D. S.; Reich, P. B.; Naumburg, E. S.; Koch, G. W.; Kubiske, M. E.; Smith, S. D. (2004) Photosynthesis, 

carboxylation and leaf nitrogen response of 16 species to elevated pCO2 across four free-air CO2 enrichment 

experiments in forest, grassland and desert. Global Change Biol. 10: 2121-2138.. Global Change Biol. 10: 

2121-2138. 

Elvira, S.; Bermejo, V.; Manrique, E.; Gimeno, B. S. (2004) On the response of two populations of Quercus 

coccifera to ozone and its relationship with ozone uptake. Atmos. Environ. 38: 2305-2311. 

Emberson, L.; Ashmore, M. R.; Cambridge, H. M.; Simpson, D.; Tuovinen, J. P. (2000a) Modelling stomatal ozone 

flux across Europe. Environ. Pollut. 109: 403-413. 

Emberson, L. D.; Wieser, G.; Ashmore, M. R. (2000b) Modelling of stomatal conductance and ozone flux of 

Norway spruce: comparison with field data. Environ. Pollut. 109: 393-402. 

Enders, G. (1992) Deposition of ozone to a mature spruce forest: measurements and comparison to models. Environ. 

Pollut. 75: 61-67. 

Engle, R. L.; Gabelman, W. H. (1966) Inheritance and mechanism for resistance to ozone damage in onion, Allium 

cepa L. Proc. Am. Soc. Hortic. Sci. 89: 423-430. 

Ennis, C. A.; Lazrus, A. L.; Kok, G. L.; Zimmerman, P. R.; Monson, R. K. (1990) A branch chamber system and 

techniques for simultaneous pollutant exposure experiments and gaseous flux determinations. Tellus Ser. B 

42: 170-182.

Ernst, D.; Schraudner, M.; Langebartels, C.; Sandermann, H. (1992) Ozone-induced changes of messenger RNA 

levels of beta-1,3-glucanase, chitinase and pathogenesis-related protein 1B in tobacco plants. Plant. Mol. Biol. 

20: 673-682.

Espelie, K. E.; Franceschi, V. R.; Kolattukudy, P. E. (1986) Immunocytochemical localization and time course of 

appearance of an anionic peroxidase associated with suberization in wound-healing potato tuber tissue. Plant 

Physiol. 81: 487-492. 

Estes, B. L.; Enebak, S. A.; Chappalka, A. H. (2004) Loblolly pine seedling growth after inoculation with plant 

growth-promoting rhizobacteria and ozone exposure. Can. J. For. Res. 34: 1410-1416. 

Evans, P. A.; Ashmore, M. R. (1992) The effects of ambient air on a semi-natural grassland community. Agric. 

Ecosyst. Environ. 38: 91-97. 

Ewart, F.; Pleijel, H. (1999) Phenological development, leaf emergence, tillering and leaf area index, and duration of 

spring wheat across Europe in response to CO2 and ozone. Eur. J. Agron. 10: 171-184. 

Ewart, F.; Van Oijen, M.; Porter, J. R. (1999) Simulation of growth and development processes of spring wheat in 

response to CO2 and ozone for different sites and years in Europe using mechanistic crop simulation models. 

Eur. J. Agron. 10: 231-247. 

Fagnano, M.; Merola, G.; Forlani, A.; Postiglione, L.; Fuhrer, J. (2004) Monitoring yield loss from ozone pollution 

in a Mediterranean environment: a comparison of methods. Water Air Soil Pollut. 155: 383-398. 

AX9-349
Fangmeier, A.; Stein, W.; Jäger, H.-J. (1992) Advantages of an open-top chamber plant exposure system to assess 

the impact of atmospheric trace gases on vegetation. Angew Bot. 66: 97-105. 

Fangmeier, A.; Brockerhoff, U.; Grüters, U.; Jäger, H. J. (1994a) Growth and yield responses of spring wheat 

(Triticum aestivum L cv Turbo) grown in open-top chambers to ozone and water stress. Environ. Pollut. 

83: 317-325.

Fangmeier, A.; Brunschön, S.; Jäger, H. J. (1994b) Time course of oxidant stress biomarkers in flag leaves of wheat 

exposed to ozone and drought stress. New Phytol. 126: 63-69. 

Fangmeier, A.; Grüters, U.; Vermehren, B.; Jäger, H. J. (1996) Responses of some cereal cultivars to CO2 

enrichment and tropospheric ozone at different levels of nitrogen supply. J. Appl. Bot. 70: 12-18. 

Fangmeier, A.; De Temmerman, L.; Mortensen, L.; Kemp, K.; Burke, J.; Mitchell, R.; Van Oijen, M.; Weigel, H. J. 

(1999) Effects on nutrients and on grain quality in spring wheat crops grown under elevated CO2 

concentrations and stress conditions in the European, multiple-site experiment 'ESPACE-wheat'. Eur. J. 

Agron. 10: 215-229. 

Fangmeier, A.; De Temmerman, L.; Black, C.; Persson, K.; Vorne, V. (2002) Effects of elevated CO2 and/or ozone 

on nutrient concentrations and nutrient uptake of potatoes. Eur. J. Agron. 17: 353-368. 

Fangmeier, A.; Bender, J.; Weigel, H. J.; Jäger, H. J. (2002) Effects of pollutant mixtures. In: Bell, J. N. B.; 

Treshow, M., eds. Air Pollution and Plant Life. 2nd ed. Chichester, United Kingdom: John Wiley & Sons 

Ltd.; pp. 251-272. 

Farage, P. K.; Long, S. P. (1999) The effects of O3 fumigation during leaf development on photosynthesis of wheat 

and pea: an in vivo analysis. Photosynth. Res. 59: 1-7. 

Farage, P. K.; Long, S. P.; Lechner, E. G.; Baker, N. R. (1991) The sequence of changes within the photosynthetic 

apparatus of wheat following short-term exposure to ozone. Plant Physiol. 95: 529-535. 

Farquhar, G. D.; Sharkey, T. D. (1982) Stomatal conductance and photosynthesis. Annu. Rev. Plant Physiol. 

33: 317-345.

Farquhar, G. D.; Caemmerer, S. V.; Berry, J. A. (1980) A biochemical model of photosynthetic CO2 assimilation in 

leaves of C3 species. Planta 149: 78-90. 

Federal Land Manager's Air Quality Related Values Workgroup (FLAG) (2000) Phase I report. Denver, CO: 

National Park Service, Air Resources Division; Available: http://www2.nature.nps.gov/ard/flagfree/ [11 

March, 2003]. 

Federal Register. (1997) National ambient air quality standards for ozone; final rule. F.R. 62: 38855-38896. 

Fenn, M. E.; De Bauer, L. I. (1999) Nitogen and sulfur deposition and forest nutrient status in the Valley of Mexico. 

Water Air Soil Pollut. 113: 155-174. 

Fenn, M. E.; Dunn, P. H. (1989) Litter decomposition across an air-pollution gradient in the San Bernardino 

Mountains. Soil Sci. Soc. Am. J. 53: 1560-1567. 

Fenn, M. E.; Poth, M. A. (1998) Indicators of nitrogen status in California forests. U. S. D. A. Forest Service Gen. 

Tech. Rep. PSW-GTR-166. 

Fenn, M. E.; Dunn, R. H.; Wilborn, R. (1990) Black stain root disease in ozone-stressed ponderosa pine. Plant Dis. 

74: 426-430.

Fenn, M. E.; Poth, M. A.; Johnson, D. W. (1996) Evidence for nitrogen saturation in the San Bernardino Mountains 

in southern California. For. Ecol. Manage. 82: 211-230. 

Fenn, M. E.; Poth, M. A.; Aber, J. D.; Baron, J. S.; Bormann, B. T.; Johnson, D. W.; Lemly, A. D.; McNulty, S. G.; 

Ryan, D. F.; Stottlemyer, R. (1998) Nitrogen excess in North American ecosystems: predisposing factors, 

ecosystem responses, and management strategies. Ecol. Appl. 8: 706-733. 

Fenn, M. E.; Poth, M. A.; Schilling, S. L.; Grainger, D. B. (2000) Throughfall and fog deposition of nitrogen and 

sulfur at an N-limited and N-saturated site in the San Bernardino Mountains, southern California. Can. J. For. 

Res. 30: 1476-1488. 

Fenn, M. E.; De Bauer, L. I.; Hernández-Tejeda, T. (2002) Urban air pollution and forests: resources at risk in the 

Mexico City air basin. New York, NY: Springer-Verlag (Ecological Studies Vol. 156). 

Fenn, M. A.; Haeuber, R.; Tonnesen, G. S.; Baron, J. S.; Grossman-Clarke, S.; Hope, D.; Jaffe, D. A.; Copeland, S.; 

Geiser, L.; Rueth, H. M.; Sickman, J. O. (2003a) Nitrogen emissions, deposition, and monitoring in the 

western United States. BioScience 53: 391-403. 

Fenn, M. E.; Baron, J. S.; Allen, E. B.; Rueth, H. M.; Nydick, K. R.; Geiser, L.; Bowman, W. D.; Sickman, J. O.; 

Meixner, T.; Johnson, D. W.; Neitlich, P. (2003b) Ecological effects of nitrogen deposition in the western 

United States. BioScience 53: 404-420. 

Ferdinand, J. A.; Fredericksen, T. S.; Kouteric, K. B.; Skelly, J. M. (2000) Leaf morphology and ozone sensitivity of 

two open pollinated genotypes of black cherry (Prunus serotina) seedlings. Environ. Pollut. 108: 297-302. 

AX9-350
Findlay, S.; Jones, C. G. (1990) Exposure of cottonwood plants to ozone alters subsequent decomposition. 

Oecologia 82: 248-250. 

Findlay, S.; Carreiro, M.; Jones, C. G. (1991) Ozone stress and mite damage to cottonwood plants alter subsequent 

litter decomposition: the afterlife effect. Bull. Ecol. Soc. Am. 72(suppl.): 113. 

Findlay, S.; Carreiro, M.; Krischik, V.; Jones, C. G. (1996) Effects of damage to living plants on leaf litter quality. 

Ecol. Appl. 6: 269-275. 

Findley, D. A.; Keever, G. J.; Chappelka, A. H.; Eakes, D. J.; Gillian, D. J. (1997) Differential responses of buddleia 

(Buddleia davidii Franch.) to ozone. Environ. Pollut. 98: 105-111. 

Finkelstein, R. R.; Gibson, S. I. (2002) ABA and sugar interactions regulating development: cross-talk or voices in a 

crowd? Curr. Opin. Plant Biol. 5: 26-32. 

Finnan, J. M.; Burke, J. L.; Jones, M. B. (1997) An evaluation of indices that describe the impact of ozone on the 

yield of spring wheat (Triticum aestivum L). Atmos. Environ. 31: 2685-2693. 

Finnan, J. M.; Donnelly, A.; Burke, J. I.; Jones, M. B. (2002) The effects of elevated concentrations of carbon 

dioxide and ozone on potato (Solanum tuberosum L) yield. Agric. Ecosyst. Environ. 88: 11-22. 

Finnan, J. M.; Jones, M. B.; Burke, J. I. (1996) A time-concentration study on the effects of ozone on spring wheat 

(Triticum aestivum L)1 Effects on yield. Agric. Ecosyst. Environ. 57: 159-167. 

Finnan, J. M.; Jones, M. B.; Burke, J. I. (1996) A time-concentration study on the effects of ozone on spring wheat 

(Triticum aestivum L) 2 A comparison of indices. Agric. Ecosyst. Environ. 57: 169-177. 

Fiscus, E. L.; Reid, C. D.; Miller, J. E.; Heagle, A. S. (1997) Elevated CO2 reduces O3 flux and O3-induced yield 

losses in soybeans: possible implications for elevated CO2 studies. J. Exp. Bot. 48: 307-313. 

Fiscus, E. L.; Miller, J. E.; Booker, F. L.; Heagle, A. S.; Reid, C. D. (2002) The impact of ozone and other 

limitations on the crop productivity response to CO2. Technology 8: 181-192. 

Fiscus, E. L.; Booker, F. L.; Burkey, K. O. (2005) Crop responses to ozone: uptake, modes of action, carbon 

assimilation and partitioning. Plant Cell Environ. 28: 997-1011. 

Fishman, J.; Watson, C. E.; Larsen, J. C.; Logan, J. A. (1990) Distribution of tropospheric ozone determined from 

satellite data. J. Geophys. Res. 95: 3617. 

Flagler, R. B. (1998) Recognition of air pollution injury to vegetation: A pictorial atlas; second edition. Pittsburgh, 

PA: Air & Waste Management Association. 

Flückiger, W.; Braun, S.; Hiltbrunner, E. (2002) Effects of air pollutants on biotic stress. In: Bell, J. N. B.; 

Treshow, M., eds. Air Pollution and Plant Life. John Wiley & Sons Ltd.; pp. 379-406. 

Fogel, R.; Cromack, K., Jr. (1977) Effect of habitat and substrate quality on Douglas fir litter decomposition in 

western Oregon. Can. J. Bot. 55: 1632-1640. 

Fontan, J. A.; Minga, A.; Lopez, A.; Druilhet, A. (1992) Vertical ozone profiles in a pine forest. Atmos. Environ. 

Part A 26A: 863-869. 

Foot, J. P.; Caporn, S. J. M.; Lee, J. A.; Ashenden, T. W. (1996) The effect of long-term ozone fumigation on the 

growth, physiology and frost sensitivity of Calluna vulgaris. New Phytol. 133: 505-511. 

Foot, J. P.; Caporn, S. J. M.; Lee, J. A.; Ashenden, T. W. (1997) Evidence that ozone exposure increases the 

susceptibility of plants to natural frosting episodes. New Phytol. 135: 369-374. 

Fortin, M.; Maufette, Y.; Albert, P. J. (1997) The effects of ozone-exposed sugar maple seedlings on the biological 

performance and the feeding preference of the forest tent catepillar (Malacosoma disstria Hbn). Environ. 

Pollut. 97: 303-309. 

Fowler, D.; Cape, J. N. (1982) Air pollutants in agriculture and horticulture. In: Unsworth, M. H.; Ormrod, D. P., 

eds. Effects of gaseous air pollution in agriculture and horticulture. London, United Kingdom: Butterworth 

Scientific; pp. 3-26. 

Fowler, D.; Cape, J. N.; Unsworth, M. H. (1989) Deposition of atmospheric pollutants on forests. Philos. Trans. R. 

Soc. London B 324: 247-265. 

Fox, S.; Mickler, R. A. (1996) Impact of air pollutants on southern pine forests. New York, NY: Springer-Verlag 

(Lange, O. L.; Mooney, H. A., eds. Ecological studies: v. 118). 

Franzaring, J.; Dueck, T. A.; Tonneijck, A. E. G. (2002) Response of two wet grassland species to ozone and water 

stress. In: Background paper: semi-natural vegetation, workshop establishing ozone critical levels II; 

November; Goteborg, Sweden. Goteborg, Sweden: Göteborg University, IVL Swedish Environmental 

Research Institute; pp. 44-49. 

Frederick, P. E.; Heath, R. L. (1970) Ozone-induced fatty acid and viability changes in chlorella. Plant Physiol. 

55: 15-19.

AX9-351 
Fredericksen, T. S.; Joyce, B. J.; Skelly, J. M.; Steiner, K. C.; Kolb, T. E.; Kouterick, K. B.; Savage, J. E.; Snyder, 

K. R. (1995) Physiology, morphology, and ozone uptake of leaves of black cherry seedlings, saplings, and 

canopy trees. Environ. Pollut. 89: 273-283. 

Fredericksen, T. S.; Kolb, T. E.; Skelly, J. M.; Steiner, K. C.; Joyce, B. J.; Savage, J. E. (1996a) Light environment 

alters ozone uptake per net photosynthetic rate in black cherry trees. Tree Physiol. 16: 485-490. 

Fredericksen, T. S.; Skelly, J. M.; Snyder, K. R.; Steiner, K. C. (1996b) Predicting ozone uptake from 

meteorological and environmental variables. J. Air Waste Manage. Assoc. 46: 464-469. 

Fredericksen, T. S.; Skelly, J. M.; Steiner, K. C.; KoIb, T. E.; Kouterick, K. B. (1996c) Size-mediated foliar 

response to ozone in black cherry trees. Environ. Pollut. 91: 53-63. 

Freer-Smith, P. H.; Taylor, G. (1992) Comparative evaluation of the effects of gaseous pollutants, acidic deposition 

and mineral deficiencies on gas exchange of trees. Agric. Ecosyst. Environ. 42: 321-332. 

Friedrich, L.; Moyer, M.; Ward, E.; Ryals, J. (1991) Pathogenesis-related protein 4 is structurally homologous to the 

carboxy-terminal domains of hevein, Win-1 and Win-2. Mol. Genet. Genom. 230: 113-119. 

Froehlich, J. E.; Itoh, A.; Howe, G. A. (2001) Tomato allene oxide synthase and fatty acid hydroperoxide lyase, two 

cytochrome P450s involved in oxylipin metabolism, are targeted to different membranes of chloroplast 

envelope. Plant Physiol. 125: 306-317. 

Fuentes, J. D.; Gillespie, T. J. (1992) A gas exchange system to study the effects of leaf surface wetness on the 

deposition of ozone. Atmos. Environ. Part A 26: 1165-1173. 

Fuentes, J. D.; Gillespie, T. J.; Den Hartog, G.; Neumann, H. H. (1992) Ozone deposition onto a deciduous forest 

during dry and wet conditions. Agric. For. Meteorol. 62: 1-18. 

Fuentes, J. D.; Hayden, B. P.; Garstang, M.; Lerdau, M.; Fitzjarrald, D.; Baldocchi, D. D.; Monson, R.; Lamb, B.; 

Geron, C. (2001) New directions: VOCs and biosphere-atmosphere feedbacks. Atmos. Environ. 35: 189-191. 

Fuhrer, J. H. (1994) Effects of ozone on managed pasture. 1. Effects of open-top chambers on microclimate, ozone 

flux and plant growth. Environ. Pollut. 86: 297-305. 

Fuhrer, J.; Grimm, A. G.; Tschannen, W.; Shariat-Madari, H. (1992) The response of spring wheat (Triticum 

aestivum L) to ozone at higher elevations: II changes in yield, yield components and grain quality in response 

to ozone flux. New Phytol. 121: 211-219. 

Fuhrer, J.; Shariat-Madari, H.; Perler, R.; Tschannen, W.; Grub, A. (1994) Effects of ozone on managed pasture: II 

Yield, species composition, canopy structure, and forage quality. Environ. Pollut. 86: 307-314. 

Fuhrer, J.; Skärby, L.; Ashmore, M. R. (1997) Critical levels for ozone effects on vegetation in Europe. Environ. 

Pollution 97: 91-106. 

Fujita, E. M.; Croes, B. E.; Bennett, C. L.; Lawson, D. R.; Lurmann, F. W.; Main, H. H. (1992) Comparison of 

emission inventory and ambient concentration ratios of CO, NMOG, and NOx in California's south coast air 

basin. J. Air Waste Manage. Assoc. 42: 264-276. 

Fumagalli, I.; Mignanego, L.; Violini, G. (1997) Effects of tropospheric ozone on white clover plants exposed in 

open-top chambers or protected by the antioxidant ethylene-diurea (EDU). Agronomie 17: 271-281. 

Fumagalli, I.; Ambrogi, R.; Mignanego, L. (2001) Yield responses of plants exposed to ambient ozone in the river 

Po valley (Italy). Agronomie 21: 227-233. 

Garcia, P.; Dixon, B.; Mjelde, J.; Adams, R. (1986) Measuring the benefits of environmental change using a duality 

approach: the case of ozone and Illinois cash grain farms. J. Environ. Econ. Manage. 13: 69-80. 

Garten, C. T., Jr.; Hanson, P. J. (1990) Foliar retention of 15N-nitrate and 15N-ammonium by red maple (Acer 

rubrum) and white oak (Quercus alba) leaves from simulated rain. Environ. Exp. Bot. 30: 333-342. 

Gasper, T.; Penel, C.; Castillo, F. J.; Greppin, H. (1985) A two step control of basic and acidic peroxidases and its 

significance for growth and development. Physiol. Plant. 64: 418-423. 

Gate, I. M.; McNeill, S.; Ashmore, M. R. (1995) Effects of air pollution on the searching behaviour of an insect 

parasitoid. Water Air Soil Pollut. 85: 1425-1430. 

Gates, D. M. (1968) Transpiration and leaf temperature. Annu. Rev. Plant Physiol. 19: 211-238. 

Gaucher, C.; Costanzo, N.; Afif, D.; Mauffette, Y.; Chevrier, N.; Dizengremel, P. (2003) The impact of elevated 

ozone and carbon dioxide on young Acer saccharum seedlings. Physiol. Plant. 117: 392-402. 

Gelang, J.; Pleijel, H.; Sild, E.; Danielsson, H.; Younis, S.; Selldén, G. (2000) Rate and duration of grain filling in 

relation to flag leaf senescence and grain yield in spring wheat (Triticum aestivum) exposed to different 

concentrations of ozone. Physiol. Plant. 110: 366-375. 

Gerant, D.; Podor, M.; Grieu, P.; Afif, D.; Cornu, S.; Morabito, D.; Banvoy, J.; Robin, C.; Dizengremel, P. (1996) 

Carbon metabolism enzyme activities and carbon partitioning in Pinus halepensis mill Exposed to mild 

drought and ozone. J. Plant Physiol. 148: 142-147. 

AX9-352
Gerosa, G.; Marzuoli, R.; Bussotti, F.; Pancrazi, M.; Ballarin-Denti, A. (2003) Ozone sensitivity of Fagus sylvatica 

and Fraxinus excelsior young trees in relation to leaf structure and foliar ozone uptake. Environ. Pollut. 125: 

91-98. 

Gerosa, G.; Marzuoli, S.; Cieslik, S.; Ballarin-Denti, A. (2004) Stomatal ozone fluxes over a barley field in Italy 

"Effective exposure" as a possible link between exposure- and flux-based approaches. Atmos. Environ. 

38: 2421-2432.

Ghosh, S.; Skelly, J. M.; Innes, J. L.; Skelly, L. (1998) Temporal development of visual injury on the foliage of 

Prunus serotina–a statistical evaluation. Environ. Pollut. 102: 287-300. 

Gifford, R. M.; Morison, J. I. L. (1993) Crop responses to the global increase in atmospheric carbon dioxide 

concentration. In: Buxton, D. R., ed. International crop science I. Madison, WI: Crop Science Society of 

America; pp. 325-331. 

Gillespie, C. T.; Winner, W. E. (1989) Development of lines of radish differing in resistance to O3 and SO2. 

New Phytol. 112: 353-361. 

Gimeno, B. S.; Bermejo, V.; Reinert, R. A.; Zheng, Y.; Barnes, J. D. (1999) Adverse effects of ambient ozone on 

watermelon yield and physiology at a rural site in eastern Spain. New Phytol. 144: 245-260. 

Goldstein, A. H.; Kurpius, M. R.; Panek, J. A. (2003) Ozone uptake by ponderosa pine in the Sierra Nevada: a 

measurement perspective. In: Bytnerowicz, A.; Arbaugh, M. J.; Alonso, R., eds. Ozone air pollution in the 

Sierra Nevada: distribution and effects on forests; pp. distribution and effects on forests; pp. 83-109. 

Goodyear, S. N.; Ormrod, D. P. (1988) Tomato response to concurrent and sequential NO2 and O3 exposures. 

Environ. Pollut. 51: 315-326. 

Gordon, J. C.; Larson, P. R. (1970) Redistribution of 14C-labeled reserve food in young red pines during shoot 

elongation. For. Sci. 16: 14-20. 

Gorissen, A.; Van Veen, J. A. (1988) Temporary disturbance of translocation of assimilates in Douglas firs caused 

by low levels of ozone and sulfur dioxide. Plant Physiol. 88: 559-563. 

Gorissen, A.; Joosten, N. N.; Jansen, A. E. (1991a) Effects of ozone and ammonium sulphate on carbon partitioning 

to mycorrhizal roots of juvenile Douglas fir. New Phytol. 119: 243-250. 

Gorissen, A.; Schelling, G. C.; Van Veen, J. A. (1991b) Concentration-dependent effects of ozone on translocation 

of assimilates in Douglas fir. J. Environ. Qual. 20: 169-173. 

Gorissen, A.; Joosten, N. N.; Smeulders, S. M.; Van Veen, J. A. (1994) Effects of short-term ozone exposure and 

soil water availability on the carbon economy of juvenile Douglas-fir. Tree Physiol. 14: 647-657. 

Goulder, L. H.; Kennedy, D. (1997) Valuing ecosystem services: philosophical bases and empirical methods. In: 

Daily, G. C., ed. Nature's services: societal dependence on natural ecosystems. Washington, DC: Island Press; 

pp. 23-47. 

Gower, S. T.; Reich, P. B.; Son, Y. (1993) Canopy dynamics and aboveground production of five tree species with 

different leaf longevities. Tree Physiol. 12: 327-345. 

Grams, T. E. E.; Anegg, S.; Haberle, K. H.; Langebartels, C.; Matyssek, R. (1999) Interactions of chronic exposure 

to elevated CO2 and O3 levels in the phytosynthetic light and dark reactions of European beech 

(Fagus sylvatica). New Phytol. 144: 95-107. 

Grantz, D. A.; Farrar, J. F. (1999) Acute exposure to ozone inhibits rapid carbon translocation from source leaves of 

Pima cotton. J. Exp. Bot. 50: 1253-1262. 

Grantz, D. A.; Farrar, J. F. (2000) Ozone inhibits phloem loading from a transport pool: compartmental efflux 

analysis in Pima cotton. Aust. J. Plant Physiol. 27: 859-868. 

Grantz, D. A.; Yang, S. (2000) Ozone impacts on allometry and root hydraulic conductance are not mediated by 

source limitation nor developmental age. J. Exp. Bot. 51: 919-927. 

Grantz, D. A.; Zhang, X. J.; Massman, W. J.; Den Hartog, G.; Neumann, H. H.; Pederson, J. R. (1995) Effects of 

stomatal conductance and surface wetness on ozone deposition in field-grown grape. Atmos. Environ. 

29: 3189-3198.

Grantz, D. A.; Zhang, X. J.; Massman, W.; Delany, A.; Pederson, R. (1997) Ozone deposition to a cotton 

(Gossypium hirsutum L) field: stomatal and surface wetness effects during the California Ozone Deposition 

experiment. Agric. For. Meteorol. 85: 19-31. 

Graumlich, L. J. (1993) A 1000-year record of temperature and precipitation in the Sierra Nevada. Quaternary Res. 

39: 249-255.

Greenwood, P.; Greenhalgh, A.; Baker, C.; Unsworth, M. (1982) A computer-controlled system for exposing field 

crops to gaseous air pollutants. Atmos. Environ. 16: 2261-2266. 

Gregg, J. W.; Jones, C. G.; Dawson, T. E. (2003) Urbanization effects on tree growth in the vicinity of New York 

City [letter]. Nature (London, U. K. ) 424: 183-187. 

AX9-353
Greitner, C. S.; Winner, W. E. (1989) Effects of O3 on alder photosynthesis and symbiosis with Frankia. 

New Phytol. 111: 647-656. 

Greitner, C. S.; Pell, E. J.; Winner, W. E. (1994) Analysis of aspen foliage exposed to multiple stresses: ozone, 

nitrogen deficiency and drought. New Phytol. 127: 579-589. 

Grennfelt, P. (2004) New directions: recent research findings may change ozone control policies. Atmos. Environ. 

38: 2215-2216.

Grimes, H. D.; Perkins, K. K.; Boss, W. F. (1983) Ozone degrades into hydroxyl radical under physiological 

conditions. Plant Physiol. 72: 1016-1020. 

Grodzki, W.; McManus, M. S.; Knížek, M.; Meshkova, V.; Mihalciuc, V.; Novotny, J.; Tur...ani, M.; Slobodyan, Y. 

(2004) Occurrence of spruce bark beetles in forest stands at different levels of air pollution stress. Environ. 

Pollut. 130: 73-83. 

Grosjean, D.; Sensharma, D. K.; Cass, G. R. (1994) Fading of colorants by atmospheric pollutants: mass 

spectrometry studies. Sci. Total Environ. 152: 125-134. 

Grosjean, D.; Williams, E. L., II; Grosjean, E. (1995) Monitoring ambient ozone with a network of passive samplers: 

a feasibility study. Environ. Pollut. 88: 267-273. 

Gross, G. G.; Janse, C. (1977) Formation of NADH and hydrogen peroxide by cell wall-associated enzymes from 

forsythia xylem. Z. Pflanzenphysiol. 84: 447-452. 

Groth, J. V.; Krupa, S. V. (2000) Crop ecosystem responses to climatic change: interactive effects of ozone, 

ultraviolet-B radiation, sulphur dioxide and carbon dioxide on crops. In: Reddy, K. R.; Hodges, H. F., eds. 

Climate change and global crop productivity. New York, NY: CABI Publishing; pp. 387-405. 

Grulke, N. E. (1999) Physiological responses of ponderosa pine to gradients of environmental stressors. In: Miller, 

P. R.; McBride, J. R., eds. Oxidant Air Pollution Impacts in the Montane Forests of Southern California. 

Springer-Verlag; pp. 126-163. 

Grulke, N. E.; Balduman, L. (1999) Deciduous conifers: high N deposition and O3 exposure effects on growth and 

biomass allocation in ponderosa pine. Water Air Soil Pollut. 116: 235-248. 

Grulke, N. E.; Lee, E. H. (1997) Assessing visible ozone-induced injury in ponderosa pine. Can. J. For. Res. 

27: 1658-1668.

Grulke, N. E.; Miller, P. R. (1994) Changes in gas-exchange characteristics during the life-span of giant sequoia: 

implications for response to current and future concentrations of atmospheric ozone. Tree Physiol. 

14: 659-668.

Grulke, N. E.; Retzlaff, W. A. (2001) Changes in physiological attributes of ponderosa pine from seedling to mature 

tree. Tree Physiol. 21: 275-286. 

Grulke, N. E.; Miller, P. R.; Scioli, D. (1996) Response of giant sequoia canopy foliage to elevated concentrations of 

atmospheric ozone. Tree Physiol. 16: 575-581. 

Grulke, N. E.; Andersen, C. P.; Fenn, M. E.; Miller, P. R. (1998a) Ozone exposure and nitrogen deposition lowers 

root biomass of ponderosa pine in the San Bernardino Mountains, California. Environ. Pollut. 103: 63-73. 

Grulke, N. E.; Miller, P. R.; Leininger, T. D. (1998b) Effect of ozone exposure on seasonal gas exchange of five 

western conifers. In: Bytnerowicz, A.; Arbaugh, M. J.; Schilling, S. L., eds. Proceedings of the International 

Symposium on Air Pollution and Climate Change Effects on Forest Ecosystems; February, 1996; Riverside, 

CA. Albany, CA: USDA Forest Service, Pacific Southwest General Technical Report no. PSW-GTR 166; pp. 

229-238. 

Grulke, N. E.; Andersen, C. P.; Hogsett, W. E. (2001) Seasonal changes in above- and belowground carbohydrate 

concentrations of ponderosa pine along a pollution gradient. Tree Physiol. 21: 173-181. 

Grulke, N. E.; Preisler, H. K.; Fan, C. C.; Retzlaff, W. A. (2002a) A statistical approach to estimate O3 uptake of 

ponderosa pine in a Mediterranean climate. Environ. Pollut. 119: 163-175. 

Grulke, N. E.; Preisler, H. K.; Rose, C.; Kirsch, J.; Balduman, L. (2002b) O3 uptake and drought stress effects on 

carbon acquisition of ponderosa pine in natural stands. New Phytol. 154: 621-631. 

Grulke, N. E.; Johnson, R.; Esperanza, A.; Jones, D.; Nguyen, T.; Posch, S.; Tausz, M. (2003a) Canopy transpiration 

of Jeffrey pine in mesic and xeric microsites: O3 uptake and injury response. Trees (Berlin, Ger. ) 

17: 292-298.

Grulke, N. E.; Johnson, R.; Monschein, S.; Nikolova, P.; Tausz, M. (2003b) Variation in morphological and 

biochemical O3 injury attributes of mature Jeffrey pine within canopies and between microsites. Tree Physiol. 

23: 923-929.

Grulke, N. E.; Alonso, R.; Nguyen, T.; Cascio, C.; Dobrowolski, W. (2004) Stomata open at night in pole-sized and 

mature ponderosa pine: implications for O3 exposure metrics. Tree Physiol. 24: 1001-1010. 

AX9-354
Grünhage, L.; Jäger, H. J. (1994) Influence of the atmospheric conductivity on the ozone exposure of plants under 

ambient conditions: considerations for establishing ozone standards to protect vegetation. Environ. Pollut. 85: 

125-129. 

Grünhage, L.; Jäger, H. J. (2003) From critical levels to critical loads for ozone: a discussion of a new experimental 

and modelling approach for establishing flux-response relationships for agricultural crops and native plant 

species. Environ. Pollut. 125: 99-110. 

Grünhage, L.; Dämmgen, U.; Haenel, H. D.; Jäger, H. J.; Holl, A.; Schmitt, J.; Hanewald, K. (1993) A new potential 

air-quality criterion derived from vertical flux densities of ozone and from plant response. Angew. Bot. 

67: 9-13.

Grünhage, L.; Dämmgen, U.; Haenel, H. D.; Jäger, H. J. (1994) Response of a grassland ecosystem to air pollutants: 

III--the chemical climate: vertical flux densities of gaseous species in the atmosphere near the ground. 

Environ. Pollut. 85: 43-49. 

Grünhage, L.; Haenel, H. D. (1997) PLATIN (PLant-ATmosphere-INteraction) I: a model of plant-atmosphere 

interaction for estimating absorbed doses of gaseous air pollutants. Environ. Pollut. 98: 37-50. 

Grünhage, L.; Jäger, H. J.; Haenel, H. D.; Hanewald, K.; Krupa, S. (1997) PLATIN (plant-atmosphere interaction) 

II: co-occurrence of high ambient ozone concentrations and factors limiting plant absorbed dose. Environ. 

Pollut. 98: 51-60. 

Grünhage, L.; Jäger, H. J.; Haenel, H. D.; Löpmeier, F. J.; Hanewald, K. (1999) The European critical levels for 

ozone: improving their usage. Environ. Pollut. 105: 163-173. 

Grünhage, L.; Haenel, H. D.; Jäger, H. J. (2000) The exchange of ozone between vegetation and atmosphere: 

micrometeorological measurement techniques and models. Environ. Pollut. 109: 373-392. 

Grünhage, L.; Krause, G. H. M.; Köllner, B.; Benderc, J.; Weigel, H. J.; Jäger, H. J.; Guderian, R. (2001) A new 

flux-oriented concept to derive critical levels for ozone to protect vegetation. Environ. Pollut. 111: 355-362. 

Grünhage, L.; Krupa, S. V.; Legge, A. H.; Jäger, H. J. (2004) Ambient flux-based critical values of ozone for 

protecting vegetation: differing spatial scales and uncertainties in risk assessment. Atmos. Environ. 

38: 2433-2437.

Grunwald, C.; Endress, A. G. (1985) Foliar sterols in soybeans exposed to chronic levels of ozone. Plant Physiol. 

77: 245-247.

Grüters, U.; Fangmeier, A.; Jäger, H. J. (1995) Modelling stomatal responses of spring wheat (Triticum aestivum0 

A. L cv Turbo) to ozone and different levels of water supply. Environ. Pollut. 87: 141-149. 

Guderian, R. (1977) Discussion of the suitability of plant responses as a basis for air pollution control measures. In: 

Billings, W. D.; Golley, F.; Lange, O. L.; Olson, J. S., eds. Air pollution: phytotoxicity of acidic gases and its 

significance in air pollution control. Berlin, Federal Republic of Germany: Springer Verlag; pp. 75-97. 

Guidi, L.; Tonini, M.; Soldatini, G. F. (2000) Effects of high light and ozone fumigation on photosynthesis in 

Phaseolus vulgaris. Plant Physiol. Biochem. 38: 717-725. 

Guidi, L.; Nali, C.; Lorenzini, G.; Filippi, F.; Soladatini, G. E. (2001) Effect of chronic ozone fumigation on the 

photosynthetic process of poplar clones showing different sensitivity. Environ. Pollut. 113: 245-254. 

Gunderson, C. A.; Taylor, G. E., Jr. (1988) Kinetics of inhibition of foliar gas exchange by exogenous ethylene: 

an ultrasensitive response. New Phytol. 110: 517-524. 

Gunderson, C. A.; Taylor, G. E., Jr. (1991) Ethylene directly inhibits foliar gas exchange in Glycine max. Plant 

Physiol. 95: 337-339. 

Gunderson, C. A.; Sholtis, J. D.; Wullschleger, S. D.; Tissue, D. T.; Hanson, P. J.; Norby, R. J. (2002) 

Environmental and stomatal control of photosynthetic enhancement in the canopy of a sweetgum 

(Liquidambar styraciflua L.) plantation during 3 years of CO2 enrichment.. Plant Cell Environ. 25: 379-393. 

Guzy, M. R.; Heath, R. L. (1993) Responses to ozone of varieties of common bean (Phaseolus vulgaris L). 

New Phytol. 124: 617-625. 

Häberle, K. H.; Werner, H.; Fabian, P.; Pretzch, H.; Reiter, I.; Matyssek, R. (1999) 'Free-air' ozone fumigation of 

mature forest trees: a concept for validitating AOT40 under stand conditions. In: Fuhrer, J.; Achermann, B. 

eds. Critical levels for ozone: level II: workshop under the Convention on Long-Range Transboundary Air 

Pollution of the United Nations Economic Commission for Europe (UN/ECE); April; Gerzensee, Switzerland; 

Berne, Switzerland: Bundesamt fur Umwelt, Wald und Landschaft; pp. 133-137. 

Hacour, A.; Craigon, J.; Vandermeiren, K.; Ojanperä, K.; Pleijel, H.; Danielsson, H.; Högy, P.; Finnan, J.; Bindi, M. 

(2002) CO2 and ozone effects on canopy development of potato crops across Europe. Eur. J. Agron. 

17: 257-272.

Hakala, K.; Mela, T. (1996) The effects of prolonged exposure to elevated temperatures and elevated CO2 levels on 

the growth, yield and dry matter partitioning of field-sown meadow fescue. Agric. Food Sci. Finl. 5: 285-298. 

AX9-355
Hankamer, B.; Barber, J.; Boekema, E. J. (1997) Structure and membrane organization of photosystem II in green 

plants. Ann. Rev. Plant Physiol. Mol. Biol. 48: 641-671. 

Hanson, P. J.; Garten, C. T., Jr. (1992) Deposition of H15NO3 to white oak, red maple, and loblolly pine foliage: 

experimental observations and a generalized model. New Phytol. 122: 329-337. 

Hanson, P. J.; Samuelson, L. J.; Wullschleger, S. D.; Tabberer, T. A.; Edwards, G. (1994) Seasonal patterns of 

light-saturated photosynthesis and leaf conductance for mature and seedling Quercus rubra L foliage: 

differential sensitivity to ozone exposure. Tree Physiol. 14: 1351-1366. 

Hao, X.; Hale, B. A.; Ormrod, D. P.; Papadopoulos, A. P. (2000) Effects of pre-exposure to ultraviolet-B radiation 

on responses of tomato (Lycopersicon esculentum cv New Yorker) to ozone in ambient and elevated carbon 

dioxide. Environ. Pollut. 110: 217-224. 

Harris, M. J.; Bailey-Serres, J. N. (1994) Ozone effects on gene expression and molecular approaches to breeding for 

air pollution resistance. In: Basra, A. S., ed. Stress-induced gene expression in plants. Langhorne, PA: 

Harwood Academic Publishers; pp. 185-207. 

Harward, M.; Treshow, M. (1975) Impact of ozone on the growth and reproduction of understorey plants in the 

Aspen zone of western U.S.A. Environ. Conserv. 2: 17-23. 

Hausladen, A.; Madamanchi, N. R.; Fellows, S.; Alscher, R. G.; Amundson, R. G. (1990) Seasonal changes in 

antioxidants in red spruce as affected by ozone. New Phytol. 115: 447-458. 

Haynes, R. W.; Adams, D. M. (1992) Assessing economic impacts of air pollution damage to US forests. Asheville, 

NC: U.S. Department of Agriculture, Southeastern Forest Experiment Station; general technical report no. 

SE-75. 

Heagle, A. S. (1979) Effects of growth media, fertiliser rate and hour and season of exposure on sensitivity of four 

soybean cultivars to ozone. Environ. Pollut. 18: 313-322. 

Heagle, A. S.; Body, D. E.; Heck, W. W. (1973) An open-top field chamber to assess the impact of air pollution on 

plants. J. Environ. Qual. 2: 365-368. 

Heagle, A. S.; Philbeck, R. B.; Knott, W. M. (1979a) Thresholds for injury, growth, and yield loss caused by ozone 

on field corn hybrids. Phytopathology 69: 21-26. 

Heagle, A. S.; Spencer, S.; Letchworth, M. B. (1979b) Yield response of winter wheat to chronic doses of ozone. 

Can. J. Bot. 57: 1999-2005. 

Heagle, A. S.; Letchworth, M. B.; Mitchell, C. A. (1983) Effects of growth medium and fertilizer rate on the yield 

response of soybeans exposed to chronic doses of ozone. Phytopathology 73: 134-139. 

Heagle, A. S.; Heck, W. W.; Lesser, V. M.; Rawlings, J. O. (1987) Effects of daily ozone exposure duration and 

concentration fluctuation on yield of tobacco. Phytopathology 77: 856-862. 

Heagle, A. S.; Kress, L. W.; Temple, P. J.; Kohut, R. J.; Miller, J. E.; Heggestad, H. E. (1988) Factors influencing 

ozone dose-yield response relationships in open-top field chamber studies. In: Heck, W. W.; Taylor, O. C.; 

Tingey, D. T. eds. Assessment of crop loss from air pollutants: proceedings of an international conference; 

October, 1987; Raleigh, NC; New York, NY: Elsevier Applied Science; pp. 141-179. 

Heagle, A. S.; Rebbeck, J.; Shafer, S. R.; Blum, U.; Heck, W. W. (1989) Effects of long-term ozone exposure and 

soil moisture deficit on growth of a ladino clover-tall fescue pasture. Phytopathology 79: 128-136. 

Heagle, A. S.; McLaughlin, M. R.; Miller, J. E.; Joyner, R. L.; Spruill, S. E. (1991) Adaptation of a white clover 

population to ozone stress. New Phytol. 119: 61-68. 

Heagle, A. S.; Miller, J. E.; Rawlings, J. O.; Vozzo, S. F. (1991) Effect of growth stage on soybean response to 

chronic ozone exposure. J. Environ. Qual. 20: 562-570. 

Heagle, A. S.; Miller, J. E.; Sherrill.D.E.; Rawlings, J. O. (1993) Effects of ozone and carbon dioxide mixtures on 

two clones of white clover. New Phytol. 123: 751-762. 

Heagle, A. S.; Brandenburg, R. L.; Burns, J. C.; Miller, J. E. (1994a) Ozone and carbon dioxide effects on spider 

mites in white clover and peanut. J. Environ. Qual. 23: 1168-1176. 

Heagle, A. S.; Miller, J. E.; Sherrill, D. E. (1994b) A white clover system to estimate effects of tropospheric ozone 

on plants. J. Environ. Qual. 23: 613-621. 

Heagle, A. S.; Miller, J. E.; Chevone, B. I.; Dreschel, T. W.; Manning, W. J.; McCool, P. M.; Morrison, C. L.; 

Neely, G. E.; Rebbeck, J. (1995) Response of a white clover indicator system to tropospheric ozone at eight 

locations in the United States. Water Air Soil Pollut. 85: 1373-1378. 

Heagle, A. S.; Reinert, R. A.; Miller, J. E. (1996) Response of white clover to ozone in different environments. 

J. Environ. Qual. 25: 273-278.

Heagle, A. S.; Miller, J. E.; Booker, F. L. (1998) Influence of ozone stress on soybean response to carbon dioxide 

enrichment: I Foliar properties. Crop Sci. 38: 113-121. 

AX9-356
Heagle, A. S.; Miller, J. E.; Pursley, W. A. (1998) Influence of ozone stress on soybean response to carbon dioxide 

enrichment: III Yield and seed quality. Crop Sci. 38: 128-134. 

Heagle, A. S.; Booker, F. L.; Miller, J. E.; Fiscus, E. L.; Pursley, W. A.; Stefanski, L. A. (1999a) Influence of daily 

CO2 exposure duration and root environment on soybean resonse to elevated CO2. J. Environ. Qual. 

28: 666-675.

Heagle, A. S.; Miller, J. E.; Booker, F. L. (1999b) Ozone stress, carbon dioxide enrichment, and nitrogen fertility 

interactions in cotton. Crop Sci. 39: 731-741. 

Heagle, A. S.; Miller, J. E.; Pursley, W. A. (2000) Growth and yield responses of winter wheat to mixtures of ozone 

and carbon dioxide. Crop Sci. 40: 1656-1664. 

Heagle, A. S.; Stefanski, L. A. (2000) Relationships between ambient ozone regimes and white clover forage 

production using different ozone exposure indexes. Atmos. Environ. 34: 735-744. 

Heagle, A. S.; Miller, J. E.; Burkey, K. O.; Eason, G.; Pursley, W. A. (2002) Growth and yield responses of snap 

bean to mixtures of carbon dioxide and ozone. J. Environ. Qual. 31: 2008-2014. 

Heagle, A. S.; Miller, J. E.; Pursley, W. A. (2003) Growth and yield responses of potato to mixtures of carbon 

dioxide and ozone. J. Environ. Qual. 32: 1603-1610. 

Heath, R. L. (1980) Initial events in injury to plants by air pollutants. Annu. Rev. Plant Physiol. 31: 395-431. 

Heath, R. L. (1984) Air pollutant effects on biochemicals derived from metabolism: organic, fatty and amino acids. 

In: Koziol, M. J.; Whatley, F. R., eds. Gaseous Air Pollutants and Plant Metabolism. London, United 

Kingdom: Butterworths; pp. 275-290. 

Heath, R. L. (1987) The biochemistry of ozone attack on the plasma membrane of plant cells. Recent Adv. 

Phytochem. 21: 29-54. 

Heath, R. L. (1988) Biochemical mechanisms of pollutant stress. In: Heck, W. W. ; Taylor, O. C. ; Tingey, D. T. , 

eds. Assessment of crop loss from air pollutants: proceedings of an international conference; October, 1987; 

Raleigh, NC. London, United Kingdom: Elsevier Applied Science; pp. 259-286. 

Heath, R. L. (1994a) Alterations of plant metabolism by ozone exposure. In: Alscher, R. G.; Wellburn, A. R., eds. 

Plant Responses to the Gaseous Environment: molecular, metabolic, and physiological aspects. molecular, 

metabolic, and physiological aspects. London, United Kingdom: Chapman & Hall; pp. 121-146. 

Heath, R. L. (1994b) Possible mechanisms for the inhibition of photosynthesis by ozone. Photosynt. Res. 

39: 439-451.

Heath, R. L. (1996) The modification of photosynthetic capacity induced by ozone exposure. In: Baker, N. R., ed. 

Photosynthesis and the environment. Kluwer Academic Publishers; pp. 409-433. 

Heath, R. L.; Castillo, F. J. (1987) Membrane disturbances in response to air pollutants. In: Schulte-Hostede, S.; 

Darrall, N. M.; Blank, L. W.; Wellburn, A. R., eds. Air pollution and plant metabolism. New York, NY: 

Elsevier Applied Science; pp. 55-75. 

Heath, R. L.; Taylor, G. E., Jr. (1997) Physiological processes and plant responses to ozone exposure. In: 

Sandermann, H.; Wellburn, A. R.; Heath, R. L., eds. Forest Decline and Ozone: a comparison of controlled 

chamber and field experiments. New York, NY: Springer-Verlag; pp. 317-368. (Ecological Studies, v. 127). 

Heck, W. W.; Cowling, E. B. (1997) The need for a long term cumulative secondary ozone standard - an ecological 

perspective. EM (January): 23-33. 

Heck, W. W.; Philbeck, R. B.; Dunning, J. A. (1978) A continuous stirred tank reactor (CSTR) system for exposing 

plants to gaseous air contaminants. Principles, specifications, construction, and operation. Washington, DC: 

U.S. Department of Agriculture, Agricultural Research Service; ARS-S-81. 

Heck, W. W.; Taylor, O. C.; Adams, R.; Bingham, G.; Miller, J.; Preston, E.; Weinstein, L. (1982) Assessment of 

crop loss from ozone. J. Air Pollut. Control Assoc. 32: 353-361. 

Heck, W. W.; Heagle, A. S.; Miller, J. E.; Rawlings, J. O. (1991) A national program (NCLAN) to assess the impact 

of ozone on agricultural resources. In: Berglund, R. L.; Lawson, D. R.; McKee, D. J., eds. Tropospheric 

ozone and the environment: papers from an international conference; March 1990; Los Angeles, CA. 

Pittsburgh, PA: Air & Waste Management Association; pp. 225-254. (A&WMA transactions series no. 

TR-19). 

Heck, W. W.; Furiness, C. S.; Cowling, E. B.; Sims, C. K. (1998) Effects of ozone on crop, forest, and natural 

ecosystems: assessment of research needs. EM (October): 11-22. 

Heggestad, H. E. (1988) Reduction in soybean seed yields by ozone air pollution? JAPCA 38: 1040-1041. 

Heliövaara, K.; Väisänen, R. (1993) Insects and pollution. Boca Raton, FL: CRC Press. 

Hendrey, G. R.; Kimball, B. A. (1994) The FACE program. Agric. For. Meteorol. 70: 3-14. 

Hendrey, G. R.; Ellsworth, D. S.; Lewin, K. F.; Nagy, J. (1999) A free-air enrichment system for exposing tall forest 

trees to elevated atmospheric CO2. Global Change Biol. 5: 293-309. 

AX9-357
Hertstein, U.; Fangmeier, A.; Jäger, H. J. (1996) ESPACE-wheat (European stress physiology and climate 

experiment - Project 1: wheat): objectives, general approach, and first results. J. Appl. Bot. 70: 172-180. 

Hertstein, U.; CoIls, J.; Ewert, F.; Van Oijen, M. (1999) Climatic conditions and concentrations of carbon dioxide 

and air pollutants during 'ESPACE--wheat' experiments. Eur. J. Agron. 10: 163-169. 

Hildebrand, E.; Skelly, J. M.; Fredericksen, T. S. (1996) Foliar response of ozone-sensitive hardwood tree species 

from 1991 to 1993 in the Shenandoah National Park, Virginia. Can. J. For. Res. 26: 658-669. 

Ho, I.; Trappe, J. M. (1984) Effects of ozone exposure on mycorrhiza formation and growth of Festuca arundinacea. 

Environ. Exp. Bot. 24: 71-74. 

Hofstra, G.; Ali, A.; Wukasch, R. T.; Fletcher, R. A. (1981) The rapid inhibition of root respiration after exposure of 

bean (Phaseolus vulgaris L) plants to ozone. Atmos. Environ. 15: 483-487. 

Hogsett, W. E.; Plocher, M.; Wildman, V.; Tingey, D. T.; Bennett, J. P. (1985a) Growth response of two varieties of 

slash pine seedlings to chronic ozone exposures. Can. J. Bot. 63: 2369-2376. 

Hogsett, W. E.; Tingey, D. T.; Holman, S. R. (1985b) A programmable exposure control system for determination of 

the effects of pollutant exposure regimes on plant growth. Atmos. Environ. 19: 1135-1145. 

Hogsett, W. E.; Olszyk, D.; Ormrod, D. P.; Taylor, G. E., Jr.; Tingey, D. T. (1987a) Air pollution exposure systems 

and experimental protocols: volume 1: a review and evaluation of performance. Corvallis, OR: U.S. 

Environmental Protection Agency, Environmental Research Laboratory; EPA report no. EPA/600/3-87/037a. 

Available from: NTIS, Springfield, VA; PB88-181680. 

Hogsett, W. E.; Olszyk, D.; Ormrod, D. P.; Taylor, G. E., Jr.; Tingey, D. T. (1987b) Air pollution exposure systems 

and experimental protocols: volume 2: description of facilities. Corvallis, OR: U.S. Environmental Protection 

Agency, Environmental Research Laboratory; EPA report no. EPA/600/3-87/037b. Available from: NTIS, 

Springfield, VA; PB88-181698. 

Hogsett, W. E.; Tingey, D. T.; Lee, E. H. (1988) Ozone exposure indices: concepts for development and evaluation 

of their use. In: Heck, W. W.; Taylor, O. C.; Tingey, D. T. eds. Assessment of crop loss from air pollutants: 

proceedings of an international conference; October 1987; Raleigh, NC; New York, NY: Elsevier Applied 

Science; pp. 107-138. 

Hogsett, W. E.; Tingey, D. T.; Hendricks, C.; Rossi, D. (1989) Sensitivity of western conifers to SO2 and seasonal 

interaction of acid fog and ozone. In: Olson, R. K.; Lefohn, A. S., eds. Effects of air pollution on western 

forests [an A&WMA symposium; June; Anaheim, CA]. Air Pollution Control Association; pp. 469-491 

(APCA transactions series: no. 16). 

Hogsett, W. E.; Herstrom, A. A.; Laurence, J. A.; Lee, E. H.; Weber, J. E.; Tingey, D. T. (1995) Risk 

characterization of tropospheric ozone to forests. In: Lee, S. D.; Schneider, T. eds. Comparative risk analysis 

and priority setting for air pollution issues: proceedings of the 4th U.S.-Dutch international symposium; June 

1993; Keystone, CO; Pittsburgh, PA: Air & Waste Management Association; pp. 119-145. (AWMA 

publication no. VIP-43). 

Hogsett, W. E.; Weber, J. E.; Tingey, D.; Herstrom, A.; Lee, E. H.; Laurence, J. A. (1997) Environmental auditing: 

an approach for characterizing tropospheric ozone risk to forests. Environ. Manage. 21: 105-120. 

Holland, D. M.; Saltzman, N.; Cox, L. H.; Nychka, D. (1999) Spatial prediction of sulfur dioxide in the eastern 

United States. In: Gómez-Hernández, J.; Soares, A.; Froidevaux, R. eds. geoENV II: geostatistics for 

environmental applications, proceedings of the second European conference; November 1998; Valencia, 

Spain; Dordrecht, The Netherlands: Kluwer Academic Publishers; pp. 65-76. (Quantitative geology and 

geostatistics: v. 10). 

Holopainen, J. K.; Kainulainen, P.; Oksanen, J. (1997) Growth and reproduction of aphids and levels of free amino 

acids in Scots pine and Norway spruce in an open-air fumigation with ozone. Global Change Biol. 

3: 139-147.

Holton, M. K.; Lindroth, R. L.; Nordheim, E. V. (2003) Foliar quality influences tree-herbivore-parisitoid 

interactions: effects of elevated CO2, O3, and plant genotype. Oecologia 137: 233-244. 

Horemans, N.; Foyer, C. H.; Asard, H. (2000) Transport and action of ascorbate at the plant plasma membrane. 

Trends Plant Sci. 5: 263-267. 

Hormaza, J. I.; Pinney, K.; Polito, V. S. (1996) Correlation in the tolerance to ozone between sporophytes and male 

gametophytes of several fruit and tree nut species (Rosaceae). Sex. Plant Reprod. 9: 44-48. 

Houpis, J. L. J.; Costella, M. P.; Cowles, S. (1991) A branch exposure chamber for fumigating ponderosa pine to 

atmospheric pollution. J. Environ. Qual. 20: 467-474. 

Howe, G. A.; Lee, G. I.; Itoh, A.; Li, L.; DeRocher, A. (2000) Cytochrome P450-dependent metabolism of oxylipins 

in tomato: cloning and expression of allene oxide synthase and fatty acid hydroperoxide lyase. Plant Physiol. 

123: 711-724. 

AX9-358
Hui, D.; Sims, D. A.; Johnson, D. W.; Chang, W.; Luo, Y. (2002) Effects of gradual versus step increases in carbon 

dioxide on Plantago photosynthesis and growth in a microcosm study. Environ. Exp. Bot. 47: 51-66. 

Hummel, R. L.; Brandenburg, R. L.; Heagle, A. S.; Arellano, C. (1998) Effects of ozone on reproduction of 

two-spotted spider mite (Acari: Tetrannychidae) on white clover. Environ. Entomol. 27: 388-394. 

Huttunen, S.; Heikkilä, H.; Bucher, J.-B.; Sundberg, B.; Jarvis, P. G.; Matyssek, R. (2001) Trends in European forest 

tree physiological research. Amsterdam, The Netherlands: Kluwer, 262 pp, 262. 

Idso, K. E.; Idso, S. B. (1994) Plant responses to atmospheric CO2 enrichment in the face of environmental 

constraints: a review of the past 10 years' research. Agric. For. Meteorol. 69: 153-203. 

Innes, J. L.; Skelly, J. M.; Schaub, M. (2001) Ozone and broadleaved species A guide to the identification of 

ozone-induced foliar injury. Bern, Switzerland: Paul Haupt Publishers. 

Intergovernmental Panel on Climate Change (IPCC) (1996) Climate change 1995: the science of climate change: 

contribution of working group I to the second assessment of the Intergovernmental Panel on Climate Change. 

Cambridge, United Kingdom: Cambridge University Press. 

Intergovernmental Panel on Climate Change (IPCC) (2001) Climate change 2001: the scientific basis Contribution 

of working group I to the third assessment report of the Intergovernmental Panel on Climate Change. 

Cambridge, United Kingdom: Cambridge University Press. 

Isebrands, J. G.; Dickson, R. E.; Rebbeck, J.; Karnosky, D. F. (2000) Interacting effects of multiple stresses on 

growth and physiological processes in northern forest trees. In: Mickler, R. A.; Birsdey, R. A.; Hom, J., eds. 

Responses of northern U.S. forests to environmental change. New York, NY: Springer-Verlag; pp. 149-180. 

(Ecological studies: v. 139). 

Isebrands, J. G.; McDonald, E. P.; Kruger, E.; Hendrey, G.; Percy, K.; Pregitzer, K.; Sober, J.; Karnosky, D. F. 

(2001) Growth responses of Populus tremuloides clones to interacting carbon dioxide and tropospheric ozone. 

Environ. Pollut. 115: 359-371. 

Islam, K. R.; Mulchi, C. L.; Ali, A. A. (2000) Interactions of tropospheric CO2 and O3 enrichments and moisture 

variations on microbial biomass and respiration in soil. Global Change Biol. 6: 255-265. 

Ito, O.; Okano, K.; Kuroiwa, M.; Totsuka, T. (1985) Effects of NO2 and O3 alone or in combination on kidney bean 

plants (Phaseolus vulgaris L): growth, partitioning of assimilates and root activities. J. Exp. Bot. 36: 652-662. 

Izuta, T. (1998) Ecopysiological responses of Japanese forest tree species to ozone, simulated acid rain and soil 

acidification. J. Plant Res. 111: 471-480. 

Jackson, G. E. (1995) The effect of ozone, nitrogen dioxide and nitric oxide fumigation of cereals on the rose grain 

aphid Metopolophium dirhodum. Agric. Ecosyst. Environ. 54: 187-194. 

Jackson, D. M.; Heagle, A. S.; Eckel, R. V. W. (1999) Ovipositional response of tobacco hornworm moths 

(Lepidoptera: Sphyngidae) to tobacco plants grown under elevated levels of ozone. Environ. Entomol. 

28: 566-571.

Jackson, D. M.; Rufty, T. W.; Heagle, A. S.; Severson, R. F.; Eckel, R. W. V. (2000) Survival and development of 

tobacco hornworm larvae on tobacco plants grown under elevated levels of ozone. J. Chem. Ecol. 26: 1-18. 

Jacobson, J. S. (1977) The effects of photochemical oxidants on vegetation. In: Ozon und Begleitsubstanzen im 

photochemischen Smog: das Kolloquium [Ozone and related substances in photochemical smog: the 

colloquium]; September 1976; Düsseldorf, Federal Republic of Germany; Dusseldorf, Federal Republic of 

Germany: VDI-Verlag GmbH; pp. 163-173. (VDI-Berichte nr. 270). 

Jäger, H. J.; Herstein, U.; Fangmeier, A. (1999) The European Stress Physiology and Climate Experiment - Project 

1: wheat (ESPACE-wheat); introduction, aims and methodology. Eur. J. Agron. 10: 155-162. 

James, R. L.; Cobb, F. W., Jr.; Miller, P. R.; Parmeter, J. R., Jr. (1980) Effects of oxidant air pollution on 

susceptibility of pine roots to Fomes annosus. Phytopathology 70: 560-563. 

Jans, U.; Hoigné, J. (2000) Atmospheric water: transformation of ozone into OH-radicals by sensitized 

photoreactions or black carbon. Atmos. Environ. 34: 1069-1085. 

Janson, R.; Granat, L. (1999) A foliar rinse study of the dry deposition of nitric acid to a coniferous forest. Agric. 

For. Meteorol. 98-99: 683-696. 

Jarvis, P. G. (1976) The interpretation of the variation in leaf water potential and stomatal conductance found in 

canopies in the field. Philosoph. Trans. Royal Soc. London Series B - Biological Sciences 273: 593-610. 

Jiang, C. Z.; Rodermel, S. R.; Shibles, R. M. (1993) Photosynthesis, rubisco activity and amount, and their 

regulation by transcription in senescing soybean leaves. Plant Physiol. 101: 105-112. 

Jiménez, A.; Calvo, E.; Martín, C.; Porcuna, J. L.; Sanz, M. J. (2001) Estudio de la interacción entre el ozone y el 

desarrolio de las virosis en el cultivo del tomate. Agricola Vergel 231: 141-150. 

Johnson, D. W.; Taylor, G. E. (1989) Role of air pollution in forest decline in eastern North America. Water Air Soil 

Pollut. 48: 21-43. 

AX9-359
Johnson, B. G.; Hale, B. A.; Ormrod, D. P. (1996a) Carbon dioxide and ozone effects on growth of a legume-grass 

mixture. J. Environ. Qual. 25: 908-916. 

Johnson, J. D.; Chappelka, A. H.; Hain, F. P.; Heagle, A. S. (1996b) Interactive effects of air pollutants with abiotic 

and biotic factors on southern pine forests. In: Cox, S.; Mickler, R. A., eds. Impact of air pollutants on 

southern pine forests. Springer; pp. 281-312. 

Jones, H. G.; Lacasse, N. L.; Liggett, W. S.; Weatherford, F. (1977) Experimental air exclusion system for field 

studies of SO2 effects on crop productivity. Washington, DC: U.S. Environmental Protection Agency; 

EPA-600/7-77-122. 

Jones, C. G.; Coleman, J. S.; Findlay, S. (1994) Effects of ozone on interactions between plants, consumers and 

decomposers. In: Alscher, R. G.; Wellburn, A. R., eds. Plant responses to the gaseous environment. Chapman 

and Hall; pp. 339-363. 

Joss, U.; Graber, W. K. (1996) Profiles and simulated exchange of H2O, O3, NO2 between the atmosphere and the 

HartX Scots pine plantation. Theor. Appl. Climatol. 53: 157-172. 

Kainulainen, P.; Holopainen, J. K.; Hyttinen, H.; Oksanen, J. (1994) Effect of ozone on the biochemistry and aphid 

infestation of Scots pine. Phytochemistry 35: 39-42. 

Kainulainen, P.; Holopainen, J. K.; Holopainen, T. (2000a) Combined effects of ozone and nitrogen on secondary 

compounds, amino acids, and aphid performance in Scots pine. J. Environ. Qual. 29: 334-342. 

Kainulainen, P.; Utriainen, J.; Holopainen, J. K.; Oksanen, J.; Holopainen, T. (2000b) Influence of elevated ozone 

and limited nitrogen availability on conifer seedlings in an open-air fumigation system: effects on growth, 

nutrient content, mycorhiza, needle ultrastructure, starch and secondary compounds. Global Change Biol. 

6: 345-355.

Kainulainen, P.; Holopainen, T.; Holopainen, J. K. (2003) Decomposition of secondary compounds from needle 

litter of Scots pine grown under elevated CO2 and O3. Glob. Change Biol. 9: 295-304. 

Kangasjärvi, J.; Talvinen, J.; Utriainen, M.; Karjalainen, R. (1994) Plant defense systems induced by ozone. Plant 

Cell Environ. 17: 783-794. 

Kanoun, M.; Goulas, P.; Bassères, A.; Biolley, J. P. (2002) Ozone-induced oxidation of Rubisco: from an ELISA 

quantification of carbonyls to putative pathways leading to oxidizing mechanisms. Funct. Plant Biol. 29: 

1357-1363. 

Kärenlampi, L.; Skärby, L. (1996) Critical levels for ozone in Europe: testing and finalizing the concepts UN-ECE 

workshop report. In: Proceedings of UN-ECE convention on long-range transboundary air pollution 

workshop; April; Kuopio, Finland. Kupio, Finland: University of Kuopio, Department of Ecology and 

Environmental Science. 

Karlsson, P. E.; Medin, E. L.; Wickström, H.; Selldén, G.; Wallin, G.; Ottosson, S.; Skärby, L. (1995) Ozone and 

drought stress: interactive effects on the growth and physiology of Norway spruce (Picea abies (L) Karst). 

Water Air Soil Pollut. 85: 1325-1330. 

Karlsson, P. E.; Medin, E. L.; Wallin, G.; Selldén, G.; Skärby, L. (1997) Effects of ozone and drought stress on the 

physiology and growth of two clones of Norway spruce (Picea abies). New Phytol. 136: 265-275. 

Karlsson, P. E.; Pleijel, H.; Karlsson, G. P.; Medin, E. L.; Skärby, L. (2000) Simulations of stomatal conductance 

and ozone uptake to Norway spruce saplings in open-top chambers. Environ. Pollut. 109: 443-451. 

Karlsson, P. E.; Medin, E. L.; Selldén, G.; Wallin, G.; Ottosson, S.; Pleijel, H.; Skärby, L. (2002) Impact of ozone 

and reduced water supply on the biomass accumulation of Norway spruce saplings. Environ. Pollut. 19: 

237-244. 

Karlsson, G. P.; Karlsson, P. E.; Danielsson, H.; Pleijel, H. (2003) Clover as a tool for bioindication of phytotoxic 

ozone–5 years of experience from southern Sweden–consequences for the short-term critical levels. Sci. Total 

Environ. 301: 205-213. 

Karlsson, P. E.; Uddling, J.; Skärby, L.; Wallin, G.; Selldén, G. (2003) Impact of ozone on the growth of birch 

(Betula pendula) saplings. Environ. Pollut. 124: 485-495. 

Karlsson, P. E.; Selldén, G.; Pleijel, H. (2003) Establishing ozone critical levels II. Gothenburg, Sweden: IVL 

Swedish Environmental Research Institute; IVL Report B 1523. 

Karlsson, G. P.; Karlsson, P. E.; Soya, G.; Vandermeiren, K.; Pleijel, H. (2004a) Test of the short-term critical levels 

for acute ozone injury on plants–improvements by ozone uptake modelling and the use of an effect threshold. 

Atmos. Environ. 38: 2237-2245. 

Karlsson, P. E.; Uddling, J.; Braun, S.; Broadmeadow, M.; Elvira, S.; Gimeno, B. S.; Le Thiec, D.; Okansen, E.; 

Vandermeiren, K.; Wilkinson, M.; Emberson, L. (2004b) New critical levels for ozone effects on young trees 

based on AOT40 and simulated cumulative leaf uptake of ozone. Atmos. Environ. 38: 2283-2294. 

AX9-360
Karnosky, D. F. (1981) Changes in eastern white pine stands related to air pollution stress. Mitt. Forstl. 

Bundesversuchsanst. Wien 137: 41-45. 

Karnosky, D. F. (2003) Impacts of elevated atmospheric CO2 on forest trees and forest ecosystems: knowledge gaps. 

Environ. Int. 29: 161-169. 

Karnosky, D. F.; Gagnon, Z. E.; Dickson, R. E.; Coleman, M. D.; Lee, E. H.; Isebrands, J. G. (1996) Changes in 

growth, leaf abscission, biomass associated with seasonal tropospheric ozone exposures of Populus 

tremuloides clones and seedlings. Can. J. For. Res. 26: 23-37. 

Karnosky, D. F.; Podila, G. K.; Gagnon, Z.; Pechter, P.; Akkapeddi, A.; Coleman, M.; Dickson, R. E.; Isebrands, 

J. G. (1998) Genetic control of responses to interacting O3 and CO2 in Populus tremuloides. Chemosphere 

36: 807-812.

Karnosky, D. F.; Mankovska, B.; Percy, K.; Dickson, R. E.; Podila, G. K.; Sober, J.; Noormets, A.; Hendrey, G.; 

Coleman, M. D.; Kubiske, M.; Pregitzer, K. S.; Isebrands, J. G. (1999) Effects of tropospheric ozone on 

trembling aspen and interaction with CO2: results from an O3-gradient and a FACE experiment. Water Air 

Soil Pollut. 116: 311-322. 

Karnosky, D. F.; Gielen, B.; Ceulemans, R.; Schlesinger, W. H.; Norby, R. J.; Oksanen, E.; Matyssek, R.; Hendrey, 

G. R. (2001) FACE systems for studying the impact of greenhouse gases on forest ecosystems. In: Karnosky, 

D. F.; Ceulemans, R.; Scarascia-Mugnozza, G. E.; Innes, J. L. eds. The impact of carbon dioxide and other 

greenhouse gases on forest ecosystems, Report No. 3 of the IUFRO Task Force on Environmental Change; 

Wallingford, United Kingdom: CABI Publ.; pp. 297-324. 

Karnosky, D. F.; Ceulemans, R.; Scarascia-Mugnozza, G. E.; Ines, J. L. (2001a) The impact of carbon dioxide and 

other greenhouse gases on forest ecosystems: report no 3 of the IUFRO Task Force on Environmental 

Change. Wallingford, United Kingdom: CABI Publishing; Available: 

http://www.cabi-publishing.org/Bookshop/Readingroom/0851995519.asp [19 August, 2003]. 

Karnosky, D. F.; Oksanen, E.; Dickson, R. E.; Isebrands, J. G. (2001b) Impact of interacting greenhouse gases on 

forest ecosystems. In: Karnosky, D. F.; Ceulemans, R.; Scarascia-Mugnozza, G. E.; Innes, J. L., eds. The 

impact of carbon dioxide and other greenhouse gases of forest ecosystems: report no. 3 of the IUFRO Task 

Force on Environmental Change. Wallingford, United Kingdom: CABI Publishing; pp. 253-267 (IUFRO 

research series no. 8). 

Karnosky, D. F.; Percy, K. E.; Xiang, B.; Callan, B.; Noormets, A.; Mankovska, B.; Hopkin, A.; Sôber, J.; Jones, 

W.; Dickson, R. E.; Isebrands, J. G. (2002) Interacting elevated CO2 and tropospheric O3 predisposes aspen 

(Populus tremuloides Michx) to infection by rust (Melampsora medusae f sp tremuloidae). Global Change 

Biol. 8: 329-338. 

Karnosky, D. F.; Zak, D. R.; Pregitzer, K. S.; Awmack, C. S.; Bockheim, J. G.; Dickson, R. E.; Hendrey, G. R.; 

Host, G. E.; King, J. S.; Kopper, B. J.; Kruger, E. L.; Kubiske, M. E.; Lindroth, R. L.; Mattson, W. J.; 

McDonald, E. P.; Noormets, A.; Oksanen, E.; Parsons, W. F. J.; Percy, K. E.; Podila, G. K.; 

Riemenschneider, D. E.; Sharma, P.; Thakur, R.; Sôber, A.; Sôber, J.; Jones, W. S.; Anttonen, S.; Vapaavuori, 

E.; Mankovska, B.; Heilman, W.; Isebrands, J. G. (2003a) Tropospheric O3 moderates responses of temperate 

hardwood forests to elevated CO2: A synthesis of molecular to ecosystem results from the Aspen FACE 

project. Funct. Ecol. 17: 289-304. 

Karnosky, D. F.; Percy, K.; Mankovska, B.; Pritchard, T.; Noormets, A.; Dickson, R. E.; Jepsen, E.; Isebrands, J. G. 

(2003b) Ozone affects the fitness of trembling aspen. In: Karnosky, D. F.; Percy, K. E.; Chappelka, A. H.; 

Simpson, C. J.; Pikkarainen, J., eds. Air pollution, global change and forests in the new millenium. Elsevier 

Science. (Developments in Environmental Science, 3). 

Karnosky, D. F.; Pregitzer, K. S.; Zak, D. R.; Kubiske, M. E.; Hendrey, G. R.; Weinstein, D.; Nosal, M.; Percy, K. 

E. (2005) Scaling ozone responses of forest trees to the ecosystem level in a changing climate. Plant Cell 

Environ. 28: 965-981. 

Karpinski, S.; Wingsle, G.; Karpinska, B.; Hällgren, J. E. (1993) Molecular response to photooxidative stress in 

Pinus sylvestris (L): II Differential expression of Cu-Zn superoxide dismutases and glutathione reductase. 

Plant Physiol. 103: 1385-1391. 

Kasurinen, A. E.; Helmisaari, H. S.; Holopainen, T. (1999) The influence of elevated CO2 and O3 on fine roots and 

mycorrihizas of naturally growing young Scots pine trees during three exposure years. Global Change Biol. 5: 

771-780. 

Kats, G.; Thompson, C. R.; Kuby, W. C. (1976) Improved ventilation of open top greenhouses. J. Air Pollut. Control 

Assoc. 26: 1089-1090. 

Kats, G.; Olszyk, D. M.; Thompson, C. R. (1985a) Open top experimental chambers for trees. J. Air Pollut. Control 

Assoc. 35: 1298-1301. 

AX9-361
Kats, G.; Dawson, P. J.; Bytnerowicz, A.; Wolf, J. W.; Thompson, C. R.; Olszyk, D. M. (1985b) Effects of ozone or 

sulfur dioxide on growth and yield of rice. Agric. Ecosyst. Environ. 14: 103-117. 

Kellomäki, S.; Wang, K. Y. (1997a) Effects of elevated O3 and CO2 concentrations on photosynthesis and stomatal 

conductance in Scots pine. Plant Cell Environ. 20: 995-1006. 

Kellomäki, S.; Wang, K. Y. (1997b) Effects of elevated O3 and CO2 on chlorophyll fluorescence and gas exchange 

in Scots pine during the third growing season. Environ. Pollut. 97: 17-27. 

Kelting, D. L.; Burger, J. A.; Edwards, G. S. (1995) The effects of ozone on the root dynamics of seedlings and 

mature red oak (Quercus rubra L). For. Ecol. Manage. 79: 197-206. 

Kerstiens, G.; Lendzian, K. J. (1989) Interaction between ozone and plant cuticles: I ozone deposition and 

permeability. New Phytol. 112: 13-19. 

Khan, M. R.; Khan, M. W. (1994) Single and interactive effects of O3 and SO2 on tomato. Environ. Exp. Bot. 

34: 461-469.

Khan, M. R.; Khan, M. W. (1997) Effects of the root-knot nematode, Meloidogyne incognita, on the sensitivity of 

tomato to sulfur dioxide and ozone. Environ. Exp. Bot. 38: 117-130. 

Khan, M. R.; Khan, M. W. (1998a) Interactive effects of ozone and powdery mildew (Sphaerotheca fuliginea) on 

bottle gourd (Lagenaria siceraria). Agric. Ecosyst. Environ. 70: 109-118. 

Khan, M. R.; Khan, M. W. (1998b) Interactive effects of ozone and root-knot nematode on tomato. Agric. Ecosyst. 

Environ. 70: 97-107. 

Khan, M. R.; Khan, M. W. (1999) Effects of intermittent ozone exposures on powdery mildew of cucumber. 

Environ. Exp. Bot. 42: 163-171. 

Kickert, R. N.; Krupa, S. V. (1991) Modeling plant response to tropospheric ozone: a critical review. Environ. 

Pollut. 70: 271-383. 

Kickert, R. N.; Tonella, G.; Simonov, A.; Krupa, S. V. (1999) Predictive modeling of effects under global change. 

Environ. Pollut. 100: 87-132. 

Kim, H. J.; Helfand, G. E.; Howitt, R. E. (1998a) An economic analysis of ozone control in California's San Joaquin 

Valley. J. Agric. Resource Econ. 23: 55-70. 

Kim, J. S.; Chappelka, A. H.; Miller-Goodman, M. S. (1998b) Decomposition of blackberry and broomsedge 

bluestem as influenced by ozone. J. Environ. Qual. 27: 953-960. 

King, J. S.; Pregitzer, K. S.; Zak, D. R.; Sober, J.; Isebrands, J. G.; Dickson, R. E.; Hendrey, G. R.; Karnosky, D. F. 

(2001) Fine-root biomass and fluxes of soil carbon in young stands of paper birch and trembling aspen as 

affected by elevated atmospheric CO2 and tropospheric O3. Oecologia 128: 237-250. 

Kleier, C.; Farnsworth, B.; Winner, W. (1998) Biomass, reproductive output, and physiological responses of 

rapid-cycling Brassica (Brassica rapa) to ozone and modified root temperature. New Phytol. 139: 657-664. 

Kleier, C.; Farnsworth, B.; Winner, W. (2001) Photosynthesis and biomass allocation of radish cv "Cherry Belle" 

in response to root temperature and ozone. Environ. Pollut. 111: 127-133. 

Kliebenstein, D. J.; Monde, R. A.; Last, R. L. (1998) Superoxide dismutase in arabidopsis: an eclectic enzyme 

family with disparate regulation and protein localization. Plant Physiol. 118: 637-650. 

Kobayashi, K.; Okada, M.; Nouchi, I. (1994) A chamber system for exposing rice (Oryza sativa L) to ozone in a 

paddy field. New Phytol. 126: 317-325. 

Kobayashi, K.; Okada, M.; Nouchi, I. (1995) Effects of ozone on dry matter partitioning and yield of Japanese 

cultivars of rice (Oryza sativa L). Agric. Ecosyst. Environ. 53: 109-122. 

Koch, J. R.; Scherzer, A. J.; Eshita, S. M.; Davis, K. R. (1998) Ozone sensitivity in hybrid poplar is correlated with a 

lack of defense-gene activation. Plant Physiol. 118: 1243-1252. 

Koch, J. R.; Creelman, R. A.; Eshita, S. M.; Seskar, M.; Mullet, J. E.; Davis, K. R. (2000) Ozone sensitivity in 

hybrid poplar correlates with insensitivity to both salicylic acid and jasmonic acid The role of programmed 

cell death in lesion formation. Plant Physiol. 123: 487-496. 

Kohut, R. J.; Laurence, J. A.; Amundson, R. G. (1988) Effects of ozone and sulfur dioxide on yield of red clover and 

timothy. J. Environ. Qual. 17: 580-585. 

Kohut, R. J.; Laurence, J. A.; King, P.; Raba, R. (2000) Identification of bioindicator species for ozone and 

assessment of the responses to ozone of native vegetation at Acadia National Park. Denver, CO: U.S. 

National Park Service, Air Resources Division; Final Report Acadia D-175. Technical Report 

NPS/BSO-RNR/NRTR/OO-13. 

Kolb, T. E.; Fredericksen, T. S.; Steiner, K. C.; Skelly, J. M. (1997) Issues in scaling tree size and age responses to 

ozone: a review. Environ. Pollut. 98: 195-208. 

Kolb, T. E.; Matyssek, R. (2001) Limitations and perspectives about scaling ozone impacts in trees. Environ. Pollut. 

115: 373-393. 

AX9-362
Köllner, B.; Krause, G. H. M. (2003) Effects of two different ozone exposure regimes on chlorophyll and sucrose 

content of leaves and yield parameters of sugar beet (Beta Vulgaris L.) and rape (Brassica napus L.). 

Water Air Soil Pollut. 144: 317-332. 

Kolomiets, M. V.; Chen, H.; Gladon, R. J.; Braun, E. J.; Hannapel, D. J. (2000) A leaf lipoxygenase of potato 

induced specifically by pathogen infection. Plant Physiol. 124: 1121-1130. 

Kopp, R. J.; Vaughn, W. J.; Hazilla, M.; Carson, R. (1985) Implications of environmental policy for US agriculture: 

the case of ambient ozone standards. J. Environ. Manage. 20: 321-331. 

Kopper, B. J.; Lindroth, R. L. (2003a) Responses of trembling aspen (Populus tremuloides) phytochemistry and 

aspen blotch leafminer (Phyllonorycter tremuloidiella) performance to elevated levels of atmospheric CO2 

and O3. Agric. For. Entomol. 5: 17-26. 

Kopper, B. J.; Lindroth, R. L. (2003b) Effects of elevated carbon dioxide and ozone on the phytochemistry of aspen 

and performance of an herbivore. Oecologia 134: 95-103. 

Kostka-Rick, R.; Manning, W. J. (1992a) Partitioning of biomass and carbohydrates in field-grown radish under 

ambient concentrations of ozone, and treated with the anti-ozonant ethylene-diurea (EDU). New Phytol. 121: 

187-200. 

Kostka-Rick, R.; Manning, W. J. (1992b) Effects and interactions of ozone and the anti-ozonant EDU at different 

stages of radish (Raphanus sativus L) development. J. Exp. Bot. 43: 1621-1631. 

Kostka-Rick, R.; Manning, W. J. (1993) Dynamics of growth and biomass partitioning in field-grown bush bean 

(Phaseolus vulgaris L), treated with the antiozonant ethylenediurea (EDU). Agric. Ecosyst. Environ. 

47: 195-214.

Kostka-Rick, R.; Manning, W. J.; Buonaccorsi, J. P. (1993) Dynamics of biomass partitioning in field-grown radish 

varieties, treated with ethylenediurea. Environ. Pollut. 80: 133-145. 

Kouterick, K. B.; Skelly, J. M.; Fredericksen, T. S.; Steiner, K. C.; KoIb, T. E.; Ferdinand, J. A. (2000) Foliar injury, 

leaf gas exchange and biomass responses of black cherry (Prunus serotina Ehrh) half-sibling families to 

ozone exposure. Environ. Pollut. 107: 117-126. 

Kozio», M. J.; Whatley, F. R. (1984) Gaseous air pollutants and plant metabolism. London, United Kingdom: 

Butterworths Scientific. 

Kozlowski, T. T.; Pallardy, S. G. (1997) Physiology of woody plants. San Diego, CA: Academic Press. 

Kraft, M.; Weigel, H. J.; Mejer, G. J.; Brandes, F. (1996) Reflectance measurements of leaves for detecting visible 

and non-visible ozone damage to crops. J. Plant Physiol. 148: 148-154. 

Krapp, A.; Hofmann, B.; Schäfer, C.; Stitt, M. (1993) Regulation of the expression of rbcS and other photosynthetic 

genes by carbohydrates: a mechanism for the "sink regulation" of photosynthesis? Plant J. 3: 817-828. 

Krupa, S. V.; Groth, J. V. (2000) Global climate change and crop responses: uncertainties associated with the current 

methodologies. In: Agrawal, S. B.; Agrawal, M., eds. Environmental pollution and plant responses. 

Boca Raton, FL: Lewis Publishers; pp. 1-18. 

Krupa, S. V.; Jäger, H. J. (1996) Adverse effects of elevated levels of ultraviolet (UV)-B radiation and ozone (O3) 

on crop growth and productivity. In: Bazzaz, F.; Sombroek, W., eds. Global climate change and agricultural 

production. Chichester, United Kingdom: John Wiley & Sons; pp. 141-169. Available: 

http://www.fao.org/docrep/W5183E/w5183e00.htm [28 August, 2003]. 

Krupa, S. V.; Grünhage, L.; Jäger, H. J.; Nosal, M.; Manning, W. J.; Legge, A. H.; Hanewald, K. (1995) Ambient 

ozone (O3) and adverse crop response: a unified view of cause and effect. Environ. Pollut. 87: 119-126. 

Krupa, S. V.; Kickert, R. N.; Jäger, H. J. (1998) Elevated ultraviolet (UV)-B radiation and agriculture. New York, 

NY: Springer-Verlag Environmental Intelligence Unit. 

Krupa, S. V.; Nosal, M.; Peterson, D. L. (2001) Use of passive ozone O3 samplers in vegetation effects assessment. 

Environ. Pollut. 112: 303-309. 

Krupa, S.; Nosal, M. (2001) Relationships between passive sampler and continuous ozone (O3) measurement data in 

ecological effects research. Scientific World 1: 593-601. 

Krupa, S. V.; Nosal, M.; Ferdinand, J. A.; Stevenson, R. E.; Skelly, J. M. (2003) A multi-variate statistical model 

integrating passive sampler and meteorology data to predict the frequency distributions of hourly ambient 

ozone (O3) concentrations. Environ. Pollut. 124: 173-178. 

Kuehler, E. A.; Flagler, R. B. (1999) The effects of sodium erythorbate and ethylenediurea on photosynthetic 

function of ozone-exposed loblolly pine seedlings. Environ. Pollut. 105: 25-35. 

Kuik, O. J.; Helming, F. M.; Dorland, C.; Sparinks; F.A. (2000) The economic benefits to agriculture of a low-level 

ozone pollution in the Netherlands. Eur. Rev. Agric. Econ. 27: 75-90. 

Kull, O.; Tulva, I.; Vapaavuori, E. (2003) Influence of elevated CO2 and O3 on Betula pendula Roth crown structure. 

Ann. Bot. 91: 559-569. 

AX9-363
Kurczy1⁄2ska, E. U.; Dmuchowski, W.; Wloch, W.; Bytnerowicz.A. (1997) The influence of air pollutants on needles 

and stems of scots pine (Pinus sylvestris L) trees. Environ. Pollut. 98: 325-334. 

Kuïniak, E. (2002) Transgenic plants: an insight into oxidative stress tolerance mechanisms. Acta Physiol. Plant. 

(Acta Physiologiae Plantarum) 24: 97-113. 

Kytöviita, M. M.; Pelloux, J.; Fontaine, V.; Botton, B.; Dizengremel, P. (1999) Elevated CO2 does not ameliorate 

effects of ozone on carbon allocation in Pinus halapensis and Betula pendula in symbiosis with Paxillus 

involutus. Physiol. Plant. 106: 370-377. 

Kytöviita, M. M.; Le Thiec, D.; Dizengremel, P. (2001) Elevated CO2 and ozone reduce nitrogen acquisition by 

Pinus halepensis from its mycorrhizal symbiont. Physiol. Plant. 111: 305-312. 

Lagrimini, L. M.; Rothstein, S. (1987) Tissue specificity of tobacco peroxidase isozymes and their induction by 

wounding and tobacco mosaic virus infection. Plant Physiol. 84: 438-442. 

Laisk, A.; Kull, O.; Moldau, H. (1989) Ozone concentration in leaf intercellular air spaces is close to zero. Plant 

Physiol. 90: 1163-1167. 

Lamb, C.; Dixon, R. A. (1997) The oxidative burst in plant disease resistance. Ann. Rev. Plant Physiol. Mol. Biol. 

48: 251-275.

Landolt, W.; Günthardt-Goerg, M. S.; Pfenninger, I.; Einig, W.; Hampp, R.; Maurer, S.; Matyssek, R. (1997) Effect 

of fertilization on ozone-induced changes in the metabolism of birch (Betula pendula) leaves. New Phytol. 

137: 389-397. 

Landolt, W.; Bühlmann, U.; Bleuler, P.; Bucher, J. B. (2000) Ozone exposure-response relationships for biomass 

and root/shoot ratio of beech (Fagus sylvatica), ash (Fraxinus excelsior), Norway spruce (Picea abies) 

and scots pine (Pinus sylvestris). Environ. Pollut. 109: 473-478. 

Lange, O. L.; Heber, U.; Schulze, E. D.; Ziegler, H. (1989) 3-D atmospheric pollutants and plant metabolism. In: 

Schulze, E. D.; Lange, O. L.; Oren, R., eds. Forest decline and air pollution: a study of spruce (Picea abies) 

on acid soils. a study of spruce (Picea abies) on acid soils. Berlin, Federal Republic of Germany: 

Springer-Verlag; pp. 238-273. 

Langebartels, C.; Kerner, K.; Leonardi, S.; Schraudner, M.; Trost, M.; Heller, W.; Sandermann, H., Jr. (1991) 

Biochemical plant responses to ozone: I differential induction of polyamine and ethylene biosynthesis in 

tobacco. Plant Physiol. 95: 882-889. 

Langebartels, C.; Ernst, D.; Heller, W.; Lutz, C.; Payer, H. D.; Sandermann, H., Jr. (1997) Ozone responses of trees: 

results from controlled chamber exposures at the GSF phytotron. In: Sandermann, H.; Wellburn, A. R.; Heath, 

R. L., eds. Forest decline and ozone. New York, NY: Springer-Verlag; pp. 163-200 (Ecological studies: v. 

127). 

Larson, J. L.; Zak, D. R.; Sinsabaugh, R. L. (2002) Extracellular enzyme activity beneath temperate trees growing 

under elevated carbon dioxide and ozone. Soil Sci. Soc. Am. J. 66: 1848-1856. 

Laurence, J. A.; Andersen, C. P. (2003) Ozone and natural systems: understanding exposure, response, and risk. 

Environ. Int. 29: 155-160. 

Laurence, J. A.; Kohut, R. J.; Amundson, R. G.; Weinstein, D. A.; MacLean, D. C. (1996) Response of sugar maple 

to multiple year exposure to ozone and simulated acidic precipitation. Environ. Pollut. 92: 119-126. 

Laurence, J. A.; Amundson, R. G.; Kohut, R. J.; Weinstein, D. A. (1997) Growth and water use of red spruce (Picea 

rubens Sarg.) exposed to ozone and simulated acidic precipitation for four growing seasons. For. Sci. 

(Bethesda, Md.) 43: 355-361. 

Laurence, J. A.; Ollinger, S. V.; Woodbury, P. B. (2000) Regional impacts of ozone on forest productivity. In: 

Mickler, R. A.; Birdsey, R. A.; Hom, J., eds. Responses of northern U. Springer-Verlag; pp. 425-453. 

Laurence, J. A.; Retzlaff, W. A.; Kern, J. S.; Lee, E. H.; Hogsett, W. E.; Weinstein, D. A. (2001) Predicting the 

regional impact of ozone and precipitation on the growth of loblolly pine and yellow poplar using linked 

TREGRO and ZELG models. For. Ecol. Manage. 146: 247-263. 

Laurence, J. A.; Retzlaff, W. A.; Kern, J. S.; Lee, E. H.; Hogsett, W. E.; Weinstein, D. A. (2003) Corrigendum to 

"Predicting the regional impact of ozone and precipitation on the growth of loblolly pine and yellow-poplar 

using linked TREGRO and ZELIG models" [Forest Ecol Manage 146 (2001) 247-263]. For. Ecol. Manage. 

174: 607. 

Lawlor, D. W. (1998) Plant responses to global change: temperature and drought stress. In: De Kok, L. J.; Stulen, I., 

eds. Responses of plant metabolism to air pollution and global change. Backhuys Publishers; pp. 193-207. 

Lawlor, D. W.; Keys, A. J. (1993) Understanding photosynthetic adaptation to changing climate. In: Fowden, L.; 

Mansfield, T.; Stoddart, J., eds. Plant adaptation to environmental stress; London, United Kingdom: Chapman 

and Hall; pp. Chapman and Hall; pp. 85-106.; pp. UnitedKingdom-106. 

AX9-364
Lawson, T.; Craigon, J.; Black, C. R.; Colls, J. J.; Tulloch, A. M.; Landon, G. (2001a) Effects of elevated carbon 

dioxide and ozone on the growth and yield of potatoes (Solanum tuberosum) grown in open-top chambers. 

Environ. Pollut. 111: 479-491. 

Lawson, T.; Craigon, J.; Tulloch, A. M.; Black, C. R.; Colls, J. J.; Landon, G. (2001b) Photosynthetic responses to 

elevated CO2 and O3 in field-grown potato (Solanum tuberosum). J. Plant Physiol. 158: 309-323. 

Lawson, T.; Craigon, J.; Black, C. R.; Colls, J. J.; Landon, G.; Weyers, J. D. B. (2002) Impact of elevated CO2 and 

O3 on gas exchange parameters and epidermal characteristics in potato (Solanum tuberosum L). J. Exp. Bot. 

53: 737-746.

Le Thiec, D.; Manninen, S. (2003) Ozone and water deficit reduced growth of Aleppo pine seedlings. Plant Physiol. 

Biochem. 41: 55-63. 

Lee, E. H.; Bennett, J. H. (1982) Superoxide dismutase: a possible protective enzyme against ozone injury in snap 

beans (Phaseolus vulgaris L). Plant Physiol. 69: 1444-1449. 

Lee, E. H.; Hogsett, W. E. (1999) Role of concentrations and time of day in developing ozone exposure indices for a 

secondary standard. J. Air Waste Manage. Assoc. 49: 669-681. 

Lee, J. J.; Lewis, R. A.; Body, D. E. (1975) A field experimental system for the evaluation of the bioenvironmental 

effects of sulfur dioxide. In: Proceedings of the Fort Coal Field Symposium, Vol. 5; Billings, Montana: 

Montana Academy of Science; pp. 608-620. 

Lee, E. H.; Tingey, D. T.; Hogsett, W. E. (1987) Selection of the best exposure-response model using various 7-hour 

ozone exposure statistics. Research Triangle Park, NC: U.S. Environmental Protection Agency, Office of Air 

Quality Planning and Standards. 

Lee, E. H.; Tingey, D. T.; Hogsett, W. E. (1988) Evaluation of ozone exposure indices in exposure-response 

modeling. Environ. Pollut. 53: 43-62. 

Lee, E. H.; Tingey, D. T.; Hogsett, W. E. (1989) Interrelation of experimental exposure and ambient air quality data 

for comparison of ozone exposure indices and estimating agricultural losses. Corvallis, OR: U.S. 

Environmental Protection Agency, Environmental Research Laboratory; EPA report no. EPA-600/3-89-047. 

Available from: NTIS, Springfield, VA; PB89-195036. 

Lee, E. H.; Hogsett, W. E.; Tingey, D. T. (1994a) Alternative attainment criteria for a secondary federal standard for 

ozone. In: Nriagu, J. O.; Simmons, M. S.; Nriagu, J. O., eds. Environmental oxidants. New York, NY: John 

Wiley & Sons; pp. 549-584. (Nriagu, J. O., ed. Advances in environmental sciences and technology series: v. 

28). 

Lee, E. H.; Hogsett, W. E.; Tingey, D. T. (1994b) Attainment and effects issues regarding alternative secondary 

ozone air quality standards. J. Environ. Qual. 23: 1129-1140. 

Lee, J. C.; Skelly, J. M.; Steiner, K. C.; Zhang, J. W.; Savage, J. E. (1999) Foliar response of black cherry 

(Prunus serotina) clones to ambient ozone exposure in central Pennsylvania. Environ. Pollut. 105: 325-331. 

Lee, J. C.; Steiner, K. C.; Zhang, J. W.; Skelly, J. M. (2002) Heritability of ozone sensitivity in open-pollinated 

families of black cherry (Prunus serotina Ehrh.). For. Sci. 48: 111-117. 

Lee, E. H.; Tingey, D. T.; Hogsett, W. E.; Laurence, J. A. (2003) History of tropospheric ozone for the San 

Bernardino Mountains of southern California, 1963-1999. Atmos. Environ. 37: 2705-2717. 

Lefohn, A. S.; Benedict, H. M. (1982) Development of a mathematical index that describes ozone concentration, 

frequency and duration. Atmos. Environ. 16: 2529-2532. 

Lefohn, A. S.; Foley, J. K. (1992) NCLAN results and their application to the standard-setting process: protecting 

vegetation from surface ozone exposures. J. Air Waste Manage. Assoc. 42: 1046-1052. 

Lefohn, A. S.; Runeckles, V. C. (1987) Establishing standards to protect vegetation - ozone exposure/dose 

considerations. Atmos. Environ. 21: 561-568. 

Lefohn, A. S.; Shadwick, D. S. (2000) Differences in trending estimates in the United States using several ozone 

metrics. In: Proceedings of the 93rd annual meeting of the Air & Waste Management Association; June; Salt 

Lake City, UT; Pittsburgh, PA: Air & Waste Management Association. 

Lefohn, A. S.; Hogsett, W. E.; Tingey, D. T. (1986) A method for developing ozone exposures that mimic ambient 

conditions in agricultural areas. Atmos. Environ. 20: 361-366. 

Lefohn, A. S.; Davis, C. E.; Jones, C. K.; Tingey, D. T.; Hogsett, W. E. (1987) Co-occurrence patterns of gaseous air 

pollutant pairs at different minimum concentrations in the United States. Atmos. Environ. 21: 2435-2444. 

Lefohn, A. S.; Laurence, J. A.; Kohut, R. J. (1988) A comparison of indices that describe the relationship between 

exposure to ozone and reduction in the yield of agricultural crops. Atmos. Environ. 22: 1229-1240. 

Lefohn, A. S.; Jackson, W.; Shadwick, D. S.; Knudsen, H. P. (1997) Effect of surface ozone exposures on vegetation 

grown in the southern Appalachian Mountains: identification of possible areas of concern. Atmos. Environ. 

31: 1695-1708.

AX9-365 
Lefohn, A. S.; Oltmans, S. J.; Dann, T.; Singh, H. B. (2001) Present-day variability of background ozone in the 

lower troposphere. J. Geophys. Res. [Atmos. ] 106: 9945-9958. 

Legge, A. H.; Grünhage, L.; Nosal, M.; Jäger, H. J.; Krupa, S. V. (1995) Ambient ozone and adverse crop response: 

an evaluation of North American and European data as they relate to exposure indices and critical levels. 

Angew. Bot. 69: 192-205. 

Leitao, L.; Goulas, P.; Biolley, J. P. (2003) Time-course of Rubisco oxidation in beans (Phaseolus vulgaris L) 

subjected to a long-term ozone stress. Plant Sci. 165: 613-620. 

León, J.; Lawton, M. A.; Raskin, I. (1995) Hydrogen peroxide synthesis stimulates salicylic acid biosynthesis in 

tobacco. Plant Physiol. 108: 1673-1683. 

Lesser, V. M.; Rawlings, J. O.; Spruill, S. E.; Somerville, M. C. (1990) Ozone effects on agricultural crops: 

statistical methodologies and estimated dose-response relationships. Crop Sci. 30: 148-155. 

Levitt, J. (1972) Responses of plants to environmental stress. New York, NY: Academic Press. 

Lichtenthaler, H. K. (1977) Regulation of prenylquinone synthesis in higher plants. In: Tevini, M.; Lichtenthaler, 

H. K.; Davies, B. H., eds. Lipids and lipid polymers in higher plants. New York, NY: Springer-Verlag; 

pp. 231-258. 

Lindroth, R. L.; Reich, P. B.; Tjoelker, M. G.; Volin, J. C.; Oleksyn, J. (1993) Light environment alters response to 

ozone stress in seedlings of Acer saccharum March, and hybrid Populus L. III. Consequences for 

performance of gypsy moth. New Phytol. 124: 647-654. 

Lindroth, R. L.; Kopper, B. J.; Parsons, W. F. J.; Bockheim, J. G.; Karnosky, D. F.; Hendrey, G. R.; Pregitzer, K. S.; 

Isebrands, J. G.; Sober, J. (2001) Consequences of elevated carbon dioxide and ozone for foliar chemical 

composition and dynamics in trembling aspen (Populus tremuloides) and paper birch (Betula papyrifera). 

Environ. Pollut. 115: 395-404. 

Liu, L.; Eriksson, K.-E. L.; Dean, J. F. D. (1995) Localization of hydrogen peroxide production in Pisum sativum L 

using epi-polarization microscopy to follow cerium perhydroxide disposition. Plant Physiol. 107: 501-506. 

Lloyd, A. C.; Lents, J. M.; Green, C.; Nemeth, P. (1989) Air quality management in Los Angeles: perspectives on 

past and future emission control strategies. JAPCA 39: 696-703. 

Loats, K. V.; Rebbeck, J. (1999) Interactive effects of ozone and elevated carbon dioxide on the growth and 

physiology of black cherry, green ash, and yellow poplar seedlings. Environ. Pollut. 106: 237-248. 

Long, S. P. (1991) Modification of the response of photosynthetic productivity to rising temperature by atmospheric 

CO2 concentrations: has its importance been underestimated? Plant Cell Environ. 14: 729-739. 

Lorenzini, G.; Nali, C. (1995) Analysis of vertical ozone and nitrogen oxides profiles in a Prunus cerasifera canopy. 

Int. J. Biometeorol. 39: 1-4. 

Lorenzini, G.; Medeghini Bonatti, P.; Nali, C.; Baroni Fornasiero, R. (1994) The protective effect of rust infection 

against ozone, sulphur dioxide and paraquat toxicity symptoms in broad bean. Physiol. Mol. Plant Pathol. 45: 

263-279. 

Lucas, P. W.; Cottam, D. A.; Mansfield, T. A. (1987) A large-scale fumigation system for investigating interactions 

between air pollution and cold stress on plants. Environ. Pollut. 43: 15-28. 

Lucas, P. W.; Cottam, D. A.; Sheppard, L. J.; Francis, B. J. (1988) Growth responses and delayed winter hardening 

in Sitka spruce following summer exposure to ozone. New Phytol. 108: 495-504. 

Luo, Y. (2001) Transient ecosystem response to free-air CO2 enrichment (FACE): experimental evidence and 

methods of analysis. New Phytol. 152: 3-8. 

Luo, Y.; Reynolds, J. F. (1999) Validity of extrapolating field CO2 experiments to predict carbon sequestration in 

natural ecosystems. Ecology 80: 1568-1583. 

Luo, Y.; Canadell, J.; Mooney, H. A. (1999) Interactive effects of carbon dioxide and environmental stress on plants 

and ecosystems: a synthesis. In: Luo, Y.; Mooney, H. A., eds. Carbon dioxide and environmental stress. San 

Diego, CA: Academic Press; pp. 393-408. 

Lütz, C.; Anegg, S.; Gerant, D.; Alaoi-Sossé, B.; Gérard, J.; Dizengremel, P. (2000) Beech trees exposed to high 

CO2 and to simulated summer ozone levels: effects on photosynthesis, chloroplast components and leaf 

enzyme activity. Physiol. Plant. 109: 252-259. 

Luwe, M. W. F.; Takahama, U.; Heber, U. (1993) Role of ascorbate in detoxifying ozone in the apoplast of spinach 

(Spinacia oleracea L) leaves. Plant Physiol. 101: 969-976. 

Luxmoore, R. J. (1988) Assessing the mechanisms of crop loss from air pollutants with process models. In: Heck, 

W. W.; Taylor, O. C.; Tingey, D. T., eds. Assessment of crop loss from air pollutants: proceedings of an 

international conference; October 1987; Raleigh, NC. Essex, United Kingdom: Elsevier Science Publishers, 

Ltd.; pp. 417-444. 

AX9-366
Lyons, T. M.; Barnes, J. D. (1998) Influence of plant age on ozone resistance in Plantago major. New Phytol. 

138: 83-89. 

Lyons, T. M.; Barnes, J. D.; Davison, A. W. (1997) Relationships between ozone resistance and climate in European 

populations of Plantago major. New Phytol. 136: 503-510. 

MacLean, D. C. (1990) Joint action of ozone and hydrogen fluoride on foliar senescence in maize. Environ. Pollut. 

63: 283-292.

Madamanchi, N. R.; Anderson, J. V.; Alscher, R. G.; Cramer, C. L.; Hess, J. L. (1992) Purification of multiple forms 

of glutathione reductase from pea (Pisum sativum L) seedlings and enzyme levels in ozone-fumigated pea 

leaves. Plant Physiol. 100: 138-145. 

Magan, N.; Kirkwood, I. A.; McLeod, A. R.; Smith, M. K. (1995) Effect of open-air fumigation with sulphur 

dioxide and ozone on phyllosphere and endophytic fungi of conifer needles. Plant Cell Environ. 18: 291-302. 

Maggs, R.; Ashmore, M. R. (1998) Growth and yield responses of Pakistan rice (Oryza sativa L) cultivars to O3 and 

NO2. Environ. Pollut. 103: 159-170. 

Mahoney, M. J.; Chevone, B. I.; Skelly, J. M.; Moore, L. D. (1985) Influence of mycorrhizae on the growth of 

loblolly pine seedlings exposed to ozone and sulfur dioxide. Phytopathology 75: 679-682. 

Maier-Maercker, U. (1997) Experiments on the water balance of individual attached twigs of Picea abies (L.) Karst. 

in pure and ozone-enriched air. Trees 11: 229-239. 

Maier-Maercker, U. (1998) Predisposition of trees to drought stress by ozone. Tree Physiol. 19: 71-78. 

Maier-Maercker, U.; Koch, W. (1992) Histological examination of spruce needles from a long-term gas exchange 

experiment in pure and polluted air in the field. Trees (Berlin) 6: 186-194. 

Manderscheid, R.; Jäger, H.-J.; Kress, L. W. (1992) Effects of ozone on foliar nitrogen metabolism of Pinus taeda L. 

and implications for carbohydrate metabolism. New Phytol. 121: 623-633. 

Mandl, R. H.; Weinstein, L. H.; McCune, D. C.; Keveny, M. (1973) A cylindrical, open-top chamber for the 

exposure of plants to air pollutants in the field. J. Environ. Qual. 2: 371-376. 

Mandl, R. H.; Laurence, J. A.; Kohut, R. J. (1989) Development and testing of open-top chambers for exposing 

large, perennial plants to air pollutants. J. Environ. Qual. 18: 534-540. 

MaÁkovská, B.; Godzik, B.; Badea, O.; Shparyk, Y.; Morav...ík, P. (2004) Chemical and morphological 

characteristics of key tree species of the Carpathian Mountains. Environ. Pollut. 130: 41-54. 

Manninen, A. M.; Holopainen, T.; Lyytikäinen-Saarenmaa, P.; Holopainen, J. K. (2000) The role of low-level ozone 

exposure and mycorrhizas in chemical quality and insect herbivore performance on Scots pine seedlings. 

Global Change Biol. 6: 111-121. 

Manninen, S.; Siivonen, N.; Timonen, U.; Huttunen, S. (2003) Differences in ozone response between two Finnish 

wild strawberry populations. Environ. Exp. Bot. 49: 29-39. 

Manning, W. J. (2000) Use of protective chemicals to assess the effects of ambient ozone on plants. In: Agrawal, 

S. B.; Agrawal, M., eds. Environmental Pollution and Plant Responses. Boca Raton, Florida: Lewis 

Publishers; pp. 247-258. 

Manning, W. J. (2003) Detecting plant effects is necessary to give biological significance to ambient ozone 

monitoring data and predictive ozone standards. Environ. Pollut. 126: 375-379. 

Manning, W. J.; Feder, W. A.; Papia, P. M.; Perkins, I. (1971) Influence of foliar ozone injury on root development 

and root surface fungi of pinto bean plants. Environ. Pollut. 1: 305-312. 

Manning, W. J.; Keane, K. D. (1988) Effects of air pollutants on interactions between plants, insects, and pathogens. 

In: Heck, W. W.; Taylor, O. C.; Tingey, D. T., eds. Assessment of crop loss from air pollutants: proceedings 

of an international conference; October 1987; Raleigh, NC. Essex, United Kingdom: Elsevier Science 

Publishers, Ltd.; pp. 365-386. 

Manning, W. J.; Krupa, S. V. (1992) Experimental methodology for studying the effects of ozone on crops and trees. 

In: Lefohn, A. S., ed. Surface level ozone exposures and their effects on vegetation. Chelsea, MI: Lewis 

Publishers, Inc.; pp. 93-156. 

Manning, W. J.; Feder, W. A.; Vardaro, P. M. (1974) Suppression of oxidant injury by benomyl: effects on yields of 

bean cultivars in the field. J. Environ. Qual. 3: 1-3. 

Manning, W. J.; Godzik, B.; Musselman, R. (2002) Potential bioindicator plant species for ambient ozone in forested 

mountain areas of central Europe. Environ. Pollut. 119: 283-290. 

Manning, W. J.; Flagler, R. B.; Frenkel, M. A. (2003) Assessing plant response to ambient ozone: growth of 

ozone-sensitive loblolly pine seedlings treated with ethylenediurea or sodium erythorbate. Environ. Pollut. 

126: 73-81. 

Marshall, J. D.; Waring, R. H. (1985) Predicting fine root production and turnover by monitoring root starch and soil 

temperature. Can. J. For. Res. 15: 791-800. 

AX9-367
Martin, M. J.; Host, G. E.; Lenz, K. E.; Isebrands, J. G. (2001) Simulating the growth response of aspen to elevated 

ozone: a mechanistic approach to scaling a leaf-level model of ozone effects on photosynthesis to a complex 

canopy architecture. Environ. Pollut. 115: 425-436. 

Massman, W. J. (1993) Partitioning ozone fluxes to sparse grass and soils and the inferred resistances to dry 

deposition. Atmos. Environ. Part A 27A: 167-174. 

Massman, W. J. (2004) Toward an ozone standard to protect vegetation based on effective dose: a review of 

deposition resistance and a possible metric. Atmos. Environ. In press. 

Massman, W. J.; Grantz, D. A. (1995) Estimating canopy conductance to ozone uptake from observations of 

evapotranspiration at the canopy scale and at the leaf scale. Global Change Biol. 1: 183-198. 

Massman, W. J.; Musselman, R. C.; Lefohn, A. S. (2000) A conceptual ozone dose-response model to develop a 

standard to protect vegetation. Atmos. Environ. 34: 745-759. 

Matsumura, H. (2001) Impacts of ambient ozone and/or acid mist on the growth of 14 tree species: an open-top 

chamber study conducted in Japan. Water Air Soil Pollut. 130: 959-964. 

Matsuyama, T.; Tamaoki, M.; Nakajima, N.; Aono, M.; Kubo, A.; Moriya, S.; Ichihara, T.; Suzuki, O. S. (2002) 

cDNA microarray assessment for ozone-stressed Arabidopsis thaliana. Environ. Pollut. 117: 191-194. 

Matters, G. L.; Scandalios, J. G. (1987) Synthesis of isozymes of superoxide dismutase in maize leaves in response 

to O3, SO2, and elevated O2. J. Exp. Bot. 38: 842-852. 

Matyssek, R.; Günthardt-Goerg, M. S.; Landolt, W.; Keller, T. (1993) Whole-plant growth and leaf formation in 

ozonated hybrid poplar (Populus x euramericana). Environ. Pollut. 81: 207-212. 

Matyssek, R.; Günthardt-Goerg, M. S.; Maurer, S.; Keller, T. (1995) Nighttime exposure to ozone reduces 

whole-plant production in Betula pendula. Tree Physiol. 15: 159-165. 

Matyssek, R.; Günthardt-Goerg, M. S.; Maurer, S.; Christ, R. (2002) Tissue structure and respiration of stems of 

Betula pendula under contrasting ozone exposure and nutrition. Trees 16: 375-385. 

Matyssek, R.; Wieser, G.; Nunn, A. J.; Kozovits, A. R.; Reiter, I. M.; Heerdt, C.; Winkler, J. B.; Baumgarten, M.; 

Häberle, K. H.; Grams, T. E. E.; Werner, H.; Fabian, P.; Havranek, W. M. (2004) Comparison between 

AOT40 and ozone uptake in forest trees of different species, age and site conditions. Atmos. Environ. 

38: 2271-2281.

Maurer, S.; Matyssek, R. (1997) Nutrition and the ozone sensitivity of birch (Betula pendula) II Carbon balance, 

water-use efficiency and nutritional status of the whole plant. Trees 12: 11-20. 

Mazarura, U. (1997) Effects of fluctuating levels of ozone and nitrogen dioxide alone and in sequence on plant 

growth and photosynthesis [dissertation]. University of British Columbia; Vancouver, British Columbia, 

Canada; 165 pp. 

McAinsh, M. R.; Evans, N. G.; Montgomery, L. T. (2002) Calcium signalling in stomatal responses to pollutants. 

New Phytol. 153: 441-447. 

McBride, J. R.; Laven, R. D. (1976) Scars as an indicator of fire frequency in the San Bernardino Mountains, 

California. J. For. 74: 439-442. 

McBride, J. R.; Laven, R. D. (1999) Impact of oxidant air pollutants on forest succession in the mixed conifer forests 

of the San Bernardino Mountains. In: Miller, P. R.; McBride, J. E., eds. Oxidant air pollution impacts in the 

montane forests of southern California: a case study of the San Bernardino Mountains. New York, NY: 

Springer; pp. 338-352. (Ecological Studies, v. 134). 

McCarl, B. A.; Alig, R.; Adams, D. M. (1998) Impacts of incorporating land exchanges between forestry and 

agriculture in sector models. J. Agric. Appl. Econ. 30: 389-401. 

McCool, P. M.; Menge, J. A. (1983) Influence of ozone on carbon partitioning in tomato: potential role of carbon 

flow in regulation of the mycorrhizal symbiosis under conditions of stress. New Phytol. 94: 241-247. 

McCool, P. M.; Menge, J. A.; Taylor, O. C. (1982) Effect of ozone injury and light stress on response of tomato to 

infection by the vesicular-arbuscular mycorrhizal fungus, Glomus fasciculatus. J. Am. Soc. Hortic. Sci. 107: 

839-842. 

McCool, P. M.; Musselman, R. C.; Younglove, T.; Teso, R. R. (1988) Response of kidney bean to sequential ozone 

exposures. Environ. Exp. Bot. 28: 307-313. 

McCrady, J. K.; Andersen, C. P. (2000) The effect of ozone on below-ground carbon allocation in wheat. Environ. 

Pollut. 107: 465-472. 

McDonald, E. P.; Kruger, E. L.; Riemenschneider, D. E.; Isebrands, J. G. (2002) Competitive status influences 

tree-growth responses to elevated CO2 and O3 in aggrading aspen stands. Funct. Ecol. 16: 792-801. 

McGartland, A. M. (1987) The implications of ambient ozone standards for US agriculture: a comment and some 

further evidence. J. Environ. Manage. 24: 139-146. 

AX9-368
McKee, I. F.; Farage, P. K.; Long, S. P. (1995) The interactive effects of elevated CO2 and O3 concentration on 

photosynthesis in spring wheat. Photosyn. Res. 45: 111-119. 

McKee, I. F.; Bullimore, J. F.; Long, S. P. (1997a) Will elevated CO2 concentrations protect the yield of wheat from 

O3 damage? Plant Cell Environ. 20: 77-84. 

McKee, I. F.; Eiblmeier, M.; Polle, A. (1997b) Enhanced ozone-tolerance in wheat grown at an elevated CO2 

concentration: ozone exclusion and detoxification. New Phytol. 137: 275-284. 

McKee, I. F.; Mulholland, B. J.; Craigon, J.; Black, C. R.; Long, S. P. (2000) Elevated concentrations of 

atmospheric CO2 protect against and compensate for O3 damage to photosynthetic tissues of field-grown 

wheat. New Phytol. 146: 427-435. 

McLaughlin, S. B.; Downing, D. J. (1995) Interactive effects of ambient ozone and climate measured on growth of 

mature forest trees. Nature (London) 374: 252-254. 

McLaughlin, S. B.; Downing, D. J. (1996) Interactive effects of ambient ozone and climate measured on growth of 

mature loblolly pine trees. Can. J. For. Res. 670-681. 

McLaughlin, S. B.; McConathy. (1983) Effects of SO2 and O3 on alocation of 14C-labeled photosynthate in 

Phaseolus vulgaris. Plant Physiol. 73: 630-635. 

McLaughlin, S.; Percy, K. (1999) Forest health in North America: some perspectives on actual and potential roles of 

climate and air pollution. Water Air Soil Pollut. 116: 151-197. 

McLaughlin, S. B.; McConathy, R. K.; Duvick, D.; Mann, L. K. (1982) Effects of chronic air pollution stress on 

photosynthesis, carbon allocation, and growth of white pine trees. For. Sci. 28: 60-70. 

McLaughlin, S. B.; Layton, P. A.; Adams, M. B.; Edwards, N. T.; Hanson, P. J.; O'Neill, E. G.; Roy, W. K. (1994) 

Growth-responses of 53 open-pollinated loblolly-pine families to ozone and acid-rain. J. Environ. Qual. 

23: 247-257.

McLaughlin, S. B.; Shortle, W. C.; Smith, K. T. (2002) Dendroecological applications in air pollution and 

environmental chemistry: research needs. Dendrochronologia 20: 133-157. 

McLaughlin, S. B.; Wullschleger, S. D.; Nosal, M. (2003) Diurnal and seasonal changes in stem increment and 

water use by yellow poplar trees in response to environmental stress. Tree Physiol. 23: 1125-1136. 

McLeod, A. R.; Fackrell, J. E.; Alexander, K. (1985) Open-air fumigation of field crops: criteria and design for a 

new experimental system. Atmos. Environ. 19: 1639-1649. 

McLeod, A. R.; Shaw, P. J. A.; Holland, M. R. (1992) The Liphook Forest Fumigation Project: studies of sulphur 

dioxide and ozone effects on coniferous trees. For. Ecol. Manage. 51: 121-127. 

Mearns, L. O.; Rosenzweig, C.; Goldberg, R. (1997) Mean and variance change in climate scenarios: methods, 

agricultural applications and measures of uncertainty. Clim. Change 35: 367-396. 

Meentemeyer, V. (1978) Macroclimate and lignin control of litter decomposition rates. Ecology 59: 465-475. 

Mehdy, M. C. (1994) Active oxygen species in plant defense against pathogens. Plant Physiol. 105: 467-472. 

Mehlhorn, H.; O'Shea, J. M.; Wellburn, A. R. (1991) Atmospheric ozone interacts with stress ethylene formation by 

plants to cause visible plant injury. J. Exp. Bot. 42: 17-24. 

Mehlhorn, H.; Wellburn, A. R. (1987) Stress ethylene formation determines plant sensitivity to ozone. Nature 

(London) 327: 417-418. 

Mehlhorn, H.; Wenzel, A. A. (1996) Manganese deficiency enhances ozone toxicity in bush beans (Phaseolus 

vulgaris L cv Saxa). J. Plant Physiol. 148: 155-159. 

Mehlhorn, H.; Tabner, B. J.; Wellburn, A. R. (1990) Electron spin resonance evidence for the formation of free 

radicals in plants exposed to ozone. Physiol. Plant. 79: 377-383. 

Meier, S.; Grand, L. F.; Schoeneberger, M. M.; Reinert, R. A.; Bruck, R. I. (1990) Growth, ectomycorrhizae and 

nonstructural carbohydrates of loblolly pine seedlings exposed to ozone and soil water deficit. Environ. 

Pollut. 64: 11-27. 

Meinhard, M.; Rodriguez, P. L.; Grill, E. (2002) The sensitivity of ABI2 to hydrogen peroxide links the abscisic 

acid-response regulator to redox signalling. Planta 214: 775-782. 

Mendelsohn, R.; Nordhaus, W.; Shaw, D. (1994) The impact of global warming on agriculture: a Ricardian analysis. 

Am. Econ. Rev. 84: 753-771. 

Miglietta, F.; Peressotti, A.; Vaccari, F. P.; Zaldei, A.; Deangelis, P.; Scarascia-Mugnozza, G. (2001) Free-air CO2 

enrichment (face) of a poplar plantation: the popface fumigation system. New Phytol. 150: 465-476. 

Mikkelsen, T. N.; Ro-Poulsen, H.; Hovmand, M. F.; Jensen, N. O.; Pilegard, K.; Egelov, A. H. (2004) Five-year 

measurements of ozone fluxes to a Danish Norway spruce canopy. Atmos. Environ. 38: 2361-2371. 

Miller, P. L. (1973) Oxidant-induced community change in a mixed conifer forest. In: Naegele, J. A. , ed. Air 

pollution damage to vegetation. Washington, DC: American Chemical Society; pp. 101-117. (Advances in 

chemistry series: v. 122). 

AX9-369
Miller, P. R.; Elderman, M. J. (1977) Photochemical oxidant air pollutant effects on a mixed conifer forest 

ecosystem: a progress report, 1976. Corvallis, OR: U.S. Environmental Protection Agency, Corvallis 

Environmental Research Laboratory; EPA report no. EPA-600/3-77-104. Available from: NTIS, Springfield, 

VA; PB-274531. 

Miller, P. R.; McBride, J. R. (1999a) Oxidant air pollution impacts in the Montane forests of southern California: 

a case study of the San Bernardino Mountains. (Ecological Studies, v. 134). 

Miller, P. R.; McBride, J. R. (1999b) Assessment of ecological risks and implications for policy and management in 

the San Bernardino Mountains. In: Miller, P. R.; McBride, J. R., eds. Oxidant air pollution impacts in the 

Montane forests of southern California: a case study of the San Bernardino Mountains. New York, NY: 

Springer; pp. 397-404. (Ecological Studies, v. 134). 

Miller, P. R.; Millecan, A. A. (1971) Extent of oxidant air pollution damage to some pines and other conifers in 

California. Plant Dis. Rep. 55: 555-559. 

Miller, J. E.; Pursley, W. A.; Heagle, A. S. (1994) Effects of ethylenediurea on snap bean at a range of ozone 

concentrations. J. Environ. Qual. 23: 1082-1089. 

Miller, P. R.; Rechel, J. (1999) Temporal changes in crown condition indices, needle litterfall, and collateral needle 

injuries of Ponderosa and Jeffrey pines. In: Miller, P. R.; McBride, J. R., eds. Oxidant air pollution impacts in 

the Montane forests of southern California: a case study of the San Bernardino Mountains. New York, NY: 

Springer; pp. 164-178. 

Miller, P. R.; Parmeter, J. R., Jr.; Taylor, O. C.; Cardiff, E. A. (1963) Ozone injury to the foliage of Pinus 

ponderosa. Phytopathology 53: 1072-1076. 

Miller, P. R.; Parmeter, J. R., Jr.; Flick, B. H.; Martinez, C. W. (1969) Ozone dosage response of ponderosa pine 

seedlings. J. Air Pollut. Control Assoc. 19: 435-438. 

Miller, P. R.; McCutchan, M. H.; Milligan, H. P. (1972) Oxidant air pollution in the Central Valley, Sierra Nevada 

foothills, and Mineral King Valley of California. Atmos. Environ. 6: 623-633. 

Miller, P. R.; Taylor, O. C.; Wilhour, R. G. (1982) Oxidant air pollution effects on a western coniferous forest 

ecosystem. Corvallis, OR: U.S. Environmental Protection Agency, Environmental Research Laboratory; EPA 

report no. EPA-600/D-82-276. Available from: NTIS, Springfield, VA; PB83-189308. 

Miller, P. R.; McBride, J. R.; Schilling, S. L.; Gomez, A. P. (1989) Trend of ozone damage to conifer forests 

between 1974 and 1988 in the San Bernardino Mountains of southern California. In: Olson, R. K.; Lefohn, A. 

S., eds. Effects of air pollution on western forests [an A&WMA symposium; June; Anaheim, CA]. Air and 

Waste Management Association; pp. 309-323. (APCA transactions series, no. 16). 

Miller, J. E.; Booker, F. L.; Fiscus, E. L.; Heagle, A. S.; Pursley, W. A.; Vozzo, S. F.; Heck, W. W. (1994) 

Ultraviolet-B radiation and ozone effects on growth, yield, and photosynthesis of soybean. J. Environ. Qual. 

23: 83-91.

Miller, P. R.; De Bauer, L. I.; Quevedo, A.; Hernandez-Tejeda, T. (1994) Comparison of ozone exposures in 

forested regions near Mexico City and Los Angeles. Atmos. Environ. 28: 141-148. 

Miller, J. E.; Heagle, A. S.; Pursley, W. A. (1998) Influence of ozone stress on soybean response to carbon dioxide 

enrichment: II Biomass and development. Crop Sci. 38: 122-128. 

Miller, J. D.; Arteca, R. N.; Pell, E. J. (1999) Senescence-associated gene expression during ozone-induced leaf 

senescence in Arabidopsis. Plant Physiol. 120: 1015-1024. 

Mills, G. (2002) Modification of plant response by environmental conditions. In: Bell, J. N. B.; Treshow, M., eds. 

Air pollution and plant life. 2nd ed. Chichester, United Kingdom: John Wiley & Sons Ltd.; pp. 343-358. 

Mills, G.; Ball, G.; Hayes, F.; Fuhrer, J.; Skarby, L.; Gimeno, B.; De Temmerman, L.; Heagle, A.; Members of the 

ICP Vegetation programme. (2000) Development of a multi-factor model for predicting the effects of ambient 

ozone on the biomass of white clover. Environ. Pollut. 109: 533-542. 

Minnich, R. A. (1999) Vegetation, fire regimes, and forest dynamics. In: Oxidant air pollution impacts in the 

montane forests southern California: a case study of the San Bernardino mountains. New York, NY: Springer; 

pp. 44-80 (Ecological studies: v. 134). 

Minnich, R. A.; Padgett, P. E. (2003) Geology, climate, and vegetation of the Sierra Nevada and the mixed-conifer 

zone: an introduction to the ecosystem. In: Bytnerowicz, A.; Arbaugh, M. J.; Alonso, R., eds. Ozone Air 

Pollution in the Sierra Nevada: Distribution and Effects on Forests. New York, NY: Elsevier; pp. 1-31. 

Minnich, R. A.; Barbour, M. G.; Burk, J. H.; Fernau, R. F. (1995) Sixty years of change in California conifer forests 

of the San Bernardino mountains. Conserv. Biol. 9: 902-914. 

Mira, H.; Mártinez, N.; Peñarrubia, L. (2002) Expression of a vegetative-storage-protein gene form Arabidopsis is 

regulated by copper, senescence and ozone. Planta 214: 939-946. 

AX9-370
Moeder, W.; Barry, C. S.; Tauriainen, A. A.; Betz, C.; Tuomainen, J.; Ultrainen, M.; Grierson, D.; Sandermann, H.; 

Langebartels, C.; Kangasjarvi, J. (2002) Ethylene synthesis regulated by biphasic induction of 

1-aminocyclopropane-1-carboxylic acid synthase and 1-aminocyclopropane-1-carboxylic acid oxidase genes 

is required for hydrogen peroxide accumulation and cell death in ozone-exposed tomato. Plant Physiol. 130: 

1918-1926. 

Moldau, H. (1998) Hierarchy of ozone scavenging reactions in the plant cell wall. Physiol. Plant. 104: 617-622. 

Moldau, H.; Bichele, I. (2002) Plasmalemma protection by the apoplast as assessed from above zero ozone 

concentration in leaf intercellular air spaces. Planta 214: 484-487. 

Momen, B.; Helms, J. A. (1996) Effects of simulated acid rain and ozone on foliar chemistry of field-grown Pinus 

ponderosa seedlings and mature trees. Environ. Pollut. 91: 105-111. 

Momen, B.; Helms, J. A.; Criddle, R. S. (1996) Foliar metabolic heat rate of seedlings and mature trees of Pinus 

ponderosa exposed to simulated acid rain and ozone. Plant Cell Environ. 19: 747-753. 

Momen, B.; Anderson, P. D.; Helms, J. A.; Houpis, J. L. J. (1997) Acid rain and ozone effects on gas exchange of 

Pinus ponderosa: a comparison between trees and seedlings. Int. J. Plant Sci. 158: 617-621. 

Momen, B.; Anderson, P. D.; Helms, J. A. (1999) Temperature dependency of acid-rain effect on photosynthesis of 

Pinus ponderosa. For. Ecol. Manage. 113: 223-230. 

Momen, B.; Anderson, P. D.; Houpis, J. L. J.; Helms, J. A. (2002) Growth of ponderosa pine seedlings as affected 

by air pollution. Atmos. Environ. 36: 1875-1882. 

Monteith, J. L.; Elston, J. (1993) Climatic constraints on crop production. In: Fowden, L.; Mansfield, T.; Stoddart, J., 

eds. Plant adaptation to environmental stress. London, United Kingdom: Chapman and Hall; pp. 3-18. 

Morgan, P. B.; Ainsworth, E. A.; Long, S. P. (2003) How does elevated ozone impact soybean? A meta-analysis of 

photosynthesis, growth and yield. Plant Cell Environ. 26: 1317-1328. 

Morgan, P. B.; Bernacchi, C. J.; Ort, D. R.; Long, S. P. (2004) An in vivo analysis of the effect of season-long 

open-air elevation of ozone to anticipated 2050 levels on photosynthesis in soybean. Plant Physiol. 135: 

2348-2357. 

Morgan, P. B.; Mies, T. A.; Bollero, G. A.; Nelson, R. L.; Long, S. P. (2006) Season-long elevation of ozone 

concentration to projected 2050 levels under fully open-air conditions substantially decreases the growth and 

production of soybean. New Phytol.: in press. 

Morison, J. I. L.; Lawlor, D. W. (1999) Interactions between increasing CO2 concentration and temperature on plant 

growth. Plant Cell Environ. 22: 659-682. 

Mortensen, L. M. (1998) Growth responses of seedlings of six Betula pubescens Ehrh Provenances to six ozone 

exposure regimes. Scand. J. For. Res. 13: 189-196. 

Mortensen, L. M. (1999) Foliar injuries caused by ozone in Betula pubescens Ehrh and Phleum pratense L as 

influenced by climatic conditions before and during O3 exposure. Acta Agric. Scand. B 49: 44-49. 

Mortensen, L.; Bastrup-Birk, A. (1996) Critical levels for O3 for biomass production in white clover. In: Kärenlampi, 

L.; Skärby, L., eds. Critical levels for ozone in Europe: testing and finalizing the concepts: UN-ECE 

workshop report. testing and finalizing the concepts: UN-ECE workshop report. UN-ECE Convention on 

Long-Range Transboundary Air Pollution; pp. Kuopio,Finland-Kuopio, Finland. 

Mudd, J. B. (1996) Biochemical basis for the toxicity of ozone. In: Yunus, M.; Iqbal, M., eds. Plant response to air 

pollution. John Wiley & Sons; pp. 267-283. 

Mudd, J. B.; Kozlowski, T. T., eds. (1975) Responses of plants to air pollution. New York, NY: Academic Press. 

(Physiological ecology, a series of monographs, texts, and treatises). 

Mudd, J. B.; Leavitt, R.; Ongun, A.; McManus, T. T. (1969) Reaction of ozone with amino acids and proteins. 

Atmos. Environ. 3: 669-681. 

Mudd, J. B.; McManus, T. T.; Ongun, A. (1971) Inhibition of lipid metabolism in chloroplasts by ozone. In: 

Englund, H. M.; Beery, W. T. eds. Proceedings of the second clean air congress; New York, NY: Academic 

Press; pp. 256-260. 

Mudd, J. B.; Dawson, P. J.; Adams, J. R.; Wingo, J.; Santrock, J. (1996) Reaction of ozone with enzymes of 

erythrocyte membranes. Arch. Biochem. Biophys. 335: 145-151. 

Mudd, J. B.; Dawson, P. J.; Santrock, J. (1997a) Ozone does not react with human erythrocyte membrane lipids. 

Arch. Biochem. Biophys. 341: 251-258. 

Mudd, J. B.; Dawson, P. J.; Tseng, S.; Liu, F. P. (1997b) Reaction of ozone with protein tryptophans: band III, 

serum albumin, and cytochrome C. Arch. Biochem. Biophys. 338: 143-149. 

Mueller, S. F. (1994) Characterization of ambient ozone levels in the Great Smoky Mountains National Park. 

J. Appl. Meteorol. 33: 465-472.

Mukammal, E. I. (1965) Ozone as a cause of tobacco injury. Agric. Meteorol. 2: 145-165. 

AX9-371
Mulchi, C. L.; Slaughter, L.; Saleem, M.; Lee, E. H.; Pausch, R.; Rowland, R. (1992) Growth and physiological 

characteristics of soybean in open-top chambers in response to ozone and increased atmospheric CO2. Agric. 

Ecosyst. Environ. 38: 107-118. 

Mulchi, C.; Rudorff, B.; Lee, E.; Rowland, R.; Pausch, R. (1995) Morphological responses among crop species to 

full-season exposures to enhanced concentrations of atmospheric CO2 and O3. Water Air Soil Pollut. 

85: 1379-1386.

Mulholland, B.; Craigon, J.; Black, C.; Atherton, J.; Landon, G. (1997a) Effects of elevated carbon dioxide and 

ozone on the growth and yield of spring wheat (Triticum aestivum L). J. Exper. Bot. 48: 113-122. 

Mulholland, B. J.; Craigon, J.; Black, C. R.; Colls, J. J.; Atherton, J.; Landon, G. (1997b) Impact of elevated 

atmospheric CO2 and O3 on gas exchange and chlorophyll content in spring wheat (Triticum aestivum L). 

J. Exp. Bot. 48: 1853-1863.

Mulholland, B. J.; Craigon, J.; Black, C. R.; Colls, J. J.; Atherton, J.; Landon, G. (1998a) Growth, light interception 

and yield responses of spring wheat (Triticum aestivum L) grown under elevated CO2 and O3 in open-top 

chambers. Global Change Biol. 4: 121-130. 

Mulholland, B. J.; Craigon, J.; Black, C. R.; Colls, J. J.; Atherton, J.; Landon, G. (1998b) Effects of elevated CO2 

and O3 on the rate and duration of grain growth and harvest index in spring wheat (Triticum aestivum L). 

Global Change Biol. 4: 627-635. 

Muntifering, R. B.; Crosby, D. D.; Powell, M. C.; Chappelka, A. H. (2000) Yield and quality characteristics of 

bahiagrass (Paspalum notatum) exposed to ground-level ozone. Anim. Feed Sci. Technol. 84: 243-256. 

Murphy, J. J.; Deluki, M. A.; McCubbin, D. R.; Kim, H. J. (1999) The cost of crop damage caused by ozone air 

pollution from motor vehicles. J. Environ. Manage. 55: 273-289. 

Musselman, R. C.; Hale, B. A. (1997) Methods for controlled and field ozone exposures of forest tree species in 

North America. In: Sanderman, H.; Wellburn, A. R.; Heath, R. L., eds. Forest decline and ozone. Berlin, 

Germany: Springer Verlag; pp. 277-315 (Ecological Studies v. 127). 

Musselman, R. C.; Massman, W. J. (1999) Ozone flux to vegetation and its relationship to plant response and 

ambient air quality standards. Atmos. Environ. 33: 65-73. 

Musselman, R. C.; Minnick, T. J. (2000) Nocturnal stomatal conductance and ambient air quality standards for 

ozone. Atmos. Environ. 34: 719-733. 

Musselman, R. C.; Oshima, R. J.; Gallavan, R. E. (1983) Significance of pollutant concentration distribution in the 

response of 'red kidney' beans to ozone. J. Am. Soc. Hortic. Sci. 108: 347-351. 

Musselman, R. C.; Huerta, A. J.; McCool, P. M.; Oshima, R. J. (1986) Response of beans to simulated ambient and 

uniform ozone distributions with equal peak concentration. J. Am. Soc. Hortic. Sci. 111: 470-473. 

Musselman, R. C.; McCool, P. M.; Younglove, T. (1988) Selecting ozone exposure statistics for determining crop 

yield loss from air pollutants. Environ. Pollut. 53: 63-78. 

Musselman, R. C.; Younglove, T.; McCool, P. M. (1994) Response of Phaseolus vulgaris L to differing ozone 

regimes having indentical total exposure. Atmos. Environ. 28: 2727-2731. 

Muzika, R. M.; Guyette, R. P.; Zielonka, T.; Liebhold, A. M. (2004) The influence of O3, NO2, and SO2 on growth 

of Picea abies and Fagus sylvatica in the Carpathian Mountains. Environ. Pollut. 130: 65-71. 

Nali, C.; Pucciariello, C.; Lorenzini, G. (2002) Mapping ozone critical levels for vegetation in central Italy. Water 

Air Soil Pollut. 141: 337-347. 

Nash, T. H., III; Sigal, L. L. (1999) Epiphytic lichens in the San Bernardino mountains in relation to oxidant 

gradients. In: Miller, P. R.; McBride, J. R., eds. Oxidant air pollution impacts in the montane forests of 

southern California: a case study of the San Bernardino Mountains. New York, NY: Springer-Verlag; 

pp 223-234. (Ecological studies: v. 134). 

Neill, S. J.; Desikan, R.; Clarke, A.; Hurst, R. D.; Hancock, J. T. (2002) Hydrogen peroxide and nitric oxide as 

signalling molecules in plants. J. Exper. Bot. 53: 1237-1247. 

Neufeld, H. S.; Renfro, J. R.; Hacker, W. D.; Silsbee, D. (1992) Ozone in Great Smoky Mountains National Park: 

dynamics and effects on plants. In: Berglund, R. L. ed. Tropospheric ozone and the environment II: effects, 

modeling and control [papers from an international specialty conference]; November; Atlanta, GA; 

Pittsburgh, PA: Air & Waste Management Association; pp. 594-617. (A&WMA transactions series no. 20). 

Neufeld, H. S.; Lee, E. H.; Renfro, J. R.; Hacker, W. D.; Yu, B. H. (1995) Sensitivity of seedlings of black cherry 

(Prunus serotina Ehrh) to ozone in Great Smoky Mountains National Park I Exposure-response curves for 

biomass. New Phytol. 130: 447-459. 

Neufeld, H. S.; Lee, E. H.; Renfro, J. R.; Hacker, W. D. (2000) Seedling insensitivity to ozone for three conifer 

species native to Great Smoky Mountains National Park. Environ. Pollut. 108: 141-151. 

AX9-372
Niewiadomska, E.; Gaucher-Veilleux, C.; Chevrier, N.; Mauffette, Y.; Dizengremel, P. (1999) Elevated CO2 does 

not provide protection against ozone considering the activity of several antioxidant enzymes in the leaves of 

sugar maple. J. Plant Physiol. 155: 70-77. 

Nijs, I.; Teughels, H.; Blum, H.; Hendrey, G.; Impens, I. (1996) Simulation of climate change with infrared heaters 

reduces the productivity of Lolium perenne L in summer. Environ. Exp. Bot. 36: 271-280. 

Nikolov, N.; Zeller, K. (2003) Modeling coupled interactions of carbon, water, and ozone exchange between 

terrestrial ecosystems and the atmosphere. Environ. Pollut. 124: 231-246. 

Noormets, A.; McDonald, E. P.; Dickson, R. E.; Kruger, E. L.; Sôber, A.; Isebrands, J. G.; Karnosky, D. F. (2001a) 

The effect of elevated carbon dioxide and ozone on leaf- and branch-level photosynthesis and potential 

plant-level carbon gain in aspen. Trees 15: 262-270. 

Noormets, A.; Sôber, A.; Pell, E. J.; Dickson, R. E.; Posila, G. K.; Sôber, J.; Isebrands, J. G.; Karnosky, D. F. 

(2001b) Stomatal and non-stomatal limitation to photosynthesis in two trembling aspen (Populus tremuloides 

Michx) clones exposed to elevated CO2 and/or O3. Plant Cell Environ. 24: 327-336. 

Norby, R. J.; Weerasuriya, Y.; Hanson, P. J. (1989) Induction of nitrate reductase activity in red spruce needles by 

NO2 and HNO3 vapor. Can. J. For. Res. 19: 889-896. 

Norby, R. J.; Wullschleger, S. D.; Gunderson, C. A.; Johnson, D. W.; Ceulemans, R. (1999) Tree responses to rising 

CO2 in field experiments: implications for the future forest. Plant Cell Environ. 22: 683-714. 

Norton, B. G. (1987) Why preserve natural variety? Princeton, NJ: Princeton University Press. 

Nouchi, I.; Toyama, S. (1988) Effects of ozone and peroxyacetyl nitrate on polar lipids and fatty acids in leaves of 

morning glory and kidney bean. Plant Physiol. 87: 638-646. 

Novak, K.; Schaub.M.; Kräuchi, N.; Hug, C.; Landolt, W.; Bleuler, P. (2003) Ozone air pollution and foliar injury 

development on native plants of Switzerland. Environ. Pollut. 125: 41-52. 

Nunn, A. J.; Reiter, I. M.; Haberle, K. H.; Werner, H.; Langebartels, C.; Sandermann, H.; Heerdt, C.; Fabian, P.; 

Matyssek, R. (2002) "Free-air" ozone canopy fumigation in an old-growth mixed forest: concept and 

observations in beech. Phyton 42: 105-119. 

Nussbaum, S.; Fuhrer, J. (2000) Difference in ozone uptake in grassland species between open-top chambers and 

ambient air. Environ. Pollut. 109: 463-471. 

Nussbaum, S.; Geissmann, M.; Fuhrer, J. (1995a) Effects of nitric oxide and ozone on spring wheat (Triticum 

aestivum). Water Air Soil Pollut. 85: 1449-1454. 

Nussbaum, S.; Geissmann, M.; Fuhrer, J. (1995b) Ozone exposure-response relationships for mixtures of perennial 

ryegrass and white clover depend on ozone exposure patterns. Atmos. Environ. 29: 989-995. 

Nussbaum, S.; Ammann, M.; Fuhrer, J. (1999) Foliar absorption and use of airborne oxidized nitrogen by terrestrial 

plants. In: Srivastava, H. S.; Singh, R. P., eds. Nitrogen nutrition and plant growth. New Delhi, India: Oxford 

& IBH Publishing Co. Pvt. Ltd; pp. 101-172. 

Nussbaum, S.; Bungener, P.; Geissmann, M.; Fuhrer, J. (2000a) Plant-plant interactions and soil moisture might be 

important in determining ozone impacts on grasslands. New Phytol. 147: 327-335. 

Nussbaum, S.; Geissman, M.; Saurer, M.; Siegwolf, R.; Fuhrer, J. (2000b) Ozone and low concentrations of nitric 

oxide have similar effects on carbon isotope discrimination and gas exchange in leaves of wheat (Triticum 

aestivum L). J. Plant Physiol. 156: 741-745. 

Nussbaum, S.; Remund, J.; Rihm, B.; Mieglitz, K.; Gurtz, J.; Fuhrer, J. (2003) High-resolution spatial analysis of 

stomatal ozone uptake in arable crops and pastures. Environ. Int. 29: 385-392. 

Odum, E. P. (1963) Ecology. New York, NY: Holt, Rinehart and Winston. (Modern biology series). 

O'Gara, P. J. (1922) Sulfur dioxide and fume problems and their solution. In: Olsen, J. C., ed. Proceedings of the 

14th semiannual meeting of the American Institute of Chemical Engineers; June; Niagara Falls, Canada. 

Proceedings and papers of the meeting summarized in J. Ind. Eng. Chem. 14: 744-745 by J. C. Olsen. 

Ojanperä, K.; Pätsikkä, E.; Yläranta, T. (1998) Effects of low ozone exposure of spring wheat on net CO2 uptake, 

Rubisco, leaf senescence and grain filling. New Phytol. 138: 451-460. 

Oksanen, E. J. (2001) Increasing tropospheric ozone level reduced birch (Betula pendula) dry mass within a five 

years period. Water Air Soil Pollut. 130: 947-952. 

Oksanen, E. (2003a) Responses of selected birch (Betula pendula Roth) clones to ozone change over time. Plant Cell 

Environ. 26: 875-886. 

Oksanen, E. (2003b) Physiological responses of birch (Betula pendula) to ozone: a comparison between 

open-soil-grown trees exposed for six growing seasons and potted seedlings exposed for one season. 

Tree Physiol. 23: 603-614. 

Oksanen, E.; Holopainen, T. (2001) Responses of two birch (Betula pendula Roth) clones to different ozone profiles 

with similar AOT40 exposure. Atmos. Environ. 35: 5245-5254. 

AX9-373
Oksanen, E.; Rousi, M. (2001) Differences in Betula origins in ozone sensitivity based open-field fumigation 

experiment over two growing seasons. Can. J. For. Res. 31: 804-811. 

Oksanen, E.; Saleem, A. (1999) Ozone exposure results in various carry-over effects and prolonged reduction in 

biomass in birch (Betula pendula Roth). Plant Cell Environ. 22: 1401-1411. 

Oksanen, E.; Amores, A.; Kokko, H.; Santamaria, J. M.; Kärenlampi, L. (2001a) Genotypic variation in growth and 

physiological responses of Finnish hybrid aspen (Populus tremuloides x P tremula) to elevated tropospheric 

ozone concentration. Tree Physiol. 21: 1171-1181. 

Oksanen, E.; Sober, J.; Karnosky, D. F. (2001b) Interactions of elevated CO2 and ozone in leaf morphology of aspen 

(Populus tremuloides) and birch (Betula papyrifera) in aspen FACE experiment. Environ. Pollut. 

115: 437-446. 

Ollerenshaw, J. H.; Lyons, T. (1999) Impacts of ozone on the growth and yield of field-grown winter wheat. 

Environ. Pollut. 106: 67-72. 

Ollerenshaw, J. H.; Lyons, T.; Barnes, J. D. (1999) Impacts of ozone on the growth and yield of field-grown winter 

oilseed rape. Environ. Pollut. 104: 53-59. 

Ollinger, S. V.; Aber, J. D.; Reich, P. B. (1997) Simulating ozone effects on forest productivity: interactions among 

leaf-, canopy-, and stand-level processes. Ecol. Appl. 7: 1237-1251. 

Ollinger, S. V.; Aber, J. D.; Federer, C. A. (1998) Estimating regional forest productivity and water yield using an 

ecosystem model linked to a GIS. Landscape Ecol. 13: 323-334. 

Olszyk, D. M.; Bytnerowicz, A.; Kats, G.; Dawson, P. J.; Wolf, J.; Thompson, C. R. (1986a) Crop effects from air 

pollutants in air exclusion systems vs field chambers. J. Environ. Qual. 15: 417-422. 

Olszyk, D. M.; Kats, G.; Dawson, P. J.; Bytnerowicz, A.; Wolf, J.; Thompson, C. R. (1986b) Characteristics of air 

exclusion systems vs. chambers for filed air pollution studies. J. Environ. Qual. 15: 326-334. 

Olszyk, D. M.; Takemoto, B. K.; Kats, G.; Dawson, P. J.; Morrison, C. L.; Preston, J. W.; Thompson, C. R. (1992) 

Effects of open-top chambers on 'Valencia' orange trees. J. Environ. Qual. 21: 128-134. 

Olszyk, D. M.; Tingey, D. T.; Watrud, L.; Seidler, R.; Andersen, C. (2000) Interactive effects of O3 and CO2: 

implications for terrestrial ecosystems. In: Singh, S. N., ed. Trace gas emissions and plants. Dordrecht, 

The Netherlands: Kluwer Academic Publishers; pp. 97-136. 

Olszyk, D. M.; Johnson, M. G.; Phillips, D. L.; Seidler, R. J.; Tingey, D. T.; Watrud, L. S. (2001) Interactive effects 

of CO2 and O3 on a ponderosa pine/litter/soil mesocosm. Environ. Pollut. 115: 447-462. 

Ommen, O. E.; Donnelly, A.; Vanhoutvin, S.; Van Oijen, M.; Manderscheid, R. (1999) Chlorophyll content of 

spring wheat flag leaves grown under elevated CO2 concentrations and other environmental stresses within 

the 'ESPACE-wheat' project. Eur. J. Agron. 10: 197-203. 

Ordin, L.; Hall, M. A.; Kindinger, J. I. (1969) Oxidant induced inhibition of enzymes involved in cell wall 

polysaccharide synthesis. Arch. Environ. Health 18: 623-626. 

Oren, R.; Ellsworth, D. S.; Johnsen, K. H.; Phillips, N.; Ewers, B. E.; Maier, C.; Schäfer, K. V. R.; McCarthy, H.; 

Hendrey, G.; McNulty, S. G.; Katul, G. G. (2001) Soil fertility limits carbon sequestration by forest 

ecosystems in a CO2-enriched atmosphere. Nature (London) 411: 469-472. 

Orendovici, T.; Skelly, J. M.; Ferdinand, J. A.; Savage, J. E.; Sanz, M.-J.; Smith, G. C. (2003) Response of native 

plants of northeastern United States and southern Spain to ozone exposures; determining exposure/response 

relationships. Environ. Pollut. 125: 31-40. 

Ormrod, D. P.; Beckerson, D. W. (1986) Polyamines as antiozonants for tomato. HortScience 21: 1070-1071. 

Ormrod, D. P.; Landry, L. G.; Conklin, P. L. (1995) Short-term UV-B radiation and ozone exposure effects on 

aromatic secondary metabolite accumulation and shoot growth of flavonoid-deficient Arabidopsis mutants. 

Physiol. Plant. 93: 602-610.

Örvar, B. L.; Ellis, B. E. (1997) Transgenic tobacco plants expressing antisense RNA for cytosolid ascorbate 

peroxidase show increased susceptibility to ozone injury. Plant J. 11: 1297-1305. 

Oshima, R. J.; Poe, M. P.; Braegelmann, P. K.; Baldwin, D. W.; Van Way, V. (1976) Ozone dosage-crop loss 

function for alfalfa: a standardized method for assessing crop losses from air pollutants. J. Air Pollut. Control 

Assoc. 26: 861-865. 

Oshima, R. J.; Braegelmann, P. K.; Baldwin, D. W.; Van Way, V.; Taylor, O. C. (1977) Responses of five cultivars 

of fresh market tomato to ozone: a contrast of cultivar screening with foliar injury and yield. J. Am. Soc. 

Hortic. Sci. 102: 286-289. 

Ottosson, S.; Wallin, G.; Skärby, L.; Karlsson, P. E.; Medin, E. L.; Räntfors, M.; Pleijel, H.; Selldén, G. (2003) 

Four years of ozone exposure at high or low phosphorus reduced biomass in Norway spruce. Trees 

(Heidelberg, Ger.) 17: 299-307. 

AX9-374
Overmyer, K.; Tuominen, H.; Kettunen, R.; Betz, C.; Langebartels, C.; Sandermann, H., Jr.; Kangasjärvi, J. (2000) 

Ozone-sensitive Arabidopsis rcd1 mutant reveals opposite roles for ethylene and jasmonate signaling 

pathways in regulating superoxide-dependent cell death. Plant Cell 12: 1849-1862. 

Overmyer, K.; Brosché, M.; Kangasjärvi, J. (2003) Reactive oxygen species and hormonal control of cell death. 

Trends Plant Sci. 8: 335-342. 

Pääkkönen, E.; Holopainen, T. (1995) Influence of nitrogen supply on the response of clones of birch (Betula 

pendula Roth) to ozone. New Phytol. 129: 595-603. 

Pääkkönen, E.; Holopainen, T.; Kärenlampi, L. (1996) Relationships between open-field ozone exposures and 

growth and senescence of birch (Betula pendula and Betula pubescens). In: Kärenlampi, L.; Skärby, L. eds. 

Critical levels for ozone in Europe: testing and finalizing the concepts: UN-ECE workshop report. UN-ECE 

Convention on Long-Range Transboundary Air Pollution; April; Kuopio, Finland. Kuopio, Finland: 

University of Kuopio, Department of Ecology and Environmental Science; pp. 298-302. 

Pääkkönen, E.; Holopainen, T.; Kärenlampi, L. (1997) Variation in ozone sensitivity among clones of Betula 

pendula and Betula pubescens. Environ. Pollut. 95: 37-44. 

Pääkkönen, E.; Günthardt-Georg, M. S.; Holopainen, T. (1998) Responses of leaf processes in a sensitive birch 

(Betula pendula Roth) clone to ozone combined with drought. Ann. Bot. 82: 49-59. 

Paludan-Müller, G.; Saxe, H.; Leverenz, J. W. (1999) Responses to ozone in 12 provenances of European beech 

(Fagus sylvatica): genotypic variation and chamber effects on photosynthesis and dry-matter partitioning. 

New Phytol. 144: 261-273. 

Panek, J. A. (2004) Ozone uptake, water loss and carbon exchange dynamics in annually drought-stressed Pinus 

ponderosa forests: measured trends and parameters for uptake modeling. Tree Physiol. 24: 277-290. 

Panek, J. A.; Goldstein, A. H. (2001) Responses of stomatal conductance to drought in ponderosa pine: implications 

for carbon and ozone uptake. Tree Physiol. 21: 337-344. 

Panek, J.; Kurpius, M. R.; Goldstein, A. H. (2002) An evaluation of ozone exposure metrics for a seasonally 

drought-stressed ponderosa pine ecosystem. Environ. Pollut. 117: 93-100. 

Panek, J. A.; Baldocchi, D. D.; Goldstein, A. H. (2003) The need for spatially and functionally integrated models of 

ozone deposition to Sierra Nevada forests. In: Bytnerowicz, A.; Arbaugh, M. J.; Alonso, R., eds. Ozone air 

pollution in the Sierra Nevada: distribution and effects on forests. New York, NY: Elsevier Science Ltd; 

pp. 325-357. (Developments in environmental science: v. 2). 

Patterson, M. T.; Rundel, P. W. (1995) Stand characteristics of ozone-stressed populations of Pinus jeffreyi 

(Pinaceae): extent, development, and physiological consequences of visible injury. Am. J. Bot. 82: 150-158. 

Pauls, K. P.; Thompson, J. E. (1980) In vitro simulation of senescence-related membrane damage by ozone-induced 

lipid peroxidation. Nature (London) 283: 504-506. 

Payne, G.; Middlesteadt, W.; Desai, N.; Williams, S.; Dincher, S.; Carnes, M.; Ryals, J. (1989) Isolation and 

sequence of a genomic clone encoding the basic form of pathogenesis-related protein-1 from Nicotiana 

tabacum. Plant Mol. Biol. 12: 595-596. 

Payne, G.; Ahl, P.; Moyer, M.; Harper, A.; Beck, J.; Meins, F., Jr.; Ryals, J. (1990) Isolation of complementary 

DNA clones encoding pathogenesis-related proteins P and Q, two acidic chitinases from tobacco. Proc. Natl. 

Acad. Sci. U. S. A. 87: 98-102. 

Pearce, D. W. (1993) Economic values and the natural world. London, United Kingdom: Earthscan Publications, 

Ltd.. 

Pearson, M.; Mansfield, T. A. (1993) Interacting effects of ozone and water stress on the stomatal resistance of 

beech (Fagus sylvatica L). New Phytol. 123: 351-358. 

Pearson, M.; Mansfield, T. A. (1994) Effects of exposure to ozone and water stress on the following season's growth 

of beech (Fagus sylvatica L). New Phytol. 126: 511-515. 

Pearson, S.; Davison, A. W.; Reiling, K.; Ashenden, T.; Ollerenshaw, J. H. (1996) The effects of different ozone 

exposures on three contrasting populations of Plantago major. New Phytol. 132: 493-502. 

Pel, Z.-M.; Murata, Y.; Benning, G.; Thomine, S.; Klüsener, B.; Allen, G. J.; Grill, E.; Schroeder, J. I. (2000) 

Calcium channels activated by hydrogen peroxide mediate abscisic acid signalling in guard cells. Nature 

(London) 406: 731-734. 

Pell, E. J.; Pearson, N. S. (1984) Ozone-induced reduction in quantity and quality of two potato cultivars. Environ. 

Pollut. Ser. A 35: 345-352. 

Pell, E. J.; Eckardt, N.; Glick, R. E. (1994) Biochemical and molecular basis for impairment of photosynthetic 

potential. Photosynth. Res. 39: 453-462. 

Pell, E. J.; Sinn, J. P.; Johansen, C. V. (1995) Nitrogen supply as a limiting factor determining the sensitivity of 

Populus tremuloides Michx to ozone stress. New Phytol. 130: 437-446. 

AX9-375
Pell, E. J.; Schlagnhaufer, C. D.; Arteca, R. N. (1997) Ozone-induced oxidative stress: mechanisms of action and 

reaction. Physiol. Plant. 100: 264-273. 

Pell, E. J.; Sinn, J. P.; Brendley, B. W.; Samuelson, L.; Vinten-Johansen, C.; Tien, M.; Skillman, J. (1999) 

Differential response of four tree species to ozone-induced acceleration of foliar senescence. Plant Cell 

Environ. 22: 779-790. 

Pellinen, R.; Palva, T.; Kangasjävi, J. (1999) Subcellular localization of ozone-induced hydrogen peroxide 

production in birch (Betula pendula) leaf cells. Plant. J. 20: 349-356. 

Pelloux, J.; Jolivet, Y.; Fontaine, V.; Banvoy, J.; Dizengremel, P. (2001) Changes in Rubisco and Rubisco activase 

gene expression and polypeptide content in Pinus halepensis M subjected to ozone and drought. Plant Cell 

Environ. 24: 123-131. 

Penel, C. (1986) The role of calcium in the control of peroxidase activity. In: Greppin, H.; Penel, C.; Gasper, T., eds. 

Molecular and physiological aspects of plant peroxidases. Universite de Genieve, Centre de Botanique; 

pp. 155-164. 

Percy, K. E.; Jensen, K. F.; McQuattie, C. J. (1992) Effects of ozone and acidic fog on red spruce needle epicuticular 

wax production, chemical composition, cuticular membrane ultrastructure and needle wettability. 

New Phytol. 122: 71-80. 

Percy, K. E.; Awmack, C. S.; Lindroth, R. L.; Kubiske, M. E.; Kopper, B. J.; Isebrands, J. G.; Pregitzer, K. S.; 

Hendry, G. R.; Dickson, R. E.; Zak, D. R.; Oksanen, E.; Sober, J.; Harrington, R.; Karnosky, D. F. (2002) 

Altered performance of forest pests under atmospheres enriched with CO2 and O3. Nature (London) 

420: 403-407. 

Persson, K.; Danielsson, H.; Seldén, G.; Pleijel, H. (2003) The effects of tropospheric ozone and elevated carbon 

dioxide on potato (Solanum tuberosum L cv Bintje) growth and yield. Sci. Total Environ. 310: 191-201. 

Peters, J. L.; Castillo, F. J.; Heath, R. L. (1988) Alteration of extracellular enzymes in pinto bean leaves upon 

exposure to air pollutants, ozone and sulfur dioxide. Plant Physiol. 89: 159-164. 

Peterson, D. L.; Arbaugh, M. J.; Wakefield, V. A.; Miller, P. R. (1987) Evidence of growth reduction in 

ozone-injured Jeffrey pine (Pinus jeffreyi Grev and Balf) in Sequoia and Kings Canyon National Parks. 

JAPCA 37: 906-912. 

Peterson, D. L.; Arbaugh, M. J.; Robinson, L. J. (1991) Regional growth changes in ozone-stressed ponderosa pine 

(Pinus ponderosa) in the Sierra Nevada, California, USA. Holocene 1: 50-61. 

Peterson, D. L.; Arbaugh, M. J.; Robinson, L. J. (1993) Effects of ozone and climate on ponderosa pine 

(Pinus ponderosa) growth in the Colorado Rocky Mountains. Can. J. For. Res. 23: 1750-1759. 

Peterson, D. L.; Silsbee, D. G.; Poth, M.; Arbaugh, M. J.; Biles, F. E. (1995) Growth response of bigcone Douglas 

fir (Pseudotsuga macrocarpa) to long-term ozone exposure in southern California. J. Air Waste Manage. 

Assoc. 45: 36-45. 

Pfirrmann, T.; Barnes, J. D. (1993) Interactive effects of CO2 and O3 on growth of radish (Raphanus sativus L). 

Vegetatio 104-105: 464-466. 

Phillips, R. L.; Zak, D. R.; Holmes, W. E.; White, D. C. (2002) Microbial community composition and function 

beneath temperate trees exposed to elevated atmospheric carbon dioxide and ozone. Oecologia 131: 236-244. 

Pimentel, D.; Wilson, C.; McCullum, C.; Huang, R.; Dwen, P.; Flack, J.; Tran, Q.; Saltman, T.; Cliff, B. (1997) 

Economic and environmental benefits of biodiversity. BioScience 47: 747-757. 

Pimm, S. L. (1997) The value of everything. Nature (London) 387: 231-232. 

Pino, M. E.; Mudd, J. B.; Bailey-Serres, J. (1995) Ozone-induced alterations in the accumulation of newly 

synthesized proteins in leaves of maize. Plant Physiol. 108: 777-785. 

Pitcher, L. H.; Zilinskas, B. A. (1996) Overexpression of copper/zinc superoxide dismutase in the cytosol of 

transgenic tobacco confers partial resistance to ozone-induced foliar necrosis. Plant Physiol. 110: 583-588. 

Pitelka, L. F. (1988) Evolutionary responses of plants to anthropogenic pollutants. Trends in Ecology and Evolution 

3: 233-236.

Pleijel, H. (1998) A suggestion of a simple transfer function for the use of ozone monitoring data in dose-response 

relationships obtained using open-top chambers. Water Air Soil Pollut. 102: 61-74. 

Pleijel, H.; Danielsson, H. (1997) Growth of 27 herbs and grasses in relation to ozone exposure and plant strategy. 

New Phytol. 135: 361-367. 

Pleijel, H.; Skärby, L.; Ojanpera, K.; Selldén, G. (1994a) Exposure of oats, Avena sativa L, to filtered and unfiltered 

air in open-top chambers: Effects on grain yield and quality. Environ. Pollut. 86: 129-134. 

Pleijel, H.; Wallin, G.; Karlsson, P. E.; Skärby, L.; Selldén, G. (1994b) Ozone deposition to an oat crop (Avena 

sativa L) grown in open-top chambers and in the ambient air. Atmos. Environ. 28: 1971-1979. 

AX9-376
Pleijel, H.; Karlsson, G. P.; Danielsson, H.; Selldén, G. (1995a) Surface wetness enhances ozone deposition to a 

pasture canopy. Atmos. Environ. 29: 3391-3393. 

Pleijel, H.; Wallin, G.; Karlsson, P.; Skärby, L.; Selldén, G. (1995b) Gradients of ozone at a forest site and over a 

field crop - consequences for the AOT4O concept of critical level. Water Air Soil Pollut. 85: 2033-2038. 

Pleijel, H.; Karlsson, G. P.; Sild, E.; Danielsson, H.; Skärby, L.; Selldén, G. (1996) Exposure of a grass-clover 

mixture to ozone in open-top chambers–effects on yield, quality and botanical composition. Agric. Ecosyst. 

Environ. 59: 55-62. 

Pleijel, H.; Ojanperä, K.; Mortensen, L. (1997) Effects of tropospheric ozone on the yield and grain protein content 

of spring wheat (Triticum aestivum L) in the nordic countries. Acta Agric. Scand. Sect. B 47: 20-25. 

Pleijel, H.; Danielsson, H.; Gelang, J.; Sild, E.; Selldén, G. (1998) Growth stage dependence of the grain yield 

response to ozone in spring wheat (Triticum aestivum L). Agric. Ecosyst. Environ. 70: 61-68. 

Pleijel, H.; Mortensen, L.; Fuhrer, J.; Ojanperä, K.; Danielsson, H. (1999a) Grain protein accumulation in relation to 

grain yield of spring wheat (Triticum aestivum L) grown in open-top chambers with different concentrations 

of ozone, carbon dioxide and water availability. Agric. Ecosyst. Environ. 72: 265-270. 

Pleijel, H.; Norberg, P. A.; Selldén, G.; Skärby, L. (1999b) Tropospheric ozone decreases biomass production in 

radish plants (Raphanus sativus) grown in rural south-west Sweden. Environ. Pollut. 106: 143-147. 

Pleijel, H.; Danielsson, H.; Karlsson, G. P.; Gelang, J.; Karlsson, P. E.; Selldén, G. (2000a) An ozone flux-response 

relationship for wheat. Environ. Pollut. 109: 453-462. 

Pleijel, H.; Gelang, J.; Sild, E.; Danielsson, H.; Younis, S.; Karlsson, P. E.; Wallin, G.; Skärby, L.; Selldén, G. 

(2000b) Effects of elevated carbon dioxide, ozone and water availability on spring wheat growth and yield. 

Physiol. Plant. 108: 61-70. 

Pleijel, H.; Danielsson, h.; Ojanperä, K.; De Temmerman, L.; Högy, P. (2002a) Relationships between ozone 

exposure and yield loss in wheat and potato - suggestions of critical levels for ozone effects on crops. In: 

Background Paper: Crops, workshop establishing ozone critical levels II; November; Göteborg, Sweden. 

Göteborg, Sweden: Göteborg University, IVL Swedish Environmental Research Institute; pp. 18-30. 

Pleijel, H.; Danielsson, H.; Vandermeiren, K.; Blum, C.; Colls, J.; Ojanperä, K. (2002b) Stomatal conductance and 

ozone exposure in relation to potato tuber yield–results from the European CHIP programme. Eur. J. Agron. 

17: 303-317.

Pleijel, H.; Danielsson, H.; Ojanpera, K.; De Temmerman, L.; Hogy, P.; Badiani, M.; Karlsson, P. E. (2004) 

Relationship between ozone exposure and yield loss in European wheat and potato–a comparison of 

concentration- and flux-based exposure indices. Atmos. Environ. 38: 2259-2269. 

Plöchl, M.; Lyons, T.; Ollerenshaw, J.; Barnes, J. (2000) Simulating ozone detoxification in the leaf apoplast 

through the direct reaction with ascorbate. Planta 210: 454-467. 

Polasky, S. ed. (2001) The economics of biodiversity conservation. Burlington, VT: Ashgate Publishers. 

Polle, A. (1998) Photochemical oxidants: uptake and detoxification mechanisms. In: De Kok, L. J.; Stulen, I., eds. 

Responses of plant metabolism to air pollution and global change. Leiden, The Netherlands: Backhuys 

Publishers; pp. 95-116. 

Polle, A.; Junkermann, W. (1994) Inhibition of apoplastic and symplastic peroxidase activity from Norway spruce 

by the photooxidant hydroxymethyl hydroperoxide. Plant Physiol. 104: 617-621. 

Polle, A.; Pell, E. J. (1999) Role of carbon dioxide in modifying the plant response to ozone. In: Luo, Y.; Mooney, 

H. A., eds. Carbon dioxide and environmental stress. Academic Press; pp. 193-213. 

Polle, A.; Rennenberg, H. (1991) Superoxide dismutase activity in needles of Scots pine and Norway spruce under 

field and changer conditions: lack of ozone effects. New Phytol. 117: 335-343. 

Polle, A.; Krings, B.; Rennenberg, H. (1989) Superoxide dismutase activity in needles of Norwegian spruce trees 

(Picea abies l). Plant Physiol. 90: 1310-1315. 

Poorter, H.; Pérez-Soba, M. (2001) The growth response of plants to elevated CO2 under non-optimal environmental 

conditions. Oecologia 129: 1-20. 

Posthumus, A. C. (1998) Air pollution and global change: significance and prospectives. In: De Kok, L. J.; Stulen, I., 

eds. Responses of plant metabolism to air pollution and global change. Leiden, The Netherlands: Backhuys 

Publishers; pp. 3-14. 

Postiglione, L.; Fagnano, M.; Merola, G. (2000) Response to ambient ozone of two white clover (Trifolium repens 

Lcv"Regal") biotypes, one resistant and one sensitive, grown in a Mediterranean environment. Environ. 

Pollut. 109: 525-531. 

AX9-377
Poth, M. A.; Fenn, M. E. (1998) Mature ponderosa pine nutrient use and allocation responses to air pollution. In: 

Bytnerowicz, A.; Arbaugh, M. J.; Schilling, S. L. Proceedings of the International Symposium on Air 

Pollution and Climate Change Effects on Forest Ecosystems; February, 1996; Riverside, CA. Albany, CA: 

USDA Forest Service, Pacific Southwest General Technical Report no. PSW-GTR 166; pp. 239-247. 

Potter, L.; Foot, J. P.; Caporn, S. J. M.; Lee, J. A. (1996) The effects of long-term elevated ozone concentrations on 

the growth and photosynthesis of Sphagnum recurvum and Polytrichum commune. New Phytol. 134: 649-656. 

Powell, M. C.; Muntifering, R. B.; Lin, J. C.; Chappelka, A. H. (2003) Yield and nutritive quality of sericea 

lespedeza (Lespedeza cuneata) and little bluestem (Schizachyrium scoparium) exposured to ground-level 

ozone. Environ. Pollut. 122: 313-322. 

Powers, R. F. (1981) Nutritional ecology of Ponderosa pine (Pinus ponderosa L.) and associated species 

[dissertation]. Berkeley, CA: University of California. 

Pronos, J.; Merrill, L.; Dahlsten, D. (1999) Insects and pathogens in a pollution-stressed forest. In: Miller, P. R.; 

McBride, J. R., eds. Oxidant air pollution impacts in the montane forests of southern California. Springer, pp. 

317-337. 

Prozherina, N.; Freiwald, V.; Rousi, M.; Oksanen, E. (2003) Interactive effect of springtime frost and elevated ozone 

on early growth, foliar injuries and leaf structure of birch (Betula pendula). New Phytol. 159: 623-636. 

Pryor, W. A. (1992) How far does ozone penetrate into the pulmonary air/tissue boundary before it reacts? Free 

Radic. Biol. Med. 12: 83-88. 

Qui, Z.; Chappelka, A. H.; Somers, G. L.; Lockaby, B. G.; Meldahl, R. S. (1993) Effects of ozone and simulated 

acidic precipitation on ectomycorrhizal formation on loblolly pine seedlings. Environ. Exp. Bot. 33: 423-431. 

Rabinowitch, E. I. (1945) Photosynthesis and related processes Vol 1: chemistry of photosynthesis, chemosynthesis 

and related processes in vitro and in vivo. New York, NY: Interscience Publishing, Inc.. 

Rafarel, C. R.; Ashenden, T. W. (1991) A facility for the large-scale exposure of plants to gaseous atmospheric 

pollutants. New Phytol. 117: 345-349. 

Randall, A. (1991) The value of biodiversity. Ambio 20: 64-68. 

Rangnekar, P. V.; Forward, D. F.; Nolan, N. J. (1969) Foliar nutrition and wood growth in red pine: the distribution 

of radiocarbon photoassimilated by individual branches of young trees. Can. J. Bot. 47: 1701-1711. 

Rantanen, L.; Palomäki, V.; Holopainen, T. (1994) Interactions between exposure to O3 and nutrient status of trees: 

effects on nutrient content and uptake, growth, mycorrhiza and needle ultrastructure. New Phytol. 

128: 679-687. 

Rao, M.; Davis, K. R. (1999) Ozone-induced cell death occurs via two distinct mechanisms in Arabidopsis: the role 

of salicylic acid. Plant J. 17: 603-614. 

Rao, M. V.; Davis, K. R. (2001) The physiology of ozone induced cell death. Planta 213: 682-690. 

Rao, M. V.; Hale, B. A.; Ormrod, D. P. (1995) Amelioration of ozone-induced oxidative damage in wheat plants 

grown under high carbon dioxide Role of antioxidant enzymes. Plant Physiol. 109: 421-432. 

Rao, M. V.; Lee, H. I.; Davis, K. R. (2002) Ozone-induced ethylene production is dependent on salicylic acid, and 

both salicylic acid and ethylene act in concert to regulate ozone-induced cell death. Plant J. 32: 447-456. 

Read, J. J.; Morgan, J. A. (1996) Growth and partitioning in Pascopyrum smithii (C3) and Bouteloua gracilis (C4) 

as influenced by carbon dioxide and temperature. Ann. Bot. (London, UK) 77: 487-496. 

Reams, G. A.; Van Deusen, P. C.; Lucier, A. A. (1995) Ambient ozone and loblolly pines. Nature (London) 

378: 449-450. 

Rebbeck, J. (1996) Chronic ozone effects on three northeastern hardwood species: growth and biomass. Can. J. For. 

Res. 26: 1788-1798. 

Rebbeck, J.; Scherzer, A. J. (2002) Growth responses of yellow-poplar (Liriodendron tulipifera L.) exposed to 

5 years of O3 alone or combined with elevated CO2. Plant Cell Environ. 25: 1527-1537. 

Rebbeck, J.; Blum, U.; Heagle, A. S. (1988) Effects of ozone on the regrowth and energy reserves of a ladino 

clover-tall fescue pasture. J. Appl. Ecol. 25: 659-681. 

Rebbeck, J.; Jensen, K. F.; Greenwood, M. S. (1993) Ozone effects on grafted mature and juvenile red spruce: 

photosynthesis, stomatal conductance, and chlorophyll concentration. Can. J. For. Res. 23: 450-456. 

Reddy, G. B.; Reinert, R. A.; Eason, G. (1991) Effect of acid rain and ozone on soil and secondary needle nutrients 

of loblolly pine. In: Wright, R. J.; Baligar, V. C.; Murrmann, R. P., eds. Plant-soil interactions at low pH: 

proceedings of the second international symposium; June 1990; Beckley, WV. Dordrecht, The Netherlands: 

Kluwer Academic Publishers; pp. 139-145. 

Reddy, G. N.; Arteca, R. N.; Dai, Y. R.; Flores, H. E.; Negm, F. B.; Pell, E. J. (1993) Changes in ethylene and 

polyamines in relation to mRNA levels of the large and small subunits of ribulose bisphosphate 

carboxylase/oxygenase in ozone-stressed potato foliage. Plant Cell Environ. 16: 819-826. 

AX9-378
Reich, P. B. (1983) Effects of low concentrations of O3 on net photosynthesis, dark respiration, and chlorophyll 

contents in aging hybrid poplar leaves. Plant Physiol. 73: 291-296. 

Reich, P. B. (1987) Quantifying plant response to ozone: a unifying theory. Tree Physiol. 3: 63-91. 

Reich, P. B. (2002) Root-shoot relations: optimality in acclimation and adaptation or the "Emperor's New Clothes"? 

In: Waisel, Y.; Eshel, A.; Kafkafi, U., eds. Plant roots - the hidden half. Marcel Dekker; pp. 205-220. 

Reich, P. B.; Lassoie, J. P. (1984) Effects of low level O3 exposure on leaf diffusive conductance and water-use 

efficiency in hybrid poplar. Plant Cell Environ. 7: 661-668. 

Reich, P. B.; Schoettle, A. W.; Stroo, H. F.; Troiano, J.; Amundson, R. G. (1985) Effects of O3, SO2, and acidic rain 

on mycorrhizal infection in northern red oak seedlings. Can. J. Bot. 63: 2049-2055. 

Reichenauer, T. G.; Goodman, B. A.; Kostecki, P.; Soja, G. (1998) Ozone sensitivity in Triticum durum and 

T. aestivum with respect to leaf injury, photosynthetic activity and free radical content. Physiol. Plant. 

104: 681-686. 

Reid, C. D.; Fiscus, E. L. (1998) Effects of elevated [CO2] and/or ozone on limitations to CO2 assimilation in 

soybean (Glycine max). J. Exp. Bot. 49: 885-895. 

Reid, C. D.; Fiscus, E. L.; Burkey, K. O. (1998) Combined effects of chronic ozone and elevated CO2 on rubisco 

activity and leaf components in soybean (Glycine max). J. Exp. Bot. 49: 1999-2011. 

Reiling, K.; Davison, A. W. (1992a) Effects of a short ozone exposure given at different stages in the development 

of Plantago major L. New Phytol. 121: 643-647. 

Reiling, K.; Davison, A. W. (1992b) The response of native, herbaceous species to ozone: growth and fluorescence 

screening. New Phytol. 120: 29-37. 

Reiling, K.; Davison, A. W. (1992c) Spatial variation in ozone resistance of British populations of Plantago major 

L. New Phytol. 122: 699-708.

Reiling, K.; Davison, A. W. (1994) Effects of exposure to ozone at different stages in the development of Plantago 

major L on chlorophyll fluorescence and gas exchange. New Phytol. 128: 509-514. 

Reiner, S.; Wiltshire, J. J. J.; Wright, C. J.; Colls, J. J. (1996) The impact of ozone and drought on the water relations 

of ash trees (Fraxinus excelsior L). J. Plant Physiol. 148: 166-171. 

Reinert, R. A.; Ho, M. C. (1995) Vegetative growth of soybean as affected by elevated carbon dioxide and ozone. 

Environ. Pollut. 89: 89-96. 

Reinert, R. A.; Eason, G.; Barton, J. (1997) Growth and fruiting of tomato as influenced by elevated carbon dioxide 

and ozone. New Phytol. 137: 411-420. 

Renaud, J. P.; Allard, G.; Mauffette, Y. (1997) Effects of ozone on yield, growth, and root starch concentrations of 

two alfalfa (Medicago sativa L) cultivars. Environ. Pollut. 95: 273-281. 

Rennenberg, H.; Herschbach, C.; Polle, A. (1996) Consequences of air pollution on shoot-root interactions. J. Plant 

Physiol. 148: 296-301. 

Retzlaff, W. A.; Weinstein, D. A.; Laurence, J. A.; Gollands, B. (1996) Simulated root dynamics of a 160-year-old 

sugar maple (Acer saccharum Marsh) tree with and without ozone exposure using the TREGRO model. Tree 

Physiol. 16: 915-921. 

Retzlaff, W. A.; Williams, L. E.; DeJong, T. M. (1997) Growth and yield response of commercial bearing-age 

'Casselman' plum trees to various ozone partial pressures. J. Environ. Qual. 26: 858-865. 

Retzlaff, W. A.; Arthur, M. A.; Grulke, N. E.; Weinstein, D. A.; Gollands, B. (2000) Use of a single-tree simulation 

model to predict effects of ozone and drought on growth of a white fir tree. Tree Physiol. 20: 195-202. 

Ribas, A.; Peñuelas, J. (2000) Effects of ethylene diurea as a protective antiozonant on beans (Phaseolus vulgaris cv 

lit) exposed to different tropospheric ozone doses in Catalonia (NE Spain). Water Air Soil Pollut. 117: 

263-271. 

Rich, S.; Ames, R.; Zuckel, J. W. (1974) 1,4-oxanthin derivatives protect plants against ozone. Plant Dis. Rep. 

58: 162-164.

Ritchie, G. A.; Dunlap, J. R. (1980) Root growth potential: its development and expression in forest tree seedlings. 

N. Z. J. For. Sci. 10: 218-248.

Robinson, J. M.; Britz, S. J. (2000) Tolerance of a field grown soybean cultivar to elevated ozone level is concurrent 

with higher leaflet ascorbic acid level, higher ascorbate-dehydroascorbate redox status, and long term 

photosynthetic productivity. Photosynth. Res. 64: 77-87. 

Robinson, M. F.; Heath, J.; Mansfield, T. A. (1998) Disturbance in stomatal behaviour caused by air pollutants. 

J. Exp. Bot. 49: 461-469.

AX9-379 
Rogers, A.; Allen, D. J.; Davey, P. A.; Morgan, P. B.; Ainsworth, E. A.; Bernacchi, C. J.; Cornic, G.; Dermody, 

O. C.; Heaton, E. A.; Mahoney, J.; Zhu, X. G.; DeLucia, E. H.; Ort, D. R.; Long, S. P. (2004) Leaf 

photosynthesis and carbohydrate dynamics of soybean grown throughout their life-cycle under free-air carbon 

dioxide enrichment. Plant Cell Environ. 28: 449-458. 

Roitsch, T. (1999) Source-sink regulation by sugar and stress. Curr. Opin. Plant Biol. 2: 198-206.

Rondön, A.; Johansson, C.; Granat, L. (1993) Dry deposition of nitrogen-dioxide and ozone to coniferous forests.

J. Geophys. Res. [Atmos. ] 98(D3): 5159-5172. 

Rosen, P. M.; Runeckles, V. C. (1976) Interaction of ozone and greenhouse whitefly in plant injury. Environ. 

Conserv. 3: 70-71. 

Roth, D. R.; Fahey, T. J. (1998) The effects of acid precipitation and ozone on the ectomycorrhizae of red spruce 

saplings. Water Air Soil Pollut. 103: 263-276. 

Rötter, R.; Van De Geijn, S. C. (1999) Climate change effects on plant growth, crop yield and livestock. Clim. 

Change 43: 651-681. 

Rowland-Bamford, A. J. (2000) Plant responses to changing carbon dioxide and temperature. In: Singh, S. N., ed. 

Trace gas emissions and plants. Dordrecht, The Netherlands: Kluwer Academic Publishers; pp. 63-74. 

Rudorff, B. F. T.; Mulchi, C. L.; Daughtry, C. S. T.; Lee, E. H. (1996a) Growth, radiation use efficiency, and canopy 

reflectance of wheat and corn grown under elevated ozone and carbon dioxide atmospheres. Remote Sens. 

Environ. 55: 163-173. 

Rudorff, B. F. T.; Mulchi, C. L.; Fenny, P.; Lee, E. H.; Rowland, R. (1996b) Wheat grain quality under enhanced 

tropospheric CO2 and O3 concentrations. J. Environ. Qual. 25: 1384-1388. 

Rudorff, B. F. T.; Mulchi, C. L.; Lee, E. H.; Rowland, R.; Pausch, R. (1996c) Effects of enhanced O3 and CO2 

enrichment on plant characteristics in wheat and corn. Environ. Pollut. 94: 53-60. 

Rudorff, B. F. T.; Mulchi, C. L.; Lee, E. H. (2000) Plant responses to elevated CO2 and interactions with O3. In: 

Singh, S. N., ed. Trace gas emissions and plants. Dordrecht, The Netherlands: Kluwer Academic Publishers; 

pp. 155-179. 

Runeckles, V. C. (1974) Dosage of air pollutants and damage to vegetation. Environ. Conserv. 1: 305-308. 

Runeckles, V. C. (2002) Air pollution and climate change. In: Bell, J. N. B.; Treshow, M., eds. Air pollution and 

plant life. Chichester, United Kingdom: John Wiley & Sons Ltd.; pp. 431-454. 

Runeckles, V. C.; Bowen, P. A. (2000) The use of calibrated passive monitors to assess crop loss due to ozone in 

rural locations. In: Agrawal, S. B.; Agrawal, M., eds. Environmental pollution and plant response. Boca 

Raton, Florida: Lewis Publishers; pp. 307-316. 

Runeckles, V. C.; Krupa, S. V. (1994) The impact of UV-B radiation and ozone on terrestrial vegetation. Environ. 

Pollut. 83: 191-213. 

Runeckles, V. C.; Palmer, K. (1987) Pretreatment with nitrogen dioxide modifies plant response to ozone. Atmos. 

Environ. 21: 717-719. 

Runeckles, V. C.; Vaartnou, M. (1997) EPR evidence for superoxide anion formation in leaves during exposure to 

low levels of ozone. Plant Cell Environ. 20: 306-314. 

Runeckles, V. C.; Wright, E. F. (1996) Delayed impact of chronic ozone stress on young Douglas-fir grown under 

field conditions. Can. J. For. Res. 26: 629-638. 

Runeckles, V. C.; Wright, E. F.; White, D. (1990) A chamberless field exposure system for determining the effects 

of gaseous air pollutants on crop growth and yield. Environ. Pollut. 63: 61-77. 

Rusch, H.; Laurence, J. A. (1993) Interactive effects of ozone and powdery mildew on pea seedlings. 

Phytopathology 83: 1258-1263. 

Sah, D. N.; Von Tiedemann, A.; Fehrmann, H. (1993) Effects of ozone exposure on resistance of wheat genotypes to 

Phrenophora tritici-repentis. J. Phytopathol. 138: 249-256. 

Saitanis, C. J.; Karandinos, M. G. (2001) Instrumental recording and biomonitoring of ambient ozone in the Greek 

countryside. Chemosphere 44: 813-821. 

Sakaki, T.; Kondo, N.; Sugahara, K. (1983) Breakdown of photosynthetic pigments and lipids in spinach leaves with 

ozone fumigation: role of active oxygens. Physiol. Plant. 59: 28-34. 

Sakaki, T.; Ohnishi, J.-I.; Kondo, N.; Yamada, M. (1985) Polar and neutral lipid changes in spinach leaves with 

ozone fumigation: triacylglyerol synthesis from polar lipids. Plant Cell Physiol. 26: 253-262. 

Sakaki, T.; Saito, K.; Kawaguchi, A.; Kondo, N.; Yamada, M. (1990) Conversion of monogalactosylidacyglyerol to 

triacylglycerols in ozone-fumigated spinach leaves. Plant Physiol. 94: 766-772. 

Salardino, D. H.; Carroll, J. J. (1998) Correlation between ozone exposure and visible foliar injury in ponderosa and 

Jeffrey pines. Atmos. Environ. 32: 3001-3010. 

AX9-380
Saleem, A.; Loponen, J.; Pihlaja, K.; Oksanen, E. (2001) Effects of long-term open-field ozone exposure on leaf 

phenolics of European silver birch (Betula pendula Roth). J. Chem. Ecol. 27: 1049-1062. 

Samuelson, P. A. (1948) Economics: an introductory analysis. New York, NY: McGraw-Hill. 

Samuelson, L. J. (1994) Ozone-exposure responses of black cherry and red maple seedlings. Environ. Exp. Bot. 

34: 355-362.

Samuelson, L. J.; Edwards, G. S. (1993) A comparison of sensitivity to ozone in seedlings and trees of Quercus 

rubra L. New Phytol. 125: 373-379. 

Samuelson, L. J.; Kelly, J. M. (1996) Carbon partitioning and allocation in northern red oak seedlings and mature 

trees in response to ozone. Tree Physiol. 16: 853-858. 

Samuelson, L. J.; Kelly, J. M. (1997) Ozone uptake in Prunus serotina, Acer rubrum and Quercus rubra forest trees 

of different sizes. New Phytol. 136: 255-264. 

Samuelson, L.; Kelly, J. M. (2001) Scaling ozone effects from seedlings to forest trees Tansley review no 21. 

New Phytol. 149: 21-41. 

Samuelson, L. J.; Kelly, J. M.; Mays, P. A.; Edwards, G. S. (1996) Growth and nutrition of Quercus rubra L 

seedlings and mature trees after three seasons of ozone exposure. Environ. Pollut. 91: 317-323. 

Sandermann, H., Jr. (1996) Ozone and plant health. Annu. Rev. Phytopathol. 34: 347-366. 

Sandermann, H., Jr. (1998) Ozone: an air pollutant acting as a plant-signaling molecule. Naturwissenschaften 

85: 369-375.

Sandermann, H., Jr. (2000) Ozone/biotic disease interactions: molecular biomarkers as a new experimental tool. 

Environ. Pollut. 108: 327-332. 

Sasek, T. W.; Richardson, C. J.; Fendick, E. A.; Bevington, S. R.; Kress, L. W. (1991) Carryover effects of acid rain 

and ozone on the physiology of multiple flushes of loblolly pine seedlings. For. Sci. 37: 1078-1098. 

Saxe, H.; Ellsworth, D. S.; Heath, J. (1998) Tree and forest functioning in an enriched CO2 atmosphere. New Phytol. 

139: 395-436. 

Sayre, R. G.; Fahey, T. J. (1999) Effects of rainfall acidity and ozone on foliar leaching in red spruce Picea rubens. 

Can. J. For. Res. 29: 486-496. 

Scagel, C. F.; Andersen, C. P. (1997) Seasonal changes in root and soil respiration of ozone-exposed ponderosa pine 

(Pinus ponderosa) grown in different substrates. New Phytol. 136: 627-643. 

Scandalios, J. G. (1993) Oxygen stress and superoxide dismutases. Plant Physiol. 101: 7-12. 

Schaub, M.; Skelly, J. M.; Steiner, K. C.; Davis, D. D.; Pennypacker, S. P.; Zhang, J.; Ferdinand, J. A.; Savage, 

J. E.; Stevenson, R. E. (2003) Physiological and foliar injury responses of Prunus serotina, Fraxinus 

americana, and Acer rubrum seedlings to varying soil moisture and ozone. Environ. Pollut. 124: 307-320. 

Scherzer, A. J.; Rebbeck, J.; Boerner, R. E. J. (1998) Folia nitrogen dynamics and decomposition of yellow-poplar 

and eastern white pine during four seasons of exposure to elevated ozone and carbon dioxide. For. Ecol. 

Manage. 109: 355-366. 

Schmieden, U.; Wild, A. (1995) The contribution of ozone to forest decline. Physiol. Plant. 94: 371-378. 

Schnitzler, J. P.; Langebartels, C.; Heller, W.; Liu, J.; Lippert, M.; Dohring, T.; Bahnweg, G.; Sandermann, H. 

(1999) Ameliorating effect of UV-B radiation on the response of Norway spruce and Scots pine to ambient 

ozone concentrations. Global Change Biol. 5: 83-94. 

Schnug, E. (1998) Response of plant metabolism to air pollution and global change - impact on agriculture. In: 

De Kok, L. J. ; Stulen, I. , eds. Responses of plant metabolism to air pollution and global change. Leiden, The 

Netherlands: Backhuys Publishers; pp. 15-22.: De-22. 

Schopfer, P. (1994) Histochemical demonstration and localization of H2O2 in organs of higher plants by tissue 

printing on nitrocellulose paper. Plant Physiol. 104: 1269-1275. 

Schraudner, M.; Ernst, D.; Langebartels, C.; Sandermann, H., Jr. (1992) Biochemical plant responses to ozone III 

Activation of the defense-related proteins $-1,3-glucanase and chitinase in tobacco leaves. Plant Physiol. 99: 

1321-1328. 

Schraudner, M.; Langebartels, C.; Sandermann, H., Jr. (1996) Plant defence systems and ozone. Biochem. Soc. 

Trans. 24: 456-461. 

Schraudner, M.; Langebartels, C.; Sandermann, H. (1997) Changes in the biochemical status of plant cells linduced 

by the environmental pollutant ozone. Physiol. Plant. 100: 274-280. 

Schraudner, M.; Moeder, W.; Wiese, C.; Van Camp, W.; Inzé, D.; Langebartels, C.; Sandermann, H., Jr. (1998) 

Ozone-induced oxidative burst in the ozone biomonitor plant, tobacco Bel W3. Plant J. 16: 235-245. 

Schulte-Hosted, S.; Darrall, N. M.; Blank, L. W.; Wellburn, A. R. (1988) Air pollution and plant metabolism. 

New York, NY: Elsivier Applied Science. 

AX9-381
Schweizer, B.; Arndt, U. (1990) CO2/H2O gas exchange parameters of one- and two-year old needles of spruce and 

fir. Environ. Pollut. 68: 275-292. 

Segerson, K. (1987) Impacts of ozone and acid rain: discussion. Am. J. Agric. Econ. 69: 970-971. 

Segerson, K. (1991) Air pollution and agriculture: a review and evaluation of policy interactions. In: Just, R. E.; 

Bockstael, N., eds. Commodity and resource policies in agricultural systems. Springer-Verlag; pp. 349-367. 

Seufert, G.; Arndt, U. (1985) Open-top-Kammern als Teil eines Konzepts zur ökosystemaren Untersuchung der 

neuartigen Waldschaden. Alg. Forstz. 40: 13-18. 

Shafer, S. R. (1988) Influence of ozone and simulated acidic rain on microorganisms in the rhizosphere of Sorghum. 

Environ. Pollut. 51: 131-152. 

Shan, Y.; Feng, Z.; Izuta, T.; Aoki, M.; Totsuka, T. (1996) The individual and combined effects of ozone and 

simulated acid rain on growth, gas exchange and water-use efficiency of Pinus armandi Franch. Environ. 

Pollut. 91: 355-361. 

Sharma, Y. K.; Davis, K. R. (1995) Isolation of a novel Arabidopsis ozone-induced cDNA by differential display. 

Plant Mol. Biol. 29: 91-98. 

Sharma, P.; Sôber, A.; Sôber, J.; Podila, G. K.; Kubiske, M. E.; Mattson, W. J.; Isebrands, J. G.; Karnosky, D. F. 

(2003) Moderation of [CO2]-induced gas exchange responses by elevated tropospheric O3 in trembling aspen 

and sugar maple. Ekologia (Bratislava): in press. 

Shaw, M. R.; Zavaleta, E. S.; Chiariello, N. R.; Cleland, E. E.; Mooney, H. A.; Field, C. B. (2002) Grassland 

responses to global environmental changes suppressed by elevated CO2. Science (Washington, DC, U. S.) 

298: 1987-1990. 

Sheng, Y.; Podila, G. K.; Karnosky, D. F. (1997) Differences in O3-induced superoxide dismutase and glutathione 

antioxidant expression in O3 tolerant and sensitive trembling aspen (Populus tremuloides Michx) clones. For. 

Gen. 4: 25-33. 

Sheppard, L. J.; Cape, J. N. (1999) Forest growth responses to the pollution climate of the 21st century: proceedings 

of the 18th IUFRO workshop on air pollution stress; 1998; Edinburgh, Scotland. Water Air Soil Pollut. 

116(1-2 special issue). 

Sheppard, L. J.; Smith, R. I.; Cannell, M. G. R. (1989) Frost hardiness of Picea rubens growing in spruce decline 

regions of the Appalachians. Tree Physiol. 5: 25-37. 

Shinshi, H.; Mohnen, D.; Meins, F., Jr. (1987) Regulation of a plant pathogenesis-related enzyme: inhibition of 

chitinase and chitinase mRNA accumulation in cultured tobacco tissues by auxin and cytokinin. Proc. Natl. 

Acad. Sci. U. S. A. 84: 89-93. 

Shinshi, H.; Wenzler, H.; Neuhaus, J. M.; Felix, G.; Hofsteenge, J.; Meins, F., Jr. (1988) Evidence for N- and 

C-terminal processing of a plant defense-related enzyme: primary structure of tobacco 

prepro-$-1,3-glucanase. Proc. Natl. Acad. Sci. U. S. A. 85: 5541-5545. 

Sild, E.; Pleijel, H.; Selldén, G. (1999) Response of subterranean clover (Trifolium subterraneum) to ozone in 

relation to plant age and light conditions during exposure. New Phytol. 143: 315-321. 

Simini, M.; Skelly, J. M.; Davis, D. D.; Savage, J. E.; Comrie, A. C. (1992) Sensitivity of four hardwood species to 

ambient ozone in north central Pennsylvania. Can. J. For. Res. 22: 1789-1799. 

Simmons, G. L.; Kelly, J. M. (1989) Influence of O3, rainfall acidity, and soil Mg status on growth and 

ectomycorrhizal colonization of loblolly pine roots. Water Air Soil Pollut. 44: 159-171. 

Simon-Plas, F.; Rustérucci, C.; Milat, M. L.; Humbert, C.; Montillet, J. L.; Blein, J. P. (1997) Active oxygen species 

production in tobacco cells elicited by cryptogein. Plant Cell Environ. 20: 1573-1579. 

Simpson, D.; Tuovinen, J.-P.; Emberson, L.; Ashmore, M. R. (2001) Characteristics of an ozone deposition module. 

Water Air Soil Pollut. Focus 1: 253-262. 

Simpson, D.; Tuovinen, J.-P.; Emberson, L.; Ashmore, M. R. (2003) Characteristics of an ozone deposition module 

II: Sensitivity analysis. Water Air Soil Pollut. 143: 123-137. 

Singh, S. N. ed. (2000) Trace gas emissions and plants. Dordrecht, The Netherlands: Kluwer Academic Publishers, 

368 pp. 

Sivasankar, S.; Sheldrick, B.; Rothstein, S. J. (2000) Expression of allene oxide synthase determines defense gene 

activation in tomato. Plant Physiol. 122: 1335-1342. 

Skärby, L.; Ro-Poulsen, H.; Wellburn, F. A. M.; Sheppard, L. J. (1998) Impacts of ozone on forests: a European 

perspective. New Phytol. 139: 109-122. 

Skärby, L.; Ottosson, S.; Karlsson, P. E.; Wallina, G.; Selldén, G.; Medina, E. L.; Pleijel, H. (2004) Growth of 

Norway spruce (Picea abies) in relation to different ozone exposure indices: a synthesis. Atmos. Environ. 38: 

2225-2236. 

AX9-382
Skeffington, R. A.; Sutherland, P. M. (1995) The effects of SO2 and O3 fumigation on acid deposition and foliar 

leaching in the Liphook forest fumigation experiment. Plant Cell Environ. 18: 247-261. 

Skelly, J. M. (1988) Forest decline versus tree decline the pathological considerations. Environ. Monit. Assess. 

12: 23-27.

Skelly, J. M. (2003) Native plants as bioindicators of air pollutants: contributed papers to a symposium held in 

conjunction with the 34th air pollution workshop. Environ. Pollut. 125: 1-2. 

Smeulders, S. M.; Gorissen, A.; Joosten, N. N.; Van Veen, J. A. (1995) Effects of short-term ozone exposure on the 

carbon economy of mature and juvenile Douglas firs [Pseudotsuga menziesii (Mirb.) Franco]. New Phytol. 

129: 45-53. 

Smirnoff, N.; Conklin, P. L.; Loewus, F. A. (2001) Biosynthesis of ascorbate acid in plants: a renaissance. Annu. 

Rev. Plant Physiol. Mol. Biol. 52: 437-467. 

Smith, W. H. (1992) Air pollution effects on ecosystem processes. In: Barker, J. R.; Tingey, D. T., eds. Air pollution 

effects on biodiversity. New York, NY: Van Nostrand Reinhold; pp. 234-260. 

Smith, W. B. (2002) Forest inventory and analysis: a national inventory and monitoring program. Environ. Pollut. 

116: S233-S242. 

Smith, S. E.; Read, D. J. (1997) Mycorrhizal symbiosis. San Diego, CA: Academic Press. 

Smith, G.; Coulston, J.; Jepsen, E.; Prichard, T. (2003) A national ozone biomonitoring program - results from field 

surveys of ozone sensitive plants in northeastern forests (1994-2000). Environ. Monit. Assess. 87: 271-291. 

Soja, G.; Eid, M.; Gangl, H.; RedI, H. (1997) Ozone sensitivity of grapevine (Vitis vinifera L): evidence for a 

memory effect in a perennial crop plant? Phyton (Austria) 37: 265-270. 

Soja, G.; Barnes, J. D.; Posch, M.; Vandermeiren, K.; Pleijel, H.; Mills, G. (2000) Phenological weighting of ozone 

exposures in the calculation of critical levels for wheat, bean and plantain. Environ. Pollut. 109: 517-524. 

Soja, G.; Reichenauer, T. G.; Eid, M.; Soja, A. M.; Schaber, R.; Gangl, H. (2004) Long-term ozone exposure and 

ozone uptake of grapevines in open-top chambers. Atmos. Environ. 38: 2313-2321. 

Somers, G. L.; Chappelka, A. H.; Rosseau, P.; Renfro, J. R. (1998) Empirical evidence of growth decline related to 

visible ozone injury. For. Ecol. Manage. 104: 129-137. 

Spash, C. L. (1997) Assessing the economic benefits to agriculture from air pollution control. J. Econ. Surv. 11: 

47-70. 

Spence, R. D.; Rykiel, E. J., Jr.; Sharpe, P. J. H. (1990) Ozone alters carbon allocation in loblolly pine: assessment 

with carbon-11 labeling. Environ. Pollut. 64: 93-106. 

Stan, H. J.; Schicker, S. (1982) Effect of repetitive ozone treatment on bean plants - stress ethylene production and 

leaf necrosis. Atmos. Environ. 16: 2267-2270. 

Stan, H. J.; Schicker, S.; Kassner, H. (1981) Stress ethylene evolution of bean plants - a parameter indicating ozone 

pollution. Atmos. Environ. 15: 391-395. 

Stark, R. W.; Cobb, F. W., Jr. (1969) Smog injury, root diseases and bark beetle damage in ponderosa pine. Calif. 

Agric. 23: 13-15. 

Stark, R. W.; Miller, P. R.; Cobb, F. W., Jr.; Wood, D. L.; Parmeter, J. R., Jr. (1968) Photochemical oxidant injury 

and bark beetle (Coleoptera: Scolytidae) infestation of ponderosa pine I. Incidence of bark beetle infestation 

in injured trees. Hilgardia 39: 121-126. 

Staszak, J.; Grulke, N. E.; Prus-Glowacki, W. (2004) Genetic differences of Pinus ponderosa (Dougl Ex Laws) trees 

tolerant and sensitive to ozone. Water Air Soil Pollut. 153: 3-14. 

Stewart, C. A. (1998) Impact of ozone on the reproductive biology of Brassica campestris L. and Plantago major L 

[dissertation]. Leicestershire, United Kingdom: University of Loughborough, Department of Geology. 

Stewart, C. A.; Black, V. J.; Black, C. R.; Roberts, J. A. (1996) Direct effects of ozone on the reproductive 

development of Brassica species. J. Plant Physiol. 148: 172-178. 

Stitt, M. (1996) Plasmodesmata play an essential role in sucrose export from leaves: a step towards an integration of 

metabolic biochemistry and cell biology. Plant Cell 8: 565-571. 

Stolzy, L. H.; Taylor, O. C.; Dugger, W. M., Jr.; Mersereau, J. D. (1964) Physiological changes in and ozone 

susceptibility of the tomato plant after short periods of inadequate oxygen diffusion to the roots. Soil Sci. Soc. 

Am. Proc. 28: 305-308. 

Stow, T. K.; Allen, H. L.; Kress, L. W. (1992) Ozone impacts on seasonal foliage dynamics of young loblolly pine. 

For. Sci. 38: 102-119. 

Stribley, G. H.; Ashmore, M. R. (2002) Quantitative changes in twig growth pattern of young woodland beech 

(Fagus sylvatica L.) in relation to climate and ozone pollution over 10 years. For. Ecol. Manage. 

157: 191-204. 

AX9-383
Strohm, M.; Eiblmeier, M.; Langebartels, C.; Jouanin, L.; Polle, A.; Sandermann, H.; Rennenberg, H. (1999) 

Responses of transgenic poplar (Populus tremula x P. alba) overexpressing glutathione synthetase of 

glutathione reductase to acute ozone stress: visible injury and leaf gas exchange. J. Exp. Bot. 50: 363-372. 

Strohm, M.; Eiblmeier, M.; Langebartels, C.; Jouanin, L.; Polle, A.; Sandermann, H.; Rennenberg, H. (2002) 

Responses of antioxidative systems to acute ozone stress in transgenic poplar (Populus tremula x P. alba) 

over-expressing glutathione synthetase or glutathione reductase. Trees (Heidelberg, Ger.) 16: 262-273. 

Stroo, H. F.; Reich, P. B.; Schoettle, A. W.; Amundson, R. G. (1988) Effects of ozone and acid rain on white pine 

(Pinus strobus) seedlings grown in five soils II Mycorrhizal infection. Can. J. Bot. 66: 1510-1516. 

Suter, G. W., II (1990) Endpoints for regional ecological risk assessments. Environ. Manage. (N. Y. ) 14: 9-23. 

Swank, W. T.; Vose, J. M. (1990) Watershed-scale responses to ozone events in a Pinus strobus L plantation. Water 

Air Soil Pollut. 54: 119-133. 

Szaro, R.; Bytnerowicz, A.; Oszlányi, J. (2002) Effects of air pollution on forest health and biodiversity in forests of 

the Carpathian Mountains. Amsterdam, The Netherlands: IOS Press. 

Taiz, L.; Zeiger, E. (2002) Plant physiology, 3rd ed. Sunderland, MA: Sinauer Associates, Publishers. 

Takemoto, B. K.; Olszyk, D. M.; Johnson, A. G.; Parada, C. R. (1988) Yield responses of field-gown crops to acidic 

fog and ambient ozone. J. Environ. Qual. 17: 192-197. 

Takemoto, B. K.; Bytnerowicz, A.; Dawson, P. J.; Morrison, C. L.; Temple, P. J. (1997) Effects of ozone on 

Pinus ponderosa seedlings: comparison of responses in the first and second growing seasons of exposure. 

Can. J. For. Res. 27: 23-30. 

Takemoto, B. K.; Bytnerowicz, A.; Fenn, M. E. (2001) Current and future effects of ozone and atmospheric nitrogen 

deposition on California's mixed conifer forests. For. Ecol. Manage. 144: 159-173. 

Takeuchi, Y.; Kubiske, M. E.; Isebrands, J. G.; Pregitzer, K. S.; Hendrey, G.; Karnosky, D. F. (2001) 

Photosynthesis, light and nitrogen relationships in a young deciduous forest canopy under open-air CO2 

enrichment. Plant Cell Environ. 24: 1257-1268. 

Tamaoki, M.; Nakajima, N.; Kubo, A.; Aono, M.; matsuyama, T.; Saji, H. (2003) Transcriptome analysis of 

O3-exposed Arabidopsis reveals that multiple signal pathways act mutually antagonistically to induce gene 

expression. Plant Mol. Biol. 53: 443-456. 

Tappel, A. L. (1972) Vitamin E and free radical peroxidation of lipids. In: Nair, P. P.; Kayden, H. J., eds. Vitamin E 

and its role in cellular metabolism. New York, NY: New York Academy of Sciences; pp. 12-28. (Annals of 

the New York Academy of Sciences: v. 203). 

Tausz, M.; Bytnerowicz, A.; Arbaugh, M. J.; Weidner, W.; Grill, D. (1999a) Antioxidants and protective pigments 

of Pinus ponderosa needles at gradients of natural stresses and ozone in the San Bernardino Mountains in 

California. Free Rad. Res. 31(suppl.): S113-S120. 

Tausz, M.; Bytnerowicz, A.; Weidner, W.; Arbaugh, M. J.; Padgett, P.; Grill, D. (1999b) Changes in free-radical 

scavengers describe the susceptibility of Pinus ponderosa to ozone in southern Californian forests. Water Air 

Soil Pollut. 116: 249-254. 

Tausz, M.; Bytnerowicz, A.; Weidner, W.; Arbaugh, M. J.; Padgett, P.; Grill, D. (1999c) Pigments and 

phytoprotection in needles of Pinus ponderosa trees with and without symptoms of ozone injury. Phyton 

(Horn, Austria) 39: 219-224. 

Tausz, M.; Bytnerowicz, A.; Arbaugh, M. J.; Wonisch, A.; Grill, D. (2001) Multivariate patterns of biochemical 

responses of Pinus ponderosa trees at field plots in the San Bernardino Mountains, southern California. Tree 

Physiol. 21: 329-336. 

Tausz, M.; Herbinger, K.; Posch, S.; Grulke, N. E. (2002) Antioxidant status of Pinus jeffreyi needles from mesic 

and xeric microsites in early and late summer. Phyton (Horn, Austria) 42: 201-207. 

Taylor, G. E., Jr.; Hanson, P. J. (1992) Forest trees and tropospheric ozone: role of canopy deposition and leaf 

uptake in developing exposure-response relationships. Agric. Ecosyst. Environ. 42: 255-273. 

Taylor, G. E., Jr.; Pitelka, L. F. (1992) Genetic diversity of plant populations and the role of air pollution. In: Barker, 

J. R.; Tingey, D. T., eds. Air pollution effects on biodiversity. Van Nostrand Reinhold; pp. 111-130. 

Taylor, G. E., Jr.; McLaughlin, S. B.; Shriner, D. S. (1982) Effective pollutant dose. In: Unsworth, M. H.; Ormrod, 

D. P., eds. Effects of gaseous air pollution in agriculture and horticulture. London, United Kingdom: 

Butterworth Scientific; pp. 458-460. 

Taylor, G. E., Jr.; Tingey, D. T.; Ratsch, H. C. (1982) Ozone flux in Glycine max (L) Merr: sites of regulation and 

relationship to leaf injury. Oecologia 53: 179-186. 

Taylor, G. E., Jr.; Hanson, P. J.; Baldocchi, D. D. (1988a) Pollutant deposition to individual leaves and plant 

canopies: sites of regulation and relationship to injury. In: Heck, W. W.; Taylor, O. C.; Tingey, D. T., eds. 

Assessment of crop loss from air pollutants. New York, NY: Elsevier Applied Science; pp. 227-257. 

AX9-384
Taylor, G. E., Jr.; Ross-Todd, B. M.; Gunderson, C. A. (1988b) Action of ozone on foliar gas exchange in Glycine 

max L Merr: a potential role for endogenous stress ethylene. New Phytol. 110: 301-307. 

Taylor, G. E.; Owens, J. G.; Grizzard, T.; Selvidge, W. J. (1993) Atmosphere x canopy interactions of nitric acid 

vapor in loblolly-pine grown in open-top chambers. J. Environ. Qual. 22: 70-80. 

Taylor, G. E., Jr.; Johnson, D. W.; Andersen, C. P. (1994) Air pollution and forest ecosystems: a regional to global 

perspective. Ecol. Appl. 4: 662-689. 

Taylor, G.; Ceulemans, R.; Ferris, R.; Gardner, S. D. L.; Shao, B. Y. (2001) Increased leaf area expansion of hybrid 

poplar in elevated CO2 from controlled environments to open-top chambers and to FACE. Environ. Pollut. 

115: 463-472. 

Tegischer, K.; Tausz, M.; Wieser, G.; Grill, D. (2002) Tree-age and needle-age dependent variations of antioxidants 

and photoprotective pigments in spruce needles at the alpine timberline. Tree Physiol. 22: 591-596. 

Temple, P. J. (1988) Injury and growth of Jeffrey pine and giant sequoia in response to ozone and acidic mist. 

Environ. Exp. Bot. 28: 323-333. 

Temple, P. J. (1999) Effects of ozone on understory vegetation in the mixed conifer forest. In: Miller, P. R.; 

McBride, J. R., eds. Oxidant air pollution impacts in the montane forests of southern California: a case study 

of the San Bernardino Mountains. New York, NY: Springer; pp. 208-222. (Ecological studies: v. 134). 

Temple, P. J.; Riechers, G. H.; Miller, P. R. (1992) Foliar injury responses of ponderosa pine seedlings to ozone, wet 

and dry acidic deposition, and drought. Environ. Exp. Bot. 32: 101-113. 

Temple, P. J.; Riechers, G. H.; Miller, P. R.; Lennox, R. W. (1993) Growth responses of ponderosa pine to 

long-term exposure to ozone, wet and dry acidic deposition, and drought. Can. J. For. Res. 23: 59-66. 

Teramura, A. H.; Sullivan, J. H.; Lyden, J. (1990) Effects of UV-b radiation on soybean yield and seed quality: a 

6-year field study. Physiol. Plant. 80: 5-11. 

Ter-Mikaelian, M. T.; Korzukhin, M. D. (1997) Biomass equations for sixty-five North American tree species. For. 

Ecol. Manage. 97: 1-24. 

Teskey, R. O.; Dougherty, P. M.; Wiselogel, A. E. (1991) Design and performance of branch chambers suitable for 

long-term ozone fumigation of foliage in large trees. J. Environ. Qual. 20: 591-595. 

Thalmair, M.; Bauw, G.; Thiel, S.; Döhring, T.; Langebartels, C.; Sandermann, H., Jr. (1996) Ozone and ultraviolet 

B effects on the defense-related proteins $-1,3-glucanase and chitinase in tobacco. J. Plant Physiol. 

148: 222-228. 

Thom, M.; Maretzki, A. (1985) Evidence for a plasmalemma redox system in sugarcane. Plant Physiol. 77: 873-876. 

Tiedemann, A. V.; Firsching, K. H. (1993) Effects of ozone exposure and leaf age of wheat on infection processes of 

Septoria nodorum Berk. Plant Pathol. 42: 287-293. 

Tiedemann, A. V.; Firsching, K. H. (2000) Interactive effects of elevated ozone and carbon dioxide on growth and 

yield of leaf rust-infected versus non-infected wheat. Environ. Pollut. 108: 357-363. 

Tiedemann, A. V.; Weigel, H. J.; Jäger, H. J. (1991) Effects of open-top chamber fumigations with ozone on three 

fungal leaf diseases of wheat and the mycoflora of the phyllosphere. Environ. Pollut. 72: 205-224. 

Tingey, D. T.; Blum, U. (1973) Effects of ozone on soybean nodules. J. Environ. Qual. 2: 341-342. 

Tingey, D. T.; Taylor, G. E., Jr. (1982) Variation in plant response to ozone: a conceptual model of physiological 

events. In: Unsworth, M. H.; Ormrod, D. P., eds. Effects of gaseous air pollution in agriculture and 

horticulture. London, United Kingdom: Butterworth Scientific; pp. 113-138. 

Tingey, D. T.; Fites, R. C.; Wickliff, C. (1975) Activity changes in selected enzymes from soybean leaves following 

ozone exposure. Physiol. Plant. 33: 316-320. 

Tingey, D. T.; Standley, C.; Field, R. W. (1976a) Stress ethylene evolution: a measure of ozone effects on plants. 

Atmos. Environ. 10: 969-974. 

Tingey, D. T.; Wilhour, R. G.; Standley, C. (1976b) The effect of chronic ozone exposures on the metabolite content 

of ponderosa pine seedlings. For. Sci. 22: 234-241. 

Tingey, D. T.; Hogsett, W. E.; Lee, E. H.; Herstrom, A. A.; Azevedo, S. H. (1991) An evaluation of various 

alternative ambient ozone standards based on crop yield loss data. In: Berglund, R. L.; Lawson, D. R.; 

McKee, D. J. eds. Tropospheric ozone and the environment: papers from an international conference; March 

1990; Los Angeles, CA; Pittsburgh, PA: Air & Waste Management Association (A&WMA transactions serie 

s no. TR-19); pp. 272-288. 

Tingey, D. T.; McVeety, B. D.; Waschmann, R.; Johnson, M. G.; Phillips, D. L.; Rygiewicz, P. T.; Olszyk, D. M. 

(1996) A versatile sun-lit controlled-environment facility for studying plant and soil processes. J. Environ. 

Qual. 25: 614-625. 

AX9-385
Tingey, D. T.; Laurence, J. A.; Weber, J. A.; Greene, J.; Hogsett, W. E.; Brown, S.; Lee, E. H. (2001) Elevated CO2 

and temperature alter the response of Pinus ponderosa to ozone: A simulation analysis. Ecol. Appl. 11: 1412-1424. 

Tingey, D. T.; Rodecap, K. D.; Lee, E. H.; Hogsett, W. E.; Gregg, J. W. (2002) Pod development increases the 

ozone sensitivity of Phaseolus vulgaris. Water Air Soil Pollut. 139: 325-341. 

Tingey, D. T.; Hogsett, W. E.; Lee, E. H.; Laurence, J. A. (2004) Stricter ozone ambient air quality standard has 

beneficial effect on Ponderosa pine in California. Environ. Manage. 34: 397-405. 

Tissue, D. T.; Griffin, K. L.; Ball, T. (1999) Photosynthetic adjustment in field-grown ponderosa pine trees after six 

years of exposure to elevated CO2. Tree Physiol. 19: 221-228. 

Tjoelker, M. G.; Luxmoore, R. J. (1991) Soil nitrogen and chronic ozone stress influence physiology, growth and 

nutrient status of Pinus taeda L. and Liriodendron tulipifera L. seedlings. New Phytol. 119: 69-81. 

Tjoelker, M. G.; Volin, J. C.; Oleksyn, J.; Reich, P. B. (1993) Light environment alters response to ozone stress in 

seedlings of Acer saccharum Marsh and hybrid Populus L. I. In situ net photosynthesis, dark respiration and 

growth. New Phytol. 124: 627-636. 

Tjoelker, M. G.; Volin, J. C.; Oleksyn, J.; Reich, P. B. (1994) An open-air system for exposing forest-canopy 

branches to ozone pollution. Plant Cell Environ. 17: 211-218. 

Tjoelker, M. G.; Volin, J. C.; Oleksyn, J.; Reich, P. B. (1995) Interaction of ozone pollution and light effects on 

photosynthesis in a forest canopy experiment. Plant Cell Environ. 18: 895-905. 

Tognetti, R.; Cherubini, P.; Innes, J. L. (2000) Comparative stem-growth rates of Mediterranean trees under 

background and naturally enhanced ambient CO2 levels. New Phytol. 146: 59-74. 

Tokarska-Schlattner, M.; Fink, A.; Castillo, F. J.; Crespi, P.; Crevecoeur, M.; Greppin, H.; Tacchini, P. (1997) 

Effects of ozone on the plasma membrane proteins in Arabidopsis thaliana (L.) leaves. Plant Cell Environ. 

20: 1205-1211.

Tomlinson, H.; Rich, S. (1968) The ozone resistance of leaves as related to their sulfhydryl and adenosine 

triphosphate content. Phytopathology 58: 808-810. 

Tomlinson, H.; Rich, S. (1971) Effect of ozone on sterols and sterol derivatives in bean leaves. Phytopathology 

61: 1404-1405.

Tonneijck, A. E. G. (1994) Effects of various ozone exposures on the susceptibility of bean leaves (Phaseolus 

vulgaris L) to Botrytis cinerea. Environ. Pollut. 85: 59-65. 

Tonneijck, A. E. G.; Leone, G. (1993) Changes in susceptibility of bean leaves (Phaseolus vulgaris) to Sclerotinia 

sclerotiorum and Botrytis cinerea by pre-inoculative ozone exposures. Neth. J. Plant Pathol. 99: 313-322. 

Tonneijck, A. E. G.; Van Dijk, C. J. (1994) Combined effects of ozone and ammonia on growth and yield of bean 

(Phaseolus vulgaris L cv Pros). In: Fuhrer, J.; Achermann, B. eds. Critical levels for ozone: a UN-ECE 

workshop report; November 1993; Bern, Switzerland. Liebefeld-Bern, Switzerland: Swiss Federal Research 

Station for Agricultural Chemistry and Environmental Hygiene; pp. 282-286 (Schriftenreihe der FAC 

Liebefeld no. 16). 

Tonneijck, A. E. G.; Van Dijk, C. J. (1997) Effects of ambient ozone on injury and yield of Phaseolus vulgaris at 

four rural sites in The Netherlands as assessed by using ethylenediurea (EDU). New Phytol. 135: 93-100. 

Tonneijck, A. E. G.; Van Dijk, C. J. (1998) Responses of bean (Phaseolus vulgaris L cv Pros) to chronic ozone 

exposure at two levels of atmospheric ammonia. Environ. Pollut. 99: 45-51. 

Tonneijck, A. E. G.; Van Dijk, C. J. (2002a) Assessing effects of ambient ozone on injury and yield of bean with 

ethylenediurea (EDU): three years of plant monitoring at four sites in The Netherlands. Environ. Monit. 

Assess. 77: 1-10. 

Tonneijck, A. E. G.; Van Dijk, C. J. (2002b) Injury and growth response of subterranean clover to ambient ozone as 

assessed by using ethylenediurea (EDU): Three years of plant monitoring at four sites in The Netherlands. 

Environ. Exp. Bot. 48: 33-41. 

Topa, M. A.; Vanderklein, D. W.; Corbin, A. (2001) Effects of elevated ozone and low light on diurnal and seasonal 

carbon gain in sugar maple. Plant Cell Environ. 24: 663-677. 

Torsethaugen, G.; Pitcher, L. H.; Zilinskas, B. A.; Pell, E. J. (1997) Overproduction of ascorbate peroxidase in the 

tobacco chloroplast does not provide protection against ozone. Plant Physiol. 114: 529-537. 

Treshow, M.; Stewart, D. (1973) Ozone sensitivity of plants in natural communities. Biol. Conserv. 5: 209-214. 

Trevathan, L. E.; Moore, L. D.; Orcutt, D. M. (1979) Symptom expression and free sterol and fatty acid composition 

of flue-cured tobacco plants exposed to ozone. Phytopathology 69: 582-585. 

Tuovinen, J.-P.; Ashmore, M. R.; Emberson, L. D.; Simpson, D. (2004) Testing and improving the EMEP ozone 

deposition module. Atmos. Environ. 38: 2372-2385. 

Turcsányi, E.; Lyons, T.; Plöchl, M.; Barnes, J. (2000) Does ascorbate in the mesophyll cell walls form the first line 

of defence against ozone? Testing the concept using broad bean (Vicia faba L.). J. Exp. Bot. 51: 901-910. 

AX9-386
U.S. Congress (2002) Food security and rural investment act of 2002, PL 107-171, May 13, 2002. Washington, DC: 

U.S. Government Printing Office. 

U.S. Environmental Protection Agency. (1978) Air quality criteria for ozone and other photochemical oxidants. 

Research Triangle Park, NC: Office of Health and Environmental Assessment, Environmental Criteria and 

Assessment Office; report no. EPA-600/8-78-004. Available from: NTIS, Springfield, VA; PB80-124753. 

U.S. Environmental Protection Agency. (1986) Air quality criteria for ozone and other photochemical oxidants. 

Research Triangle Park, NC: Office of Health and Environmental Assessment, Environmental Criteria and 

Assessment Office; report nos. EPA-600/8-84-020aF-eF. 5v. Available from: NTIS, Springfield, VA; 

PB87-142949. 

U.S. Environmental Protection Agency. (1992) Summary of selected new information on effects of ozone on health 

and vegetation: supplement to 1986 air quality criteria for ozone and other photochemical oxidants. Research 

Triangle Park, NC: Office of Health and Environmental Assessment, Environmental Criteria and Assessment 

Office; report no. EPA/600/8-88/105F. Available from: NTIS, Springfield, VA; PB92-235670. 

U.S. Environmental Protection Agency. (1996a) Review of national ambient air quality standards for ozone: 

assessment of scientific and technical information OAQPS staff paper. Research Triangle Park, NC: Office of 

Air Quality Planning and Standards; report no. EPA/452/R-96/007. Available from: NTIS, Springfield, VA; 

PB96-203435/XAB. Available: http://www.epa.gov/ttn/naaqs/standards/ozone/s_O3_pr_sp.html (9 September 

2003). 

U.S. Environmental Protection Agency. (1996b) Air quality criteria for ozone and related photochemical oxidants. 

Research Triangle Park, NC: Office of Research and Development; report nos. EPA/600/AP-93/004aF-cF. 3v. 

Available from: NTIS, Springfield, VA; PB96-185582, PB96-185590, and PB96-185608. Available: 

www.epa.gov/ncea/ozone.htm. 

U.S. National Park Service (2003) Ozone sensitive plant species on National Park Service and U.S. Fish and 

Wildlife Service lands: results of a June 24-25, 2003 workshop, Baltimore, Maryland. Denver, CO: U.S. 

Department of the Interior, National Park Service Air Resources Division, U.S. Fish and Wildlife Service Air 

Quality Branch; Natural Resource Report NPS/NRARD/NRR-2003/01. Available at 

www2.nature.nps.gov/ard/pubs/index.htm. 

United Nations Economic Commission for Europe (UNECE). (1988) ECE Critical Levels Workshop; Bad Harzburg, 

Germany [final report]. Geneva, Switzerland: United Nations Economic Commission for Europe. 

United Nations Environment Programme (UNEP) (1993) Environmental data report 1993-94. 4th ed. Oxford, United 

Kingdom: Blackwell Publishers. 

United Nations Environment Programme (UNEP) (1999) Geo-2000: global environmental outlook 2000. London, 

United Kingdom: Earthscan Publishers; Available: http://www.unep.org/geo2000/english/index.htm [26 July 

2002]. 

Unsworth, M. H.; Hogsett, W. E. (1996) Combined effects of changing CO2, temperature, UV-B radiation and O3 on 

crop growth. In: Bazzaz, F.; Sombroek, W., eds. Global climate change and agricultural production. John 

Wiley & Sons; pp. 171-197. 

Unsworth, M. H.; Heagle, A. S.; Heck, W. W. (1984a) Gas exchange in open-top field chambers - I measurement 

and analysis of atmospheric resistances to gas exchange. Atmos. Environ. 18: 373-380. 

Unsworth, M. H.; Heagle, A. S.; Heck, W. W. (1984b) Gas exchange in open-top field chambers - II resistances to 

ozone uptake by soybeans. Atmos. Environ. 18: 381-385. 

Utriainen, J.; Holopainen, T. (2001) Nitrogen availability modifies the ozone responses of Scots pine seedlings 

exposed in an open-field system. Tree Physiol. 21: 1205-1213. 

Utriainen, J.; Janhunen, S.; Helmisaari, H. S.; Holopainen, T. (2000) Biomass allocation, needle structural 

characteristics and nutrient composition in Scots pine seedlings exposed to elevated CO2 and O3 

concentrations. Trees (Berlin, Ger. ) 14: 475-484. 

Vacek, S.; Bastl, M.; Leps, J. (1999) Vegetation changes in forests of the Krkonose Mountains over a period of air 

pollution stress (1980-1995). Plant Ecol. 143: 1-11. 

Vahala, J.; Schlagnhaufer, C. D.; Pell, E. J. (1998) Induction of an ACC synthase cDNA by ozone in light-grown 

Arabidopsis thaliana leaves. Physiol. Plant. 103: 45-50. 

Vahala, J.; Keinänen, M.; Schützendübel, A.; Polle, A.; Kangasjärvi, J. (2003a) Differential effects of elevated 

ozone on two hybrid aspen genotypes predisposed to chronic ozone fumigation Role of ethylene and salicylic 

acid. Plant Physiol. 132: 196-205. 

Vahala, J.; Ruonala, R.; Keinanen, M.; Tuominen, H.; Kangasjarvi, J. (2003b) Ethylene insensitivity modulates 

ozone-induced cell death in birch. Plant Physiol. 132: 185-195. 

AX9-387
Van Camp, W.; Willekens, H.; Bowler, C.; Van Montagu, M.; Inzé, D.; Reupold-Popp, P.; Sandermann, H., Jr.; 

Langebartels, C. (1994) Elevated levels of superoxide dismutase protect transgenic plants against ozone 

damage. Biotechnology 12: 165-168. 

Van De Staaij, J. W. M.; Lenssen, G. M.; Stroetenga, M.; Rozema, J. (1993) The combined effects of elevated CO2 

levels and UV-B radiation on growth characteristics of Elymus athericus (= E. pycnanthus). Vegetatio 

104/105: 433-439. 

Van De Staaij, J. W. M.; Tonneijck, A. E. G.; Rozema, J. (1997) The effect of reciprocal treatments with ozone and 

ultraviolet-B radiation on photosynthesis and growth of perennial grass Elymus athericus. Environ. Pollut. 97: 

281-286. 

Van Den Driessche, R. (1978) Seasonal changes in root growth capacity and carbohydrates in red pine and white 

spruce nursery seedlings. In: Riedacker, A.; Gagnaire-Michard, J., eds. Root physiology and symbiosis: 

proceedings of the IUFRO symposium; September; Nancy, France. Washington, DC: International Union of 

Forestry Research Organizations; pp. 6-19. 

Van Den Driessche, R. (1991) New root growth of Douglas-fir seedlings at low carbon dioxide concentration. Tree 

Physiol. 8: 289-295. 

Van Den Driessche, R.; Langebartels, C. (1994) Foliar symptoms, ethylene biosynthesis and water use of young 

Norway spruce (Picea abies (L) Karst) exposed to drought and ozone. Water Air Soil Pollut. 78: 153-168. 

Van Oijen, M.; Ewart, F. (1999) The effects of climatic variation in Europe on the yield response of spring wheat cv 

Minaret to elevated CO2 and O3: an analysis of open-top chamber experiments by means of two crop growth 

simulation models. Eur. J. Agron. 10: 249-264. 

Van Oijen, M.; Schapendonk, A. H. C. M.; Jansen, M. J. H.; Pot, C. S.; Maciorowski, R. (1999) Do open-top 

chambers overestimate the effects of rising CO2 on plants? An analysis using spring wheat. Global Change 

Biol. 5: 411-421. 

Vanacker, H.; Carver, T. L. W.; Foyer, C. H. (2000) Early H2O2 accumulation in mesophyll cells leads to induction 

of glutathione during the hyper-sensitive response in the barley-powdery mildew interaction. Plant Physiol. 

123: 1289-1300. 

VanderHeyden, D.; Skelly, J.; Innes, J.; Hug, C.; Zhang, J.; Landolt, W.; Bleuler, P. (2001) Ozone exposure 

thresholds and foliar injury on forest plants in Switzerland. Environ. Pollut. 111: 321-331. 

Vandermeiren, K.; De Temmerman, L.; Hookham, N. (1995) Ozone sensitivity of Phaseolus vulgaris in cultivar 

differences, growth stage and growing conditions. Water Air Soil Pollut. 85: 1455-1460. 

Vandermeiren, K.; Black, C.; Lawson, T.; Casanova, M. A.; Ojanperä, K. (2002) Photosynthetic and stomatal 

responses of potatoes grown under elevated CO2 and/or O3–results from the European CHIP-programme. Eur. 

J. Agron. 17: 337-352.

Vann, D. R.; Strimbeck, G. R.; Johnson, A. H. (1995) Effects of mist acidity and ambient ozone removal on montane 

red spruce. Tree Physiol. 15: 639-647. 

Volin, J. C.; Reich, P. B.; Givnish, T. J. (1998) Elevated carbon dioxide ameliorates the effects of ozone on 

photosynthesis and growth: species respond similarly regardless of photosynthetic pathway or plant functional 

group. New Phytol. 138: 315-325. 

Volk, M.; Geissmann, M.; Blatter, A.; Contat, F.; Fuhrer, J. (2003) Design and performance of a free-air exposure 

system to study long-term effects of ozone on grasslands. Atmos. Environ. 37: 1341-1350. 

Vollenweider, P.; Woodcock, H.; Kelty, M. J.; Hofer, R.-M. (2003a) Reduction of stem growth and site dependency 

of leaf injury in Massachusetts black cherries exhibiting ozone symptoms. Environ. Pollut. 125: 467-480. 

Vollenweider, P.; Ottiger, M.; Günthardt-Goerg, M. S. (2003b) Validation of leaf ozone symptoms in natural 

vegetation using microscopical methods. Environ. Pollut. 124: 101-118. 

Von Gunten, U. (2003) Ozonation of drinking water: part I Oxidation kinetics and product formation. Water Res. 

37: 1443-1467.

Vorne, V.; Ojanperä, K.; De Temmerman, L.; Bindi, M.; Högy, P.; Jones, M. B.; Lawson, T.; Persson, K. (2002) 

Effects of elevated carbon dioxide and ozone on potato tuber quality in the European multiple-site experiment 

'CHIP-project'. Eur. J. Agron. 17: 369-381. 

Vozzo, S. F.; Miller, J. E.; Pursley, W. A.; Heagle, A. S. (1995) Effects of ozone and water deficit on field-grown 

soybean: I. Leaf gas exchange. J. Environ. Qual. 24: 663-670. 

Waite, C. E.; DeHayes, D. H.; Rebbeck, J.; Schier, G. A.; Johnson, A. H. (1994) The influence of elevated ozone on 

freezing tolerance of red spruce seedlings. New Phytol. 126: 327-335. 

Walcek, C. J.; Yuan, H. H.; Stockwell, W. R. (1997) The influence of aqueous-phase chemical reactions on ozone 

formation in polluted and nonpolluted clouds. Atmos. Environ. 31: 1221-1237. 

AX9-388
Wallin, G.; Skärby, L.; Selldén, G. (1992) Long-term exposure of Norway spruce, Picea abies (L.) Karst, to ozone in 

open-top chambers 3 Effects on the light response of net photosynthesis in shoots of different ages. 

New Phytol. 121: 387-394. 

Wallin, G.; Karlsson, P. E.; Selldén, G.; Ottosson, S.; Medin, E.-L.; Pleijel, H.; Skärby, L. (2002) Impact of four 

years exposure to different levels of ozone, phosphorus and drought on chlorophyll, mineral nutrients, and 

stem volume of Norway spruce, Picea abies. Physiol. Plant. 114: 192-206. 

Ward, E. R.; Uknes, S. J.; Williams, S. C.; Dincher, S. S.; Wiederhold, D. L.; Alexander, D. C.; Ahi-Goy, P.; 

Métraux, J. P.; Ryals, J. A. (1991) Coordinate gene activity in response to agents that induce systemic 

acquired resistance. Plant Cell 3: 1084-1094. 

Warwick, K. R.; Taylor, G. (1995) Contrasting effects of tropospheric ozone on five native herbs which coexist in 

calcareous grassland. Global Change Biol. 1: 143-151. 

Weber, J.; Grulke, N. E. (1995) Effects of air pollution on stem growth. In: Gartner, B., ed. Stems and trunks in plant 

form and function. San Diego, CA: Academic Press; pp. 343-364. 

Weber, J. A.; Clark, C. S.; Hogsett, W. E. (1993) Analysis of the relationships among O3 uptake, conductance, and 

photosynthesis in needles of Pinus ponderosa. Tree Physiol. 13: 157-172. 

Weiner, J.; Fredro-Boniecki, S.; Reed, D.; Maclean, A.; Strong, M. (1997) NiepoLomice forest-A GIS analysis of 

ecosystem response to industrial pollution. Environ. Pollut. 3: 381-388. 

Weinstein, D. A.; Yanai, R. D. (1994) Integrating the effects of simultaneous multiple stresses on plants using the 

simulation model TREGRO. J. Environ. Qual. 23: 418-428. 

Weinstein, D. A.; Samuelson, L. J.; Arthur, M. A. (1998) Comparison of the response of red oak (Quercus rubra) 

seedlings and mature trees to ozone exposure using simulation modeling. Environ. Pollut. 102: 307-320. 

Weinstein, D. A.; Gollands, B.; Retzlaff, W. A. (2001) The effects of ozone on a lower slope forest of the Great 

Smoky Mountain National Park: simulations linking an individual tree model to a stand model. For. Sci. 

47: 29-42.

Weinstein, D. A.; Woodbury, P. B.; Gollands, B.; King, P.; Lepak, L.; Pendleton, D. (2002) Assessment of effects of 

ozone on forest resources in the southern Appalachian mountains. Asheville, NC: Southern Appalachian 

Mountains Initiative. Available: http://cycas.cornell.edu/ebp/pmg.html [19 March, 2003]. 

Welfare, K.; Flowers, T. J.; Taylor, G.; Yeo, A. R. (1996) Additive and antagonistic effects of ozone and salinity on 

the growth, ion contents and gas exchange of five varieties of rice (Oryza sativa L). Environ. Pollut. 92: 

257-266. 

Wellburn, A. R. (1990) Why are atmospheric oxides of nitrogen usually phytotoxic and not alternative fertilizers? 

New Phytol. 115: 395-429. 

Wellburn, A. (1994) Air pollution and climate change: the biological impact. Harlow, Essex, United Kingdom: 

Longman Scientific & Technical. 

Wellburn, F. A. M.; Wellburn, A. R. (1994) Atmospheric ozone affects carbohydrate allocation and winter hardiness 

of Pinus halepensis (Mill). J. Exp. Bot. 274: 607-614. 

Wellburn, F. A. M.; Wellburn, A. R. (1996) Variable patterns of antioxidant protection but similar ethene emission 

differences in several ozone-sensitive and ozone-tolerant plant selections. Plant Cell Environ. 19: 754-760. 

Werner, H.; Fabian, P. (2002) Free-air fumigation of mature trees A novel system for controlled ozone enrichment in 

grown-up beech and spruce canopies. Environ. Sci. Pollut. Res. 9: 117-121. 

Wesely, M. L. (1989) Parameterization of surface resistances to gaseous dry deposition in regional-scale numerical 

models. Atmos. Environ. 23: 1293-1304. 

Wesely, M. L.; Hicks, B. B. (2000) A review of the current status of knowledge on dry deposition. Atmos. Environ. 

34: 2261-2282.

Westenbarger, A. D.; Frisvold, G. B. (1994) Agricultural exposure to ozone and acid precipitation. Atmos. Environ. 

28: 2895-2907.

Westman, W. E. (1979) Oxidant effects on Californian coastal sage scrub. Science (Washington, DC) 

205: 1001-1003. 

Westman, W. E. (1981) Factors influencing the distribution of species of Californian coastal sage scrub. Ecology 

62: 439-455.

Weubbles, D. J.; Jain, A.; Edmonds, J.; Harvey, D.; Hayhoe, K. (1999) Global change: state of the science. Environ. 

Pollut. 100: 57-86. 

Wezyk, P.; Guzik, M. (2002) The use of "photogrammetry-GIS" (P-GIS) for the analysis of changes in the Tatra 

Mountains' natural environment. In: Widacki, W.; Bytnerowicz, A.; Riebau, A., eds. A Message from the 

Tatra – geographical information systems and remote sensing in mountain environmental research. Krakow, 

Poland: Jagiellonian University Press. 

AX9-389
Wheeler, T. R.; Hong, T. D.; Ellis, R. H.; Batts, G. R.; Morrison, J. I. L.; Hadley, P. (1996) The duration and rate of 

grain growth, and harvest index, of wheat (Triticum aestivum L) in response to temperature and CO2. J. 

Exper. Bot. 47: 623-630. 

Whitfield, C. P.; Davison, A. W.; Ashenden, T. W. (1996a) Interactive effects of ozone and soil volume on Plantago 

major. New Phytol. 134: 287-294. 

Whitfield, R. G.; Biller, W. F.; Jusko, M. J.; Keisler, J. M. (1996b) A probabilistic assessment of health risks 

associated with short-term exposure to tropospheric ozone. Argonne, IL: Argonne National Laboratory, 

Decision and Information Sciences Division; report no. ANL/DIS-3. 

Whitfield, C. P.; Davison, A. W.; Ashenden, T. W. (1997) Artificial selection and heritability of ozone resistance in 

two populations of Plantago major. New Phytol. 137: 645-655. 

Whitfield, C. P.; Davison, A. W.; Ashenden, T. W. (1998) The effects of nutrient limitation on the response of 

Plantago major to ozone. New Phytol. 140: 219-230. 

Whittaker, J. B. (1994) Interactions between insects and air pollutants. In: Alscher, R. G.; Wellburn, A. R., eds. 

Plant responses to the gaseous environment. London, United Kingdom: Chapman and Hall; pp. 365-384. 

Wickman, B. E. (1992) Forest Health in the Blue Mountains: the influence of insects and diseases. Portland OR: 

U.S. Department of Agriculture, Forest Service, Pacific Northwest Research Station; general technical report 

PNW-GTR-295. (Quigley, T. M., ed. Forest health in the Blue Mountains: science perspectives). 

Widacki, W.; Bytnerowicz, A.; Riebau, A. (2002) A message from the Tatra – geographical information systems and 

remote sensing in mountain environmental reseach. Krakow, Poland: Jagiellonian University Press. 

Wieser, G. (1997) Ozone impact on photosynthetic capacity of mature and young Norway spruce (Picea abies (L) 

Karst): External versus internal exposure. Phyton (Austria) 37: 297-302. 

Wieser, G.; Emberson, L. (2004) Evaluation of the stomatal and conductance formulation in the EMEP ozone 

deposition model for Picea abies. Atmos. Environ. 38: 2339-2348. 

Wieser, G.; Havranek, W. M. (1993) Ozone uptake in the sun and shade crown of spruce: quantifying the 

physiological effects of ozone exposure. Trees (Berlin, Ger. ) 7: 227-232. 

Wieser, G.; Havranek, W. M. (1995) Environmental control of ozone uptake in Larix decidua Mill: a comparison 

between different altitudes. Tree Physiol. 15: 253-258. 

Wieser, G.; Häsler, R.; Götz, B.; Koch, W.; Havranek, W. M. (1999) Seasonal ozone uptake of mature evergreen 

conifers at different altitudes. Phyton (Austria) 39: 233-240. 

Wieser, G.; Tegischer, K.; Tausz, M.; Häberle, K. H.; Grams, T. E. E.; Matyssek, R. (2002) Age effects on Norway 

spruce (Picea abies) susceptibility to ozone uptake: a novel approach relating stress avoidance to defense. 

Tree Physiol. 22: 583-590. 

Wilbourn, S.; Davison, A. W.; Ollerenshaw, J. H. (1995) The use of an unenclosed field fumigation system to 

determine the effects of elevated ozone on a grass-clover mixture. New Phytol. 129: 23-32. 

Willekens, H.; Van Camp, M.; Van Montagu, M.; Inze, D.; Langebartels, C.; Sandermann, H., Jr. (1994) Ozone, 

sulfur dioxide and ultraviolet B have similar effects on mRNA accumulation of fantioxidant genes in 

Nicotiana plumbaginifolia L. Plant Physiol. 106: 1007-1014. 

Williams, T. P. W. (1994) Use of passive monitors to assess plant bioindicators for ground-level ozone [master's 

thesis]. University of New Brunswick; 131 pp. 

Williams, J.; Bulman, M. P.; Neill, S. J. (1994) Wilt-induced aba biosynthesis, gene expression and down regulation 

of rbcS mRNA levels in Arabidopsis thaliana. Physiol. Plant. 91: 177-182. 

Wiltshire, J. J. J.; Wright, C. J.; Unsworth, M. H. (1992) A new method for exposing mature trees to ozone episodes. 

For. Ecol. Manage. 51: 115-120. 

Wiltshire, J. J. J.; Unsworth, M. H.; Wright, C. J. (1994) Seasonal changes in water use of ash trees exposed to 

ozone episodes. New Phytol. 127: 349-354. 

Wiltshire, J. J. J.; Wright, C. J.; Colls, J. J.; Craigon, J.; Unsworth, M. H. (1996) Some foliar characteristics of ash 

trees (Fraxinus exelsior) exposed to ozone episodes. New Phytol. 134: 623-630. 

Winner, W. E.; Lefohn, A. S.; Cotter, I. S.; Greitner, C. S.; Nellessen, J.; McEvoy, L. R., Jr.; Olson, R. L.; Atkinson, 

C. J.; Moore, L. D. (1989) Plant responses to elevational gradients of O3 exposures in Virginia. Proc. Natl. 

Acad. Sci. U. S. A. 86: 8828-8832. 

Winner, W. E.; Coleman, J. S.; Gillespie, C.; Mooney, H. A.; Pell, E. J. (1991) Consequences of evolving reistance 

to air pollutants. In: Taylor, G. E.; Pitelka, L. F.; Clegg, M. T., eds. Ecological genetics and air pollution. 

New York, NY: Springer-Verlag; pp. 177-202. 

Wolf, J.; Van Oijen, M. (2003) Model simulation of effects of changes in climate and atmospheric CO2 and O3 on 

tuber yield potential of potato (cv Bintje) in the European Union. Agric. Ecosyst. Environ. 94: 141-157. 

AX9-390
Wolff, K.; Morgan-Richards, M.; Davison, A. W. (2000) Patterns of molecular genetic variation in Plantago major 

and P. intermedia in relation to ozone resistance. New Phytol. 145: 501-509. 

Wöllmer, H.; Kottke, I. (1990) Fine root studies in situ and in the laboratory. Environ. Pollut. 68: 383-407. 

Wolters, V. (1998) Functional aspects of animal diversity in soil: introduction and overview. Appl. Soil Ecol. 

10: 185-190.

Wonisch, A.; Müller, M.; Tausz, M.; Soja, G.; Grill, D. (1998) Stress-physiological investigations and chromosomal 

analyses on cloned Norway spruce trees exposed to various levels of ozone in open-top chambers. 

Chemosphere 36: 709-714. 

Wonisch, A.; Müller, M.; Tausz, M.; Soja, G.; Grill, D. (1999) Simultaneous analyses of chromosomes in root 

meristems and of the biochemical status of needle tissues of three different clones of Norway spruce trees 

challenged with moderate ozone levels. Eur. J. For. Pathol. 29: 281-294. 

Woodbury, P. B.; Laurence, J. A.; Hudler, G. W. (1994) Chronic ozone exposure increases the susceptibility of 

hybrid Populus to disease caused by Septoria musiva. Environ. Pollut. 86: 109-114. 

Wu, Y. X.; Tiedemann, A. V. (2002) Impact of fungicides on active oxygen species and antioxidant enzymes in 

spring barley (Hordeum vulgare L.) exposed to ozone. Environ. Pollut. 116: 37-47. 

Wulff, A.; Hänninen, O.; Tuomainen, A.; Kärenlampi, L. (1992) A method for open-air exposure of plants to ozone. 

Ann. Bot. Fennici 29: 253-262. 

Wullschleger, S. D.; Hanson, P. J.; Edwards, G. S. (1996) Growth and maintenance respiration in leaves of northern 

red oak seedlings and mature trees after 3 years of ozone exposure. Plant Cell Environ. 19: 577-584. 

Wustman, B. A.; Oksanen, E.; Karnosky, D. F.; Noormets, A.; Isebrands, J. G.; Pregitzer, K. S.; Hendrey, G. R.; 

Sober, J.; Podila, G. K. (2001) Effects of elevated CO2 and O3 on aspen clones varying in O3 sensitivity: can 

CO2 ameliorate the harmful effects of O3? Environ. Pollut. 115: 473-481. 

Yalpani, N.; Enyedi, A. J.; León, J.; Raskin, I. (1994) Ultraviolet light and ozone stimulate accumulation of 

salicyclic acid, pathogenesis-related proteins and virus resistance in tobacco. Planta 193: 372-376. 

Yang, S. F.; Hoffman, N. E. (1984) Ethylene biosynthesis and its regulation in higher plants. Ann. Rev. Plant 

Physiol. 35: 155-189. 

Yonekura, T.; Dokiya, Y.; Fukami, M.; Izuta, T. (2001) Effects of ozone and/or soil water stress on growth and 

photosynthesis of Fagaus crenata seedlings. Water Air Soil Pollut. 130: 965-970. 

Yoshida, L. C.; Gamon, J. A.; Andersen, C. P. (2001) Differences in above- and below-ground responses to ozone 

between two populations of a perennial grass. Plant Soil 233: 203-211. 

Young, T. F.; Sanzone, S., eds. (2002) A framework for assessing and reporting on ecological condition: an SAB 

report. Washington, DC: U.S. Environmental Protection Agency, Science Advisory Board; report no. 

EPA-SAB-EPEC-02-009. Available: http://www.epa.gov/sab/pdf/epec02009.pdf [9 December, 2003]. 

Younglove, T.; McCool, P. M.; Musselman, R. C.; KahI, M. E. (1994) Growth-stage dependent crop yield response 

to ozone exposure. Environ. Pollut. 86: 287-295. 

Yun, S. C.; Laurence, J. A. (1999a) The response of sensitive and tolerant clones of Populus tremuloides to dynamic 

ozone exposure under controlled environmental conditions. New Phytol. 143: 305-313. 

Yun, S. C.; Laurence, J. A. (1999b) The response of clones of Populus tremuloides differing in sensitivity to ozone 

in the field. New Phytol. 141: 411-421. 

Yunus, M.; Iqbal, M. e. (1996) Plant response to air pollution. Chichester, United Kingdom: John Wiley & Sons Ltd. 

Yuska, D. E.; Skelly, J. M.; Ferdinand, J. A.; Stevenson, R. E.; Savage, J. E.; Mulki, J.; Hines, A. (2003) Use of 

bioindicators and passive sampling devices to evaluate ambient ozone concentrations in north central 

Pennsylvania. Environ. Pollut. 125: 71-80. 

Zambrano, A.; Nash, T. H., III. (2000) Lichen responses to short-term transplantation in Desierto de los Leones, 

Mexico City. Environ. Pollut. 107: 407-412. 

Zeller, K.; Nikolov, N. (2000) Quantifying simultaneous fluxes of ozone, carbon dioxide, and water vapor above a 

subalpine forest ecosystem. Environ. Pollut. 107: 20. 

Zhang, X.; Zhang, L.; Dong, F.; Gao, J.; Galbraith, D. W.; Song, C. P. (2001) Hydrogen peroxide is involved in 

abscisic acid-induced stomatal closure in Vicia faba. Plant Physiol. 126: 1438-1448. 

Zinser, C.; Jungblut, T. H.; Seidlitz, H. K.; Schnitzler, J. P.; Ernst, D.; Sandermann, H., Jr. (2000) The effect of 

ozone in Scots pine (Pinus sylvestris L): gene expression, biochemical changes and interactions with UV-B 

radiation. Plant Cell Environ. 23: 975-982. 

Ziska, L.; Bunce, J. A. (1997) The role of temperature in determining the stimulation of CO2 assimilation at elevated 

carbon dioxide concentrations in soybean seedlings. Physiol. Plant. 100: 126-132. 

AX9-391
Please make all necessary changes in the below label, PRESORTED STANDARD EPA detach copy or copy, and return to the address in the upper POSTAGE & FEES PAID 

left-hand corner. EPA 

United States PERMIT No. G-35 

Environmental Protection If you do not wish to receive these reports CHECK HERE ; 

detach copy or copy, and return to the address in the Agency upper left-hand corner. 

National Center for 

Environmental Assessment 

Research Triangle Park, NC 27711 

Official Business 

Penalty for Private Use 

$300 

EPA 600/R-05/004cF 

February 2006