Chapter 9. Environmental Effects: Ozone 

Effects on Vegetation and Ecosystems
 

9.1. In tro d u c tio n 

1 This chapter synthesizes and evaluates the most policy-relevant science to help form the 

2 scientific foundation for the review of a vegetation- and ecologically-based secondary NAAQS for 

3 O3. The secondary NAAQS are based on welfare effects. The Clean Air Act (CAA) definition of 

4 welfare effects includes, but is not limited to, effects on soils, water, wildlife, vegetation, visibility, 

5 weather, and climate, as well as effects on materials, economic values, and personal comfort and 

6 well-being. The effects of O3 as a greenhouse gas and its direct effects on climate are discussed in 

7 Chapter 10 of this document. 

8 The intent of the ISA, according to the CAA, is to “accurately reflect the latest scientific 

9 knowledge expected from the presence of [a] pollutant in ambient air” (42U.S.C.7408 (1990, 

10 080701) 42U.S.C.7409 (1990, 037658)). This chapter of the ISA includes scientific research from 

11 biogeochemistry, soil science, plant physiology, and ecology conducted at multiple scales (e.g., 

12 organ, individual, population, community, ecosystem). Key information and judgments formerly 

13 found in the AQCDs regarding O3 effects on vegetation and ecosystems are found in this chapter. 

14 This chapter of the O3 ISA serves to update and revise Chapter 9 and AX9 of the 2006 O3 AQCD 

15 (U.S. EPA, 2006, 088089). 

16 Numerous studies of the effects of O3 on vegetation and ecosystems were reviewed in the 

17 2006 O3 AQCD. That document concluded that the effects of ambient O3 on vegetation and 

18 ecosystems appear to be widespread across the U.S., and experimental studies demonstrated 

19 plausible mechanisms for these effects. Ozone effect studies published from 2005 to September 2010 

20 are reviewed in this document in the context of the previous O3 AQCDs (U.S. EPA, 2006, 

21 088089)(U.S. EPA, 1996, 080827)(U.S. EPA, 1984, 029711)(U.S. EPA, 1978, 040586). From 2005 

22 to 2010, some areas have had very little new research published and the reader is referred back to 

23 sections of the 2006 O3 AQCD for a more comprehensive treatment of those subjects. This chapter is 

24 focused on studies of vegetation and ecosystems that occur in the U.S. and that report endpoints or 

25 processes most relevant to the review of the secondary standard. Many studies have been published 

26 about vegetation and ecosystems outside of the U.S. and North America, largely in Europe and Asia. 

27 This document includes discussion of studies of vegetation and ecosystems outside of North America 

28 only if those studies contribute to the general understanding of O3 effects across species and 

29 ecosystems. For example, studies outside North America are discussed that consider physiological 

Note: Hyperlinks to the reference citations throughout this document will take you to the NCEA HERO database (Health and 

Environmental Research Online) at http://epa.gov/hero. HERO is a database of scientific literature used by U.S. EPA in the process of 

developing science assessments such as the Integrated Science Assessments (ISAs) and the Integrated Risk Information System (IRIS). 

March 2011 9-1 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

1 and biochemical processes that contribute to the understanding of effects of O3 across species. Also, 

2 ecosystem studies outside of North America that contribute to the understanding of O3 effects on 

3 general ecosystem processes are discussed in the chapter. 

4 This chapter is organized in sections that discuss exposure methods, followed by effects on 

5 vegetation and ecosystems at various scales and ends with policy-relevant discussions of exposure 

6 indices and exposure-response. First, Section 9.2 presents summary information and conclusions 

7 based on the literature assessed in this chapter and the assessments and conclusions of the previous 

8 O3 AQCDs. A brief overview of various methodologies that have been, and continue to be, central to 

9 quantifying O3 effects on vegetation is provided in Section 9.3 (AX9.1 of the 2006 O3 AQCD for 

10 more detailed discussion) (U.S. EPA, 2006, 088089). Sections 9.4 through 9.6 begin with a 

11 discussion of effects at the cellular and subcellular level followed by consideration of the whole 

12 plant and finally, O3 impacts on ecosystem-level processes (Figure 9-1). In Section 9.4, research is 

13 reviewed from the molecular to the biochemical and physiological levels in impacted plants, offering 

14 insight into the mode of action of O3. Section 9.5 provides a review of the effects of O3 exposure on 

15 major endpoints at the whole plant scale including growth, reproduction, visible foliar injury and leaf 

16 gas exchange in woody and herbaceous plants in the U.S., as well as a brief discussion of O3 effects 

17 on agricultural crop yield and quality. The response of plants to O3 as influenced by numerous 

18 environmental biotic and abiotic factors is also discussed Section 9.5. In Section 9.6, available 

19 research for assessing the effect of O3 on ecosystems is reviewed, along with data potentially 

20 available for estimating the loss of various ecosystem services. The development of indices of O3 

21 exposure and dose modeling is discussed in Section 9.7. Finally, exposure-response relationships for 

22 a number of tree species, native vegetation, and crop species and cultivars are reviewed, tabulated, 

23 and compared in Section 9.8 to form the basis for an assessment of the potential risk to vegetation 

24 from current ambient levels of O3. 

9.2. S u mm a ry a n d In te g ra tio n 

9.2.1. Introd uc tion 

25 The subsequent sections of this chapter will present the most policy-relevant information 

26 related to this review of the NAAQS for the effects of O3 on vegetation and ecosystems. This section 

27 integrates the key findings from the disciplines evaluated in this current assessment of the O3 

28 scientific literature, which includes plant physiology, biochemistry, whole plant biology, ecosystems 

29 and exposure-response. 

30 Ozone effects at small scales, such as the leaf of an individual plant, can result in effects at a 

31 continuum of larger scales. Figure 9-1is a simplified diagram of the major pathway through which 

32 O3 enters plants and the major endpoints O3 may affect from small to large scales. The subsequent 

33 sections in this chapter are organized around this paradigm of effects at the cellular and subcellular 

34 level followed by consideration of the whole plant and finally, O3 impacts on ecosystem-level 

March 2011 9-2 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

1

2

3

4

5

6

7

processes. Ozone enters leaves through stomata, and can alter stomatal conductance and disrupt CO2 

fixation (Section 9.4). These effects can change rates of leaf gas exchange, growth and reproduction 

at the individual plant level (Section 9.5). Those O3-induced effects can translate from the individual 

plant level to the ecosystem level, and cause changes in ecosystem services, such as C storage, water 

production, nutrient cycling, and community composition (Section 9.6). The EPA framework for 

causal determinations described in Chapter 1 has been applied to the body of scientific evidence to 

collectively examine effects attributed to O3 exposure (See Table 9-1). 

March 2011 9-3 DRAFT – DO NOT CITE OR QUOTE 
 

Figure 9-1. The effects of ozone at leaf, plant and ecosystem scales. 

March 2011 9-4 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

 

 

 

 

 

Table 9-1 Summary of ozone causal determinations for vegetation and ecosystem effects 

Vegetation and Ecosystem Effects Causality Determination 

Reduced Vegetation Growth Causal 

Alteration of Vegetation Reproduction Causal 

Visible Foliar Injury Effects on Vegetation Causal 

Alteration of Leaf Gas Exchange in Vegetation Causal 

Reduced Yield and Quality of Agricultural Crops Causal 

Reduced Productivity in Terrestrial Ecosystems Causal 

Reduced Carbon (C) Sequestration in Terrestrial Ecosystems Likely Causal 

Alteration of Terrestrial Ecosystem Water Cycling Likely Causal 

Alteration of Below-ground Biogeochemical Cycles Causal 

Alteration of Terrestrial Community Composition Likely Causal 

9.2.2. Me ch a nis ms Go ve rnin g Res p on s e 

1 Section 9.4 focuses on the effects of O3 stress on plants and their responses to that stress on the 

2 molecular, biochemical and physiological levels. Many of the studies focus on the molecular 

3 mechanisms that underlie the observed biochemical and physiological changes observed in many 

4 plant species in response to O3 exposure. The results support and strengthen those reported in the 

5 2006 O3 AQCD. The most significant change in this section from the 2006 O3 AQCD is the emphasis 

6 on molecular mechanisms as new techniques, such as those used in evaluating transcriptomes (total 

7 set of RNA transcripts in a particular cell at a particular time) and proteomes (total set of proteins 

8 expressed in a particular cell at a particular time), have been utilized to perform very comprehensive 

9 analyses of changes in gene transcription and protein expression in plants exposed to O3. These 

10 newer molecular studies not only provide very important and wide-ranging information regarding 

11 the many mechanisms of plant responses to O3, they also allow for the analysis of interactions 

12 between various biochemical pathways which are induced in response to O3. However, many of 

13 these studies are conducted in artificial conditions with model plants which are typically exposed to 

14 very high, short doses of O3. Therefore, additional work remains to elucidate whether these plant 

15 responses are transferable to other plant species exposed to more realistic ambient conditions. 

16 Ozone is taken up into leaves through open stomata. Once inside the substomatal cavity, O3 is 

17 thought to rapidly react with the aqueous layer surrounding the cell (apoplast) to form breakdown 

) and peroxy · ), hydroxyl radicals (HO ־ 18 products such as hydrogen peroxide (H2O2), superoxide (O2 

19 radicals (HO2

·

). These radicals may play a role in signaling processes and may also interact with 

20 sensitive molecules both outside and inside the cell to cause damage. This process was very 

March 2011 9-5 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

1

2

3

4

5

6

7

8

9

10

11

12

13

14

15

16

17

18

19

20

21

22

23

24

25

26

27

28

29

30

31

32

33

34

35

36

37

38

39

comprehensively described in the 2006 O3 AQCD and is only summarized in this document in 

Section 9.4.2. 

As plants have been shown to respond to O3 exposure very rapidly, this response must result 

from a plant's ability to sense the presence of O3 and/or its breakdown products and then 

communicate that information to the nucleus to initiate changes in gene expression. While it remains 

unclear what the exact mechanism is by which the plant senses the presence of O3, whether there are 

multiple simultaneous mechanisms by which O3 can be sensed, and how much variation exists in O3 

sensing between species and exposure conditions, some progress has been made in the understanding 

of this process since the 2006 O3 AQCD. Experimental evidence described in Section 9.4.3.1 

suggests that O3 and/or its breakdown products may be directly sensed by apoplastic receptor 

proteins (although they have not yet been identified). Additionally, a change in cellular redox state 

due to plant exposure to O3 could be the manner in which plants sense the presence of the pollutant. 

Once the plant has sensed the presence of the pollutant, there is much evidence to suggest that 

mitogen-activated protein kinases (MAPK) play an important role in communicating signals to the 

nucleus that result in gene expression changes in response to O3. Calcium has also been implicated to 

play a role in the signal transduction processes. To summarize, the evidence to date suggests there 

may be several mechanisms by which plants sense the presence of O3 and then communicate this 

signal to the nucleus to induce changes in gene expression. 

New technologies have allowed for the evaluation of changes in the entire transcriptome and 

proteome, rather than analyzing the modification of the expression of individual genes and proteins; 

the results of these studies are presented in Section 9.4.3.2. While transcriptome and proteome 

analyses per se were not previously addressed, the 2006 O3 AQCD did provide much information 

regarding changes in gene expression and protein quantity of individual genes and proteins in 

O3-treated plants. In the transcriptome and proteome studies described here, O3 exposure conditions 

(concentration, duration of exposure), plant species and sampling times vary significantly; however, 

functional classification of the genes and proteins that are either up- or down-regulated by plant 

exposure to O3 exhibit common trends. In summary, genes involved in plant defense, signaling, and 

those associated with the synthesis of plant hormones and secondary metabolism are generally upregulated in plants exposed to O3, while those related to photosynthesis and general metabolism are 

typically down-regulated. Proteome studies support these results by demonstrating concomitant 

increases or decreases in the proteins encoded by these genes. The transcriptome and proteome 

results support and enhance the findings of the 2006 O3 AQCD. 

The 2006 O3 AQCD included a discussion on the role of phytohormones, including salicylic 

acid, ethylene and jasmonic acid, in plant response to O3. Many additional studies using microarray 

technology (used to determine changes in the transcriptome) and a variety of Arabidopsis mutants 

are described in Section 9.4.3.3 and support the conclusions from the 2006 O3 AQCD. Transcriptome 

analysis has also illuminated the complex interactions that exist between these hormones to better 

define plant response to O3. To summarize, the results indicate that while ethylene and salicylic acid 

are needed to develop O3-induced leaf lesions, jasmonic acid acts antagonistically to ethylene and 

March 2011 9-6 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

1

2

3

4

5

6

7

8

9

10

11

12

13

14

15

16

17

18

19

20

21

22

23

24

25

26

27

28

29

30

31

32

33

34

35

36

37

38

39

salicylic acid to limit the spread of the lesions. Abscisic acid, in addition to its role in regulating 

stomatal aperture, may also act antagonistically to the jasmonic acid signaling pathway. Nitric oxide 

has also been proposed to play a role in regulating O3-induced changes in gene expression; however, 

its role is not yet well defined. Changes in phytohormones and the interactions between them reveal 

some of the complexity of plant responses to an oxidative stressor such as O3. 

Antioxidant metabolites, such as ascorbate and glutathione, and the enzymes that regenerate 

them are a critical part of plant defense responses to oxidative stress. The role of ascorbate, which is 

located in several cellular compartments and also in the apoplast, was comprehensively evaluated in 

the 2006 O3 AQCD as a first line of defense (due to its location in the apoplast) against oxidative 

stress. Ascorbate has also been the focus of studies investigating differences in O3 tolerance between 

plant species or cultivars/genotypes within one species. While the studies evaluated for the current 

document support the important role of ascorbate, several studies suggest that ascorbate quantity, 

especially in the apoplast, is not the primary factor in determining plant tolerance to O3. In summary, 

antioxidant metabolites and enzymes increase in quantity in plants exposed to O3. In most cases, 

there is a correlation between the degree to which these defensive systems are induced and the 

ability of the plant to tolerate exposure to O3. This up-regulation of antioxidant defenses and the 

need to keep antioxidant metabolites in a reduced state requires a significant shift in C metabolism 

away from growth and reproduction to sustain the energy needs of the plant for defense. 

While declines in C fixation as a result of plant exposure to O3 were extensively described in 

the 2006 O3 AQCD, some recent studies (described in Section 9.4.5.1) of O3-induced declines in 

photosynthesis have focused also on O3 effects on the light reactions. Declines in the Fv/Fm ratio (a 

measure of the maximum efficiency of the light reactions of photosynthesis) were observed in 

several studies using a variety of plant species and exposure conditions. Additionally, O3 increased 

the coefficient of non-photochemical quenching in several species, an indication that defense and 

repair mechanisms of a non-photochemical nature are activated in these plants while less absorbed 

light is being used to drive photosynthesis. This indicates a shift away from photosynthesis to 

defense, resulting in negative impacts on growth and reproduction. 

Section 9.4.5.2 evaluates the effects of O3 on respiration. While C assimilation declines in O3 

exposed plants, respiration is generally up-regulated. These increases in respiration are thought to 

result from a plant's greater energy needs for defense (maintaining its antioxidant metabolites in a 

reduced state) and repair. The increased energy needs will negatively impact plant growth and 

reproduction. 

Secondary metabolism is most often up-regulated in a variety of species exposed to either 

acute or chronic O3 exposures as a part of a generalized plant defense mechanism. Changes in gene 

expression, quantity and activity of enzymes associated with secondary metabolism and alterations 

in secondary metabolite quantity have been documented in plants exposed to O3. Some secondary 

metabolites, such as flavonoids and polyamines, are of particular interest as they are known to have 

antioxidant properties. Investigations on the importance of isoprenes in plant response to O3 have 

revealed conflicting results; however, there is some evidence to suggest that they may play a 

March 2011 9-7 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

1 protective role. In summary, secondary metabolites increase in quantity in O3-treated plants as part of 

2 a generalized plant defense response. Some secondary metabolites are of particular importance in 

3 O3-treated plants as they may have antioxidant functions. Increased synthesis of secondary 

4 metabolites represents a large energy investment of the plant into defense responses and away from 

5 growth and reproduction. 

6 Section 9.4.6 focuses on O3-induced changes in stomatal function. Stomata play a critical role 

7 in limiting O3 uptake into the plant by reducing stomatal aperture. Declines in stomatal conductance 

8 in response to O3 have been documented for many plant species, and much evidence suggests that 

9 this results from increases in intercellular CO2 concentration due to reductions in C fixation. 

10 Additionally, sensitivity of some plants to O3 has been related to a sluggish stomatal response, in 

11 which plants are unable to close their stomata rapidly in response to O3. To summarize, stomatal 

12 response to O3 can help to determine plant sensitivity to the pollutant, and the decreases in stomatal 

13 conductance are thought to be related to declines in C fixation rates. Reduced stomatal conductance 

14 will decrease rates of C assimilation and lead to diminished growth and reproduction in plants. 

9.2.3. Na ture o f Effec ts on Veg eta tio n 

15 Ambient O3 concentrations have long been known to cause visible foliar injury, decreases in 

16 photosynthetic rate, decreases in growth, and decreases in the quality and yield of some plant species 

17 (U.S. EPA, 2006, 088089)(U.S. EPA, 1996, 080827)(U.S. EPA, 1984, 029711)(U.S. EPA, 1978, 

18 040586). Numerous studies have related O3 exposure to plant responses, with most research effort 

19 focused on the growth of tree seedlings and the yield of crops as endpoints. The response of a plant 

20 species to O3 exposure depends upon many factors, including genetic characteristics, biochemical 

21 and physiological status, and previous and current exposure to other stressors. The associated 

22 sections in Section 9.5 focus mainly on studies published since the release of the 2006 O3 AQCD 

23 (U.S. EPA, 2006, 088089). However, because much O3 research was conducted prior to the 2006 O3 

24 AQCD, the conclusions presented below are collectively based on this ISA as well as the 1978, 

25 1986, 1996, and 2006 AQCDs (U.S. EPA, 2006, 088089)(U.S. EPA, 1996, 080827)(U.S. EPA, 1984, 

26 029711)(U.S. EPA, 1978, 040586). 

9.2.3.1. Effe c ts o n Wo ody a nd He rba c e ous Ve ge ta tio n 

Growth and Biomass Allocation 

27 The previous O3 AQCDs concluded that there is strong and consistent evidence that ambient 

28 concentrations of O3 decrease growth in numerous plant species across the U.S. Studies published 

29 since the last review continue to support that conclusion (Section 9.5.2.1). 

30 In a recently published meta-analysis of 263 studies, Wittig et al. (2009, 191631) reported that 

31 current ambient O3 concentrations (~40 ppb) significantly decreased annual total biomass growth of 

32 forest species by an average of 7%, with potentially greater decreases (11-17%) in areas that have 

March 2011 9-8 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

5 

 

 

 

10 

15 

 

 

 

20

 

25 

 

30 

 

 

35

1 higher O3 concentrations and as background O3 increases in the future. This meta-analysis 

2 demonstrates the coherence of O3 effects across numerous studies and species using a variety of 

3 experimental techniques. In a study conducted on mature forest trees, McLaughlin et al. (2007, 

4 090348) reported that the cumulative effects of ambient levels of O3 decreased seasonal stem growth 

by 30-50% for most of the species in a high O3 year in comparison to a low O3 year. 

6 Since the 2006 O3 AQCD, several studies were published based on the Aspen free-air carbon7 dioxide/ozone enrichment (FACE) experiment using “free air”, O3, and CO2 exposures in a forest in 

8 Wisconsin. It was found that O3 caused reductions in total biomass relative to the control in aspen, 

9 paper birch, and sugar maple communities during the first seven years of stand development. 

Overall, the studies at the Aspen FACE experiment were consistent with many of the open-top 

11 chamber (OTC) studies that were the foundation of previous O3 NAAQS reviews. These results 

12 strengthen our understanding of O3 effects on forests and demonstrate the relevance of the 

13 knowledge gained from trees grown in open-top chamber studies. 

14 In recent studies, O3 was shown to have either negative, non-significant, or positive effects on 

root biomass and root:shoot ratio. While the findings of individual studies were mixed, recent meta16 analyses have generally indicated that O3 reduced C allocated to roots (Grantz et al., 2006, 

17 191545)(Wittig et al., 2009, 191631). 

18 For some annual species, particularly crops, the endpoint for an assessment of the risk of O3 

19 exposure can be defined as yield or growth, e.g., production of grain. For plants grown in mixtures 

such as hayfields, and natural or semi-natural grasslands (including native nonagricultural species), 

21 endpoints other than production of biomass may be important. Such endpoints include biodiversity 

22 or species composition, and measures of plant quality. Effects may also result from competitive 

23 interactions among plants in mixed-species communities. Most of the available data on non-crop 

24 herbaceous species are for grasslands with many of the recent studies conducted in Europe. 

Evidence is sufficient to conclude that there is a causal relationship between O3 exposure 

26 and reduced growth of woody and herbaceous vegetation. 

Reproduction 

27 Studies during recent decades have demonstrated O3 effects on different stages of plant 

28 reproduction (Section 9.5.2.2). Several recent studies published since the 2006 O3 AQCD further 

29 demonstrate the effects of O3 on reproductive processes in herbaceous and woody plant species. 

The impacts of O3 on reproductive development can occur by influencing (1) age at time of 

31 initial flowering, particularly in long-lived trees that often have long juvenile periods of early growth 

32 without flower and seed production; (2) flower bud initiation and development; (3) pollen 

33 germination and pollen tube growth; and (4) seed, fruit, or cone yields and seed quality. 

34 Evidence is sufficient to conclude that there is a causal relationship between O3 exposure 

and changes in reproduction of vegetation. 

March 2011 9-9 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

5

 

 

10

 

15

 

 

 

20

25 

 

 

 

30

 

 

35

Visible Foliar Injury 

1 Visible foliar injury resulting from exposure to O3 has been well characterized and 

2 documented over several decades of research on many tree, shrub, herbaceous, and crop species 

3 (U.S. EPA, 2006, 088089)(U.S. EPA, 1996, 080827)(U.S. EPA, 1984, 029711)(U.S. EPA, 1978, 

4 040586)(Section 9.5.2.3). Ozone-induced visible foliar injury symptoms on certain bioindicator plant 

species are considered diagnostic as they have been verified experimentally in exposure-response 

6 studies, using exposure methodologies such as continuous stirred tank reactors (CSTRs), OTCs, and 

7 free-air fumigation. Experimental evidence has clearly established a consistent association of visible 

8 injury with O3 exposure, with greater exposure often resulting in greater and more prevalent injury. 

9 Since the 2006 O3 AQCD, several multiple-year field surveys of O3-induced visible foliar injury 

have been conducted at National Wildlife Refuges in Maine, Michigan, New Jersey, and South 

11 Carolina. New sensitive species showing visible foliar injury continue to be identified from field 

12 surveys and verified in controlled exposure studies. 

13 The use of biological indicators in field surveys to detect phytotoxic levels of O3 is a 

14 longstanding and effective methodology. The USDA Forest Service through the Forest Health 

Monitoring (FHM) Program (1990-2001) and currently the Forest Inventory and Analysis (FIA) 

16 Program has been collecting data regarding the incidence and severity of visible foliar injury on a 

17 variety of O3 sensitive plant species throughout the U.S. The network has provided evidence that O3 

18 concentrations were high enough to induce visible symptoms on sensitive vegetation. From repeated 

19 observations and measurements made over a number of years, specific patterns of areas experiencing 

visible O3 injury symptoms can be identified. 

21 In addition, a study by Kohut (2007, 093289) assessed the risk of O3-induced visible foliar 

22 injury on bioindicator plants (NPS, 2006, 677536) in 244 national parks in support of the National 

23 Park Service's Vital Signs Monitoring Network (NPS, 2007, 677537). Kohut (2007, 093289) 

24 concluded that the risk of visible foliar injury was high in 65 parks (27%), moderate in 46 parks 

(19%), and low in 131 parks (54%). Some of the well-known parks with a high risk of O3-induced 

26 visible foliar injury include Gettysburg, Valley Forge, Delaware Water Gap, Cape Cod, Fire Island, 

27 Antietam, Harpers Ferry, Manassas, Wolf Trap Farm Park, Mammoth Cave, Shiloh, Sleeping Bear 

28 Dunes, Great Smoky Mountains, Joshua Tree, Sequoia and Kings Canyon, and Yosemite. 

29 Evidence is sufficient to conclude that there is a causal relationship between ambient O3 

exposure and the occurrence of O3-induced visible foliar injury on sensitive vegetation across 

31 the U.S. 

Gas Exchange 

32 There is strong experimental evidence over several decades of research that exposure to O3 

33 reduces photosynthesis and alters stomatal conductance in a wide variety of plant species. The mode 

34 of action, as characterized in Section 9.4 and in previous reviews, provides biological plausibility for 

O3 effects on leaf gas exchange. 

March 2011 9-10 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

1 In compiling more than 55 studies, Wittig et al. (2007, 191695) reported that current O3 

2 concentrations in the northern hemisphere are decreasing photosynthesis (11%) and stomatal 

3 conductance (13%) across tree species. It was also found that younger trees (<4 year) were affected 

4 less by O3 than older trees. Further, the authors also found that decreases in photosynthesis are 

5 consistent with the cumulative uptake of O3 into the leaf. In contrast, several studies reported that O3 

6 exposure may result in loss of stomatal control, incomplete stomatal closure at night and a 

7 decoupling of photosynthesis and stomatal conductance, which may have implications for whole8 plant water use (Section 9.6.3). 

9 Evidence is sufficient to conclude that there is a causal relationship between O3 exposure 

10 and the alteration of leaf gas exchange in vegetation. 

9.2.3.2. Agric ultura l Cro ps 

Yield and Crop Quality 

11 The detrimental effect of O3 on crop production has been recognized since the 1960’s and a 

12 large body of research has subsequently stemmed from those initial findings. Previous O3 AQCDs 

13 have extensively reviewed this body of literature (U.S. EPA, 2006, 088089). Recent experimental 

14 studies of O3 effects on crops are discussed in Section 9.5.3 and summarized in Table 9-3 and 9-16 

15 Current O3 concentrations across the U.S. are high enough to cause yield loss for a variety of 

16 agricultural crops including, but not limited to, soybean, wheat, cotton, potato, watermelon, beans, 

17 turnip, onion, lettuce, and tomato. Continued increases in O3 concentration may further decrease 

18 yield in these sensitive crops while also initiating yield losses in less sensitive crops. Despite the 

19 well-documented yield losses due to increasing O3 concentration, there is still a knowledge gap 

20 pertaining to the exact mechanism of O3-induced yield loss. Research has linked increasing O3 

21 concentration to decreased photosynthetic rates and accelerated senescence, which are related to 

22 yield. 

23 Recent modeling research has correlated satellite air-column observations with direct air24 sampling O3 data and modeled the yield-loss due to O3 over the continuous tri-state area of Illinois, 

25 Iowa and Wisconsin. This modeling data correlates well with the previous results from FACE-type 

26 experiments and OTC experiments. 

27 New research is beginning to consider the mechanism of damage caused by long, lower O3 

28 concentration (so-called chronic exposure) compared to short, very high O3 concentration (so-called 

29 acute exposure). Both types of O3 exposure cause damage to agricultural crops, but through very 

30 different mechanisms. Until recently, most research on the mechanism of O3 damage has used acute 

31 exposure studies. It has become clear that the same cellular and biochemical processes involved in 

32 the response to acute O3 exposure are not involved in response to chronic O3 exposure, yet both 

33 cause yield-loss in agriculturally important crops. 

34 In addition, new research has highlighted the effects of O3 on crop quality. Increasing O3 

35 concentration decreases nutritive quality of grasses, decreases macro- and micro-nutrient 

March 2011 9-11 DRAFT – DO NOT CITE OR QUOTE 
 

 

5 

 

 

 

10

 

15 

 

 

20 

 

 

 

25 

 

 

30 

 

 

1 concentrations in fruits and vegetable crops, and decreases cotton fiber quality. These areas of
 

2 research require further investigation to determine the mechanism and dose-responses.
 

3 Evidence is sufficient to conclude that there is a causal relationship between O3 exposure
 

4 and reduced yield and quality of agricultural crops.
 

9.2.3.3. Fa c tors Tha t Mo dify Fu nc tio na l a nd Gro wth Re s pons e 

Many biotic and abiotic factors, including insects, pathogens, root microbes and fungi, 

6 temperature, water and nutrient availability, and other air pollutants, as well as elevated CO2, 

7 influence or alter plant response to O3. These modifying factors were comprehensively reviewed in 

8 the 2006 O3 AQCD. A limited number of studies published since 2006 provide further support for 

9 our understanding of the role of these interactions in modifying O3-induced plant responses and are 

discussed in Section 9.5.4. 

9.2.4. Ecos ys tems and Services 

11 Ozone has been found to alter plant physiological processes such as growth, biomass 

12 allocation, reproduction and gas exchange (Section 9.5). Those O3-induced effects at the individual 

13 plant scale have the potential to translate to effects at the ecosystem level, and cause changes in 

14 biogeochemical cycling and community composition. Information presented in the associated section 

(Section 9.6) was collected at multiple scales, ranging from responses at the population level to the 

16 ecosystem level. The effects of O3 on ecosystem productivity, C sequestration, water cycling, 

17 nutrient cycling, and community composition are reviewed. 

9.2.4.1. Productivity and Carbon Se ques tration 

18 During the previous NAAQS reviews, there were very few studies that investigated the effect 

19 of O3 exposure on ecosystem productivity and C sequestration. Recent studies from long-term FACE 

experiments provided evidence of the association of O3 exposure and reduced productivity at the 

21 ecosystem level. Elevated O3 reduced stand-level biomass by 13-23% at Aspen FACE after 7 years 

22 of O3 exposure (King et al., 2005, 191701), and annual volume growth by 9.5 m3/ha at the Kranzberg 

23 Forest (Germany) FACE (Pretzsch et al., 2010, 580435). Studies at the leaf and plant scales showed 

24 that O3 reduced photosynthesis and plant growth, which provided coherence and biological 

plausibility for the decrease in ecosystem productivity. Results across different ecosystem models 

26 were consistent with the FACE experimental evidence, which showed that O3 reduced ecosystem 

27 productivity. 

28 Although O3 generally causes negative effects on plant growth, the magnitude of the response 

29 varies among plant communities. For example, O3 had little impact on white fir, but greatly reduced 

growth of ponderosa pine in southern California (Weinstein et al., 2005, 179965). Ozone decreased 

31 net primary production (NPP) of most forest types by 7-8% in Mid-Atlantic region, but had small 

32 impacts on spruce-fir forest, which was decreased by only 1% (Pan et al., 2009, 596032). Among 

March 2011 9-12 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

5

 

10 

 

15

 

20

 

 

25

 

 

30

 

35 

1 crop species, the estimated yield loss for wheat (7-12%) and soybean (6-16%) were higher than rice 

2 (3-4%) and maize (3-5%) (Van Dingenen et al., 2009, 199765). 

3 In addition to plant growth, other indicators that are typically estimated by model studies 

4 include net ecosystem CO2 exchange (NEE), C sequestration, and crop yield. Model simulations 

consistently found that O3 exposure caused negative impacts on those indicators (Section 9.6.2, 

6 Table 9-5), but the severity of these impacts was influenced by multiple interactions of biological 

7 and environmental factors. For example, the largest O3-induced crop yield losses occurred in high8 production areas exposed to high O3 concentrations, such the Midwest and the Mississippi Valley 

9 regions of the U.S. (Van Dingenen et al., 2009, 199765). 

The suppression of ecosystem C sinks results in more CO2 accumulation in the atmosphere. 

11 Globally, the indirect radiative forcing, reported in Watts/square meter (W/m2), caused by O3 

12 exposure through lowering ecosystem C sink (0.62-1.09 W/m2 ) could have an even greater impact 

13 on global warming than the direct radiative forcing of O3 (0.89 W/m2) (Sitch et al., 2007, 093294). 

14 Ozone could also affect regional C budgets through interacting with multiple factors, such as N 

deposition, elevated CO2 and land use history. Model simulations suggested that O3 partially offset 

16 the growth stimulation caused by elevated CO2 and N deposition in both Northeast- and Mid17 Atlantic-region forest ecosystems of the U.S. (Ollinger et al., 2002, 180189)(Pan et al., 2009, 

18 596032). 

19 The evidence is sufficient to infer that there is a causal relationship between O3 exposure 

and reduced productivity, and a likely causal relationship between O3 exposure and reduced 

21 carbon sequestration in terrestrial ecosystems. 

9.2.4.2. Wa te r Cyc ling 

22 Although the evidence was from a limited number of field and modeling studies, these 

23 findings showed an association of O3 exposure and the alteration of water cycle at the ecosystem 

24 level. Field studies conducted by McLaughlin et al. (2007, 090348)(2007, 090347) suggested that 

peak hourly O3 exposure increased the rate of water loss from several tree species, and led to a 

26 reduction in the late-season modeled stream flow in three forested watersheds in eastern Tennessee. 

27 Evidence of sluggish stomatal responses during O3 exposure was found in their study and several 

28 other studies (Section 9.6.3), which provided biological plausibility for the observed higher water 

29 loss at the ecosystem level. However, many experiments, mostly based on short-term O3 exposure, 

found that O3 generally reduced stomatal conductance. The O3-induced reduction in stomatal 

31 aperture is the biological assumption for most process-based models. Therefore, results of those 

32 models normally found that O3 reduced water loss. For example, Felzer (2009, 191460) found that 

33 O3 damage and N limitation together reduced evapotranspiration and increase runoff. 

34 Although the direction of the response differed among studies, the evidence is sufficient to 

conclude that there is likely to be a causal relationship between O3 exposure and the alteration 

36 of ecosystem water production. 

March 2011 9-13 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

5 

 

 

10 

 

 

 

15 

 

20 

 

 

25

 

30 

 

9.2.4.3. Below-Ground Processes 

1 Since the 2006 O3 AQCD, more evidence has shown that although the responses are often 

2 species specific, O3 altered the quality and quantity of C input to soil, microbial community 

3 composition, and C and nutrient cycling. Biogeochemical cycling of below-ground processes is 

4 driven by C input from plants. Studies at the leaf and plant level have provided biologically plausible 

mechanisms, such as reduced photosynthetic rates, increased metabolic cost, and reduced root C 

6 allocation (Section 9.6.4) for the association of O3 exposure and the alteration of below-ground 

7 processes. 

8 Results from Aspen FACE and other experimental studies consistently found that O3 reduced 

9 litter production and altered C chemistry, such as soluble sugars, soluble phenolics, condensed 

tannins, lignin, and macro/micro nutrient concentration in litter (Liu et al., 2005, 187005)(Parsons et 

11 al., 2008, 191853)(Kasurinen et al., 2006, 191269). The changes in substrate quality and quantity 

12 could alter microbial metabolism under elevated O3, and therefore soil C and nutrient cycling. 

13 Several studies indicated that O3 generally suppressed soil enzyme activities (Chung et al., 2006, 

14 191729)(Pritsch et al., 2009, 626808). However, the impact of O3 on litter decomposition was 

inconsistent and varied among species, sites and exposure length. Ozone had small impact on 

16 dynamics of micro and macro nutrients, except for N. Ozone was found to reduce N release from 

17 leaf litter and decrease gross N mineralization, which could potentially decrease N availability to 

18 plants (Holmes et al., 2006, 191372)(Liu et al., 2007, 093286). 

19 Studies from the Aspen FACE experiment suggested that the response of below-ground 

C cycle to O3 exposure, such as litter decomposition, soil respiration and soil C content, changed 

21 over time. For example, in the early part of the experiment (1998-2003), O3 had no impact on soil 

22 respiration but reduced the formation rates of total soil C under elevated CO2. However, after 10­

23 11 yr of exposure, O3 was found to increase soil respiration but have no significant impact on soil 

24 C formation under elevated CO2 (Section 9.6.4.3). 

The evidence is sufficient to infer that there is a causal relationship between O3 exposure 

26 and the alteration of below-ground biogeochemical cycles. 

9.2.4.4. Community Compos ition 

27 In the 2006 O3 AQCD, the impact of O3 exposure on species competition and community 

28 composition was assessed. Ozone was found to cause a significant decline in ponderosa and Jeffrey 

29 pine in the San Bernardino Mountains in southern California. Ozone exposure also tended to shift 

the grass-legume mixtures in favor of grass species (U.S. EPA, 2006, 088089). Since the 2006 O3 

31 AQCD, more evidence has shown that O3 exposure changed the competitive interactions and led to 

32 loss of O3 sensitive species or genotypes. Studies at plant level found that the severity of O3 damage 

33 on growth, reproduction and foliar injury varied among species (Section 9.6.5), which provided the 

34 biological plausibility for the alteration of community composition. Additionally, research since the 

March 2011 9-14 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

5 

 

 

 

10

 

 

15

 

20 

 

25

 

 

 

30

 

35

1 last review has shown that O3 can alter community composition and diversity of soil microbial 

2 communities. 

3 The decline of conifer forests under O3 exposure was continually observed in several regions. 

4 Ozone damage was believed to be an important causal factor in the dramatic decline of sacred fir in 

the valley of Mexico (de Lourdes de Bauer and Hernandez-Tejeda, 2007, 196891), as well as 

6 cembran pine in southern France and Carpathian Mountains (Wieser et al., 2006, 191391). Results 

7 from the Aspen FACE site indicated that O3 could alter community composition of broadleaf forests 

8 as well. At the Aspen FACE site, O3 reduced growth and increased mortality of a sensitive aspen 

9 clone, while the O3 tolerant clone emerged as the dominant clone in the pure aspen community. In 

the mixed aspen-birch and aspen-maple communities, O3 reduced the competitive capacity of aspen 

11 compared to birch and maple (Kubiske et al., 2007, 191336). 

12 The tendency for O3-exposure to shift the biomass of grass-legume mixtures in favor of grass 

13 species, was reported in the 2006 O3 AQCD (U.S. EPA, 2006, 088089) and has been generally 

14 confirmed by recent studies. However, in a high elevation mature/species-rich grass-legume pasture, 

O3 fumigation showed no significant impact on community composition (Bassin et al., 2007, 

16 191534). 

17 Ozone exposure not only altered community composition of plant species, but also 

18 microorganisms. The shift in community composition of bacteria and fungi has been observed in 

19 both natural and agricultural ecosystems, although no general patterns could be identified (Kanerva 

et al., 2008, 191264)(Morsky et al., 2008, 191507)(Kasurinen et al., 2005, 191245). 

21 The evidence is sufficient to conclude that there is likely a causal relationship between O3 

22 exposure and the alteration of community composition. 

9.2.5. Air Qu ality Indic es 

23 Exposure indices are metrics that quantify exposure as it relates to measured plant damage 

24 (i.e., reduced growth). They are summary measures of monitored ambient O3 concentrations over 

time intended to provide a consistent metric for reviewing and comparing exposure-response effects 

26 obtained from various studies. No new information is available since 2006 that alters the basic 

27 conclusions put forth in the 2006 and 1996 O3 AQCDs (U.S. EPA, 1996, 080827)(U.S. EPA, 2006, 

28 088089). These AQCDs focused on the research used to develop various exposure indices to help 

29 quantify effects on growth and yield in crops, perennials, and trees (primarily seedlings). The 

performance of indices was compared through regression analyses of earlier studies designed to 

31 support the estimation of predictive O3 exposure-response models for growth and/or yield of crops 

32 and tree (seedling) species. 

33 The main conclusions from the 1996 and 2006 O3 AQCDs (U.S. EPA, 1996, 

34 080827)(U.S. EPA, 2006, 088089) regarding an index based on ambient exposure are still valid. 

These key conclusions can be restated as follows: 

36  O3 effects in plants are cumulative; 

March 2011 9-15 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

5

 

 

 

10

 

15 

 

 

 

20 

 

25 

 

 

30 

 

 

 

1  higher O3 concentrations appear to be more important than lower concentrations in 

2 eliciting a response;
 

3  plant sensitivity to O3 varies with time of day and plant development stage; and 

4  exposure indices that cumulate hourly O3 concentrations and preferentially weight the 

higher concentrations have better statistical fits to growth/yield response data than do the 

6 mean and peak indices. 

7 Various weighting functions have been used, including threshold-weighted (e.g., SUM06) and 

8 continuous sigmoid-weighted (e.g., W126) functions. Based on statistical goodness-of-fit tests, these 

9 cumulative, concentration-weighted indices could not be differentiated from one another using data 

from previous exposure studies. Additional statistical forms for O3 exposure indices have been 

11 discussed in Lee et al. (1988, 042136) and in Section 9.7 of this ISA. The majority of studies 

12 published since the 2006 O3 AQCD (2006, 088089) do not change earlier conclusions, including the 

13 importance of peak concentrations, and the duration and occurrence of O3 exposures in altering plant 

14 growth and yield. 

Given the current state of knowledge and the best available data, exposure indices that 

16 cumulate and differentially weight the higher hourly average concentrations and also include the 

17 mid-level values continue to offer the most defensible approach for use in developing response 

18 functions and comparing studies, as well as for defining future indices for vegetation protection. 

9.2.5.1. Mo de le d Ozone De p os ition or "Flux" 

19	 Another approach for improving risk assessment of vegetation response to ambient O3 is based 

on determining the O3 concentration from the atmosphere that enters the leaf (i.e., flux or 

21 deposition). Interest has been increasing in recent years, particularly in Europe, in using 

22 mathematically tractable flux models for O3 assessments at the regional, national, and European 

23 scale (Matyssek et al., 2008, 191262)(Paoletti and Manning, 2007, 180174)(Emberson et al., 2000, 

24 040350)(Emberson et al., 2000, 042537)(ICP M&M, 2004, 677471). While some efforts have been 

made in the U.S. to calculate O3 flux into leaves and canopies (Grantz et al., 1997, 026664)(Grantz 

26 et al., 1995, 026659)(Grulke et al., 2004, 042646)(Turnipseed et al., 2009, 588752)(Uddling et al., 

27 2009, 596219)(Bergweiler et al., 2008, 191656)(Hogg et al., 2007, 199349), little information has 

28 been published relating these fluxes to effects on vegetation. There is also concern that not all O3 

29 stomatal uptake results in a yield reduction, which depends to some degree on the amount of internal 

detoxification occurring with each particular species. Those species having high amounts of 

31 detoxification potential may, in fact, show little relationship between O3 stomatal uptake and plant 

32 response (Musselman and Massman, 1999, 040706). The lack of data in the U.S. and the lack of 

33 understanding of detoxification processes have made this technique less viable for vulnerability and 

34 risk assessments in the U.S. 

March 2011	 9-16 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

5 

 

 

10 

 

 

15

 

 

20

 

 

25

30

 

 

35

9.2.5.2. Night-Time Exp os ures 

1 A 12-h (8:00 a.m. to 8:00 p.m.) daylight period for a W126 cumulating exposure was proposed 

2 in 2007 and 2009 following the release of the 2006 O3 AQCD and was based primarily on evidence 

3 that the conditions for uptake of O3 into the plant occur mainly during the daytime hours. Plants have 

4 the highest stomatal conductance during the daytime and atmospheric turbulent mixing is greatest 

then as well (U.S. EPA, 2006, 088089)(Uddling et al., 2010, 387073). Recent reviews of the 

6 literature reported that a large number of species had varying degrees of nocturnal stomatal 

7 conductance (Caird et al., 2007, 199337)(Dawson et al., 2007, 670381)(Musselman and Minnick, 

8 2000, 011612). In general, stomatal conductance at night is at a much lower rate compared to 

9 daytime conductance (Caird et al., 2007, 199337). For significant nocturnal stomatal flux and O3 

effects to occur, specific conditions must exist. A susceptible plant with nocturnal stomatal 

11 conductance and low defense must be growing in an area with relatively high night-time O3 and 

12 appreciable nocturnal turbulence. It is unclear how many areas there are in the U.S. where these 

13 conditions occur. More information is needed in these locations in order to assess the local O3 

14 patterns, micrometeorology and responses of potentially vulnerable plant species. 

9.2.6. Expos ure -Response 

None of the information on effects of O3 on vegetation published since the 2006 O3 AQCD has 

16 modified the assessment of quantitative exposure-response relationships that was presented in that 

17 document (U.S. EPA, 2006, 088089). This assessment updates the 2006 exposure-response models 

18 by computing them using the W126 metric, cumulated over 90 days. Almost all of the experimental 

19 research on the effects of O3 on growth or yield of plants published since 2006 used only two levels 

of exposure. In addition, hourly O3 concentration data that would allow calculations of exposure 

21 using the W126 scale are generally unavailable. However, two long-term experiments, one with a 

22 crop species (soybean), one with a tree species (aspen), have produced data that can be used to 

23 validate the exposure-response models presented in the 2006 O3 AQCD, and methodology used to 

24 derive them. 

Quantitative characterization of exposure-response in the 2006 O3 AQCD was based on 

26 experimental data generated for that purpose by the National Crop Loss Assessment Network 

27 (NCLAN) and EPA National Health and Environmental Effects Research Laboratory, Western 

28 Ecology Division (NHEERL-WED) projects, using OTCs to expose crops and trees seedling to O3. 

29 In recent years, yield and growth results for two of the species that had provided extensive exposureresponse information in those projects have become available from studies that used FACE 

31 technology, which is intended to provide conditions much closer to natural environments (Dickson et 

32 al., 2000, 628220; Morgan et al., 2004, 072764; Morgan et al., 2006, 079186; Pregitzer et al., 2008, 

33 191677). The robust methods that were used previously with exposure measured as SUM06 were 

34 applied to the NCLAN and NHEERL-WED data with exposure measured as W126, in order to 

derive single-species median models for soybean and aspen from studies involving different 

March 2011 9-17 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

1 genotypes, years, and locations. The resulting models were used to predict the change in yield of 

2 soybean and biomass of aspen between the two levels of exposure reported in current FACE 

3 experiments. Results from these new experiments were exceptionally close to predictions from the 

4 models. The accuracy of model predictions for two widely different plant species provides support 

5 for the validity of the corresponding multiple-species models for crops and trees in the NCLAN and 

6 NHEERL-WED projects. However, variability among species in those projects indicates that the 

7 range of sensitivity is likely quite wide. This was confirmed by a recent experiment with cottonwood 

8 in a naturally occurring gradient of exposure (Gregg et al., 2006, 186961), which established the 

9 occurrence of species with responses substantially more severe under currently existing conditions 

10 than are predicted by the median model for multiple species. 

11 Results from several meta-analyses have provided approximate values for responses of yield 

12 of soybean, wheat, rice and other crops under broad categories of exposure, relative to charcoal13 filtered air (Ainsworth, 2008, 191646; Feng et al., 2008, 191453; Morgan et al., 2003, 055527). 

14 Likewise, Feng and Kobayashi (2009, 199223) have summarized yield data for six crop species 

15 under various broad comparative exposure categories, while Wittig et al. (2009, 191631) reviewed 

16 263 studies that reported effects on tree biomass. However, these analyses have proved difficult to 

17 compare with exposure-response models, especially given that exposure was not expressed on the 

18 same W126 scale. 

9.3. Exp e rime n ta l Expo s u re Me th o do lo gie s 

9.3.1. Introd uc tion 

19 A variety of methods for studying plant response to O3 exposures have been developed over 

20 the last several decades. Methodological advancements since 2006 have not fundamentally altered 

21 our understanding of O3 effects on plants or ecosystems. The majority of methodologies currently 

22 used have been discussed in detail in the 1996 O3 AQCD (Section 5.2; U.S. EPA, 1996, 080828) and 

23 2006 O3 AQCD (Section AX9.1; U.S. EPA, 2006, 088089). This section will serve as a short 

24 overview of the methodologies and the reader is referred to the previous O3 AQCDs for more in25 depth discussion. 

9.3.2. “Ind o or,” Con trolled En viro n men t, an d Gre en ho u s e Ch a mb ers 

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

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

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

29 and locations of these chambers have been numerous. Hogsett et al. (1987, 043465) have 

30 summarized the construction and performance of more elaborate and better instrumented chambers 

31 since the 1960s, including those installed in greenhouses (with or without some control of 

32 temperature and light intensity). 

March 2011 9-18 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

5 

 

 

 

 

10

 

15 

 

20

 

 

25 

 

30 

 

 

35 

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

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

3 reactors (CSTRs) first described by Heck et al. (1978, 037673). Although originally developed to 

4 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 (Heagle et al., 1994, 

6 026699)(Loats and Rebbeck, 1999, 029709)(Rao et al., 1995, 030221)(Reinert and Ho, 1995, 

7 030247)(Reinert et al., 1997, 030252), and validation of visible foliar injury on a variety of plant 

8 species (Kline et al., 2009, 196918)(Orendovici et al., 2003, 049080). In many cases, supplementary 

9 lighting and temperature control of the surrounding structure have been used to control or modify the 

environmental conditions (Heagle et al., 1994, 026699). 

11 Many investigations have utilized commercially available controlled environment chambers 

12 and walk-in rooms adapted to permit the introduction of a flow of O3 into the controlled air-volume. 

13 Such chambers continue to find use in genetic screening and in physiological and biochemical 

14 studies aimed primarily at improving our understanding of modes of action. For example, some of 

the studies of the O3 responses of common plantain (Plantago major) populations have been 

16 conducted in controlled environment chambers (Reiling and Davison, 1994, 035373)(Whitfield et 

17 al., 1996, 055544). 

18 More recently, some researchers have been interested in direct O3 effects on reproductive 

19 processes, separate from the effects on vegetative processes (Black et al., 2010, 625575). For this 

purpose, controlled exposure systems have been employed to expose the reproductive structures of 

21 annual plants to gaseous pollutants independently of the vegetative component (Black et al., 2010, 

22 625575)(Stewart et al., 1996, 036384). 

9.3.3. Field Chambers 

23 In general, field chamber studies are dominated by the use of various versions of the open top 

24 chamber (OTC) design, first described by Heagle et al. (1973, 038348) and Mandl et al. (1973, 

039967). The OTC method continues to be a widely used technique in the U.S. and Europe for 

26 exposing plants to varying levels of O3. Most of the new information confirms earlier conclusions 

27 and provides additional support for OTC use in assessing plant species and in developing exposure28 response relationships. Chambers are generally ~3 m in diameter with 2.5-m-high walls. Hogsett et 

29 al. (1987, 043780) described in detail many of the various modifications to the original OTC designs 

that appeared subsequently, e.g., the use of larger chambers for exposing small trees (Kats et al., 

31 1985, 055511) or grapevines (Mandl et al., 1989, 043987), the addition of a conical baffle at the top 

32 to improve ventilation (Kats et al., 1976, 039799), a frustum at the top to reduce ambient air 

33 incursions, and a plastic rain-cap to exclude precipitation (Hogsett et al., 1985, 039383). All versions 

34 of OTCs included the discharge of air via ports in annular ducting or interiorly perforated doublelayered walls at the base of the chambers to provide turbulent mixing and the upward mass flow of 

36 air. 

March 2011 9-19 DRAFT – DO NOT CITE OR QUOTE 
 

1

2

3 

4

5 

6

7

8 

9

10

11 

12 

13 

14 

15

16 

17

18

19 

20 

21 

22 

23

24

25

26 

27

28 

29

30 

31

32 

33 

34 

35 

36

37

38 

Chambered systems, including OTCs, have several advantages. For instance, they can provide 

a range of treatment levels including charcoal-filtered (CF), clean-air control, and several above 

ambient concentrations for O3 experiments. Depending on experimental intent, a replicated, clean-air 

control treatment is an essential component in many experimental designs. The OTC can provide a 

consistent, definable exposure because of the constant wind speed and delivery systems. Statistically 

robust concentration-response (C-R) functions can be developed using such systems for evaluating 

the implications of various alternative air quality scenarios on vegetation response. Nonetheless, 

there are several characteristics of the OTC design and operation that can lead to exposures that 

might differ from those experienced by plants in the field. First, the OTC plants are subjected to 

constant air flow turbulence, which, by lowering the boundary layer resistance to diffusion, may 

result in increased uptake. This may lead to an overestimation of effects relative to areas with less 

turbulence (Krupa et al., 1995, 038895)(Legge et al., 1995, 040689). Conversely, however, other 

research has found that OTC's may slightly change vapor pressure deficit (VPD) in a way that may 

decrease the uptake of O3 into leaves (Piikki et al., 2008, 199265). As with all methods that expose 

vegetation to modified O3 concentrations in chambers, OTCs create internal environments that differ 

from ambient air. This so-called “chamber effect” refers to 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, 055549)(Manning and Krupa, 

1992, 044155). However, in at least one case where canopy resistance was 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., 1984, 

041239)(Unsworth et al., 1984, 040024). Because of the standardized methodology and protocols 

used in National Crop Loss Assessment Network (NCLAN) and other programs, the database can be 

assumed to be internally consistent. 

While it is clear that OTCs can alter some aspects of the microenvironment and plant growth, 

it is important to establish whether or not these differences affect the relative response of a plant to 

O3. As noted in the 1996 O3 AQCD (U.S. EPA, 1996, 080828), evidence from a number of 

comparative studies of OTCs and other exposure systems suggested that responses were essentially 

the same regardless of exposure system used and chamber effects did not significantly affect 

response. For example, a 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, 042660). The response found in OTCs was the same as in ambient plots. 

Another type of field chamber called a “terracosm” has been developed and used in recent 

studies (Lee et al., 2009, 595904). 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, 055536) 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. 

March 2011 9-20 DRAFT – DO NOT CITE OR QUOTE 
 

 

5 

 

 

10

 

15 

 

 

20 

 

 

 

25

 

 

 

30 

 

 

35 

 

 

1 Other researchers have recently published studies using another type of out-door chamber 

2 called recirculating Outdoor Plant Environment Chambers (OPECs) (Flowers et al., 2007, 191852). 

3 These closed chambers are approximately 2.44 m×1.52 m with a growth volume of approximately 

4	 3.7 m3 in each chamber. These chambers admit 90% of full sunlight and control temperature,
 

humidity and vapor pressure (Fiscus et al., 1999, 672964).
 

9.3.4. Plu me an d FACE-Type Sys te ms 

6 Plume systems are chamberless exposure facilities in which the atmosphere surrounding plants 

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

8 multiple orifices spaced to permit diffusion and turbulence, so as to establish relatively homogeneous 

9 conditions as the individual plumes disperse and mix with the ambient air. They can only be used to 

increase the O3 levels in the ambient air. 

11 The most common plume system used in the U.S. is a modification of the free-air carbon12 dioxide/ozone enrichment (FACE) system (Hendrey and Kimball, 1994, 040397)(Hendrey et al., 

13 1999, 042641). Although originally designed to provide chamberless field facilities for studying the 

14 CO2 effects of climate change, FACE systems have been adapted to include the dispensing of O3 

(Karnosky et al., 1999, 035307). This method has been employed in Illinois (SoyFACE) to study 

16 soybeans (Morgan et al., 2004, 072764)(Rogers et al., 2004, 079201) and in Wisconsin (Aspen 

17 FACE) to study trembling aspen (Populus tremuloides), birch (Betula papyrifera) and maple (Acer 

18 saccharum)(Karnosky et al., 1999, 035307). Volk et al. (2003, 055568) also described a similar 

19 system for exposing grasslands that uses 7-m diameter plots. FACE systems discharge the pollutant 

gas (O3 and/or CO2) through orifices spaced along an annular ring (or torus) or at different heights on 

21 a ring of vertical pipes. Computer-controlled feedback from the monitoring of gas concentration 

22 regulates the feed rate of enriched air to the dispersion pipes. Feedback of wind speed and direction 

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

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

the arrays and their height (25-30 m) in some FACE systems requires large throughputs of enriched 

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

27 of enrichment treatments, although Hendrey et al. (1999, 042641) argued that the cost on an enriched 

28 volume basis is comparable to that of chamber systems. 

29 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, 

31 sizes, and orientations of the discharge orifices to avoid “hot-spots” resulting from the direct 

32 impingement of jets of pollutant-enriched air on plant foliage (Werner and Fabian, 2002, 053040). 

33 Because mixing is unassisted and completely dependent on wind turbulence and diffusion, local 

34 gradients are inevitable especially in large-scale systems. FACE systems have provisions for shutting 

down under low wind speed or calm conditions and for an experimental area that is usually defined 

36 within a generous border in order to strive for homogeneity of the exposure concentrations within the 

37 treatment area. They are also dependent upon continuous computer-controlled feedback of the O3 

March 2011	 9-21 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

5 

 

10

 

 

15 

 

 

20 

 

 

 

25 

 

 

 

 

30

 

1 concentrations in the mixed treated air and of the meteorological conditions. Plume and FACE
 

2 systems also are unable to reduce O3 levels below ambient in areas where O3 concentrations are
 

3 phytotoxic.
 

9.3.5. Ambient Gradients 

4 Ambient O3 gradients that occur in the U.S. hold potential for the examination of plant 

responses over multiple levels of exposure that are occurring. However, few such gradients can be 

6 found that meet the rigorous statistical requirements for comparable site characteristics such as soil 

7 type, temperature, rainfall, radiation, and aspect (Manning and Krupa, 1992, 044155); although with 

8 small plants, soil variability can be avoided by the use of plants in large pots. The use of soil 

9 monoliths transported to various locations along natural O3 gradients is another possible approach to 

overcome differences in soils; however, this approach is also limited to small plants. 

11 Studies in the 1970s used the natural gradients occurring in southern California to assess yield 

12 losses of alfalfa and tomato) (Oshima et al., 1976, 038475)(Oshima et al., 1977, 038938). A transect 

13 study of the impact of O3 on the growth of white clover and barley in the U.K. was confounded by 

14 differences in the concurrent gradients of SO2 and NO2 pollution (Ashmore et al., 1988, 037038). 

Studies of forest tree species in national parks in the eastern U.S. (Winner et al., 1989, 043403) 

16 revealed increasing gradients of O3 and visible foliar injury with increased elevation. 

17 Several studies have used the San Bernardino Mountains Gradient Study in southern 

18 California to study the effects of O3 and N deposition on forests dominated by ponderosa and Jeffrey 

19 pine (Arbaugh et al., 2003, 052925)(Miller and Elderman, 1977, 038488)(Grulke, 1999, 

052983)(Jones and Paine, 2006, 191301). However, it is difficult to separate the effects of N and O3 

21 in some instances in these studies (Arbaugh et al., 2003, 052925). An O3 gradient in Wisconsin has 

22 been used to study foliar injury in a series of trembling aspen clones (Populus tremuloides) differing 

23 in O3 sensitivity (Karnosky et al., 1999, 035307)(Maňkovská et al., 2005, 672965). 

24 More recently, some studies have been published that have used natural gradients to study a 

variety of endpoints and species. For example, Gregg et al. (2003, 046996) studied cottonwood 

26 saplings grown in an urban to rural gradient of O3 in the New Yor k City area. The secondary nature 

27 of the reactions of O3 formation and NOX titration reactions within the city center resulted in 

28 significantly higher cumulative O3 exposures in the rural sites. The results of this gradient study were 

29 similar to those of a parallel OTC study. Also, the U.S. forest service Forest Inventory and Analysis 

(FIA) program uses large-scale O3 exposure patterns across the continental U.S. to study occurrences 

31 of foliar injury due to O3 exposure (Smith et al., 2003, 044183)( Section 9.5.2.3). Finally, 

32 McLaughlin et al. (2007, 090348)(2007, 090347) used spatial and temporal O3 gradients to study 

33 forest growth and water use in the southern Appalachians. These studies found varying O3 exposures 

34 between years and between sites. 

March 2011 9-22 DRAFT – DO NOT CITE OR QUOTE 
 

1 

2

3

4

5

6

7 

8

9

10

11 

12

13

14 

15 

16

17 

18 

19

20 

21 

22

23 

24 

25

26 

27 

28 

29 

30

31

32 

33 

34

35

36

37

9.3.6. Comparative Studies 

All experimental approaches used to expose plants to O3 have shortcomings. The use of 

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

micrometeorology. In contrast, plume, FACE and gradient systems suffer from limited exposure 

levels, few replicates and an inability to reduce O3 levels below ambient in areas where O3 

concentrations are phytotoxic. 

While it is clear that chambers can alter some aspects of plant growth, it is important to 

establish whether or not these differences affect plant response to O3. As noted in the 1996 O3 AQCD 

(U.S. EPA, 1996, 080828), 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. (1986, 055530) 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, 043559) 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 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, 042660). 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 response of yield 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, 039329)(Heagle et al., 1983, 039372). 

A few recent studies have compared results of O3 experiments between OTCs, FACE 

experiments, and gradient studies. For example, a series of studies undertaken at Aspen FACE 

(Isebrands et al., 2000, 044174)(Isebrands et al., 2001, 036345) showed that O3-symptom expression 

was generally similar in OTCs, FACE, and ambient-O3 gradient sites, and supported the previously 

observed variation among trembling aspen clones using OTCs (Karnosky et al., 1999, 

035307)(Maňkovská et al., 2005, 672965). In the SoyFACE experiment in Illinois, soybean (Pioneer 

93B15 cultivar) yield loss data from a two-year study was published (Morgan et al., 2006, 079186). 

This cultivar is a recent selection and, like most modern cultivars, has been selected under an already 

high current O3 exposure. It was found to have average sensitivity to O3 compared to 22 other 

cultivars tested at SoyFACE. In this experiment, ambient hourly O3 concentrations were increased by 

approximately 20% and measured yields were decreased by 15% in 2002 as a result of the increased 

O3 exposure (Morgan et al., 2006, 079186). To compare these results to chamber studies, Morgan et 

al. (2006, 079186) calculated the expected yield loss from a linear relationship constructed from 

chamber data using 7-h seasonal averages (Ashmore, 2002, 672967). They calculated an 8% 

expected yield loss from the 2002 O3 exposure using that linear relationship. In another study, Gregg 

et al. (2003, 046996)(2006, 186961) found similar O3 effects on cottonwood sapling biomass growth 

March 2011 9-23 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

1 and physiology along an ambient O3 gradient in the New Yor k City area and a parallel OTC study. 

2 Additionally, Section 9.8.3 of this document presents comparisons of exposure-response in from
 

3 OTC studies in trees and crops with results from more recent FACE experiments.
 

9.4.	 Me c h a n is ms Go ve rn in g Veg e ta tio n Re s p o n s e to

Ozone 

9.4.1.	 Introd uc tion 

4 This section focuses on the effects of O3 stress on plants and their responses to that stress on 

5 the molecular, biochemical and physiological levels. First, the pathway of O3 uptake into the leaf and 

6 the initial chemical reactions occurring in the substomatal cavity and apoplast will be described 

7 (Section 9.4.2). Once O3 has entered the substomatal cavity and apoplast, it is thought that the cell 

8 must be able to sense the presence of O3 or its breakdown products in order to initiate the rapid 

9 changes in gene expression that have been measured in O3-treated plants. While an “O3 sensor” still 

10 remains elusive, much progress has been made in examining several different mechanisms that may 

11 contribute both to sensing the presence of O3 and its breakdown products, and also transducing a 

12 signal to the nucleus to initiate changes in gene transcription, which will be described in Section 

13 9.4.3.1. The next section focuses on changes in gene expression in response to O3 exposure, with 

14 particular emphasis on results from transcriptome and proteome analyses (Section 9.4.3.2). 

15 Subsequently, the role of phytohormones such as salicylic acid (SA), ethylene (ET), jasmonic acid 

16 (JA), and abscisic acid (ABA) and their interactions in both signal transduction and determining 

17 plant response to O3 is discussed in Section 9.4.3.3. After O3 uptake and sensing, plants can respond 

18 to the oxidative stress to minimize damage. These mechanisms of detoxification, with particular 

19 emphasis on antioxidant enzymes and metabolites, are reviewed in Section 9.4.4. The next section 

20 focuses on the effects of O3 on primary and secondary metabolism in plants, looking at 

21 photosynthesis, respiration and several secondary metabolites, some of which may also act as 

22 antioxidants and protect the plant from oxidative stress (Section 9.4.5). The last section focuses on 

23 the mechanisms underlying changes in stomatal function (Section 9.4.6). For many of these topics, 

24 information from the 2006 O3 AQCD (U.S. EPA, 2006, 088089) has been summarized, as this 

25 information is still valid and supported by more recent findings. For other topics, such as genomics 

26 and proteomics, which have arisen due to the availability of new technologies, the information is 

27 based solely on new publications with no reference to the 2006 O3 AQCD. 

28 As Section 9.4 focuses on mechanisms underlying vegetation response to O3, the conditions 

29 that are used to study these mechanisms are often artificial and do not necessarily reflect conditions 

30 that a plant may be exposed to in an agricultural setting or natural ecosystem. The goal of many of 

31 these studies is to elicit a plant response to O3 in a relatively short period of time and not always to 

32 simulate ambient O3 exposures. Therefore, plants are often exposed to unrealistically high O3 

33 concentrations for several hours or days (acute exposure), and only in a few cases to ambient or 

March 2011	 9-24 DRAFT – DO NOT CITE OR QUOTE 
 

1 

2

3

4 

5

6

7

8

9 

10

11

12

13 

14 

15 

16 

17 

18 

19 

20

21

22 

23 

24

25 

26 

27 

28 

29 

30 

31

32 

33 

34

35

slightly elevated O3 concentrations for longer time periods (chronic exposure). Additionally, the 

plant species utilized in these studies are often not agriculturally important or commonly found as 

part of natural ecosystems. Model organisms such as Arabidopsis thaliana are frequently used as 

they are easy to work with, and mutants or transgenic plants are easy to develop or have already been 

developed. Furthermore, the Arabidopsis genome has been sequenced, and much is known about the 

molecular basis of many biochemical and cellular processes. 

Many of the studies described in this section focus on changes in the expression of genes in 

O3-treated plants. However, changes in gene expression (i.e., either up- or down-regulation of gene 

expression) do not always translate into changes in protein quantity and/or activity, as there are many 

levels of post-transcriptional and post-translational modifications which impact protein quantity and 

activity. Frequently, these studies do not evaluate whether the observed changes in gene expression 

lead to changes at the protein level and, therefore, it is not always clear how relevant the changes in 

gene expression are in determining plant response to O3. 

The advent of new technologies, such as those employed in genomics and proteomics, has 

allowed for a more comprehensive analysis of the many molecular and biochemical mechanisms of 

plant response to O3 and how all these responses interact with or affect each other to determine the 

ultimate response of plants to a stressor such as O3. While the studies of transcriptome changes are 

very valuable, further work needs to be done to evaluate whether the transcriptome changes result in 

concomitant changes in the proteome. A few recent studies have evaluated proteome changes in 

response to O3, and those results are also discussed in Section 9.4.3.2. 

The most significant change in this section in relation to the 2006 O3 AQCD (U.S. EPA, 2006, 

088089) is the increased emphasis on the understanding of molecular mechanisms underlying plant 

responses to O3, as a significant number of the newer studies that were evaluated for this section 

focus on changes in gene expression in plants exposed to elevated O3. Conclusions from the 2006 O3 

AQCD have been supported by these new studies, and the advent of new technologies has allowed 

for a more comprehensive understanding of the mechanisms governing plant response to O3. 

In summary, the goal of many of these new studies reported on in this section was to increase 

knowledge of the mechanisms of plant response to O3 by using artificial exposure conditions and 

model organisms. This information adds to the understanding of the basic biology of plant response 

to oxidative stress in the absence of any other potential stressors. The results of these studies are 

important and valid, even though they may not always directly translate into effects observed in 

other plants under more realistic exposure conditions. They represent one step in the process of 

comprehensively understanding plant responses to oxidative stress, which then need to be followed 

up with additional experiments using other plant species exposed to O3 under more natural 

conditions. 

March 2011 9-25 DRAFT – DO NOT CITE OR QUOTE 
 

1

2

3

4

5 

6

7

8 

9

10

11 

12

13 

14

15 

16 

17

18

19 

20 

21 

9.4.2. Ozone Uptake into the Leaf 

AX9.2.3 of the 2006 O3 AQCD clearly described the process by which O3 enters plant leaves 

through open stomata (U.S. EPA, 2006, 088089). This information continues to be valid and is only 

summarized here. 

Ozone moves into the leaf interior by diffusing through open stomata, and environmental 

conditions which promote high rates of gas exchange will favor the uptake of the pollutant by the 

leaf. Factors that may limit uptake include boundary layer resistance and the size of the stomatal 

aperture (Figure 9-2) (U.S. EPA, 2006, 088089). Once inside the substomatal cavity, O3 is thought to 

rapidly react with the aqueous apoplast to form breakdown products known as reactive oxygen 

species (ROS), such as hydrogen peroxide (H2O2), superoxide (O2 ), hydroxyl radicals (HO·

) and 

peroxy radicals (HO2

·

) (Figure 9-3). Hydrogen peroxide is not only a toxic breakdown product of O3, 

but has been shown to function as a signaling molecule, which is activated in response to both biotic 

and abiotic stressors. The role of H2O2 in signaling was described in detail in the 2006 O3 AQCD 

(U.S. EPA, 2006, 088089). Additional organic molecules present in the apoplast or cell wall, such as 

those containing double bonds or sulfhydryls that are sensitive to oxidation, could also be converted 

to oxygenated molecules after interacting with O3 (Figure 9-4). These reactions are not only pH 

dependent but are also influenced by the presence of other molecules in the apoplast (U.S. EPA, 

2006, 088089). The 2006 O3 AQCD provided a comprehensive summary of what is known about the 

possible interactions of O3 with other biomolecules (U.S. EPA, 2006, 088089). It is in the apoplast 

that initial detoxification reactions by antioxidant metabolites and enzymes take place, and these 

initial reactions are critical to reduce concentrations of the oxidative breakdown products of O3; 

these reactions are described in more detail in Section 9.4.4 of this document. 

March 2011 9-26 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

Figure 9-2. 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 ozone enter through the

stomata on the lower (abaxial) leaf surface, while water vapor exits through the

stomata (transpiration). 

Figure 9-3. Possible reactions of ozone within water. ﷒(a) Ozone reacts at the double bonds to 

form carbonyl groups. (b) Under certain circumstances, peroxides are generated. 

March 2011 9-27 DRAFT – DO NOT CITE OR QUOTE 
 

 

Source: Adapted from Mudd (1996, 052795) 

Figure 9-4. The Crigee mechanism of ozone 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 ozone. 

March 2011 9-28 DRAFT – DO NOT CITE OR QUOTE 
 

1

2

3

4 

5

6 

7 

8 

9 

10 

11 

12

13

14

15 

16 

17

18 

19 

20

21

22 

23 

24 

25 

26 

27

28

29

30

31

32 

33 

34 

35 

9.4.3. Cellular to Sys temic Res pons es 

9.4.3.1. Ozone Sens ing a nd Signal Trans duc tio n 

New technologies allowing for large-scale analysis of oxidative stress-induced changes in 

gene expression have facilitated the study of signal transduction processes associated with the 

perception and integration of responses to the stress. Many of these studies have been conducted 

using Arabidopsis or tobacco plants, for which a variety of mutants are available and/or which can be 

easily genetically modified to generate either loss-of-function or over-expressing genotypes. Several 

comprehensive review articles provide an overview of what is known of O3-induced signal 

transduction processes and how they may help to explain differential sensitivity of plants to the 

pollutant (Kangasjarvi et al., 2005, 180341)(Ludwikow and Sadowski, 2008, 191426)(Baier et al., 

2005, 186866). Additionally, analysis of several studies of transcriptome changes has also allowed 

for the compilation of these data to determine an initial time-course for O3-induced activation of 

various signaling compounds (Kangasjarvi et al., 2005, 180341). 

A number of different mechanisms for plant sensing of O3 have been proposed; however, there 

is still much that is not known about this process. Ozone and/or the ROS breakdown products could 

be sensed by an apoplastic receptor protein, which would either be directly modified by O3 or ROS 

or which might sense O3 or ROS modifications in other apoplastic components (Kangasjarvi et al., 

2005, 180341)(Baier et al., 2005, 186866). Some of the earliest events that occur in plant response to 

O3 have been described in the guard cells of stomata. Reactive oxygen species were observed in the 

chloroplasts of guard cells in the O3 tolerant Col-0 Arabidopsis thaliana ecotype plants within 

5 minutes of plant exposure to 350 ppb O3 (Joo et al., 2005, 191307). Reactive oxygen species from 

the breakdown of O3 in the apoplast are believed to activate GTPases (G-proteins), which, in turn, 

activate several intracellular sources of ROS, including ROS derived from the chloroplasts. 

G-proteins are also believed to play a role in activating membrane-bound NADPH oxidases to 

produce ROS and, as a result, propagate the oxidative burst to neighboring cells (Joo et al., 2005, 

191307). Therefore, G-proteins are recognized as important molecules involved in plant responses to 

O3 and may play a role in perceiving ROS from the breakdown of O3 in the apoplast (Booker et al., 

2004, 020581)(Kangasjarvi et al., 2005, 180341). 

A change in the redox state of the plant and the oxidation of sensitive molecules in itself may 

represent a means of perception and signaling of oxidative stress in plants. Disulfide-thiol 

conversions in proteins and the redox state of the glutathione pool are likely to be important 

components of redox sensing and signal transduction (Foyer and Noctor, 2005, 191555)(Foyer and 

Noctor, 2005, 631198). 

Calcium (Ca2+) has also been implicated in the transduction of signals to the nucleus in 

response to oxidative stress. The influx of Ca2+ from the apoplast into the cell occurs early in plant 

response to O3, and it is thought to play a role in regulating the activity of protein kinases, which are 

discussed below (Hamel et al., 2005, 191214)(Baier et al., 2005, 186866). Calcium channel blockers 

March 2011 9-29 DRAFT – DO NOT CITE OR QUOTE 
 

1 

2

3 

4

5

6

7 

8 

9

10 

11 

12 

13

14 

15

16

17 

18

19 

20 

21

22 

23

24 

25 

26

27 

28 

29

30

31

32

33 

34

35 

36

37 

38 

39 

inhibited O3-induced activation of protein kinases in tobacco suspension cells exposed to 500 ppb O3 

for 10 minutes, indicating that the opening of Ca2+ channels is an important upstream signaling event 

or that the as yet unknown upstream process has a requirement for Ca2+ (Samuel et al., 2000, 

625706). 

Once ROS are generated by NADPH oxidase, signals are further transmitted to the nucleus to 

initiate changes in gene expression associated with plant defense responses. Integral to these signal 

cascades are mitogen-activated protein kinases (MAPK), which phosphorylate proteins and activate 

various cellular responses (Hamel et al., 2005, 191214). Mitogen-activated protein kinases are 

induced in several different plant species in response to O3 exposure, including tobacco (Samuel et 

al., 2005, 199316), Arabidopsis (Ludwikow et al., 2004, 595939), the shrub Phillyrea latifolia 

(Paolacci et al., 2007, 191422) and poplar (Hamel et al., 2005, 191214). In tobacco, the MAPK that 

is induced by plant exposure to 500 ppb O3 is a salicylic acid (SA) induced protein kinase (SIPK), 

which was found to positively regulate O3-induced ethylene (ET) production and negatively regulate 

SA accumulation (Samuel et al., 2005, 199316). In poplar suspension cells exposed to 500 ppb O3, 

O3-induced activation of two distinct MAP kinases was dependent on ROS formation and the 

activity of Ca2+ channels (Hamel et al., 2005, 191214). Arabidopsis mutants with suppressed activity 

of MAPK3 and MAPK6 (an ortholog of SIPK in tobacco) were more susceptible to exposure to 

500 ppb O3, as evidenced by tissue death resembling the hypersensitive response (Miles et al., 2005, 

191648). Similar results of increased susceptibility to 500 ppb O3 were found in tobacco lines either 

over-expressing SIPK or with suppressed SIPK function. The authors concluded that these tobacco 

lines were unable to cope with increased oxidative stress due to the alteration of the normal 

O3-induced MAPK signal transduction process (Samuel and Ellis, 2002, 625703). The Arabidopsis 

MAPK3 and MAPK6 are also differentially induced in Col-0 and the O3-sensitive Arabidopsis 

mutant radical induced cell death (rcd1) exposed to 250-300 ppb O3 for 6 hours, indicating a role for 

these kinases in plant response to oxidative stress (Overmyer et al., 2005, 191596). Similarly, 

components of the MAPK cascades were down-regulated in the sensitive Wassilewskija (Ws) 

Arabidopsis ecotype after exposure to 300 ppb O3 for 6 h, suggesting that the sensitive ecotypes may 

not activate these defense responses (Mahalingam et al., 2006, 191221). Expression of MAPK5 was 

also reduced in the Ws ecotype in response to chronic O3 exposures; Ws Arabidopsis plants were 

exposed to O3 concentrations 20-25% above ambient for 8-12 days in the SoyFACE site (Li et al., 

2006, 191332). 

The cysteine-rich RLKs (CRKs), which are part of the receptor-like/Pelle kinase (RLKs) 

group, are thought to be involved in the regulation of defense responses and cell death in Arabidopsis 

(Wrzaczek et al., 2010, 644190). CRKs were up-regulated by treatments that resulted in apoplastic 

ROS production, such as O3 (250 ppb O3 for 6 hours) and pathogen treatments, but either remained 

unchanged or were down-regulated by treatments resulting in ROS production in other cellular 

compartments, such as the mitochondria or chloroplasts. Although their function remains unclear, it 

has been postulated that a conserved cysteine motif in the CRKs could serve as a sensor for redox 

modifications in the cell resulting from ROS production (Wrzaczek et al., 2010, 644190). 

March 2011 9-30 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

5 

 

 

10

 

 

15 

 

 

 

20 

 

 

25 

 

 

30

 

 

35 

 

1 In conclusion, experimental evidence suggests that there may be several different mechanisms 

2 involved in sensing the presence of O3 or its breakdown products. These mechanisms may vary by 

3 species or developmental stage of the plant or may co-exist and be activated by different exposure 

4 conditions. Calcium and protein kinases are likely involved in the transduction of the initial signal to 

the nucleus and other cellular compartments to initiate the changes in gene transcription discussed in 

6 Section 9.4.3.2. 

9.4.3.2. Gene Expres s ion Changes in Res pons e to Ozone 

7 The advent of DNA microarray technology has allowed for the study of gene expression in 

8 cells on a large scale. Rather than assessing changes in gene expression of individual genes, DNA 

9 microarrays facilitate the evaluation of entire transcriptomes, providing a comprehensive picture of 

alterations in gene expression. In addition, these studies have provided more insight into the complex 

11 interactions between molecules and signal pathways, which result in the regulation of plant 

12 responses to stresses such as O3 (Ludwikow and Sadowski, 2008, 191426). Transcriptome analysis 

13 of O3-treated plants has been performed in several species, including Arabidopsis thaliana (Tosti et 

14 al., 2006, 191425)(Heidenreich et al., 2005, 191260)(Li et al., 2006, 191332)(Mahalingam et al., 

2005, 191693)(Tamaoki et al., 2003, 080053), pepper (Capsicum annuum) (Lee and Yun, 2006, 

16 191592), clover (Medicago truncatula) (Puckette et al., 2008, 191698), Phillyrea latifolia (Paolacci 

17 et al., 2007, 191422), and European beech (Fagus sylvatica) (Olbrich et al., 2005, 191697)(Olbrich 

18 et al., 2009, 596020)(Olbrich et al., 2010, 625424). In some cases, researchers compared 

19 transcriptomes of two or more cultivars, ecotypes or mutants that differed in their sensitivity to O3 

(Lee and Yun, 2006, 191592)(Tamaoki et al., 2003, 080053)(Li et al., 2006, 191332)(Puckette et al., 

21 2008, 191698)(Rizzo et al., 2007, 191447). Species, O3 exposure conditions (concentration, duration 

22 of exposure) and sampling times varied significantly in these studies. However, functional 

23 classification of the genes that were either up- or down-regulated by plant exposure to O3 exhibited 

24 common trends. Genes involved in plant defense, signaling and those associated with the synthesis 

of plant hormones and secondary metabolism were generally up-regulated, while those related to 

26 photosynthesis and general metabolism were generally down-regulated in O3-treated plants (Tosti et 

27 al., 2006, 191425)(Olbrich et al., 2005, 191697)(Tamaoki et al., 2003, 080053)(Lee and Yun, 2006, 

28 191592)(Puckette et al., 2008, 191698)(Li et al., 2006, 191332). 

29 Analysis of the transcriptome has been used to evaluate differences in gene expression 

between O3 sensitive and tolerant plants. In pepper, 67% of the 180 genes studied that were affected 

31 by O3 were differentially regulated in the sensitive and tolerant cultivars. At both 0 hours and 

32 48 hours after a 3-day exposure at 150 ppb, O3 responsive genes were either up- or down-regulated 

33 more markedly in the sensitive than in the tolerant cultivar (Lee and Yun, 2006, 191592). 

34 Transcriptome analysis also revealed differences in timing and magnitude of changes in gene 

expression between sensitive and tolerant clovers. Acute exposure (300 ppb O3 for 6 hours) led to the 

36 production of an oxidative burst in both clovers (Puckette et al., 2008, 191698). However, the 

37 sensitive Jemalong cultivar exhibited a sustained ROS burst and a concomitant down-regulation of 

March 2011 9-31 DRAFT – DO NOT CITE OR QUOTE 
 

1 

2

3 

4 

5 

6

7 

8

9

10 

11 

12 

13 

14

15 

16 

17 

18 

19

20

21 

22 

23 

24

25 

26

27 

28 

29 

30 

31

32

33 

34 

35 

36 

37 

38

39 

defense response genes at 12 hours after the onset of exposure, while the tolerant JE 154 accession 

showed much more rapid and large-scale transcriptome changes than the Jemalong cultivar (Puckette 

et al., 2008, 191698). 

Arabidopsis ecotypes WS and Col-0 were exposed to 1.2 × ambient O3 concentrations for 

8-12 days at the SoyFACE site (Li et al., 2006, 191332). The sensitive WS ecotype showed a far 

greater number of changes in gene expression in response to this low-level O3 exposure than the 

tolerant Col-0 ecotype. Exposure of the WS ecotype to 300 ppb O3 for 6 hours showed a rapid 

induction of genes leading to cell death, such as proteases, and down-regulation or inactivation of 

cell signaling genes, demonstrating an ineffective defense response in this O3 sensitive ecotype 

(Mahalingam et al., 2006, 191221). 

The temporal response of plants to O3 exposure was evaluated in the Arabidopsis Col-0 

ecotype during a 6-h exposure at 350 ppb O3 and for 6 hours after the exposure was completed. 

Results of this study, shown in Figure 9-5, indicate that genes associated with signal transduction and 

regulation of transcription were in the class of early up-regulated genes, while genes associated with 

redox homeostasis and defense/stress response were in the class of late up-regulated genes 

(Mahalingam et al., 2005, 191693). 

A few studies have been conducted to evaluate transcriptome changes in response to longer 

term chronic O3 exposures in woody plant species. Longer term exposures resulted in the upregulation of genes associated with secondary metabolites, including isoprenoids, polyamines and 

phenylpropanoids in 2-year-old seedlings of the Mediterranean shrub Phillyrea latifolia exposed to 

110 ppb O3 for 90 days (Paolacci et al., 2007, 191422). In 3-year-old European beech saplings 

exposed to O3 for 20 months, with monthly average twice ambient O3 concentrations ranging from 

11 to 80 ppb, O3-induced changes in gene transcription were similar to those observed for 

herbaceous species, including genes associated with plant stress response, primary metabolism, 

hormone synthesis, cell structure and premature senescence (Olbrich et al., 2009, 596020). In 

another study, the magnitude of these transcriptional changes was far greater in beech saplings than 

in adult trees exposed to the same O3 concentrations for the same time period, indicating that adult 

trees may be less responsive to this stressor than younger plants (Olbrich et al., 2010, 625424). 

These results have been substantiated by results from proteome analysis in rice, poplar, wheat, 

and soybean. Exposure of soybean to 120 ppb O3 for 12 h/day for 3 days in growth chambers 

resulted in decreases in the quantity of proteins associated with photosynthesis, while proteins 

involved with antioxidant defense and C metabolism increased (Ahsan et al., 2010, 644189). Young 

poplar plants exposed to 120 ppb O3 in a growth chamber for 35 days also showed significant 

changes in proteins involved in C metabolism (Bohler et al., 2007, 199408). Declines in enzymes 

associated with C fixation, the Calvin cycle and photosystem II were measured, while ascorbate 

peroxidase and enzymes associated with glucose catabolism increased in abundance. Two-week-old 

rice seedlings exposed to varying levels of O3 (4, 40, 80, 120 ppb) in a growth chamber for 9 days 

showed reductions in expression of proteins associated with photosynthesis and energy metabolism, 

and increases in some antioxidant and defense related proteins (Feng et al., 2008, 191626). A 

March 2011 9-32 DRAFT – DO NOT CITE OR QUOTE 
 

1 

2

3

4

5

6

7 

8 

9 

10 

11

12

13 

14 

15

16

17 

18

19 

20 

21

22 

23

24

25 

26

subsequent study of O3-treated rice seedlings (exposed to 200 ppb O3 for 24 h) focusing on the 

integration of transcriptomics and proteomics, supported and further enhanced these results (Cho et 

al., 2008, 603254). The authors found that of the 22,000 genes analyzed from the rice genome, 1,535 

were differentially regulated by O3. Those genes were functionally categorized as transcription 

factors, MAPK cascades, those encoding for enzymes involved in the synthesis of JA, ET, shikimate, 

tryptophan and lignin, and those involved in glycolysis, citric acid cycle, oxidative respiration and 

photosynthesis. The authors determined that the proteome and metabolome analysis supported the 

results of the transcriptome changes described above (Cho et al., 2008, 603254). This type of study, 

which ties together results from changes in gene expression, protein quantity and activity, and 

metabolite levels, provides the most complete picture of the molecular and biochemical changes 

occurring in plants exposed to a stressor such as O3. Sarkar et al. (2010, 657214) compared two 

cultivars of wheat grown in OTCs at several O3 concentrations, including filtered air, ambient O3 

(mean concentration 47 ppb), ambient + 10 ppb and ambient + 20 ppb for 5 h/day for 50 days. 

Declines in the rate of photosynthesis and stomatal conductance were related to decreases in proteins 

involved in C fixation and electron transport, and evidence of increased proteolysis of photosynthetic 

proteins such as the large subunit of ribulose-1,6-bisphosphate carboxylase/oxygenase (Rubisco). 

Enzymes that take part in energy metabolism, such as ATP synthesis, were also down-regulated, 

while defense/stress related proteins were induced with O3 treatment. In comparing the two wheat 

cultivars, Sarkar et al. (2010, 657214) found that while the qualitative changes in protein expression 

between the two cultivars was similar, the magnitude of these changes differed between the sensitive 

and tolerant wheat cultivars. 

All of these studies describe common trends for changes in gene and protein expression which 

occur in a variety of plant species in response to O3. While genes associated with C assimilation and 

general metabolism are down-regulated, genes associated with signaling, catabolism, and defense are 

up-regulated. The magnitude of these changes in gene and protein expression appears to be related to 

plant sensitivity or tolerance to O3. 

March 2011 9-33 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

 

 

Source: Used with permission from Springer, Mahalingam et al. (2005, 191693). 

Figure 9-5 Composite diagram of major themes in the temporal evolution of the genetic

response to ozone stress. ﷒(A) Temporal profile of the oxidative stress response to

ozone. The biphasic ozone-induced oxidative burst is represented in black, with the

ROS control measurements shown as a broken line. Average transcript profiles are 

shown for early up-regulated genes (yellow, peaks at 0.5-1 hours), and the 3 hours

(blue), 4.5 hours (red) and 9-12 hours (green) late up-regulated genes and for the 

down-regulated genes coding for photosynthesis proteins (brown). (B)

Diagrammatic representation of redox regulation of the oxidative stress response. 

9.4.3.3. Role of P hytoh ormo ne s in Pla n t Re s p ons e to Ozone 

1 Many studies of O3 effects on plants have analyzed the importance of plant hormones such as 

2 SA, ET and JA in determining plant response to O3; some of the roles of these hormones were 

3 described in the 2006 O3 AQCD (U.S. EPA, 2006, 088089). Transcriptome analysis and the use of a 

4 variety of mutants have allowed for further elucidation of the complex interactions between SA, ET, 

5 JA and the role of abscisic acid (ABA) in mediating plant response to O3 (Ludwikow and Sadowski, 

6 2008, 191426). In addition to their roles in signaling pathways, phytohormones also appear to 

March 2011 9-34 DRAFT – DO NOT CITE OR QUOTE 
 

1 

2 

3 

4

5 

6 

7

8

9 

10

11 

12 

13 

14 

15

16 

17 

18 

19 

20

21 

22 

23

24 

25

26

27 

28 

29 

30 

31

32

33

34

35 

36 

37 

38

regulate, and be regulated by, the MAPK signaling cascades described previously. Most evidence 

suggests that while ET and SA are needed to develop O3-induced leaf lesions, JA acts 

antagonistically to SA and ET to limit the lesions (Figure 9-6) (Kangasjarvi et al., 2005, 180341). 

The rapid production of ET in response to O3 has been described in many plant species and 

has been further characterized through the use of a variety of mutants that either over-produce or are 

insensitive to ET. Production of stress ET in O3-treated plants, which is thought to be a wounding 

response, was found to be correlated to the degree of injury development in leaves (U.S. EPA, 2006, 

088089). More recent studies have supported these conclusions and have also focused on the 

interactions occurring between several oxidative-stress induced phytohormones. Yoshida et al. 

(2009, 191385) determined that ET likely amplifies the oxidative signal generated by ROS, thereby 

promoting lesion formation. By analyzing the O3-induced transcriptome of several Arabidopsis 

mutants of the Col-0 ecotype, Tamaoki et al. (2003, 080053) determined that at 12 hours after 

initiating the O3 exposure (200 ppb for 12 hours), the ET and JA signaling pathways were the main 

pathways used to activate plant defense responses, with a lesser role for SA. The authors also 

demonstrated that low levels of ET production could stimulate the expression of defense genes, 

rather than promoting cell death when ET production is high. Tosti et al. (2006, 191425) supported 

these findings by showing that O3 not only activates the biosynthetic pathways of ET, JA and SA, but 

also increases the expression of genes related to the signal transduction pathways of these 

phytohormones in O3-treated Arabidopsis plants (300 ppb O3 for 6 hours). Conversely, in the O3 

sensitive Ws ecotype, its sensitivity may, in part, be due to intrinsically high ET levels leading to SA 

accumulation, and the high ET and SA may act to repress JA-associated genes, which would serve to 

inhibit the spread of lesions (Mahalingam et al., 2006, 191221). Ogawa et al. (2005, 191653) found 

that an O3-induced increase in SA leads to the formation of leaf lesions in tobacco plants exposed to 

200 ppb O3 for 6 hours. Furthermore, several genes encoding for enzymes in the biosynthetic 

pathway of SA were suppressed in transgenic tobacco plants with reduced levels of O3-induced ET 

production, suggesting that SA levels are controlled by ET in the presence of O3. 

Exposure of the Arabidopsis mutant rcd1 to acute doses of O3 (250 ppb O3 for 8 h/day for 

3 days) resulted in programmed cell death (PCD) and the formation of leaf lesions. Salicylic acid 

was required for the initiation of PCD in these mutants, and increased ET synthesis resulted in 

amplification of cell death, while JA was thought to contain the spreading of lesions (Overmyer et 

al., 2000, 036361). In addition, the O3-treated rcd1 mutants showed higher SA and JA accumulation 

as compared to the Col-0 ecotype. High SA levels could be involved in amplifying cell death in these 

mutants. Jasmonic acid, which is thought to accumulate as a direct result of cell death, may represent 

an autocatalytic mechanism for limiting cell death, by which the increased cell death observed in the 

rcd1 mutant would determine the high amount of JA needed to contain lesion formation (Overmyer 

et al., 2005, 191596). In cotton plants exposed to a range of O3 concentrations (0-120 ppb) and 

methyl jasmonate (MeJA), Grantz et al. (2010, 625419) determined that endogenous applications of 

MeJA did not protect plants from chronic O3 exposure. 

March 2011 9-35 DRAFT – DO NOT CITE OR QUOTE 
 

1

2 

3 

4 

5

6 

7

8 

9 

10

11 

12

13 

14

15 

16 

17

18 

19

20 

21

22

23 

24

25 

26

27

28 

29 

30

There are distinct patterns of gene expression when evaluating early and later O3 responsive 

genes (Mahalingam et al., 2005, 191693). While ET and JA pathways were stimulated rapidly after 

Arabidopsis Col-0 exposure to O3, at 48 hours postexposure, JA and ET synthesis in O3-treated 

Arabidopsis were reduced, while the synthesis of SA was stimulated (D'Haese et al., 2006, 191448). 

Abscisic acid has been investigated for its role in regulating stomatal aperture and also for its 

contribution to signaling pathways in the plant. The role of ABA and the interaction between ABA 

and H2O2 in O3-induced stomatal closure was described in the 2006 O3 AQCD (U.S. EPA, 2006, 

088089). More recently, it was determined that synthesis of ABA was induced in O3-treated 

Arabidopsis plants (250-350 ppb O3 for 6 hours), with a more pronounced induction in several O3 

sensitive rcd mutants as compared to the wildtype Col-0 (Overmyer et al., 2008, 191283). Ludwikow 

et al. (2009, 199440) used Arabidopsis ABI1td mutants, in which a key negative regulator of ABA 

action (abscisic acid insensitive1 protein phosphatase 2C) has been knocked out, to examine O3 

responsive genes in this mutant compared to the Arabidopsis Col-0. Results of this study indicate a 

role for ABI1 in negatively regulating the synthesis of both ABA and ET in O3-treated plants 

(350 ppb O3 for 9 hours). Additionally, ABI1 may stimulate JA-related gene expression, providing 

evidence for an antagonistic interaction between ABA and JA signaling pathways (Ludwikow et al., 

2009, 199440). 

Nitric oxide (NO) has also been shown to play a role in regulating O3-induced gene expression 

in plants. However, little is known to date about NO and its role in the complex interactions of 

molecules in response to O3. Exposure of tobacco to O3 (150 ppb for 5 hours) stimulated NO and 

NO-dependent ET production, while NO production itself did not depend on the presence of ET 

(Ederli et al., 2006, 191479). Analysis of O3-treated Arabidopsis indicated the possibility of a dual 

role for NO in the initiation of cell death and later lesion containment (Ahlfors et al., 2009, 191533). 

While much work remains to be done to better elucidate how plants sense O3 and how signals 

are communicated to the nucleus to generate plant responses to oxidative stress, it is clear that the 

mechanism for O3 sensing and signal transduction is very complex. Many of the phytohormones and 

other signaling molecules thought to be involved in these processes are interactive and depend upon 

a variety of other factors, which could be either internal or external to the plant. This results in a 

highly dynamic and complex system, capable of generating a variety of plant responses to oxidative 

stress. 

March 2011 9-36 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

 

 

 

Source: Used with permission from Blackwell Publishing Ltd., Kangasjarvi et al. (2005, 180341). 

Figure 9-6. The oxidative cell death cycle. ﷒Ozone-derived radicals induce endogenous ROS

production (1) which results in salicylic acid (SA) accumulation and programmed

cell death; (2) Cell death triggers ethylene (ET) production, which is required for the

continuing ROS production responsible for the propagation of cell death; (3)

Jasmonates counteract the progression of the cycle by antagonizing the cell death 

promoting function of SA and ET; (4) Abscisic acid (ABA) antagonizes ET function in

many situations and might also have this role in ozone-induced cell death; (5) 

Mutually antagonistic interactions between ET, SA and jasmonic acid (JA) are

indicated with red bars. 

9.4.4. De toxific a tio n 

9.4.4.1. Overview o f Ozone -In duc e d De fe ns e Me c ha nis ms 

1 Plants are exposed to an oxidizing environment on a continual basis, and many reactions that 

2 are part of the basic metabolic processes, such as photosynthesis and respiration, generate ROS. As a 

3 result, there is an extensive and complex mechanism in place to detoxify these oxidizing radicals, 

4 including both enzymes and metabolites, which are located in several locations in the cell and also in 

5 the apoplast of the cell. As O3 enters the leaf through open stomata, the first point of contact of O3 

6 with the plant is thought to be in the apoplast, where it breaks down to form oxidizing radicals such 

7 as H2O2, O2

-

, HO· and HO2. Another source of oxidizing radicals is an oxidative burst, generated by 

8 a membrane-bound NADPH oxidase enzyme, which is thought to be part of the plant’s defense 

9 system against pathogens (Schraudner et al., 1998, 056358). Antioxidant metabolites and enzymes 

10 located in the apoplast are thought to form a first line of defense by detoxifying O3 and/or the ROS 

11 that are formed as breakdown products of O3 (Section 9.4.2.). However, even with the presence of 

12 several antioxidants, including ascorbate, the redox buffering capacity of the apoplast is far less than 

March 2011 9-37 DRAFT – DO NOT CITE OR QUOTE 
 

 

5 

 

 

10 

 

15

 

 

 

20

 

 

25

 

 

30 

 

 

 

35 

 

1 that of the cytoplasm, as it lacks the regeneration systems necessary to retain a reduced pool of 

2 antioxidants (Foyer and Noctor, 2005, 631198). 

3 Redox homeostasis is regulated by the presence of a pool of antioxidants, which are typically 

4 found in a reduced state and detoxify ROS produced by oxidases or electron transport components. 

As ROS increase due to environmental stress such as O3, the antioxidant pool will no longer be able 

6 to maintain its reduced state (Foyer and Noctor, 2005, 631198). As such, not only the quantity and 

7 types of antioxidant enzymes and metabolites present, but also the cellular ability to regenerate those 

8 antioxidants are important considerations in mechanisms of plant tolerance to oxidative stress 

9 (Dizengremel et al., 2008, 191587). Molecules such as glutathione (GSH), thioredoxins and NADPH 

play very important roles in this regeneration process; additionally, alterations in C metabolism will 

11 be necessary to supply the needed reducing power for antioxidant regeneration (Dizengremel et al., 

12 2008, 191587). Increases in the activity of catabolic pathways allow the cell to generate more 

13 NADPH (Section 9.4.5). 

9.4.4.2. Role of An tioxida n ts in Pla nt De fe ns e Re s po n s e s 

14	 Ascorbate has been the focus of many different studies as an antioxidant metabolite that 

protects plants from exposure to O3. It is found in several cellular locations, including the 

16 chloroplast, the cytosol and the apoplast (Noctor and Foyer, 1998, 657213). Ascorbate is synthesized 

17 in the cell and transported to the apoplast. Apoplastic ascorbate can be oxidized to dehydroascorbate 

18 (DHA) with exposure to O3 and is then transported back to the cytoplasm. Here, DHA is reduced to 

19 ascorbate by the enzyme dehydroascorbate reductase (DHAR) and reduced GSH, which is part of the 

ascorbate-glutathione cycle (Noctor and Foyer, 1998, 657213). Many studies have focused on 

21 evaluating whether ascorbate is the determining factor in differential sensitivity of plants to O3. 

22 Cheng et al. (2007, 191499) exposed two soybean cultivars to elevated O3 (77 ppb) and filtered air 

23 for 7 h/day for 6 days. The differences in sensitivity between the two cultivars could not be 

24 explained by differential O3 uptake or by the fraction of reduced ascorbate present in the apoplast. 

However, total antioxidant capacity of the apoplast was twofold higher in the tolerant Essex cultivar 

26 as compared to the sensitive Forrest cultivar, indicating that there may be other compounds in the 

27 leaf apoplast that scavenge ROS. D'Haese et al. (2005, 191551) exposed the NC-S (sensitive) and 

28 NC-R (resistant) clones of white clover (Trifolium repens) to 60 ppb O3 for 7 h/day for 5 days in 

29 environmental chambers. Surprisingly, the NC-S clone had a higher constitutive concentration of 

apoplastic ascorbate with a higher redox status than the NC-R clone. However, the redox status of 

31 symplastic GSH was higher in NC-R, even though the concentration of GSH was not higher than in 

32 NC-S. In addition, total symplastic antioxidative capacity was not a determining factor in differential 

33 sensitivity between these two clones. Severino et al. (2007, 199293) also examined the role of 

34 antioxidants in the differential sensitivity of the two white clover clones by growing them in the field 

for a growing season and then exposing them to elevated O3 (100 ppb for 8 h/day for 10 days) in 

36 OTC at the end of the field season. The NC-R clone had greater quantities of total ascorbate and total 

37 antioxidants than the NC-S clone at the end of the experiment. While the second study indicates a 

March 2011	 9-38 DRAFT – DO NOT CITE OR QUOTE 
 

1 

2 

3 

4 

5 

6 

7 

8 

9 

10 

11 

12 

13

14

15 

16

17

18 

19

20 

21

22 

23 

24 

25 

26

27

28 

29

30 

31 

32

33 

34 

35 

36

37 

38

39 

possible relationship between O3 tolerance and ascorbate levels, the first study indicates that there 

are other factors besides ascorbate that determine plant tolerance to O3. In snap bean, plants of the O3 

tolerant Provider cultivar had greater total ascorbate and more ascorbate in the apoplast than the 

sensitive S156 cultivar after exposure to 71 ppb O3 for 10 days in OTC (Burkey et al., 2003, 

630251). While most of the apoplastic ascorbate was in the oxidized form, the ratio of reduced 

ascorbate to total ascorbate was higher in Provider than S156, indicating that Provider is better able 

to maintain this ratio to maximize plant protection from oxidative stress. 

While the quantities of antioxidant metabolites such as ascorbate are an important indicator of 

plant tolerance to O3, the ability of the plant to recycle oxidized ascorbate efficiently plays a large 

role in determining the plant’s ability to deal with a sustained exposure to oxidative stress. Tobacco 

plants over-expressing DHAR were better protected from exposure to either chronic (100 ppb O3 

4 h/day for 30 days) or acute (200 ppb O3 for 2 hours) conditions than control plants and those with 

reduced expression of DHAR. The DHAR over-expressing plants exhibited an increase in guard cell 

ascorbic acid, leading to a decrease in stomatal responsiveness to O3 and an increase in stomatal 

conductance and O3 uptake. Despite this, the presence of higher levels of ascorbic acid led to a lower 

oxidative load and a higher level of photosynthetic activity in the DHAR over-expressing plants 

(Chen and Gallie, 2005, 191465). A subsequent study with tobacco plants over-expressing DHAR 

confirmed some of these results. Levels of ascorbic acid were higher in the transgenic tobacco 

plants, and they exhibited greater tolerance to O3 exposure (200 ppb O3) as demonstrated by higher 

photosynthetic rates in the transgenic plants as compared to the control plants (Eltayeb et al., 2006, 

191377). Over-expression of monodehydroascorbate reductase (MDAR) in tobacco plants also 

showed enhanced stress tolerance in response to O3 exposure (200 ppb O3), with higher rates of 

photosynthesis and higher levels of reduced ascorbic acid as compared to controls (Eltayeb et al., 

2007, 191553). Results of these studies show the importance of ascorbic acid as a detoxification 

mechanism, but more importantly emphasize that the recycling of oxidized ascorbate and 

maintenance of a reduced pool of ascorbate is critical in determining plant tolerance to oxidative 

stress. 

The roles of other antioxidant metabolites and enzymes, including GSH, catalase (CAT), and 

superoxide dismutase (SOD), were comprehensively reviewed in the 2006 O3 AQCD (U.S. EPA, 

2006, 088089). Additional studies have supported the findings reported in that document. Superoxide 

dismutase (SOD) and peroxidase (POD) activities were measured in both the tolerant Bel B and 

sensitive Bel W3 tobacco cultivars exposed to ambient O3 concentrations for 2 weeks 3 times 

throughout a growing season (Borowiak et al., 2009, 191247). Tulip poplar (Liriodendron tulipifera) 

trees exposed to increasing O3 concentrations (from 100 to 300 ppb O3 during a 2-week period) 

showed increases in activities of SOD, ascorbate peroxidase (APX), glutathione reductase (GR), 

MDAR, DHAR, CAT and POD in the 2-week period, although individual enzyme activities 

increased at different times during the 2-week period (Ryang et al., 2009, 191267). 

Longer, chronic O3 exposures in trees revealed some distinctive patterns of increases in SOD 

and APX activity that were measured in Quercus mongolica after 45 days of plant exposure to 

March 2011 9-39 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

1 80 ppb O3, which were followed by declines in the activities and quantities of these enzymes after 

2 75 days of exposure (Yan et al., 2010, 628514). Similarly, activities of SOD, APX, DHAR, MDAR, 

3 and GR increased in Gingko biloba trees during the first 50 days of exposure to 80 ppb O3, followed 

4 by decreases in activity below control values after 50 days of exposure (He et al., 2006, 199430). 

5 Soybean plants exposed to 70 or 100 ppb O3 for 4 h/day over the course of a growing season showed 

6 elevated POD activity and a decrease in CAT activity at 40 and 60 days after germination (Singh et 

7 al., 2010, 386275). 

8 Antioxidant enzymes appear to increase in quantity in O3-treated plants as a defense 

9 mechanism against oxidative stress. However, it appears that plants cannot maintain these elevated 

10 levels of antioxidant enzymes for an extended period of time, likely due to the high metabolic costs 

11 involved. Therefore, plants exposed to chronic O3 conditions may show more symptoms of exposure 

12 as defensive mechanisms are down-regulated over time. 

9.4.5. Effe cts on P rimary a n d S ec on d ary Meta bolis m 

9.4.5.1. Lig ht a nd Da rk Re a c tio ns o f Ph o tos yn the s is 

13 Declines in the rate of photosynthesis and stomatal conductance in O3-treated plants have been 

14 documented for many different plant species (U.S. EPA, 2006, 088089)(Booker et al., 2009, 191569) 

15 (Wittig et al., 2007, 191695). Much of the literature regarding O3 effects on photosynthesis has 

16 focused on C assimilation; however, more recently, attention has also been focused on deleterious 

17 effects on the light reactions. Chlorophyll fluorescence provides a useful measure of changes to the 

18 photosynthetic process from exposure to oxidative stress. Decreases in the Fv/Fm ratio (a measure of 

19 the maximum efficiency of Photosystem II) in dark adapted leaves indicate a decline in the 

20 efficiency of the PSII photosystems and a concomitant increase in non-photochemical quenching 

21 (Guidi and Degl'lnnocenti, 2008, 191571)(Scebba et al., 2006, 191219). Changes in these parameters 

22 have been correlated to differential sensitivity of plants to the pollutant. In a study to evaluate the 

23 response of 4 maple species to O3 (exposed to an 8-h avg of 51 ppb for ambient and 79 ppb for 

24 elevated treatment in OTC), the 2 species which were most sensitive based on visible injury and 

25 declines in CO2 assimilation also showed the greatest decreases in Fv/Fm in symptomatic leaves. In 

26 asymptomatic leaves, CO2 assimilation decreased significantly but there was no significant decline 

27 in Fv/Fm (Calatayud et al., 2007, 191411). Degl 'Innocenti et al. (2007, 191350) measured 

28 significant decreases in Fv/Fm in young and symptomatic leaves of a resistant tomato genotype (line 

29 93.1033/1) in response to O3 exposure (150 ppb O3 for 3 hours in a growth chamber), but only minor 

30 decreases in asymptomatic leaves with no associated changes in net photosynthetic rate. In the O3 

31 sensitive tomato cultivar Cuor Di Bue, the Fv/Fm ratio did not change, while the photosynthetic rate 

32 declined significantly in asymptomatic leaves (Degl'Innocenti et al., 2007, 191350). In two soybean 

33 cultivars, Fv/Fm also declined significantly with plant exposure to O3 (Singh et al., 2009, 199427). It 

March 2011 9-40 DRAFT – DO NOT CITE OR QUOTE 
 

 

5

10 

 

 

 

15 

 

20

 

 

25 

 

30 

 

 

 

35 

1 appears that in asymptomatic leaves, photoinhibition, as indicated by a decrease in Fv/Fm, is not the 

2 main reason for a decline in photosynthesis. 

3 An evaluation of photosynthetic parameters of two white clover (Trifolium repens cv. Regal) 

4 clones that differ in their O3 sensitivity revealed that O3 (40-110 ppb O3 for 7 h/day for 5 days) 

increased the coefficient of non-photochemical quenching (qNP) in both the resistant (NC-R) and 

6 sensitive (NC-S) clones, however qNP was significantly lower for the sensitive clone (Crous et al., 

7 2006, 199321). Sensitive Acer clones had a lower coefficient of non-photochemical quenching, 

8 while exposure to O3 increased qNP in both sensitive and tolerant clones (Calatayud et al., 2007, 

9 191411). While exposure to O3 also increased qNP in tomato, there were no differences in the 

coefficient of photochemical quenching between cultivars thought to be differentially sensitive to O3. 

11 (Degl'Innocenti et al., 2007, 191350). Higher qNP as a result of exposure to O3 indicates a reduction 

12 in the proportion of absorbed light energy being used to drive photochemistry. A lower coefficient of 

13 non-photochemical quenching in O3 sensitive plants could indicate increased vulnerability to ROS 

14 generated during exposure to oxidative stress (Crous et al., 2006, 199321). 

Several measures of the light reactions of photosynthesis are sensitive to exposure to O3, 

16 however, photosynthetic C assimilation is generally considered to be more affected by pollutant 

17 exposure, resulting in an overall decline in photosynthesis (Heath, 2008, 195632)(Fiscus et al., 2005, 

18 079155)(Guidi and Degl'lnnocenti, 2008, 191571). Loss of C assimilation capacity has been shown 

19 to result primarily from declines in the quantity of Rubisco (Calatayud et al., 2007, 191411)(Singh et 

al., 2009, 199427). Experimental evidence suggests that both decreases in Rubisco synthesis and 

21 enhanced degradation of the protein contribute to the measured reduction in its quantity (U.S. EPA, 

22 2006, 088089). Reduced C assimilation has been linked to reductions in biomass and yield (Keutgen 

23 et al., 2005, 191295)(He et al., 2007, 199789)(Novak et al., 2007, 194630)(Gregg et al., 2006, 

24 186961), (Wang et al., 2009, 199303). 

Most of the research on O3 effects on photosynthesis has focused on C3 (Calvin cycle) plants 

26 because C4 (Hatch-Slack) plants have lower stomatal conductance and therefore assumed to be less 

27 sensitive to O3 stress. However, a few studies have been conducted to evaluate the effects of O3 on 

28 C4 photosynthesis. In older maize leaves, Leitao et al. (2007, 191456)(2007, 191263) found that the 

29 activity, quantity and transcript levels of both Rubisco and phosphoenolpyruvate carboxylase (PEPc) 

decreased as a function of rising O3 concentration. In younger maize leaves, the quantity, activity, 

31 and transcript levels of the carboxylases were either increased or unaffected in plants exposed to 

32 40 ppb O3 for 7 h/day for 28-33 days, but decreased at 80 ppb (Leitao et al., 2007, 191263)(Leitao et 

33 al., 2007, 199379). 

9.4.5.2. Re s pira tio n a n d Da rk Re s pira tion 

34	 While much research emphasis regarding O3 effects on plants has focused on the negative 

impacts on C assimilation, other studies have measured impacts on catabolic pathways such as 

36 respiration and photorespiration. Generally, respiration has been found to increase in plants exposed 

37 to O3. Bean plants exposed to ambient (average 12-h mean 43 ppb) and twice ambient (average 12-h 

March 2011	 9-41 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

5 

 

 

 

 

10 

 

 

15

 

 

20 

 

 

25

 

30

 

 

 

35 

1 mean 80 ppb) O3 showed increases in respiration. When mathematically partitioned, the maintenance 

2 coefficient of respiration was significantly increased by O3 exposure, while the growth coefficient of 

3 respiration was not affected (Amthor, 1988, 041870). Loblolly pines were exposed to ambient (12-h 

4 daily mean was 45 ppb) and twice ambient (12 hours daily mean was 86 ppb) O3 for 12 h/day for 

approximately seven months per year for 3 and 4 years. While photosynthetic activity declined with 

6 the age of the needles and increasing O3 concentration, enzymes associated with respiration showed 

7 higher levels of activity with increasing O3 concentration (Dizengremel et al., 1994, 187217). In 

8 their review on the role of metabolic changes in plant redox status after O3 exposure, Dizengremel et 

9 al. (2009, 199424) summarized multiple studies in which several different tree species were exposed 

to O3 concentrations ranging from ambient to 200 ppb O3 for at least several weeks. In all cases, the 

11 activity of enzymes, including phosphofructokinase, pyruvate kinase and fumarase, which are part of 

12 several catabolic pathways, were increased in response to O3 exposure. 

13 Photorespiration is a light-stimulated process which consumes O2 and releases CO2. While it 

14 has been regarded as a wasteful process, more recent evidence suggests that it may play a role in 

photoprotection during photosynthesis (Bagard et al., 2008, 191593). The few studies that have been 

16 conducted on O3 effects on photorespiration suggest that rates of photorespiration decline 

17 concomitantly with rates of photosynthesis. Soybean plants were exposed to ambient (daily averages 

18 43-58 ppb) and 1.5 ambient O3 (daily averages 63-83 ppb) O3 in OTCs for 12 h/day for 4 months. 

19 Rates of photosynthesis and photorespiration and photorespiratory enzyme activity declined only at 

the end of the growing season and did not appear to be very sensitive to O3 exposure (Booker et al., 

21 1997, 026425). Young hybrid poplars exposed to 120 ppb O3 for 13 h/day for 35 days in phytotron 

22 chambers showed that effects on photorespiration and photosynthesis were dependent upon the 

23 developmental stage of the leaf. While young leaves were not impacted, reductions in photosynthesis 

24 and photorespiration were measured in fully expanded leaves (Bagard et al., 2008, 191593). 

9.4.5.3. Se c on da ry Me ta bolis m 

Transcriptome analysis of Arabidopsis plants has revealed modulation of several genes 

26 involved in plant secondary metabolism (Ludwikow and Sadowski, 2008, 191426). Phenylalanine 

27 ammonia lyase (PAL) has been the focus of many studies involving plant responses to O3 due to its 

28 importance in linking the phenylpropanoid pathway of plant secondary metabolism to primary 

29 metabolism in the form of the shikimate pathway. Genes encoding several enzymes of the 

phenylpropanoid pathway and lignin biosynthesis were up-regulated in transcriptome analysis of 

31 Arabidopsis plants (Col-0) exposed to 350 ppb O3 for 6 hours, while 2 genes involved in flavonoid 

32 biosynthesis were down-regulated (Ludwikow et al., 2004, 595939). Exposure of Arabidopsis 

33 (Col-0) to lower O3 concentrations (150 ppb for 8 h/day for 2 days) resulted in the induction of 11 

34 transcripts involved in flavonoid synthesis. In their exposure of 2-year-old Mediterranean shrub 

Phillyrea latifolia to 110 ppb O3 for 90 days, Paolacci et al. (2007, 191422) identified four clones 

36 that were up-regulated and corresponded to genes involved in the synthesis of secondary 

37 metabolites, such as isoprenoids, polyamines and phenylpropanoids. Up-regulation of genes 

March 2011 9-42 DRAFT – DO NOT CITE OR QUOTE 
 

1 

2 

3 

4

5 

6 

7

8

9 

10 

11 

12 

13 

14 

15

16 

17 

18 

19 

20 

21 

22

23 

24 

25

26 

27

28 

29 

30 

31 

32 

33 

34 

35 

36 

37 

38

involved in isoprene synthesis was also observed in Medicago trunculata exposed to 300 ppb O3 for 

6 hours, while genes encoding enzymes of the flavonoid synthesis pathway were either up- or downregulated (Puckette et al., 2008, 191698). Exposure of red clover to 1.5 × ambient O3 (average 

concentrations of 32.4 ppb) for up to 9 weeks in an open field exposure system resulted in increases 

in leaf total phenolic content. However, the types of phenolics that were increased in response to O3 

exposure differed depending upon the developmental stage of the plant. While almost all of the 31 

different phenolic compounds measured increased in quantity initially during the exposure, after 

3 weeks the quantity of isoflavones decreased while other phenolics increased (Saviranta et al., 2010, 

102177). Exposure of beech saplings to ambient and 2 × ambient O3 concentrations over 2 growing 

seasons resulted in the induction of several enzymes which contribute to lignin formation, while 

enzymes involved in flavonoid biosynthesis were down-regulated (Olbrich et al., 2009, 596020). 

Exposure of tobacco Bel W3 to 160 ppb O3 for 5 hours showed up-regulation of almost all genes 

encoding for enzymes which are part of the prechorismate pathway (Janzik et al., 2005, 191581). 

Isoprenoids can serve as antioxidant compounds in plants exposed to oxidative stress (Paolacci et al., 

2007, 191422). 

The prechorismate pathway is the pathway leading to the formation of chorismate, a precursor 

to the formation of the aromatic amino acids tryptophan, tyrosine and phenylalanine. These amino 

acids are precursors for the formation of many secondary aromatic compounds, and, therefore, the 

prechorismate pathway represents a branch-point in the regulation of metabolites into either primary 

or secondary metabolism (Janzik et al., 2005, 191581). Exposure of the O3 sensitive Bel W3 tobacco 

cultivar at 160 ppb for 5 hours showed an increase in transcript levels of most of the genes encoding 

enzymes of the prechorismate pathway. However, shikimate kinase (SK) did not show any change in 

transcript levels and only one of three isoforms of DAHPS (3-deoxy-D-arabino-heptulosonat-7­

phosphate synthase), the first enzyme in this pathway, was induced by O3 exposure (Janzik et al., 

2005, 191581). Differential induction of DAHPS isoforms was also observed in European beech 

after 40 days of exposure to 150-190 ppb O3. At this time point in the beech experiment, transcript 

levels of shikimate pathway enzymes, including SK, were generally strongly induced after an only 

weak initial induction after the first 40 days of exposure. Both soluble and cell-wall bound phenolic 

metabolites showed only minimal increases in response to O3 for the duration of the exposure period 

(Alonso et al., 2007, 199289). Total leaf phenolics decreased with leaf age in Populus nigra exposed 

to 80 ppb O3 for 12 h/day for 14 days. Ozone increased the concentration of total leaf phenolics in 

newly expanded leaves, with the most significant increases occurring in compounds such as 

quercitin glycoside, which has a high antioxidant capacity (Fares et al., 2010, 628522). While O3 

exposure induced the activity of several phenylpropanoid pathway enzymes, the degree of induction 

differed in the two poplar clones exposed to 60 ppb for 5 h/day for 15 days. In the tolerant I-214 

clone, PAL activity increased ninefold in O3-treated plants as compared to controls, while there was 

no significant difference in PAL activity in the sensitive Eridano clone (Di Baccio et al., 2008, 

199850). 

March 2011 9-43 DRAFT – DO NOT CITE OR QUOTE 
 

1 

2

3 

4

5

6 

7

8 

9 

10

11 

12

13

14 

15

16

17

18 

19

20 

21 

22

23

24

25 

26 

27 

28 

29 

30 

31 

32

33 

34 

35

36 

37 

38 

39

Polyamines such as putrescine, spermidine and spermine play a variety of roles in plants and 

have been implicated in plant defense responses to both abiotic and biotic stresses. They exist in both 

a free form and conjugated to hydroxycinnamic acids. Investigations on the role of polyamines have 

found that levels of putrescine increase in response to oxidative stress. This increase stems largely 

from the increase in the activity of arginine decarboxylase (ADC), a key enzyme in the synthesis of 

putrescine (Groppa and Benavides, 2008, 191616). Langebartels et al. (1991, 043506) described 

differences in putrescine accumulation in O3-treated tobacco plants exposed to several O3 

concentrations, ranging from 0-400 ppb for 5-7 hours. A large and rapid increase in putrescine 

occurred in the tolerant Bel B cultivar and only a small increase in the Bel W3 cultivar, which 

occurred only after the formation of necrotic leaf lesions. Van Buuren et al. (2002, 631199) further 

examined the role of polyamines in these two tobacco cultivars during an acute (130 ppb O3 for 7 h 

in a growth chamber) exposure. They found that while free putrescine accumulated in undamaged 

tissue of both cultivars, conjugated putrescine predominantly accumulated in tissues undergoing cell 

death after plant exposure to O3 (van Buuren et al., 2002, 631199). The authors suggest that while 

free putrescine may not play a role in conferring tolerance in the Bel B cultivar, conjugated 

putrescine may play a role in O3-induced programmed cell death in Bel W3 plants. 

Isoprene is emitted by some plant species and represents the predominant biogenic source of 

hydrocarbon emissions in the atmosphere (Guenther et al., 2006, 607080). In the atmosphere, the 

oxidation of isoprene by hydroxyl radicals can enhance O3 formation in the presence of NOX, 

thereby impacting the O3 concentration that plants are exposed to. While isoprene emission varies 

widely between species, and it has been proposed to stabilize membranes and provide those plant 

species that produce it with a mechanism of thermotolerance (Sharkey et al., 2008, 191492). It has 

also been suggested that isoprene may act as an antioxidant compound to scavenge O3 (Loreto and 

Velikova, 2001, 657212). Recent studies using a variety of plant species have shown conflicting 

results in trying to understand the effects of O3 on isoprene emission. Acute doses of O3 (300 ppb for 

3 h) stimulated isoprene emissions in detached leaves of Phragmites australis (Velikova et al., 2005, 

199410). Similarly, isoprene emissions were stimulated in Populus nigra after exposure to 100 ppb 

O3 for 5 days continuously (Fares et al., 2008, 191683). Isoprene emission in attached leaves of 

Populus alba, which were exposed to 150 ppb O3 for 11 h/day for 30 days inside cuvettes, was 

inhibited, while isoprene emission and transcript levels of isoprene synthase mRNA were increased 

in the leaves exposed to ambient O3 (40 ppb), which were located above the leaves enclosed in the 

exposure cuvettes (Fares et al., 2006, 191455). Exposure of 2 genotypes of hybrid poplar to 120 ppb 

O3 for 6 h/day for 8 days resulted in a significant reduction in isoprene emission in the O3-sensitive 

but not the tolerant genotype (Ryan et al., 2009, 191299). Similarly, O3 treatment (80 ppb 12 h/day 

for 14 days) of Populus nigra showed that isoprene emission was reduced in the treated plants 

relative to the control plants (Fares et al., 2010, 628522). Based on results of this and other studies, 

Fares et al. (2010, 628522) concluded that the isoprenoid pathway may be induced in plants exposed 

to acute O3 doses, while at lower doses isoprene emission may be inhibited. Vickers et al. (2009, 

191497) developed transgenic tobacco plants with the isoprene synthase gene from Populus alba and 

March 2011 9-44 DRAFT – DO NOT CITE OR QUOTE 
 

 

5 

 

 

10

 

 

 

15

 

 

20 

 

 

25 

 

 

 

30 

 

35 

 

1 exposed them to 120 ppb O3 for 6 h/day for 2 days. They determined that the wildtype plants showed 

2 significantly more O3 damage, including the development of leaf lesions and a decline in 

3 photosynthetic rates, than the transgenic, isoprene-emitting plants. Transgenic plants also 

4 accumulated less H2O2 and had lower levels of lipid peroxidation following exposure to O3 than the 

wildtype plants (Vickers et al., 2009, 191497). These results indicate that isoprene may have a 

6 protective role for plants exposed to oxidative stress. 

9.4.6. Changes in Stomatal Function 

7 There has been some debate as to whether O3-induced reductions in photosynthesis result from 

8 changes in stomatal conductance or direct effects on C assimilation. A review of the literature 

9 suggests there may be species-specific differences in how O3 affects photosynthesis through either 

direct or indirect mechanisms. Recent studies utilizing a new simultaneous O3 exposure/gas 

11 exchange device have demonstrated that exposure of Arabidopsis plants to 150 ppb O3 resulted in a 

12 60-70% decline in stomatal conductance within 9-12 minutes of beginning the exposure. Twenty to 

13 thirty minutes later, stomatal conductance had returned to its initial value, even with continuing 

14 exposure to O3, indicating a rapid direct effect of O3 on stomatal function (Kollist et al., 2007, 

191539). The contributions of stomatal versus non-stomatal factors in reducing net photosynthesis 

16 have been investigated in a number of tree species. Ginkgo biloba trees exposed to ambient O3 

17 (40 ppb) and elevated O3 (80 ppb) in OTCs for 90 days were investigated over the course of seven 

18 months. Initially, stomatal limitation restricted C assimilation and, along with higher rates dark 

19 respiration, served to protect against oxidative stress. However, as the season progressed, declines in 

C assimilation resulted from direct oxidative damage to the photosynthetic apparatus (He et al., 

21 2007, 199789). In other studies focusing on O3 effects on adult trees, including Quercus mongolica 

22 (exposed to 80 ppb O3 for 9 h/day for 4 months in OTCs) and Fagus sylvatica (exposed to the 

23 2 × ambient concentrations, with a mean of 57 ppb O3 for 6 months in a free air fumigation system) 

24 O3-induced reductions in stomatal conductance were accompanied by a decrease in intercellular CO2 

concentration, indicating that stomatal closure did not result from high internal CO2 but rather from a 

26 direct effect of O3 on the guard cells (Wang et al., 2009, 199303)(Kitao et al., 2009, 191331). In a 

27 related study, Populus deltoides grown in rural areas showed a greater O3-induced decline in biomass 

28 allocation than those grown in urban areas (with means in O3 concentrations ranging from 20-36 ppb 

29 in urban areas and 36-47 ppb in rural areas). This decline in biomass was determined to result from 

significantly higher rates of stomatal conductance in the rural trees, indicating a loss of stomatal 

31 control (Gregg et al., 2006, 186961). An investigation of the differences between O3-sensitive and 

32 insensitive cutleaf coneflower (Rudbeckia laciniata var. digitata) revealed differences in stomatal 

33 behavior (Grulke et al., 2007, 186963). Cutleaf coneflower plants growing in Great Smoky Mountain 

34 National Park (exposed to mean O3 concentration of 42 ppb from mid-May through mid-September) 

showed variable stomatal responses to changes in environmental conditions, such as light and vapor 

36 pressure deficit. It is likely that the O3-sensitive coneflower plants have a set of traits, such as a 

March 2011 9-45 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

 

 

 

 

 

 

1 sluggish stomatal response to changes in light intensity, which predispose them to being more 

2 sensitive to O3 exposure (Grulke et al., 2007, 186963)(Paoletti and Grulke, 2010, 628561). 

3 Feng et al. (2007, 191641) determined that reductions in net photosynthesis in O3-exposed 

4 winter wheat (Triticum aestivum), exposed to a mean concentration of either 52 or 105 ppb O3 for 

5 6 weeks in OTCs, likely occurred as a result of both stomatal and non-stomatal factors. Other studies 

6 suggest that O3 impacts net photosynthesis only through non-stomatal factors. A direct effect of O3 

7 on C fixation as described above would lead to a buildup of internal CO2, resulting in a reductions in 

8 stomatal conductance through stomatal closure. In Acer opalus, a significant reduction in light9 saturated photosynthesis was accompanied by significant increases in internal CO2 concentration and 

10 decreases in water use efficiency without appreciable changes in stomatal conductance (Calatayud et 

11 al., 2007, 191411). A similar decrease in photosynthesis, accompanied by an increase in internal CO2 

12 concentration, was measured in soybean plants exposed to O3. Singh et al. (2009, 199427) attributed 

13 the declines in photosynthesis to direct damage to the photosynthetic apparatus rather than to 

14 stomatal limitations. 

9.5. Na tu re o f Effe c ts o n Ve g e ta tio n 

9.5.1. Introd uc tion 

15 Ambient O3 concentrations have long been known to cause visible symptoms, decreases in 

16 photosynthetic rates, decreases in growth, and decreases in the yield of plants (U.S. EPA, 1978, 

17 040586)(U.S. EPA, 1986, 017607)(U.S. EPA, 1996, 080828)(U.S. EPA, 2006, 088089). Numerous 

18 studies have related O3 exposure to plant responses, with most effort focused on the yield of crops 

19 and the growth of tree seedlings. Most experiments exposed individual plants grown in pots or soil 

20 under controlled conditions to known concentrations of O3 for a segment of daylight hours for some 

21 portion of the plant’s life span (Section 9.3). This section focuses on the responses of plants to 

22 seasonal or multi-year exposures to known amounts of O3. Quantitative responses include changes in 

23 growth and biomass allocation, changes in reproduction, onset of visible foliar injury, and changes in 

24 leaf gas exchange. The response of a plant species or variety to O3 exposure depends upon many 

25 factors, including biochemical and physiological status (Section 9.4), genetic characteristics, and 

26 previous and current exposure to other stressors (Section 9.5.4.). Because of the available 

27 information, most of this section focuses on the response of individual plants, especially tree 

28 seedlings and crops, with limited discussion of mixtures of herbaceous species. Responses at the 

29 ecosystem scale are discussed in Section 9.6. 

30 This section will focus mainly on studies published since the release of the 2006 O3 AQCD 

31 (U.S. EPA, 2006, 088089). However, because much O3 research was conducted prior to the 2006 O3 

32 AQCD, the present discussion of vegetation response to O3 exposure is largely based on the 

33 conclusions of the 1978, 1986, 1996, and 2006 O3 AQCDs (U.S. EPA, 1978, 040586)(U.S. EPA, 

34 1986, 017607)(U.S. EPA, 1996, 080828)(U.S. EPA, 2006, 088089). 

March 2011 9-46 DRAFT – DO NOT CITE OR QUOTE 
 

 

1 

2 

3

4 

5 

6

7 

8 

9 

10 

11

12

13

14 

15

16

17 

18

19

20

21 

22 

23 

24

25

26

27 

28

29 

30

31 

32

33 

34

35 

9.5.2. Effe cts on Wo od y an d Herb ac e ou s Ve ge ta tio n 

9.5.2.1. Growth and Bio ma s s Alloc a tion 

The previous O3 AQCDs concluded that there is strong evidence that exposure to O3 decreases 

growth in numerous plant species. Studies published since the last review support those conclusions 

and are summarized below. 

In a recently published meta-analysis, Wittig et al. (2009, 191631) quantitatively compiled 

peer reviewed studies from the past 40 years on the effect of current and future O3 exposures on the 

physiology and growth of forest species. Wittig et al. (2009, 191631) reported that current ambient 

O3 concentrations (~40 ppb) significantly decreased annual total biomass growth (7%) across 263 

studies. However, this effect could be greater (11 to 17%) in areas that have higher O3 concentrations 

and as background O3 increases in the future (Wittig et al., 2009, 191631). This meta-analysis 

demonstrates the coherence of O3 effects across numerous studies and species using a variety of 

experimental techniques. 

In two companion papers, McLaughlin et al. (2007, 090348)(2007, 090347) investigated the 

effects of ambient O3 on tree growth and hydrology at forest sites in the southern Appalachian 

Mountains. The authors reported the cumulative effects of ambient levels of O3 decreased seasonal 

stem growth by 30-50% for most trees species in a high O3 year in comparison to a low O3 year 

(McLaughlin et al., 2007, 090348). The authors also report that high ambient O3 concentrations can 

disrupt whole-tree water use and in turn reduce late-season streamflow (McLaughlin et al., 2007, 

090347); see Section 9.6.3 for more on water cycling. 

Since the 2006 O3 AQCD, several new studies based on the Aspen FACE “free air” O3 and 

CO2 exposure experiment in a forest in Wisconsin were published (Darbah et al., 2007, 

093288)(Darbah et al., 2008, 196890)(King et al., 2005, 191701)(Kubiske et al., 2006, 

093284)(Kubiske et al., 2007, 191336)(Riikonen et al., 2008, 191258). King et al. (2005, 191701) 

found that O3 fumigation over the first seven years of stand development reduced total biomass 

relative to the control by 23, 13, and 14% in the aspen, aspen-birch, and aspen-maple communities, 

respectively. Over the same time period, Kubiske et al. (2006, 093284) observed that elevated O3 

decreased tree heights, diameters, and main stem volumes in the aspen community by 11, 16, and 

20%, respectively. In addition, Kubiske et al. (2007, 191336) reported that elevated O3 may change 

the intra- and inter-species competition. For example, O3 treatments increased the rate of conversion 

from a mixed aspen-birch community to a birch dominated community. In another study at this site, 

Percy et al. (2007, 093287) suggested that negative growth effects were seen below the previous 8-h 

O3 standard level of 0.084 ppm, but the informativeness of the study was diminished by severe 

methodological problems. 

Several studies at the Aspen FACE site also considered other growth-related effects of elevated 

O3. Darbah et al. (2007, 093288)(2008, 196890) reported that O3 treatments decreased paper birch 

seed weight and seed germination and that this would likely lead to a negative impact of 

March 2011 9-47 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

5

 

 

 

10 

 

 

15

 

 

20 

 

25

 

 

 

30 

 

35 

 

 

1 regeneration for that species. Riikonen et al. (2008, 191258) found that elevated O3 decreased the 

2 amount of starch in birch buds by 16%, and reduced aspen bud size, which may have been related to 

3 the observed delay in spring leaf development. The results suggest that elevated O3 concentrations 

4 have the potential to alter C metabolism of overwintering buds, which may have carry-over effects in 

the subsequent growing season (Riikonen et al., 2008, 191258). 

6 Effects on growth of understory vegetation were also investigated at Aspen FACE. Bandeff et 

7 al. (2006, 191733) found that the effects of elevated CO2 and O3 on understory species composition, 

8 total and individual species biomass, N content, and 15N recovery were a result of overstory 

9 community responses to those treatments; however, there were no apparent direct treatment effects 

due to high variability of the data. Total understory biomass increased with increasing light and was 

11 greatest under the open canopy of the aspen/maple community, as well as the more open canopy of 

12 the elevated O3 treatments (Bandeff et al., 2006, 191733). Similarly, data from a study by Awmack et 

13 al. (2007, 191415) suggest that elevated CO2 and O3 may have indirect growth effects on red 

14 (Trifolium pratense) and white (Trifolium repens) clover in the understory via overstory community 

effects; however, no direct effects of elevated O3 were observed. 

16 Overall, the studies at the Aspen FACE experiment are consistent with many of the OTC 

17 studies that were the foundation of previous O3 NAAQS reviews. These results strengthen our 

18 understanding of O3 effects on forests and demonstrate the relevance of the knowledge gained from 

19 trees grown in open-top chamber studies. 

For some annual species, particularly crops, the endpoint for an assessment of the risk of O3 

21 exposure can be defined as yield or growth, e.g., production of grain. For plants grown in mixtures 

22 such as hayfields, and natural or semi-natural grasslands (including native nonagricultural species), 

23 endpoints other than production of biomass may be important. Such endpoints include biodiversity 

24 or species composition, and effects may result from competitive interactions among plants in mixedspecies communities. Most of the available data on non-crop herbaceous species are for grasslands 

26 with many of the recent studies conducted in Europe. See Section 9.6.5 for a review of the recent 

27 literature on O3 effects on competition and biodiversity in grasslands. 

Root Growth 

28 Although O3 does not penetrate soil, it could alter root development by decreasing 

29 C assimilation via photosynthesis (Andersen, 2003, 041673). The response of root development to 

O3 exposure depends on available photosynthate and could vary over time. Many biotic and abiotic 

31 factors, such as community dynamics and drought stress, have been found to affect root production 

32 under elevated O3. An earlier study at the AspenFACE experiment found that elevated O3 reduced 

33 coarse root and fine roots biomass in young stands of paper birch and trembling aspen (King et al., 

34 2001, 041751). However, this reduction disappeared several years later. Ozone significantly 

increased fine-root (<1.0 mm) in the aspen community (Pregitzer et al., 2008, 191677). This increase 

36 in fine root production was due to changes in community composition, such as better survival of the 

37 O3-tolerant aspen genotype, birch, and maple, rather than changes in C allocation at the individual 

March 2011 9-48 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

5 

 

 

10

 

 

 

 

15

 

 

 

 

20

25

 

30

 

 

 

35 

1 tree level (Pregitzer et al., 2008, 191677)(Zak et al., 2007, 191239). In an adult European 

2 beech/Norway spruce forest in Germany, drought was found to nullify the O3-driven stimulation of 

3 fine root growth. Ozone stimulated fine-root production of beech during the humid year, but had no 

4 significant impact on fine root production in the dry year (Nikolova et al., 2010, 626810)(Matyssek 

et al., 2010, 628553). 

6 Using a non-destructive method, Vollsnes et al. (2010, 625576) studied the in vivo root 

7 development of subterranean clover (Trifolium subterraneum) before, during and after short-term O3 

8 exposure. It was found that O3 reduced root tip formation, root elongation, the total root length, and 

9 the ratios between below- and above-ground growth within one week after exposure. Those effects 

persisted for up to three weeks; however, biomass and biomass ratios were not significantly altered 

11 at the harvest five weeks after exposure. 

12 Ozone has been shown to have negative (Jones et al., 2010, 567354), non-significant (Phillips 

13 et al., 2009, 199802)(Andersen et al., 2010, 628559) and positive effects (Pregitzer et al., 2008, 

14 191677)(Grebenc and Kraigher, 2007, 191265) on root biomass and root: shoot ratio. While the 

findings of individual studies were mixed, several recent meta-analyses have generally indicated that 

16 O3 reduced C allocated to roots. In one meta-analysis, Grantz et al. (2006, 191545) estimated the 

17 effect of O3 on the root:shoot allometric coefficient (k), the ratio between the relative growth rate of 

18 the root and shoot. The results showed that O3 reduced k by 5.6%, and the largest decline was 

19 observed in slow-growing plants. In another meta-analysis including 263 publications, Wittig et al. 

(2009, 191631) found that current O3 exposure had no significant impacts on root biomass and 

21 root:shoot ratio when compared to pre-industrial O3 exposure. However, if O3 concentrations rose to 

22 81-101 ppb (projected O3 levels in 2100), both root biomass and root:shoot ratio were found to 

23 significantly decrease. Gymnosperms and angiosperms differed in their responses, with 

24 gymnosperms being less sensitive to elevated O3. In two other meta-analyses, Wang et al. (2010, 

387478) found elevated O3 reduced biomass allocation to roots by 8.3% at ambient CO2 and 6.0% at 

26 elevated CO2, and Morgan et al. (2003, 055527) found O3 reduced root dry weight of soybean. 

9.5.2.2. Reproduction 

27 Studies during recent decades have demonstrated O3 effects on different stages of plant 

28 reproduction. The impacts of O3 on reproductive development, as reviewed by Black et al. (2000, 

29 036322), can occur by influencing (1) age at which flowering occurs, particularly in long-lived trees 

that often have long juvenile periods of early growth without flower and seed production; (2) flower 

31 bud initiation and development; (3) pollen germination and pollen tube growth; and (4) seed, fruit, or 

32 cone yields and seed quality (Table 9-2) (U.S. EPA, 2006, 088089). 

33 Several recent studies since the 2006 O3 AQCD further demonstrate the effects of O3 on 

34 reproductive processes in herbaceous and woody plant species. Rämö et al. (2007, 191441) exposed 

several meadow species to elevated O3 (40-50 ppb) and CO2 (+100 ppm), both individually and 

36 combined, over three growing seasons in ground-planted mesocosms, using OTCs. Elevated O3 

March 2011 9-49 DRAFT – DO NOT CITE OR QUOTE 
 

1

2

3 

4

5

6

7

8

9

10

11

12 

13 

14

15

16 

17 

18

19 

20 

21

22 

23

24 

delayed flowering of Campanula rotundifolia and Vicia cracca. Ozone also reduced the overall 

number of produced flowers and decreased fresh weight of individual Fragaria vesca berries. 

Black et al. (2007, 191558) exposed Brassica campestris to 70 ppb for two days during late 

vegetative growth or ten days during most of the vegetative phase. The two-day exposure had no 

effect on growth or reproductive characteristics, while the 10 day exposure reduced vegetative 

growth and reproductive site number on the terminal raceme, emphasizing the importance of 

exposure duration and timing. Mature seed number and weight per pod were unaffected due to 

reduced seed abortion, suggesting that, although O3 affected reproductive processes, indeterminate 

species such as B. campestris possess enough compensatory flexibility to avoid reduced seed 

production (Black et al., 2007, 191558). 

In the determinate species, Plantago major, Black et al. (2010, 625575) found that O3 can 

have direct effects on reproductive development in populations of differing sensitivity. Only the first 

flowering spike was exposed to 120 ppb O3 for 7 hours per day on 9 successive days (corresponding 

to flower development) while the leaves and second spike were exposed to charcoal-filtered air. 

Exposure of the first spike to O3 affected seed number per capsule on both spikes even though spike 

two was not exposed. The combined seed weight of spikes one and two was increased by 19% in the 

two resistant populations, suggesting an overcompensation for injury; whereas, a decrease of 21% 

was observed in the most sensitive population (Black et al., 2010, 625575). 

A study by Darbah et al. (2007, 093288)(2008, 196890) of paper birch (Betula papyrifera) 

trees at the Aspen FACE site in Rhinelander, WI investigated the effects of elevated O3 and/or CO2 

on reproductive fitness. Elevated O3 increased flowering, but decreased seed weight and germination 

success rate of seeds from the exposed trees. These results suggest that O3 can dramatically affect 

flowering, seed production, and seed quality of paper birch, ultimately affecting its reproductive 

fitness (Darbah et al., 2007, 093288)(Darbah et al., 2008, 196890). 

March 2011 9-50 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

Table 9-2. Ozone effects on plant reproductive processes (derived from Table AX9-22 of the 2006

ozone AQCD) 

Species Condition Measures References 

Apocynun androsaemifolium Flowering time Bergweiler and Manning (1999, 036321) 

Buddleia davidii Flowering time Findley et al. (1997, 036337) 

Rubus cuneifolius Pollen germination Chappelka (2002, 052946) 

Plantago major Pollen tube elongation Stewart (1998, 094233) 

Fragaria × ananassa Fruit yield Drogoudi and Ashmore (2001, 052959); Drogoudi and Ashmore 

(2000, 042517) 

Plantago major Seed yield Lyons and Barnes (1998, 036354); Pearson et al. (1996, 053074); 

Reiling and Davison (1992, 043591); Whitfield et al. (1997, 036397) 

Understory herbs Seed yield Harward and Treshow (1975, 038366) 

Source: Adapted from 2006 O3 AQCD (U.S. EPA, 2006, 088089) 

9.5.2.3. Vis ible Foliar Injury 

1 Visible foliar injury resulting from exposure to O3 has been well characterized and 

2 documented over several decades on many tree, shrub, herbaceous, and crop species (U.S. EPA, 

3 2006, 088089)(U.S. EPA, 1996, 080827)(U.S. EPA, 1984, 029711)(U.S. EPA, 1978, 040586). 

4 Visible foliar injury symptoms are considered diagnostic as they have been verified experimentally 

5 in exposure-response studies, using exposure methodologies such as CSTRs, OTCs, and free-air 

6 fumigation (see Section 9.3 for more detail on exposure methodologies). Several pictorial atlases and 

7 guides have been published, providing details on diagnosis and identification of O3-induced visible 

8 foliar injury on many plant species throughout North America (Penn State, 1987, 626971)(Flagler, 

9 1998, 025525) and Europe (Sánchez et al., 2001, 626980)(Sánchez et al., 2001, 626980)(Innes et al., 

10 2001, 048954). Typical visible injury symptoms on broad-leaved plants include: stippling, flecking, 

11 surface bleaching, bifacial necrosis, pigmentation (e.g., bronzing), chlorosis, and/or premature 

12 senescence. Typical visible injury symptoms for conifers include: chlorotic banding, tip burn, 

13 flecking, chlorotic mottling, and/or premature senescence of needles. Although common patterns of 

14 injury develop within a species, these foliar lesions can vary considerably between and within 

15 taxonomic groups. Furthermore, the degree and extent of visible foliar injury development varies 

16 from year to year and site to site (Chappelka et al., 2007, 093290)(Orendovici-Best et al., 2008, 

17 196940)(Smith et al., 2003, 044183), even among co-members of a population exposed to similar O3 

18 levels, due to the influence of co-occurring environmental and genetic factors. Nevertheless, 

19 Chappelka et al. (2007, 093290) reported that the average incidence of O3-induced foliar injury was 

20 73% on milkweed in the Great Smokey Mountain National Park in the years 1992-1996. 

21 Although the majority of O3-induced visible foliar injury occurrence has been observed on 

22 seedlings and small plants, many studies have reported visible injury of mature coniferous trees, 

23 primarily in the western U.S. (Arbaugh et al., 1998, 040297) and to mature deciduous trees in eastern 

24 North America (Chappelka et al., 1999, 041860; Chappelka et al., 1999, 052952; Hildebrand et al., 

March 2011 9-51 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

5 

 

 

 

10

 

 

 

 

15 

 

 

 

20

 

 

25 

 

 

30

 

35

 

 

1 1996, 042670; Schaub et al., 2005, 191608; Somers et al., 1998, 044010; Vollenweider et al., 2003, 

2 051157). 

3 It is important to note that visible foliar injury occurs only when sensitive plants are exposed 

4 to elevated O3 concentrations in a predisposing environment. A major confounding factor for O3­

induced visible foliar injury is the amount of soil moisture available to a plant during the year that 

6 the visible foliar injury is being assessed. This is because lack of soil moisture generally decreases 

7 stomatal conductance of plants and, therefore, limits the amount of O3 entering the leaf that can 

8 cause injury (Grulke et al., 2003, 042637)(Matyssek et al., 2006, 191481)(Panek, 2004, 

9 079202)(Panek and Goldstein, 2001, 030190)(Temple et al., 1992, 043349)(Temple et al., 1988, 

043237). Consequently, many studies have shown that dry periods in local areas tend to decrease the 

11 incidence and severity of O3-induced visible foliar injury; therefore, the incidence of visible foliar 

12 injury is not always higher in years and areas with higher O3, especially with co-occurring drought 

13 (Smith et al., 2003, 044183). Other factors such as leaf age influence the severity of symptom 

14 expression with older leaves showing greater injury severity (Zhang et al., 2010, 628555). 

Although visible injury is a valuable indicator of the presence of phytotoxic concentrations of 

16 O3 in ambient air, it is not always a reliable indicator of other negative effects on vegetation. The 

17 significance of O3 injury at the leaf and whole plant levels depends on how much of the total leaf 

18 area of the plant has been affected, as well as the plant’s age, size, developmental stage, and degree 

19 of functional redundancy among the existing leaf area. Previous O3 AQCDs have noted the difficulty 

in relating visible foliar injury symptoms to other vegetation effects such as individual plant growth, 

21 stand growth, or ecosystem characteristics (U.S. EPA, 2006, 088089)(U.S. EPA, 1996, 

22 080827)(U.S. EPA, 1996, 080827). As a result, it is not presently possible to determine, with 

23 consistency across species and environments, what degree of injury at the leaf level has significance 

24 to the vigor of the whole plant. However, in some cases, visible foliar symptoms have been 

correlated with decreased vegetative growth (Benoit et al., 1982, 039778; Karnosky et al., 1996, 

26 036347; Peterson et al., 1987, 042148; Somers et al., 1998, 044010) and with impaired reproductive 

27 function (Black et al., 2000, 036322; Chappelka, 2002, 052946). Conversely, the lack of visible 

28 injury does not always indicate a lack of phytotoxic concentrations of O3 or a lack of non-visible O3 

29 effects (Gregg et al., 2003, 046996; Gregg et al., 2006, 186961). 

Biomonitoring 

The use of biological indicators to detect phytotoxic levels of O3 is a longstanding and 

31 effective methodology (Chappelka and Samuelson, 1998, 093687; Manning and Krupa, 1992, 

32 044155). A plant bioindicator can be defined as a vascular or nonvascular plant exhibiting a typical 

33 and verifiable response when exposed to a plant stress such as an air pollutant (Manning, 2003, 

34 053773). 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 

36 (U.S. EPA, 2006, 088089). Such sensitive plants can be used to detect the presence of a specific air 

37 pollutant such as O3 in the ambient air at a specific location or region and, as a result of the 

March 2011 9-52 DRAFT – DO NOT CITE OR QUOTE 
 

1 

2

3 

4

5 

6 

7

8

9 

10 

11 

12

13

14 

15 

16

17

18

19

20

21 

22 

23

24

25

26 

27 

28 

29

30 

31 

32

33

34 

35

36

37 

38 

39

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 (Calatayud et al., 2007, 191568; Gombert et al., 2006, 089309; Heggestad, 

1991, 042533; Kostka-Rick and Hahn, 2005, 199428; Laffray et al., 2007, 191384; Nali et al., 2007, 

191478) or detectors, which are sensitive native plant species (e.g., tall milkweed [Asclepias 

exaltata]; Chappelka et al., 2007, 093290; Souza et al., 2006, 191658). The approach is especially 

useful in areas where O3 monitors are not operated (Manning, 2003, 053773). 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, 040430) 

may be a useful methodology (Manning, 2003, 053773). 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 2006 O3 AQCD (U.S. EPA, 2006, 088089), new sensitive plant species have been 

identified from field surveys and verified in controlled exposure studies (Kline et al., 2008, 191591; 

Kline et al., 2009, 196918). Several multiple-year field surveys have also been conducted at National 

Wildlife Refuges in Maine, Michigan, New Jersey, and South Carolina (Davis, 2007, 093291; Davis, 

2007, 093292; Davis, 2009, 199287; Davis and Orendovici, 2006, 093293). 

The USDA Forest Service through the Forest Health Monitoring Program (FHM) (1990 ­

2001) and currently the Forest Inventory and Analysis (FIA) Program has been collecting data 

regarding the incidence and severity of visible foliar injury on a variety of O3 sensitive plant species 

throughout the U.S. (Coulston et al., 2003, 041871)(Smith et al., 2003, 044183). The plots where 

these data are taken are known as biosites. These biosites are located throughout the country and 

analysis of visible foliar injury within these sites follows a set of established protocols. For more 

details, see http://www.nrs.fs.fed.us/fia/topics/ozone/ (USDA, 2011, 677550). 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, 

041871) used information gathered over a 6-year period (1994-1999) from the network to identify 

several species that were sensitive to O3 over a regional scale including sweetgum (Liquidambar 

styraciflua), loblolly pine (Pinus taeda), and black cherry (Prunus serotina). In a study of the west 

coast of the U.S, Campbell et al. (2007, 602360) reported O3 injury in 25-37% of biosites in 

California forested ecosystems from 2000-2005. 

A study by Kohut (2007, 093289) assessed the risk of O3-induced visible foliar injury on 

bioindicator plants (NPS, 2006, 677536) in 244 national parks in support of the National Park 

Service's Vital Signs Monitoring Network (NPS, 2007, 677537). The risk assessment was based on a 

simple model relating response to the interaction of the plant, the level of O3 exposure, and the 

exposure environment. Kohut (2007, 093289) concluded that the risk of visible foliar injury was 

high in 65 parks (27%), moderate in 46 parks (19%), and low in 131 parks (54%). Some of the wellknown parks with a high risk of O3-induced visible foliar injury include Gettysburg, Valley Forge, 

March 2011 9-53 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

5 

 

 

 

10

 

 

15 

 

20 

 

 

 

25

 

 

30

 

1 Delaware Water Gap, Cape Cod, Fire Island, Antietam, Harpers Ferry, Manassas, Wolf Trap Farm 

2 Park, Mammoth Cave, Shiloh, Sleeping Bear Dunes, Great Smoky Mountains, Joshua Tree, Sequoia 

3 and Kings Canyon, and Yosemite. 

9.5.2.4. Le a f Ga s Exc ha n ge 

4	 In general, there is strong experimental evidence over several decades of research that 

exposure to O3 reduces photosynthesis and alters stomatal conductance in a wide variety of plant 

6 species (Wittig et al., 2007, 191695) and these effects have been discussed in detail in previous O3 

7 AQCDs (U.S. EPA, 2006, 088089)(U.S. EPA, 1996, 080827). Recent studies related to these effects 

8 are discussed in several sections within this document; therefore, this section refers to those sections 

9 for a more in depth discussion. 

Ozone effects on photosynthesis were presented Section 9.4.5.1, with a focus on recent studies 

11 on the effects of O3 on light reactions. Changes in stomatal function in response to O3 exposure were 

12 discussed in Section 9.4.6. In addition, the implications of changes in stomatal control for water 

13 cycling are discussed in Section 9.6.3. Leaf gas exchange as it relates to effects based air quality 

14 exposure indices and dose modeling is discussed throughout Section 9.7. 

9.5.3. Ag ric ultu ral Crops 

The detrimental effect of O3 on crop production has been recognized since the 1960’s and a 

16 large body of research has stemmed from that recognition. Previous O3 AQCDs have extensively 

17 reviewed this body of literature (U.S. EPA, 2006, 088089). Table 9-3 summarizes recent 

18 experimental studies of O3 effects on agricultural crops, exclusive of growth and yield. Growth and 

19 yield results are summarized in Table 9-16. 

Ozone diffuses into the leaf apoplast via the stomata where it is rapidly converted into other 

21 ROS that signal a diverse metabolic response (Kangasjarvi et al., 2005, 180341)(Long and Naidu, 

22 2002, 038179). The mechanism of O3 sensing and cellular response is detailed in Section 9.4.3.1. 

23 Ozone stress has been characterized as either acute or chronic, depending on the O3 concentration 

24 and the exposure duration (Fiscus et al., 2005, 079155). While the actual concentration and duration 

threshold for O3 damage varies from species to species and sometimes even among genotypes of the 

26 same species (Ariyaphanphitak et al., 2005, 191349)(Biswas et al., 2008, 191428)(Dalstein and Vas, 

27 2005, 191570)(Guidi et al., 2009, 199825)(Keutgen et al., 2005, 191295)(Sawada and Kohno, 2009, 

28 199426), it is commonly accepted that acute damage results from a very high concentration of O3 

29 (>150 ppb) over a short period of time, and chronic O3 damage results from a lower concentration of 

exposure over a longer period of time. In general, acute O3 damage has been well characterized and 

31 mimics the biochemical defense response of plants to pathogen attack (Kangasjarvi et al., 2005, 

32 180341)(Overmyer et al., 2003, 053537). In contrast, the mechanism leading to chronic O3 damage 

33 is less well-characterized but hallmark physiological symptoms include: decreased photosynthetic 

34 productivity, decreased Rubisco activity and chlorophyll content, lower stomatal conductance, leaf 

March 2011	 9-54 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

5 

 

 

 

10 

 

 

 

15

 

 

20

 

 

25

 

 

30

 

 

 

35 

1 chlorosis, accelerated senescence and a general decrease in green leaf area and plant productivity 

2 (Ashmore et al., 2006, 191557). Despite the knowledge gap pertaining to the mechanism of O3 

3 damage, a number of comprehensive reviews and meta-analyses have recently been published 

4 discussing both the current understanding of the quantitative effects of O3 concentration on a variety 

of crop species and the potential focus areas for biotechnological improvement to a future growing 

6 environment that will include higher O3 concentrations (Ainsworth, 2008, 191646)(Booker et al., 

7 2009, 191569)(Feng et al., 2008, 191453)(Grantz et al., 2006, 191545)(Hayes et al., 2007, 

8 196911)(Mills et al., 2007, 180221)(Morgan et al., 2003, 055527)(Van Dingenen et al., 2009, 

9 199765). Since the 2006 O3 AQCD, exposure-response indices for a variety of crops have been 

suggested (Mills et al., 2007, 180221) and many reports have investigated the effects of O3 

11 concentration on seed or fruit quality to extend the knowledge base beyond yield quantity. This 

12 section will outline the key findings from these papers as well as highlight some of the recent 

13 research addressing the endpoints such as yields and crop quality. 

14 Genetic variability is not the only factor that determines the crop response to O3-damage. 

Ozone concentrations throughout a growing-season is not homogeneous and other environmental 

16 conditions, such as elevated CO2 concentrations, drought, cold or nutrient availability may alleviate 

17 or exacerbate the oxidative stress response to a given O3 concentration. This section will also 

18 highlight recent literature that focuses on O3 damage to crops as influenced by other environmental 

19 factors. 

9.5.3.1. Yield 

It is well known that yield is negatively impacted in many crop species in response to high O3 

21 concentrations. However the threshold for damage varies from species to species. Reproductive 

22 organs such as seeds may be particularly sensitive to injury or biomass reductions due to O3, as 

23 reviewed by Black et al. (2000, 036322). Numerous analyses of experiments conducted in OTCs and 

24 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 

26 the seed or grain-filling period (Lee et al., 1988, 594572)(Pleijel et al., 1998, 053021)(Soja et al., 

27 2000, 030388)(Younglove et al., 1994, 044162). AX9.5.4.1 of the 2006 O3 AQCD summarized many 

28 previous studies on crop yield (U.S. EPA, 2006, 088089). 

29 The effect of O3 exposure on U.S. crops remains an important area of research and several 

studies have been published on this topic since the 2006 O3 AQCD (Tables 9-3 and 9-16). For 

31 example, one study with cotton in a crop-weed interaction study (Grantz and Shrestha, 2006, 

32 191702) utilizing OTCs suggests that ambient O3 concentrations (12-h avg: 79.9 ppb) decreased 

33 cotton biomass by 25% and 1.5 × ambient O3 concentration (12-h avg: 122.7 ppb) decreased cotton 

34 biomass by 75% compared to charcoal filtered control (12-h avg: 12.8 ppb). Further, this study 

suggests that the weed, yellow nutsedge, was less sensitive to increasing O3 concentration which 

36 would increase weed competition (Grantz and Shrestha, 2006, 191702). In a study of peanuts in 

37 North Carolina, near ambient and elevated exposures of O3 reduced photosynthesis and yield 

March 2011 9-55 DRAFT – DO NOT CITE OR QUOTE 
 

1 

2

3

4

5 

6 

7 

8 

9

10 

11 

12

13

14

15 

16 

17 

18

19

20

21

22

23 

24 

25

26

27 

28 

29

30

31 

32

33

34 

35 

36

37 

38 

compared to very low O3 conditions (Booker et al., 2007, 191370)(Burkey et al., 2007, 191371). In 

another study, Grantz and Vu (2009, 195237) reported that sugarcane biomass growth significantly 

declined under O3 exposure. 

The average yield loss reported across a number of meta-analytic studies have been published 

recently for soybean (Morgan et al., 2003, 055527), wheat (Feng et al., 2008, 191453), rice 

(Ainsworth, 2008, 191646), semi-natural vegetation (Hayes et al., 2007, 196911), potato, bean and 

barley (Feng and Kobayashi, 2009, 199223). The meta-analytic technique allows for the objective 

development of a quantitative consensus of the effects of a treatment across a wide body of literature 

and therefore provides an average response ratio compiled from these sources. Further, this 

technique allows for a compilation of data across a range of O3-fumigation techniques, durations and 

concentrations in order to assemble the existing literature in a meaningful manner. 

Morgan et al. (2003, 055527) reported an average seed yield loss for soybean of 24% 

compared to charcoal filtered air across all O3 concentrations used in the 53 compiled studies. The 

decrease in seed yield appeared to be the product of nearly equal decreases (7-12%) in seed weight, 

seed number and pod number. As would be expected, the lowest O3 concentration (30-59 ppb) 

resulted in the smallest yield losses, approximately 8%, while the highest O3 concentration 

(80-120 ppb ) resulted in the largest yield losses, approximately 35% (Morgan et al., 2003, 055527). 

Further, the oil/protein ratio within the soybean seed was altered due to growth at elevated O3 

concentrations, with a decrease in oil content. The studies included in this meta-analysis all used 

enclosed fumigation systems or growth chambers which may have altered the coupling of the 

atmosphere to the lower plant canopy (McLeod and Long, 1999, 688834). Utilizing the Soybean 

Free Air gas Concentration Enrichment Facility (SoyFACE; www.soyface.illinois.edu), Morgan et al. 

(2006, 079186) report a 20% seed yield loss due to a 23% increase in average daytime O3 

concentration (56-69 ppb) within a single soybean cultivar across two growing seasons in Illinois, 

supporting the results from the meta-analysis. A further breakdown of the effects of current O3 

concentrations (AOT40 of 4.7 ppm-h) on bean seed quality (Phaseolus vulgaris) has identified that 

growth at current O3 concentrations compared to charcoal-filtered air raised total lipids, total crude 

protein and dietary fiber content (Iriti et al., 2009, 195635). An increase in total phenolics was also 

observed, however the individual phenolics compounds responded differently, with significant 

decreases in anthocyanin content. The seeds from ambient O3 exposed plants also displayed 

increased total antioxidant capacity compared to charcoal-filtered air controls (Iriti et al., 2009, 

195635). Betzelberger et al. (2010, 644183) has recently utilized the SoyFACE facility to compare 

the impact of elevated O3 concentrations across 10 soybean cultivars to investigate intraspecific 

variability of the O3 response to find physiological or biochemical markers for eventual O3 tolerance 

breeding efforts (Betzelberger et al., 2010, 644183). They report an average 17% decrease in yield 

across all 10 cultivars across two growing seasons due to a doubling of ambient O3 concentrations, 

with a the individual cultivar responses ranging from -7% to -36%. The dose-response functions 

derived for these 10 modern cultivars were similar to the response functions derived from the 

March 2011 9-56 DRAFT – DO NOT CITE OR QUOTE 
 

1

2

3

4 

5

6 

7

8

9

10

11

12 

13

14 

15

16 

17 

18

19

20

21

22

23 

24

25 

26

27

28

29 

30 

31 

32 

33 

34 

35 

36 

37 

38 

39

NCLAN studies conducted in the 1980’s (Heagle, 1989, 093985) suggesting there has not been any 

inadvertent selection for more O3-tolerant cultivars in recent history. 

A meta-analysis has also been performed on studies investigating the effects of O3 

concentrations on wheat (Feng et al., 2008, 191453). Across 23 studies included, elevated O3 

concentrations (ranging from a 7-h daily average of 31-200 ppb) decreased grain yield by 29%. 

Winter wheat and spring wheat did not differ in their responses; however the response in both 

varieties to increasing O3 concentrations resulted in successively larger decreases in yield, from a 

20% decrease in 42 ppb to 60% in 153 ppb O3. These yield losses were mainly caused by a 

combination of decreases in individual grain weight (-18%), ear number per plant (-16%), and grain 

number per ear (-11%). Further, the grain starch concentration decreased by 8% and the grain protein 

yield decreased by 18% due to growth at elevated O3 concentrations as well. However, increases in 

grain calcium and potassium levels were reported (Feng et al., 2008, 191453). 

A recent meta-analysis found that growth at elevated O3 concentrations negatively impacts 

nearly every aspect of rice performance as well (Ainsworth, 2008, 191646). While rice is not a major 

crop in the U.S., it provides a staple food for over half of the global population (IRRI, 2002, 688833) 

and the effects of rising O3 concentrations on rice yields merits consideration. On average, rice 

yields decreased 14% in 62 ppb O3 compared to charcoal-filtered air. This yield loss was largely 

driven by a 20% decrease in grain number (Ainsworth, 2008, 191646). 

Feng and Kobayashi (2009, 199223) have recently compiled yield data for six major crop 

species, potato, barley, wheat, rice, bean and soybean and grouped the O3 treatments used in those 

studies into three categories: baseline O3 concentrations (<26 ppb), current ambient 7- or 12-h daily 

O3 concentrations (31-50 ppb), and future ambient 7- or 12-h daily O3 concentrations (51-75 ppb). 

Using these categories, they have effectively characterized the effects of current O3 concentrations 

and the effects of future O3 concentrations compared to the baseline O3 concentrations. At current O3 

concentrations, which ranged from 41-49 ppb in the studies included, soybean (-7.7%), bean 

(-19.0%), barley (-8.9%), wheat (-9.7%), rice (-17.5%) and potato (-5.3%) all reported yield losses 

compared to the baseline O3 concentrations (<26 ppb). At future O3 concentrations, averaging 

63 ppb, soybean (-21.6%), bean (-41.4%), barley (-14%), wheat (-28%), rice (-17.5%) and potato 

(-11.9%) all reported significantly larger yield losses compared to the losses at current O3 

concentrations (<26 ppb) (Feng and Kobayashi, 2009, 199223). 

An extensive review of OTC literature has determined the AOT40 critical level that causes a 

5% yield reduction across a variety of agricultural and horticultural species (Mills et al., 2007, 

180221). They classify the species into three groups: sensitive, moderate and tolerant. The sensitive 

crops, including watermelon, beans, cotton, wheat, turnip, onion, soybean, lettuce, and tomato, 

respond with a 5% reduction in yield under a 3-month AOT40 of 6 ppm-h. Watermelon was the most 

sensitive with a critical level of 1.6 ppm-h. The moderately sensitive crops, including sugar beet, 

oilseed rape, potato, tobacco, rice, maize, grape and broccoli, responded with a 5% reduction in yield 

between 8.6 and 20 ppm-h. The crops classified as tolerant, including strawberry, plum and barley, 

responded with a 5% yield reduction between 62-83.3 ppm-h (Mills et al., 2007, 180221). 

March 2011 9-57 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

5

 

 

10 

15 

 

 

 

20

 

25 

30

 

 

 

35 

1 Feng and Kobayashi (2009, 199223) compared their response-exposure results to those 

2 published by Mills et al. (2007, 180221) and found the ranges of yield loss to be similar for soybean, 

3 rice and bean. However, Feng and Kobayasi reported smaller yield losses for potato and wheat and 

4 larger yield losses for barley compared to the dose-response functions published by Mills et al. 

(2007, 180221), which they attributed to their more lenient criteria for literature inclusion. 

6 While the studies investigating the impact of various O3 concentrations on yield are important 

7 and aid in determining the vulnerability of various crops to a variety of O3 concentrations, there is 

8 still uncertainty as to how these crops will respond under field conditions with interacting 

9 environmental factors such as temperature, soil moisture, CO2 concentration, and soil fertility 

(Booker et al., 2009, 191569). Further, there appears to be a distinct developmental and genotype 

11 dependant influence on plant sensitivity to O3 that has yet to be fully investigated across O3 

12 concentrations in a field setting. 

13 Because O3 is heterogeneous in both time and space and O3 monitoring stations are 

14 predominantly near urban areas, the O3 impacts on current crop yields are difficult to estimate. 

Fishman et al. (2010, 644259) have used satellite observations to estimate O3 concentrations in the 

16 contiguous tri-state area of Iowa, Illinois and Indiana and have combined that information with other 

17 measured environmental variables to model the historical impact of O3 concentrations on soybean 

18 yield across the 2002-2006 growing seasons. When soybean yield across Iowa, Indiana and Illinois 

19 was modeled as a function of seasonal temperature, soil moisture and O3 concentrations, O3 had the 

largest contribution to the variability in yield for the southern-most latitudes included in the dataset. 

21 Fishman et al. (2010, 644259) determined that O3 concentrations significantly reduced soybean yield 

22 by -0.38 to -1.63% ppb/v across the 5 years. This value is consistent with previous chamber studies 

23 (NCLAN; Heagle, 1989, 093985) and results from SoyFACE (Morgan et al., 2006, 079186). 

24 Satellite estimates of tropospheric O3 concentrations exist globally (Fishman et al., 2008, 193927), 

therefore utilizing this historical modeling approach is feasible across a wider geographical area, 

26 longer time-span and perhaps for more crop species. 

9.5.3.2. Crop Quality 

27 In general, it appears that increasing O3 concentrations above current ambient concentrations 

28 can cause species dependant biomass losses, decreases in root biomass and nutritive quality, 

29 accelerated senescence and shifts in biodiversity. A study conducted with highbush blackberry has 

demonstrated decreased nutritive quality with increasing O3 concentration despite no change in 

31 biomass among a charcoal-filtered control, an ambient O3 and a 2 × ambient O3 treatment (Ditchkoff 

32 et al., 2009, 192230). A study conducted with sedge using control (30 ppb), low (55 ppb), medium 

33 (80 ppb) and high (105 ppb) O3 treatments has demonstrated decreased root biomass and accelerated 

34 senescence in the medium and high O3 treatments (Jones et al., 2010, 567354). Alfalfa showed no 

biomass changes across two years of double ambient O3 concentrations (AOT40 of 13.9 ppm-h) 

36 using FACE fumigation (Maggio et al., 2009, 191645). However a modeling study has demonstrated 

37 that 84% of the relative feed value in high-yielding alfalfa was due to the variability in mean O3 

March 2011 9-58 DRAFT – DO NOT CITE OR QUOTE 
 

1

2

3

4

5

6 

7

8

9

10 

11 

12 

13 

concentration from 1998-2002 (Lin et al., 2007, 196925). Further, in a managed grassland FACE 

system, the reduction in total biomass harvest over five years decreased twice as fast in the elevated 

treatment (AOT40 of 13-59 ppm-h) compared to ambient (AOT40 of 1-20.7 ppm-h). Compared with 

the ambient control, loss in annual dry matter yield was 23% after 5 year. Further, there was 

significant changes in the functional categories growing in each plot with legumes showing the 

strongest negative response (Volk et al., 2006, 191434). This study suggests a shift in biodiversity 

away from nitrogen-fixers in managed grasslands. An OTC study conducted with Trifolium 

subterraneum exposed to filtered ( <15 ppb), ambient, and 40 ppb above ambient O3 demonstrates 

decreases in biomass in the highest O3 treatment as well as 10-20% decreased nutritive quality which 

was mainly attributed to accelerated senescence (Sanz et al., 2005, 196963). A study conducted with 

Eastern gamagrass and big bluestem in OTCs suggests that big bluestem is not sensitive to O3, but 

gamagrass displayed decreased nutritive quality, due to higher lignin content and decreased N, in the 

2 × ambient O3 treatment (Lewis et al., 2006, 191542). 

March 2011 9-59 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

Table 9-3. Summary of recent studies of ozone effects on crops (exclusive of growth and yield) 

Species

Facility

Location 

Exposure

Duration 

Ozone Exposurea 

(Additional treatment) Variable(s) measured 

percent change from 

CFb 

(percent change from

ambient) 

Reference 

Alfalfa (Medicago 

sativa cv. Beaver) 

Growth chambers 

1, 2 or 

4 days 

3, 5 or 7 h/day

85 ppb 

(Exposure duration) 

Relative feed value n.s *high variability among 

treatment groups (N/A) 

Muntifering et al. 

(2006, 191270) 

Bean 

(Phaseolus vulgaris I. 

cv Borlotto) 

OTC, ground-planted 

Curno, Italy 

4 months 

Seasonal AOT40: 

CF = 0.5 ppm-h;

Ambient = 4.6 ppm-h 

(N/A) 

Seed lipid, 

Protein content 

Fiber content 

+28.5 (N/A) 

+7.88 (N/A) 

+14.54 (N/A) 

Iriti et al. (2009, 

195635) 

Big Blue Stem 

(Andropogon gerardii) 

OTC 4 months 

12-h avg:

CF = 14 ppb;

Ambient = 29 ppb;

Elevated = 71 ppb Relative feed value n.s. (n.s.) Lewis et al. (2006,

191542) 

Alabama, U.S. (N/A) 

Brassica napus 

Growth chambers 

Belgium 

4 days 

CF & 176 ppb 

for 4 h/day 

(N/A) 

Glucosinolates −41 (N/A) Gielen et al. (2006,

191271) 

Brassica napus

cv. Westar 

Growth chambers 

Finland 

17-26 

days 

8-h avg:

CF & 100 ppb 

(Bt/non-Bt;

herbivory) 

VOC emissions 

−30.7 (N/A);

−34 (N/A)

Himanen et al. 

(2009, 191338) 

Eastern Gamagrass

(Tripsacum

dactyloides) 

OTC 

4 months 

12-h avg:

CF = 14 ppb;

Ambient = 29 ppb;

Elevated = 71 ppb Relative feed value −17 (−12) Lewis et al. (2006,191542) 

Alabama, U.S. (N/A) 

Lettuce 

(Lactuca sativa) 

OTC 

Carcaixent 

30 days 

12-h mean: 

CF = 10.2 ppb;

NF = 30.1 ppb;

NF+O3 = 62.7 ppb 

Lipid peroxidation; 

Root length 

+77 (+38) 

−22 (−14)

Calatayud et. al. 

(2002, 684222) 

Experimental Station, 

Spain 

(4 cultivars) 

Peanut 

(Arachis hypogaea) 

OTC 3 yr 

12-h avg:

CF = 22 ppb;

Ambient = 46 ppb;

Elevated = 75 ppb Harvest biomass −40 (−10) Booker et al. (2007, 

191370) 

Raleigh, NC; U.S. (CO2: 375 ppm; 548 ppm; 

730 ppm) 

Poa pratensis 

OTC 

Braunschweig,

Germany 

3 yr;

4-5 wk 

in the 

spring 

8-h avg:

CF+25 = 21.7 ppb;

NF+50 = 73.1 ppb 

(Competition) 

Relative feed value N/A (n.s.; −8) Bender et al. (2006,

191437) 

Potato 

(Solanum tuberosum 

cv. Bintje 

OTC 

Sweden & Finland 

2 yr 

CF = 10 ppb;

Ambient = 25 ppb);

Ambient(+) = (36 ppb);

Ambient(++) = (47 ppb) 

(N/A) 

[K], [Ca], [Mg], [P], [N] per dry weight 

of tubers *dose-response regression, 

report significant positive or negative 

slope with increasing [O3] 

[N] [P] [Ca] n.s.;

[K] & [Mg] sig + 

(N/A) 

Piikki et al. (2007,

191451) 

March 2011 9-60 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

Species

Facility

Location 

Exposure

Duration 

Ozone Exposurea 

(Additional treatment) Variable(s) measured 

percent change from 

CFb 

(percent change from

ambient) 

Reference 

Potato 

(Solanum tuberosum

cv. Indira) 

Climate chambers 

Germany 

8 wk 

CF = 10 ppb;

Ambient = 50 ppb;

2×Ambient = 100 ppb 

(CO2: 400 ppm &

700 ppm) 

Pathogen infestation using % 

necrosis +52 (n.s.) Plessl et al. (2007,196952) 

AOT40: 

Soybean 

OTC 3 yr 

CF = 0 ppm-h;

Ambient = 3.4 ppm-h; 

Elevated = 9.0 ppm-h Daily

evapotranspiration −28 (−14) Jaude et al. (2008,191222) 

Italy (Well-watered & 

water-stressed) 

Soybean 

(Glycine max

cv. 93B15) 

SoyFACE 

Urbana, IL; U.S. 

3 yr

May-Oct 

AOT40: 

Ambient = 5-22 ppm-h; 

Elevated = 20-43 ppm-h 

(CO2: 550 ppm; 

environmental variability) 

Photosynthesis in new leaves, N/A (n.s.) Bernacchi et al. 

(2006, 158001) 

Soybean 

(Glycine max

cv. 93B15) 

SoyFACE 

Urbana, IL; U.S. 

4 months 

8-h avg:

Ambient = 38.5 ppb;

Elevated = 52 ppb 

(Herbivory) 

Herbivory

defense-related 

genes 

N/A (N/A) Casteel et al. 

(2008, 191696) 

Soybean 

(Glycine max

cv. Essex) 

OTC, ground-planted 

Raleigh, NC; U.S. 

2 yr 

12-h avg:

CF = 21 ppb;

1.5×Ambient = 74 ppb 

(CO2: 370 ppm &

714 ppm) 

Post-harvest residue N/A ( −15.46) Booker et al. (2005,

079151) 

Soybean 

(Glycine max

cv. Essex) 

OTCs, 21 L pots 

Raleigh, NC; U.S. 

2×3 

months 

12-h avg:

CF = 18 ppb);

Elevated = 72 ppb) 

(CO2: 367 & 718) 

Water-use efficiency n.s. (N/A) 

Booker et al. 

(Booker et al.,

2004, 079138) 

Soybean 

(Glycine max)

10 cultivars) 

SoyFACE 

Urbana, IL; U.S. 

2 yr 

8-h avg (ppb):

Ambient = 46.3 & 37.9; 

Elevated = 82.5 & 61.3 

(Cultivar comparisons) 

Total antioxidant capacity N/A (+19) Betzelberger et al.

(2010, 644183) 

Spring Wheat 

(Triticum aestivum 

cv. Minaret; Satu; 

Drabant; Dragon) 

OTCs 

Belgium, Finland, 

& Sweden 

7 yr 

Seasonal AOT40s 

ranged from

0 to16 ppm-h 

(N/A) 

Seed protein content; 

1,000-seed weight regressed across

all experiments 

N/A (Significant negative 

correlation) 

N/A (Significant negative 

correlation) 

Piikki et al. (2008, 

199812) 

Strawberry

(Fragaria x ananassa 

Duch. Cv. Korona 

& Elsanta) 

Growth chambers 

2 months 

8-h avg:

CF = 0 ppb;

Elevated = 78 ppb 

(N/A) 

Total leaf area −16 (N/A) Keutgen et al.

(2005, 191295) 

Bonn, Germany 

Sweet Potato 

Growth Chambers 4 wk 

8-h avg:

CF = 0 ppb;

Ambient < 40 ppb;

Elevated = 255 ppb Tuber weight −14 (−11.5) Keutgen et al.

(2008, 191690) 

Bonn, Germany (N/A) 

March 2011 9-61 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

Species

Facility

Location 

Exposure

Duration 

Ozone Exposurea 

(Additional treatment) Variable(s) measured 

percent change from 

CFb 

(percent change from

ambient) 

Reference 

Tomato 

(Lycopersicon 

esculentum) 

OTC 

Valencia, Spain 

133 days 

8- mean: 

CF = 16.3 ppb;

NF = 30.1 ppb;

NF(+) = 83.2 ppb 

(Various cultivars; 

early & late harvest) 

Brix degree −7.2 (−3.6) Calvo, et al. (2005,

191570) 

Trifolium repens & 

Trifolium pretense 

Aspen FACE 

Rhinelander, WI; U.S. 

3 months 

3-mo daylight avg:

Ambient = 34.8 ppb;

1.2×Ambient = 42.23 ppb 

(CO2; 560 ppm) 

Lignin; 

Dry-matter

digestibility 

N/A (+19.3) 

N/A ( −4.2)

Muntifering et al. 

(2006, 191270) 

a

Ozone exposure in ppb unless otherwise noted.
 b

CF = Carbon-filtered air.
 

NF = Non-filtered air.
 

9.5.4. Fa c to rs th at Mo dify Fu nc tio n al a nd Gro wth Re s p on s e 

1 Many biotic and abiotic factors, including insects, pathogens, root microbes and fungi, 

2 temperature, water and nutrient availability, and other air pollutants, as well as elevated CO2, 

3 influence or alter plant response to O3. These modifying factors were comprehensively reviewed in 

4 AX9.3 of the 2006 O3 AQCD (U.S. EPA, 2006, 088089) and thus, this section serves mainly as a 

5 brief summary of the previous findings. A limited number of new studies published since the 2006 

6 O3 AQCD add to our understanding of the role of these interactions in modifying O3-induced plant 

7 responses. Many of these modifying factors and interactions are integrated into discussions 

8 elsewhere in this chapter and the reader is directed to those sections. 

9.5.4.1. Genetics 

9 It is well known that species vary greatly in their responsiveness to O3. Even within a given 

10 species, individual genotypes or populations can also vary significantly with respect to O3 sensitivity 

11 (see section AX 9.3.2 of the 2006 AQCD; U.S. EPA, 2006, 088089). Therefore, caution should be 

12 taken when considering a species' degree of sensitivity to O3. Plant response to O3 is determined by 

13 genes that are directly related to oxidant stress and to an unknown number of genes that are not 

14 specifically related to oxidants, but instead control leaf and cell wall thickness, stomatal 

15 conductance, and the internal architecture of the air spaces. It is rarely the case that single genes are 

16 responsible for O3 tolerance. Studies using molecular biological tools and transgenic plants have 

17 positively verified the role of various genes and gene products in O3 tolerance and are continuing to 

18 increase the understanding of O3 toxicity and differences in O3 sensitivity. See Section 9.4.3.2 of this 

19 document for a discussion of recent studies related to gene expression changes in response to O3. 

March 2011 9-62 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

5

 

10 

 

 

15

20 

 

 

25 

 

30 

 

 

 

 

35

9.5.4.2. Environme ntal Biological Factors 

1 As stated in the 2006 O3 AQCD, the biological factors within the plant’s environment that may 

2 directly or indirectly influence its response to O3 in a positive or negative manner encompass insects 

3 and other animal pests, diseases, weeds, and other competing plant species. Ozone may influence the 

4 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, 

6 pathogen, or weed may influence the response of the target host species to O3 (U.S. EPA, 2006, 

7 088089). Several recent studies on the effects of O3 on insects via their interactions with plants are 

8 discussed in Section 9.6.6.1. In addition, O3 has also been shown to alter soil fauna communities 

9 (Section 9.6.6.2). 

In contrast to detrimental biological interactions, there are mutually beneficial relationships or 

11 symbioses involving higher plants and bacteria or fungi. These include (1) the nitrogen-fixing 

12 species Rhizobium and Frankia that nodulate the roots of legumes and alder and (2) the mycorrhizae 

13 that infect the roots of many crop and tree species, all of which may be affected by exposure of the 

14 host plants to O3. Some discussion of mycorrhizae can be found in Section 9.6.4. 

In addition to the interactions involving animal pests, O3 also has indirect effects on higher 

16 herbivorous animals, e.g., livestock, due to O3-induced changes in feed quality. Recent studies on the 

17 effects of O3 on nutritive quality of plants are discussed in Sections 9.5.3 and 9.6.6.3. 

18 Intra- and interspecific competition are also important factors in determining vegetation 

19 response to O3. Plant competition involves the ability of individual plants to acquire the 

environmental resources needed for growth and development: light, water, nutrients, and space. 

21 Intraspecific competition involves individuals of the same species, typically in monoculture crop 

22 situations, while interspecific competition refers to the interference exerted by individuals of 

23 different species on each other when they are in a mixed culture. This topic was previously reviewed 

24 in AX9.3.3.4 of the 2006 O3 AQCD (U.S. EPA, 2006, 088089). Recent studies on competition and its 

implications for community composition are discussed in Section 9.6.5. 

9.5.4.3. Phys ical Factors 

26 Physical or abiotic factors play a large role in modifying plant response to O3, and have been 

27 extensively discussed in previous O3 AQCDs (U.S. EPA, 1996, 080828)(U.S. EPA, 2006, 088089). 

28 This section summarizes those findings as well as recent studies published since the last review. 

29 Although some studies have indicated that O3 impact significantly increases with increased 

ambient temperature (Ball et al., 2000, 026354)(Mills et al., 2000, 030098), other studies have 

31 indicated that temperature has little effect (Balls et al., 1996, 026370)(Fredericksen et al., 1996, 

32 026653). A recent study by Riikonen et al. (2009, 195664) at the Ruohoniemi open air exposure field 

33 in Kuopio, Finland found that the effects of temperature and O3 on total leaf area and photosynthesis 

34 of Betula pendula were counteractive. Elevated O3 reduced the saplings’ ability to utilize the warmer 

growth environment by increasing the stomatal limitation for photosynthesis and by reducing the 

March 2011 9-63 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

1 redox state of ascorbate in the apoplast in the combination treatment as compared to temperature 

2 alone (Riikonen et al., 2009, 195664). 

3 Temperature affects the rates of all physiological processes based on enzyme catalysis and 

4 diffusion; each process and overall growth (the integral of all processes) has a distinct optimal 

5 temperature range. It is important to note that a plant’s response to changes in temperature will 

6 depend on whether it is growing near its optimum temperature for growth or near its maximum 

7 temperature (Rowland-Bamford, 2000, 030257). However, temperature is very likely an important 

8 variable affecting plant O3 response in the presence of the elevated CO2 levels contributing to global 

9 climate change. In contrast, some evidence suggests that O3 exposure sensitizes plants to low 

10 temperature stress (Colls and Unsworth, 1992, 026469) and, also, that O3 decreases below-ground 

11 carbohydrate reserves, which may lead to responses in perennial species ranging from rapid demise 

12 to impaired growth in subsequent seasons (i.e., carry-over effects) (Andersen et al., 1997, 052923). 

13 Light, a component of the plant’s physical environment, is an essential “resource” of energy 

14 content that drives photosynthesis and C assimilation. It has been suggested that increased light 

15 intensity may increase the O3 sensitivity of light-tolerant species while decreasing that of shade16 tolerant species, but this appears to be an oversimplification with many exceptions. Several studies 

17 suggest that the interaction between O3 sensitivity and light environment is complicated by the 

18 developmental stage as well as the light environment of individual leaves in the canopy (Chappelka 

19 and Samuelson, 1998, 093687)(Topa et al., 2001, 035398)(Kitao et al., 2009, 191331). 

20 Although the relative humidity of the ambient air has generally been found to increase the 

21 adverse effects of O3 by increasing stomatal conductance (thereby increasing O3 flux into the leaves), 

22 abundant evidence also indicates that the ready availability of soil moisture results in greater O3 

23 sensitivity (Mills, 2002, 035322). The partial “protection” against the adverse effects of O3 afforded 

24 by drought has been observed in field experiments (Low et al., 2006, 191396) and modeled in 

25 computer simulations (Broadmeadow and Jackson, 2000, 021325). Conversely, O3 may enhance the 

26 negative effects of O3 on plants (Grulke et al., 2003, 052984)(Pollastrini et al., 2010, 644392). There 

27 is also some evidence that O3 can predispose plants to drought stress (Maier-Maercker, 1998, 

28 029961). Hence, the nature of the response is largely species-specific and will depend to some extent 

29 upon the sequence in which the stressors occur. 

9.5.4.4. Inte ra c tio ns with o the r Polluta nts 

Ozone-Nitrogen Interactions 

30 Elevated O3 exposure and N deposition often co-occur as major pollutant types. However, the 

31 interactions of O3 exposure and N deposition on vegetation are complex and less well understood 

32 compared to their independent effects. Consistent with the conclusion of the 2006 O3 AQCD 

33 (U.S. EPA, 2006, 088089), studies published since the last review indicated that the interactive 

34 effects of N and O3 varied among species and ecosystems (Table 9-4). This section will focus on O3 

35 and N interactions at the plant scale. Responses at the ecosystem scale are discussed in Section 9.6. 

March 2011 9-64 DRAFT – DO NOT CITE OR QUOTE 
 

1

2 

3

4 

5 

6

7

8

9

10

11 

12 

13

14 

15

16 

17 

18

19 

20 

Nitrogen deposition could stimulate relative growth rate (RGR), and lead to increased stomatal 

conductance. Therefore, plants might become more susceptible to O3 exposure. Alternatively, N 

deposition may increase the availability of photosynthates for use in detoxification and plants could 

become more tolerant to O3 (Bassin et al., 2007, 196879). Only a few recent studies have 

investigated the interactive effects of O3 and N in the U.S. Grulke et al. (2005, 199433) measured 

stomatal conductance of California black oak (Quercus kelloggii) at a long-term N-enrichment site 

located in the San Bernardino Mountains, which is accompanied by high O3 exposure (80 ppb, 

24-h avg. over a six month growing season). The authors found that N amendment led to poor 

stomatal control in full sun in midsummer of the average precipitation years, but enhanced stomatal 

control in shade leaves of California black oak. In an OTC study, Handley and Grulke (2008, 

191485) found that O3 lowered photosynthetic ability and water-use efficiency, and increased leaf 

chlorosis and necrosis of California black oak. Nitrogen fertilization tended to reduce plant 

sensitivity to O3 exposure; however, the interaction was not statistically significant. 

Studies conducted outside the U.S. are also summarized in Table 9-4. Generally, the responses 

were species specific. The O3-induced reduction in photosynthetic rate and biomass loss were greater 

in the relatively high N treatment for watermelon (Citrillus lanants) (Calatayud et al., 2006, 191482) 

and Japanese beech (Fagus crenata) seedlings (Yamaguchi et al., 2007, 191438). However, there 

was no significant interactive effect of O3 and N on biomass production for Quercus serrata 

seedlings (Watanabe et al., 2007, 191474), young Norway spruce (Picea abies) trees (Thomas et al., 

2005, 075930), and young European beech (Fagus sylvatica) trees (Thomas et al., 2006, 191362). 

March 2011 9-65 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

Table 9-4. Response of plants to the interactive effects of elevated ozone exposure and N enrichment 

Site Species Ozone exposure N addition Responses	 References 

N-amended trees had lower late summer 

San C gain and greater foliar chlorosis in the California black Bernardino 0, and 50 kg N/ drought year, and poor stomatal control Grulke et al. (2005, oak (Quercus 80 ppb Mountains, ha/yr	 and lower leaf water use efficiency and in 199433) kelloggii) U.S.	 midsummer of the average precipitation 

year. 

San California black	 N fertilization tended to reduce plant Bernardino 0, and 50 kg N/ Handley and Grulke oak (Quercus 0, 75, and 150 ppb	 sensitivity to O3 exposure; however the Mountains, ha/yr	 (2008, 191485) kelloggii)	 interaction was not statistically significant. U.S. 

Higher nitrogen levels alleviated the spruce trees Filtered (19.4-28.1 ppb); 0, 20, 40 and 80	 Thomas et al.(2005, Switzerland	 negative impact of O3 on root starch (Picea abies) ambient (37.6-47.4 ppb) kg N/ ha/yr	 075930) concentrations 

Nitrogen addition amplified the negative beech trees Filtered (19.4-28.1 ppb); 0, 20, 40 and 80	 Thomas et al.(2006, Switzerland	 effects of O3 on leaf area and shoot (Fagus sylvatica) ambient (37.6-47.4 ppb) kg N/ ha/yr	 191362) elongation. 

ambient ( AOT40 of 11.1­ The positive effects of N addition on 12.6 ppm-h); 1.2 ambient 0, 5, 10’ 25, 50 canopy greenness were counteracted by Bassin et al.(2007, Switzerland Alpine pasture (AOT40 of 15.2-29.5 ppm-h) kg N/ ha/yr accelerated leaf senescence in the highest 191534) and 1.6 ambient (28.4-64.9 O3 treatment. ppm-h) 

ambient ( AOT40 of 11.1- Only a small number of species showed 

12.6 ppm-h); 1.2 ambient significant O3 and N interactive effects on 0, 5, 10, 25, 50	 Bassin et al.(2009, Switzerland Alpine pasture (AOT40 of 15.2-29.5 ppm-h) leaf chlorophyll concentration, leaf weight kg N/ha/yr	 191333) and 1.6 ambient (28.4-64.9 and change in 18O, and the patterns were 

ppm-h) not consistent. 

ambient ( AOT40 of 11.1­ The positive effects of N addition on 12.6 ppm-h); 1.2 ambient 0, 5, 10’ 25, 50 canopy greenness were counteracted by Bassin et al.(2007, Switzerland Alpine pasture (AOT40 of 15.2-29.5 ppm-h) kg N/ ha/yr accelerated leaf senescence in the highest 191534) and 1.6 ambient (28.4-64.9 O3 treatment. ppm-h) 

O3 free (AOT40 of 0 ppm-h), High N concentration enhanced the 

watermelon ambient (AOT40 of 5.1-6.3 140, 280, and detrimental effects of O3 on Chl a Calatayud et al.(2006, Spain (Citrillus lanants) ppm-h) and elevated O3 436 kg N/ ha/yr fluorescence parameters, lipid 191482)

(AOT40 of 32.5-35.6 ppm-h) peroxidation, and the total yield. 

Ozone reduced total aerial biomass. N 

Filtered (24-h avg. of 8-22 fertilization counterbalanced O3-induced 10, 30, and 60 Sanz et al. (2007, Spain Trifolium striatum ppb); ambient (29-34 ppb), effects only when plants were exposed to kg N/ ha/yr	 199245) elevated O3 (35-56 ppb)	 moderate O3 levels (ambient) but not 

under elevated O3 concentrations. 

Filtered (24-h avg. of 10.3­ The O3-induced reduction in net Japanese beech 13.2 ppb); ambient (42.0­ 0, 20 and 50 kg photosynthesis and whole-plant dry mass Yamaguchi et al.(2007, Japan seedlings (Fagus 43.3 ppb), 1.5 ambient (62.6­ N/ ha/yr were greater in the relatively high N 191438) crenata) 63.9 ppb) and 2.0 ambient treatment than that in the low N treatment. (82.7-84.7 ppb) 

Filtered (24-h avg. of 10.3­

13.2 ppb); ambient (42.0- No significant interactive effects of O3 and Quercus serrata 0, 20 and 50 kg Watanabe et al.(2007, Japan 43.3 ppb), 1.5 ambient (62.6- N load on the growth and net seedlings N/ ha/yr	 191474) 63.9 ppb) and 2.0 ambient photosynthetic rate were detected. 

(82.7-84.7 ppb) 

Ozone-Carbon Dioxide Interactions 

1 Several decades of research has shown that exposure to elevated CO2 increases photosynthetic 

2 rates (Tissue et al., 1999, 029011)(Bernacchi et al., 2006, 158001)(Will and Ceulemans, 1997, 

3 679719)(Tissue et al., 1997, 679718)(Bernacchi et al., 2005, 679713), decreases stomatal 

4 conductance (Bernacchi et al., 2006, 158001)(Ainsworth and Rogers, 2007, 092940)(Paoletti et al., 

5 2007, 199365)(Medlyn et al., 2001, 679716)(Leakey et al., 2006, 679714) and generally increases 

6 the growth of plants(McCarthy et al., 2010, 679715)(Norby et al., 2005, 679717). This is in contrast 

7 to the decrease on photosynthesis and growth in many plants that are exposed to elevated O3. The 

March 2011	 9-66 DRAFT – DO NOT CITE OR QUOTE 
 

1

2 

3

4 

5 

6

7 

8

9 

10

11

12

13 

14

15

16 

17

18 

19 

20

21

22 

23

24 

25 

26

27

28 

29 

30 

31 

32

33 

34 

35

36

37

38 

39

interactive effects on vegetation have been the subject of research in the past two decades because 

the implications on productivity and water use of ecosystems. This area of research was covered 

thoroughly in AX9.3.8.1 of the 2006 O3 AQCD and much of the conclusions made then are still 

relevant (U.S. EPA, 2006, 088089). 

The bulk of the available evidence 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 (ambient + 200 to 400 ppm) 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 studies, 

recent data show a long-term and sustained reduction in stomatal conductance under elevated CO2 

for a number of species (Ainsworth and Long, 2005, 042647)(Ellsworth et al., 2004, 

080092)(Gunderson et al., 2002, 080097). 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, 029961)(Grulke et al., 2007, 186963)(Paoletti and 

Grulke, 2010, 628561). 

Important caveats must be raised with regard to the findings presented in published research. 

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 (globally 

2 ppm/year) and may change little over several seasons of growth. On the other hand, O3 presents a 

fluctuating stressor with considerable hour-to-hour, day-to-day and regional variability (Polle and 

Pell, 1999, 093689). Almost all of the evidence presented comes from experimentation involving 

plants subjected to an abrupt step increase to a higher, steady CO2 concentration. In contrast, the O3 

exposure concentrations usually varied from day to day. Luo and Reynolds (1999, 035319), Hui et 

al. (2002, 035288), and Luo (2001, 035318) 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. It is also important to note that the levels of elevated CO2 in many of the studies will 

not be experienced in the field for 30 or 40 years, but elevated levels of O3 can occur in several areas 

of the U.S. Therefore, the CO2 × O3 interaction studies may be less policy relevant for current 

ambient conditions. 

Another caveat concerns the interactions of O3 and CO2 with other climatic variables, such as 

temperature and precipitation. In light of the key role played by temperature in regulating 

physiological processes and modifying plant response to increased CO2 levels (Long, 1991, 

029710)(Morison and Lawlor, 1999, 094194) and the knowledge that relatively modest increases in 

temperature may lead to dramatic consequences in terms of plant development (Lawlor, 1998, 

029015), it is important to consider that studying CO2 and O3 interactions alone may not create a 

complete understanding of effects on plants under future climate change. 

March 2011 9-67 DRAFT – DO NOT CITE OR QUOTE 
 

1

2

3

4 

5

6 

7 

8 

9

10 

11 

12 

13 

14 

15

16 

17

18

19 

20 

21

22 

23

24

25 

26

27 

28 

29 

30 

31

32

33

34

35

9.6. Effects of Ozone on Ecos ys tems and Services 

9.6.1. Ecos ys tem Scale, Function, and Structure 

Information presented in this section was collected at multiple scales, ranging from the 

physiology of a given species to population, community, and ecosystem-level investigations. For this 

assessment, “ecosystem” is defined as a functional entity consisting of interacting groups of living 

organisms and their abiotic (chemical and physical) environment. Ecosystems cover a hierarchy of 

spatial scales and can comprise the entire globe, biomes at the continental scale, or small, wellcircumscribed systems such as a small pond. 

Ecosystems have both structure and function. Structure may refer to a variety of measurements 

including the species richness, abundance, community composition and biodiversity as well as 

landscape attributes. Competition among and within species and their tolerance to environmental 

stressors are key elements of survivorship. When environmental conditions are shifted, for example, 

by the presence of anthropogenic air pollution, these competitive relationships may change and 

tolerance to stress may be exceeded. “Function” refers to the suite of processes and interactions 

among the ecosystem components and their environment that involve nutrient and energy flow as 

well as other attributes including water dynamics and the flux of trace gases. Plant processes 

including photosynthesis, respiration, C allocation, nutrient uptake and evaporation, are directly 

related to functions of energy flow and C, nutrient and water cycling. The energy accumulated and 

stored by vegetation (via photosynthetic C capture) is available to other organisms. Energy moves 

from one organism to another through food webs, until it is ultimately released as heat. Nutrients and 

water can be recycled. Air pollution alters the function of ecosystems when elemental cycles or the 

energy flow are altered. This alteration can also be manifested in changes in the biotic composition 

of ecosystems. 

There are at least three levels of ecosystem response to pollutants: (1) the individual organism 

and its environment; (2) the population and its environment; and (3) the biological community 

composed of many species and their environment (Billings, 1978, 034165). Individual organisms 

within a population vary in their ability to withstand the stress of environmental change. The 

response of individual organisms within a population is based on their genetic constitution, stage of 

growth at time of exposure to stress, and the microhabitat in which they are growing (Levine and 

Pinto, 1998, 029599). The stress range within which organisms can exist and function determines the 

ability of the population to survive. Those best able to cope with environmental stressors survive and 

reproduce. Competition among different species results in succession (community change over time) 

and, ultimately, sensitive species may be progressively replaced and communities shift to favor those 

species that may have the capability to tolerate stressors such as O3 (Guderian, 1985, 

019325)(Rapport and Whitford, 1999, 004595). In the sections that follow, available information on 

individual, population and community response to O3 will be discussed. Effects of O3 on productivity 

and C sequestration, water cycling, below-ground processes, competition and biodiversity, and 

March 2011 9-68 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

1 insects and wildlife are considered below and in the context of ecosystem services where 

2 appropriate.
 

9.6.1.1. Ecos ys te m Services 

3 Ecosystem structure and function may be translated into ecosystem services. Ecosystem 

4 services identify the varied and numerous ways that ecosystems are important to human welfare. 

5 Ecosystems provide many goods and services that are of vital importance for the functioning of the 

6 biosphere and provide the basis for the delivery of tangible benefits to human society. Hassan et al. 

7 (2005, 092759) define these benefits to include supporting, provisioning, regulating, and cultural 

8 services: 

9  Supporting services are necessary for the production of all other ecosystem services. 

10 Some examples include biomass production, production of atmospheric O2, soil 

11 formation and retention, nutrient cycling, water cycling, and provisioning of habitat. 

12 Biodiversity is a supporting service that is increasingly recognized to sustain many of the 

13 goods and services that humans enjoy from ecosystems. These provide a basis for three 

14 higher-level categories of services. 

15  Provisioning services, such as products (Gitay et al., 2001, 092761), i.e., food (including 

16 game, roots, seeds, nuts and other fruit, spices, fodder), fiber (including wood, textiles), 

17 and medicinal and cosmetic products (such as aromatic plants, pigments). 

18  Regulating services that are of paramount importance for human society such as 

19 (1) C sequestration, (2) climate and water regulation, (3) protection from natural hazards 

20 such as floods, avalanches, or rock-fall, (4) water and air purification, and (5) disease and 

21 pest regulation. 

22  Cultural services that satisfy human spiritual and aesthetic appreciation of ecosystems 

23 and their components. 

9.6.1.2. As s e s s ing Ozone Effe c ts a t La rge r S pa tia l Sc a le s 

24 Ozone effects at large spatial scales start as effects on plants at smaller spatial scales. Ozone enters 

25 leaves through stomata, and has been shown to alter stomatal conductance and reduce the activity 

26 and concentration of Rubisco (Section 9.4.6). Those changes alter the rates of CO2 uptake and water 

27 loss from leaves, and therefore the rates of photosynthesis and transpiration at the plant level 

28 (Section 9.5.2). Those O3-induced effects could translate from the plant level to the ecosystem level, 

29 and cause changes in ecosystem services, such as C storage, water production, nutrient cycling, and 

March 2011 9-69 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

5 

 

 

10 

 

 

15 

 

 

 

20 

 

25

 

 

 

30 

1 community composition (Figure 9-1). Changes at the ecosystem level are difficult to evaluate 

2 directly due to the complexity and the large spatial and temporal scale of ecosystems. These 

3 assessments generally involve the extrapolation of laboratory or field results by ecological models or 

4 long-term field experiments as discussed below. 

9.6.2. Productivity and Carbon Seques tration 

During the previous NAAQS review, there were very few studies that investigated the effect of 

6 O3 exposure on ecosystem productivity and C sequestration. Recent studies from long-term FACE 

7 experiments provide new evidence of the association of O3 exposure and changes in productivity at 

8 the ecosystem level. In addition to experimental studies, several model studies also assessed the 

9 impact of O3 exposure on productivity and C sequestration from stand to global scales. Three types 

of models are most often used to study the ecological consequences of O3 exposure: (1) regression 

11 models such as dose-response function derived from the Nation Crop Loss Assessment Network 

12 (NCLAN) (Wang and Mauzerall, 2004, 179978)(Tong and Mauzerall, 2008, 621169); (2) tree growth 

13 models such as TREGRO and ECOPHYS (Hogsett et al., 2008, 191229)(Martin et al., 2001, 

14 043678); and (3) process-based ecosystem models such as PnET, Dynamic Land Ecosystem Model 

(DLEM) and Terrestrial Ecosystem Model (TEM) (Ollinger et al., 2002, 180189)(Ren et al., 2007, 

16 191366)(Felzer et al., 2009, 191460). The experimental and model studies on ecosystem productivity 

17 and C sequestration, at the stand scale as well as regionally and globally, are reviewed in the 

18 following section. 

9.6.2.1. Stand Scale 

19	 The above- and below-ground biomass and net primary production (NPP) were measured at 

the Aspen FACE site after 7-year O3 exposure. Elevated O3 caused 23, 13 and 14% reductions in 

21 total biomass relative to the control in the aspen, aspen–birch and aspen–maple communities, 

22 respectively (King et al., 2005, 191701). At the Kranzberg Forest FACE experiment, O3 reduced 

23 annual volume growth by 9.5 m3/ha in a mixed mature stand of Norway spruce and European beech 

24 (Pretzsch et al., 2010, 580435). Ozone also altered C accumulation and turnover in soil, and the 

details of these studies are discussed in Section 9.6.4. 

26 Changes in stand productivity under elevated O3 were assessed by several model studies. 

27 TREGRO is a process-based, single tree growth model and has been widely used to simulate the 

28 effects of O3 on the growth of several species in different regions in the U.S. Ozone acts within the 

29 model by reducing the maximum potential photosynthetic rate as a function of the cumulative uptake 

of O3 (Weinstein et al., 1991, 043993)(Tingey et al., 2004, 042385). Hogsett et al. (2008, 191229) 

31 used TREGRO to evaluate the effectiveness of various forms and levels of air quality standards for 

32 protecting tree growth in the San Bernardino Mountains of California. They found that O3 exposures 

33 at the Crestline site resulted in a mean 20.9% biomass reduction from 1980 to 1985 and 10.3% 

34 biomass reduction from 1995 to 2000, compared to the ‘‘background’’ O3 concentrations (O3 

March 2011	 9-70 DRAFT – DO NOT CITE OR QUOTE 
 

 

5

 

 

10 

 

 

15

 

 

20 

 

 

 

25

 

 

30 

 

 

35

1 concentration in Crook County, Oregon). The level of vegetation protection projected was different 

2 depending on the air quality scenarios under consideration. Specifically, when air quality was 

3 simulated to just meet the California 8 h average maximum of 70 ppb and the maximum 3 months 

4 12-h SUM06 of 25 ppm-h, annual growth reductions were limited to 1% or less, while air quality 

that just met a previous NAAQS (the second highest 1-h max [125 ppb]) resulted in 6-7% annual 

6 reduction in growth, resulting in the least protection relative to background O3 (Hogsett et al., 2008, 

7 191229). 

8 Combining TREGRO with ZELIG, Weinstein et al. (2005, 179965) simulated the effects of 

9 different O3 levels ( 0.5, 1.5, 1.75, and 2 times ambient) on the growth and competitive interactions 

of white fir and ponderosa pine at three sites in California: Lassen National Park, Yosemite National 

11 Park, and Crestline. Their results suggested that O3 had little impact on white fir, but greatly reduced 

12 the growth of ponderosa pine. If current O3 concentrations continue over the next century, ambient 

13 O3 exposure (SUM06 of 110 ppm-h) at Crestline was predicted to decrease individual tree C budget 

14 by 10% and decrease ponderosa pine abundance by 16%. Effects at Lassen National Park and 

Yosemite National Park sites were found to be smaller because of lower O3 exposure levels 

16 (Weinstein et al., 2005, 179965). 

17 The effects of O3 on stand productivity and dynamics were also studied by other tree growth or 

18 stand models, such as ECOPHYS, INTRAST and LINKAGES. ECOPHYS is a functional-structural 

19 tree growth model. The model used the linear relationship between the maximum capacity of 

carboxylation and O3 dose to predict the relative effect of O3 on leaf photosynthesis (Martin et al., 

21 2001, 043678). Simulations with ECOPHYS found that O3 decreased stem dry matter production, 

22 stem diameter and leaf dry matter production, induced earlier leaf abscission, and inhibited root 

23 growth (Martin et al., 2001, 043678). Simulation with INTRAST and LINKAGES showed similar 

24 adverse effects on stand growth. Linking INTRAST with LINKAGES, Hanson et al. (2005, 191461) 

found that a simulated increase O3 concentration in 2100 (a mean 20-ppb increase over the current 

26 O3 concentration) yields a 35% loss of net ecosystem C exchange (NEE) with respect to the current 

27 conditions (174 g C/m2/year). 

9.6.2.2. Regional and Global Scales 

28 Since the publication of the 2006 O3 AQCD, there is additional evidence suggesting that O3 

29 exposure alters ecosystem productivity and biogeochemical cycling at the regional and continentalscale. Most of those studies were conducted by using process-based ecosystem models (Table 9-5) 

31 and are briefly reviewed in the following sections. 

Carbon Dynamics in Natural Ecosystems 

32 Results of Reich (1987, 019314) and Tjoelker et al. (1995, 035394) indicated that O3 effects 

33 on photosynthesis can be determined as a function of O3 uptake to internal leaf surface. Pooling data 

34 from Reich (1987, 019314) and Tjoelker et al. (1995, 035394), Ollinger et al. (1997, 040707) 

derived an equation to simulate the O3 effects on photosynthesis by cumulative O3 dose (AOT40) 

March 2011 9-71 DRAFT – DO NOT CITE OR QUOTE 
 

1

2 

3

4

5

6

7

8

9 

10 

11

12 

13 

14

15 

16

17 

18 

19

20 

21 

22 

23 

24

25 

26 

27 

28

29

30 

31 

32 

33

34 

35

36 

37 

38 

and stomatal conductance. They incorporated this equation into the PnET-II model and studied the 

effect of O3 on hardwood forest productivity of 64 hardwood sites in northeastern U.S. Their model 

indicated that O3 caused a 3-16% reduction in NPP from 1987 to 1992 (Table 9-5). Ollinger et al. 

(2002, 180189) assessed the interactive effects of O3, N deposition, elevated CO2 and land use 

history on C dynamics by PnET-CN. Their results indicated that O3 offset the increase in net 

C exchange caused by elevated CO2 and N deposition by 13% (25.0 g C/m2

/year) under agriculture 

site history, and 23% (33.6 g C/m2/year) under timber harvest site history. PnET-CN was also used to 

assess changes in C sequestration of U.S. Mid-Atlantic temperate forest. Pan et al. (2009, 596032) 

designed a factorial modeling experiment to separate the effects of changes in atmospheric 

composition, historical climatic variability and land-disturbances on the C cycle. They also found O3 

acted as a negative factor, partially offsetting the growth stimulation caused by elevated CO2 and N 

deposition of U.S. Mid-Atlantic temperate forest. Ozone decreased NPP of most forest types by 7­

8%. Among all the forest types, spruce-fir forest was most resistant to O3 damage, and NPP 

decreased by only 1% (Pan et al., 2009, 596032). 

Felzer et al. (2004, 186927) developed TEM 4.3 to simulate the effects of O3 on plant growth. 

The effects of O3 on NPP and C sequestration of deciduous trees, conifers and crops in the 

conterminous U.S. were estimated by TEM. The results indicated that O3 reduced NPP and 

C sequestration in the U.S. (Table 9-5) and the largest decreases (over 13% in some locations) in 

NPP occurred in the Midwest agricultural lands during the mid-summer. TEM was also used to 

evaluate the magnitude of O3 damage at a global scale (Table 9-5) (Felzer et al., 2005, 186928). 

Simulations for the historical period (1860-1995) show that the largest reductions in NPP and net 

C exchange occurred in the mid western U.S., eastern Europe, and eastern China (Felzer et al., 2005, 

186928). DLEM was developed to simulate the detrimental effect of O3 on ecosystems, and has been 

used to examine the O3 damage on NPP and C sequestration in Great Smoky Mountains National 

Park (Zhang et al., 2007, 196983), grassland ecosystems and terrestrial ecosystems in China (Ren et 

al., 2007, 581541)(Ren et al., 2007, 191366). Results of those simulations are listed in Table 9-5. 

Instead of using AOT40 as their O3 exposure metric as PnET, TEM and DLEM did, Sitch et al. 

(2007, 093294) incorporated a different O3 metric named CUOt (cumulative stomatal uptake of O3), 

derived from Pleijel et al. (2004, 056608), into the MOSES-TRIFFID coupled model. In the CUOt 

metric, the fractional reduction of plant production is dependent on O3 uptake by stomata over a 

critical threshold for damage and this threshold level is different for different plant functional types. 

Consistent with previous studies, their model simulation indicated that O3 reduced global gross 

primary production (GPP), C exchange rate and C sequestration (Table 9-5). The largest reductions 

in GPP and land-C storage were projected over North America, Europe, China and India. In the 

model, reduced ecosystem C uptake due to O3 damage, results in additional CO2 accumulation in the 

atmosphere and an indirect radiative forcing of climate change. Their simulations indicated that the 

indirect radiative forcing caused by O3 (0.62-1.09 W/m2

) could have even greater impact on global 

warming that the direct radiative forcing of O3 (0.89 W/m2

) (Sitch et al., 2007, 093294). 

March 2011 9-72 DRAFT – DO NOT CITE OR QUOTE 
 

1

2 

3

4

5 

6 

7

8 

9 

10 

11 

Crop Yield Loss 

Two large scale field studies were conducted in the U.S. (NCLAN) and in Europe (European 

Open Top Chamber Programme, EOTCP) to assess the impact of O3 on crop production. Ozone 

exposure-response regression models derived from the two programs have been widely used to 

estimate crop yield loss (Wang and Mauzerall, 2004, 179978)(Tong and Mauzerall, 2008, 

621169)(Van Dingenen et al., 2009, 199765). Those studies found that O3 generally reduced crop 

yield and different crops showed different sensitivity to O3 pollution (Table 9-5). Ozone was 

calculated to induce a possible 45-82 million metric tons loss for wheat globally. Production losses 

for rice, maize and soybean were on the order of 17-23 million metric tons globally (Van Dingenen 

et al., 2009, 199765). The largest yield losses occur in high-production areas exposed to high O3 

concentrations, such the Midwest and the Mississippi Valley regions in the U.S., Europe, China and 

India (Van Dingenen et al., 2009, 199765)(Tong et al., 2007, 107431). 

March 2011 9-73 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

 

 

 

 

 

 

Table 9-5. The effects of ozone on primary production, C exchange, C sequestration and yield loss 

Scale Model Index Ozone Impacts	 Reference 

GPP Global MOSES- CUOta Decreased by 14-23% over the period 1901-2100 Sitch et al. (2007, 

TRIFFID 093294) 

NPP Global TEM AOT40 Decreased by 0.8% without agricultural management and a Felzer et al. (2005, 

decrease of 2.9% with optimal agricultural management 186928) 

U.S. TEM AOT40 Reduced by 2.3% without optimal N fertilization and 7.2% Felzer et al. (2005, 

with optimal N fertilization from 1983-1993 186928) 

U.S. TEM AOT40 Reduced by 2.6–6.8% during the late 1980s-early 1990s. Felzer et al. (2004, 

186927) 

northeastern PnET AOT40 A reduction of 3-16% from 1987-1992	 Ollinger et al. 

U.S.	 (1997, 040707) 

U.S. Mid- PnET AOT40 Decreased NPP of most forest types by 7-8% Pan et al. (2009, 

Atlantic 596032) 

China DLEM AOT40 Reduced NPP of grassland in China by 8.5 Tg C from Ren et al. (2007, 

1960s to 1990s 581541) 

C exchange Global TEM AOT40	 Reduced net C exchange (1950–1995) by 0.1 Pg C/yr Felzer et al. (2005, 

without agricultural management and 0.3 Pg C/yr with 186928) 

optimal agricultural management 

Global MOSES- CUOt Decreased global mean land–atmosphere C fluxes by 1.3 Sitch et al. (2007, 

TRIFFID Pg C/yr and 1.7 Pg C/yr for the ‘high’ and ‘low’ plant O3 093294)

sensitivity models, respectively 

C sequestration Global MOSES- CUOt Reduced land-C storage accumulation by between 143 Pg C/yr Sitch et al. (2007, 

TRIFFID and 263 Pg C/yr from 1900–2100 093294) 

U.S. TEM AOT40 Reduced C sequestration by 18–38 Tg C/yr from 1950 to Felzer et al. (2004, 

1995 186927) 

GSM National DLEM AOT40 Decreased the ecosystem C storage of deciduous forests Zhang et al. (2007, 

Park by 2.5% and pine forest by 1.4% from 1971 to 2001 196983) 

China DLEM AOT40 Reduced total C storage by 0.06% in 1960s and 1.6% in Ren et al. (2007, 

1990s in China’s terrestrial ecosystems 191366) 

Crop yield loss Global	 Dose– M7b; Reduced by 7.3% to 12.3% for wheat, 5.4% to 15.6% for Van Dingenen et al. 

response M12c; soybean, 2.8% to 3.7% for rice, and 2.4% to 4.1% for maize (2009, 199765)

function AOT40 in year 2000. 

U.S.	 Dose– M7; M12; Reduced by 4.1% to 4.4% for wheat, 7.1% to 17.7% for Van Dingenen et al. 

response AOT40 soybean, 2.6% to 3.2% for rice, and 2.2% to 3.6% for maize (2009, 199765)

function in year 2000. 

U.S.	 Dose– SUM06 Caused a loss of 53.8 million to 438 million bushels in Tong et al. (2007, 

response soybean production, which account for 1.7–14.2% of total 107431)

function U.S. soybean production in 2005 

East Asia	 Dose– M7; M12 Reduced the yield of wheat, rice and corn by 1–9% and Wang and 

response soybean by 23–27% in China, Japan and South Korea in Mauzerall (2004, 

function 1990 179978) 

a

CUOt is defined as the cumulative stomatal uptake of O3, using a constant O3-uptake rate threshold of t nmol/m2/s.
 b

M7 is defined as 7-h mean O3 concentration (ppb).
 c

M12 is defined as 12-h mean O3 concentration (ppb).
 d

Pg equals 1 × 1015 grams.
 

9.6.3. Wa te r Cyclin g 

1 Ozone has been shown to alter stomatal performance, which affects plant and stand 

2 transpiration and therefore hydrological cycling. However, there is not a clear consensus on stomatal 

3 response to O3 exposure. A meta-analysis found that O3 reduced stomatal conductance by 11% 

4 (Wittig et al., 2007, 191695). Stomatal closure could help protect the plant from water loss. On the 

5 other hand, a number of studies suggested that O3 exposure could impair stomatal function, which 

6 leads to greater stomatal apertures, delays stomatal closure at night and results in higher transpiration 

7 (Grulke et al., 2004, 042646)(McLaughlin et al., 2007, 090348)(McLaughlin et al., 2007, 

March 2011	 9-74 DRAFT – DO NOT CITE OR QUOTE 
 

1 

2 

3 

4 

5

6 

7

8

9

10

11

12 

13

14 

15

16 

17

18

19 

20

21

22

23 

24

25

26

27

28

29 

30 

31

32 

33 

34

35

36

37 

38

39 

090347)(Mills et al., 2009, 191272)(Wilkinson and Davies, 2009, 199758)(Wilkinson and Davies, 

2010, 598245). This O3-induced impairment of stomatal control may be more pronounced for plants 

growing under drought stress (McLaughlin et al., 2007, 090348)(McLaughlin et al., 2007, 

090347)(Wilkinson and Davies, 2010, 598245). The discrepancy regarding stomatal function has 

lead to further debate on the effects of O3 on ecosystem water production. 

Felzer et al. (2009, 191460) used TEM-Hydro to assess the interactions of O3, climate, 

elevated CO2 and N limitation on the hydrological cycle in the eastern U.S. They found that elevated 

CO2 decreased evapotranspiration by 2-4% and increased runoff by 3-7%, as compared to the effects 

of climate alone. When O3 damage and N limitation were included, evapotranspiration was reduced 

by an additional 4-7% and runoff was increased by an additional 6-11% (Felzer et al., 2009, 

191460). Based upon simulation with INTRAST and LINKAGES, Hanson et al. (2005, 191461) 

found that increasing O3 concentration by 20 ppb above the current ambient level yields a modest 

3% reduction in water use. Those ecological models were generally built on the assumption that O3 

induces stomatal closure. Therefore, results of those models normally found that O3 reduced water 

use. 

In contrast to the model simulations (Hanson et al., 2005, 191461)(Felzer et al., 2009, 

191460), field studies conducted by McLaughlin et al. (2007, 090348)(2007, 090347) indicated that 

O3 increases water use in a mixed deciduous forest in eastern Tennessee. McLaughlin et al. (2007, 

090348)(2007, 090347) found that O3, with daily maximum levels ranging from 69.2 to 82.9 ppb, 

reduced stem growth by 30-50% in the high-O3 year 2002. The decrease in growth rate was caused 

in part by amplification of diurnal cycles of water loss and recovery. Peak hourly O3 exposure 

increased the rate of water loss through transpiration as indicated by the increased stem sap flow. 

The increased canopy conductance resulted in higher water uptake as reflected in the reduced soil 

moisture in the rooting zone. The change in tree water use led to further impacts on the hydrological 

cycle at the landscape level. Increased water use under high O3 exposure was reported to reduce lateseason modeled streamflow in three forested watersheds in eastern Tennessee (McLaughlin et al., 

2007, 090347). 

In addition to the impacts on stomatal performance, O3-induced physiological changes, such as 

reduced leaf area index and accelerated leaf senescence, could alter water use efficiency. At the 

Aspen FACE experiment, stand-level water use, as indicated by sap flux per unit ground area, was 

not significantly affected by elevated O3 despite a 22% decrease in leaf area index and 20% decrease 

in basal area (Uddling et al., 2008, 191655). Several factors could contribute to the lack of negative 

effect of elevated O3 on stand water use. The maximum sap flux per unit total leaf area was 

substantially increased by elevated O3, suggesting that whole-plant hydraulic conductance per unit 

leaf area was increased (Uddling et al., 2009, 596219). Other potential contributing factors included 

the higher proportion of sun leaves, and similar or even increased fine root biomass under elevated 

O3 (Uddling et al., 2008, 191655). Elevated O3 could also affect evapotranspiration by altering tree 

crown interception of precipitation. Ozone has been shown to change branch architectural 

parameters, and the effects were species specific at the Aspen FACE experiment (Rhea et al., 2010, 

March 2011 9-75 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

1 647084). The authors found that there was a significant correlation between canopy architecture
 

2 parameters and stem flow for birch but not aspen.
 

9.6.4. Below-Ground Proces s es 

3 Above-ground and below-ground processes are tightly interconnected. Because roots and soil 

4 organisms are not exposed directly to O3, below-ground processes are affected by O3 more through 

5 altering the quality and quantity of C supply from photosynthates and litterfall (Andersen, 2003, 

6 041673). Ozone can decrease leaf C uptake by altering stomatal function, reducing the activity and 

7 concentration of Rubisco, and accelerating leaf senescence (Section 9.4). Ozone can also increase 

8 the metabolic costs by stimulating the production of chemical compounds for defense and repair 

9 processes, and increasing the synthesis of antioxidants to neutralize free radicals (see Section 9.4), 

10 which increase the consumption of carbon for above-ground processes. Therefore, O3 could 

11 significantly reduced the amount of C available for allocation to below-ground by decreasing 

12 C uptake but increasing C consumption of above-ground processes (Andersen, 2003, 041673). 

13 Since the 2006 O3 AQCD, there is additional evidence for O3 effects on below-ground 

14 processes. Ozone has been found to alter root growth, soil food web structure, decomposer activities, 

15 C turnover and nutrient flow (Figure 9-7). Ozone effects on root development and root biomass 

16 production (Section 9.5.2.1) and soil food web structure (Section 9.6.5.3) are reviewed in other 

17 sections. The focus in this section is on the response of litter input, decomposer activities, soil 

18 respiration, soil C formation and nutrient cycling. 

March 2011 9-76 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

 

 

 

 

Source: Used with permission from Andersen (2003, 041673) 

Figure 9-7. Conceptual diagram showing where ozone disrupts C flow in a tree-soil system,

including transfer between biotic and abiotic components below ground that

influence soil physical and chemical properties. ﷒Arrows denote C flux pathways that 

are affected by ozone. Dashed lines indicate where the impact of ozone is suspected

but unknown. 

9.6.4.1. Litter Carbo n Che mis try, Litter Nu trien t an d Their Ecos ys tem Bu dge ts 

1 Consistent with previous findings, recent studies show that, although the responses are often 

2 species specific, O3 tends to alter litter chemistry (U.S. EPA, 2006, 088089).Alterations in chemical 

3 parameters, such as changes in C chemistry and nutrient concentrations, were observed in both leaf 

4 and root litter (Table 9-6). 

5 At the Aspen FACE site, several studies investigated litter chemistry changes (Chapman et al., 

6 2005, 191345)(Liu et al., 2005, 187005)(Johnson and Pregitzer, 2007, 191287)(Parsons et al., 2008, 

7 191853). In both aspen and birch leaf litter, elevated O3 increased the concentrations of soluble 

8 sugars, soluble phenolics and condensed tannins (Liu et al., 2005, 187005)(Parsons et al., 2008, 

9 191853). Compared to other treatments, aspen litter under elevated O3 had the highest fiber 

10 concentration, with the lowest concentration associated with the birch litter under the same 

11 conditions (Parsons et al., 2008, 191853). Chapman et al. (2005, 191345) measured chemical 

12 changes in fine root litter and found that elevated O3 decreased lignin concentration. The O3-induced 

March 2011 9-77 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

1 chemistry changes were also reported from other experimental sites. Results from an OTC study in 

2 Finland suggested that elevated O3 increased the concentration of acid-soluble lignin, but had no 

3 significant impact on other chemicals such as total sugars, hemicelluloses, cellulose or total lignin in 

4 the litter of silver birch (Kasurinen et al., 2006, 191269). Results from the free air canopy O3 

5 exposure experiment at Kranzberg Forest, showed that O3 increased starch concentrations but had no 

6 impact on cellulose and lignin in beech and spruce leaf litter (Aneja et al., 2007, 191472). The effect 

7 of O3 on three antioxidants (ascorbate, glutathione and α-tocopherol) in fine roots of beech was also 

8 assessed at Kranzberg Forest. The results indicated that O3 had no significant effect on α-tocopherol 

9 and ascorbate concentrations, but decreased glutathione concentrations in fine roots (Haberer et al., 

10 2008, 191334). In addition to changing C chemistry, O3 also altered nutrient concentrations in green 

11 leaves and litter (Table 9-6). 

12 The combined effects of O3 on biomass productivity and chemistry changes could alter 

13 C chemicals and nutrient contents at the canopy or ecosystem level. For example, although O3 had 

14 different impacts on their concentrations, annual fluxes of C chemicals (soluble sugar, soluble 

15 phenolics, condensed tannins, lipid and hemicelluloses), macro nutrients (N, P, K and S) and micro 

16 nutrients (Mg, B, Cu and Zn) to soil were all reduced due to lower litter biomass productivity at 

17 Aspen FACE (Liu et al., 2005, 187005)(Liu et al., 2007, 093286). At the Kranzberg Forest, N 

18 content of spruce canopy in a mixed culture and Ca2+ content of beech canopy in a monoculture 

19 increased due to elevated O3 increased leaf concentrations of those nutrients although leaf production 

20 was not significantly altered by O3 (Rodenkirchen et al., 2009, 191540). 

Table 9-6. The effect of elevated ozone on leaf/litter nutrient concentrations 

Study Site Species Ozone Concentration Response Reference 

Suonenjoki Research Silver birch Ambient: 10-60 ppb Decreased the concentration of P, Mn, Zn Kasurinen et al.
 

Station, Finland Elevated: 2×ambient and B in leaf litter (2006, 191269)
 

Aspen FACE Aspen and birch Ambient: 50-60 ppb Decreased the concentrations of P, S, Ca Liu et al. (2007, 

Elevated: 1.5×ambient and Zn, but had no impact on the 093286) 

concentrations of N, K, Mg, Mn, B and Cu in 

leaf litter. 

Aspen FACE Birch Ambient: 50-60 ppb Increase N concentration in birch litter Parsons et al. (2008, 

Elevated: 1.5×ambient 191853) 

Kranzberg Forest, Germany Beech and spruce Ambient: 9-41 ppb Increased N concentration in beach leaf, but Kozovits et al. 

Elevated: 2×ambient not in spruce needle (2005, 191282) 

Kranzberg Forest, Germany Beech and spruce Ambient: 9-41 ppb 1) Had no significant effects on spruce Rodenkirchen et al. 

Elevated: 2×ambient needle chemistry; 2) increased Ca (2009, 191540) 

concentration in beech leaves in 

monoculture, but had no impacts on other 

nutrients 

9.6.4.2. De c o mp os er Meta bolis m a nd Litter Decompos ition 

21 The above- and below-ground physiological changes caused by O3 exposure cascade through 

22 the ecosystem and affect soil food webs. In the 2006 O3 AQCD, there were very few studies on the 

23 effect of O3 on the structure and function of soil food webs, except two studies conducted by Larson 

24 et al. (2002, 053015) and Phillips et al. (2002, 041768). Since the last O3 AQCD (U.S. EPA, 2006, 

March 2011 9-78 DRAFT – DO NOT CITE OR QUOTE 
 

 

5 

 

 

 

10

 

 

15 

 

20

25

 

30 

 

 

35

 

1 088089), new studies have provided more information on how O3 affects the metabolism of soil 

2 microbes and soil fauna. 

3 Chung et al.(2006, 191729) found that the activity of the cellulose-degrading enzyme 1,4-β4 glucosidase was reduced by 25% under elevated O3 at AspenFACE. The decrease in cellulosedegrading enzymatic activity was associated with the lower cellulose availability under elevated O3 

6 (Chung et al., 2006, 191729). The suppression of soil enzyme activities was also found in the 

7 rhizosphere of beech trees (Fagus sylvatica) in a lysimeter study in Germany (Esperschutz et al., 

8 2009, 595669)(Pritsch et al., 2009, 626808). Except for xylosidase, enzyme activities involved in 

9 plant cell wall degradation (cellobiohydrolase, beta-glucosidase and glucuronidase) were decreased 

in rhizosphere soil samples under elevated O3 (2 × ambient level) (Pritsch et al., 2009, 626808). 

11 Similarly, Chen et al. (2009, 191452) found O3 exposure, with a 3-month AOT40 of 

12 21.4-44.1 ppm-h, decreased the microbial metabolic capability in the rhizosphere and bulk soil of 

13 wheat, although the observed reduction in bulk soil was not significant. 

14 Ozone-induced change in soil organisms’ activities could affect litter decomposition rates. 

However, no general decomposition pattern has been identified. The responses varied among 

16 species, sites and exposure length. Parsons et al. (2008, 191853) collected litter from aspen and birch 

17 seedlings at AspenFACE site, and conducted a 23-month field litter incubation starting in 1999. They 

18 found that elevated O3 had different impacts on the decomposition of aspen and birch litter. Elevated 

19 O3 was found to reduce aspen litter decomposition. However, O3 accelerated birch litter 

decomposition under ambient CO2, but reduced it under elevated CO2 (Parsons et al., 2008, 191853). 

21 Liu et al. (2009, 191470) conducted another litter decomposition study at Aspen FACE from 2003 to 

22 2006, when stand leaf area index (LAI) reached its maximum. During the 935-day field incubation, 

23 elevated O3 was shown to reduce litter mass loss in the first year, but not in the second year. They 

24 suggested that higher initial tannin and phenolic concentrations under elevated O3 reduced microbial 

activity in the first year (Liu et al., 2009, 191470). In an OTC experiment, Kasurinen et al. (2006, 

26 191269) collected silver birch leaf litter from three consecutive growing seasons and conducted three 

27 separate litter-bag incubation experiments. Litter decomposition was not affected by O3 exposure in 

28 the first two incubations, but a slower decomposition rate was found in the third incubation. Their 

29 principle component analysis indicated that the litter chemistry changes caused by O3 (decreased 

Mn, P, B and increased C:N) might be partially responsible for the decreased mass loss of their third 

31 incubation. 

9.6.4.3. Soil re s pira tio n a n d c a rb on forma tion 

32 Ozone could reduce the availability of photosynthates for export to roots, and increase root 

33 mortality and turnover rates. Ozone has also been shown to reduce above-ground litter productivity 

34 and alter litter chemistry, which would affect the quality and quantity of the C supply to soil 

organisms (Section 9.6.6.1). The complex interactions among those changes make it difficult to 

36 predict the response of soil C cycling under elevated O3. The 2006 O3 AQCD concluded that O3 had 

37 no consistent impact on soil respiration (U.S. EPA, 2006, 088089). Ozone could increase or decrease 

March 2011 9-79 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

1 soil respiration, depending on the approach and timing of the measurements. Ozone may also alter 

2 soil C formation. However, very few experiments directly measured changes in soil organic matter 

3 content under O3 fumigation (U.S. EPA, 2006, 088089). Recent studies on soil respiration and soil 

4 C content also found mixed responses. Most importantly, new publications derived from long-term 

5 fumigation experiments, such as the Aspen FACE experiment, suggest that ecosystem response to O3 

6 exposure can change over time. Observations made during the late exposure years can be 

7 inconsistent with those during the early years, highlighting the need for caution to assess O3 effects 

8 based on short-term studies (Table 9-7). 

Table 9-7. The temporal variation of ecosystem responses to ozone exposure at AspenFACE site 

The Time of the Endpoint Response	 Reference Measurement 

Litter decomposition 1999-2001	 O3 reduced aspen litter decomposition. However, O3 accelerated birch Parsons et al. (2008, 191853)

litter decomposition under ambient CO2, but reduced it under elevated 

CO2 

2003-2006 O3 reduced litter mass loss in the first year, but not in the second year. Liu et al. (2009, 191470) 

Fine root production 1999 O3 had no significant impact on fine root biomass	 King et al. (2001, 041751) 

2002, 2005 O3 increased fine root biomass	 Pregitzer et al. (2008, 191677) 

Soil respiration 1998-1999 Soil respiration under +CO2+O3 treatment was lower than that under King et al. (2001, 041751)

+CO2 treatment 

2003-2007 Soil respiration under +CO2+O3 treatment was 5-25% higher than Pregitzer et al. (2006, 

under elevated CO2 treatment. 191676), Pregitzer et al. 

(2008, 191677) 

Soil C formation 1998-2001 O3 reduced the formation rates of total soil C by 51% and acid- Loya et al. (2003, 074380)

insoluble soil C by 48% 

2004-2008 No significant effect of O3 on the new C formed under elevated CO2 Talhelm et al. (2009, 596189) 

Soil Respiration 

9 Ozone has shown inconsistent impacts on soil respiration. A sun-lit controlled-environment 

10 chamber study found that O3 had no significant effects on soil respiration, fine root biomass or any 

11 of the soil organisms in a reconstructed ponderosa pine/soil-litter system (Tingey et al., 2006, 

12 191341). In an adult European beech/Norway spruce forest at Kranzberg Forest, the free air O3 

13 fumigation (AOT40 of 10.2-117 ppm-h) increased soil respiration under both beech and spruce 

14 during a humid year (Nikolova et al., 2010, 626810) . The increased soil respiration under beech has 

15 been accompanied by the increase in fine root biomass and ectomycorrhizal fungi diversity and 

16 turnover (Grebenc and Kraigher, 2007, 191265). The stimulating effect on soil respiration 

17 disappeared under spruce in a dry year, which was associated with a decrease in fine root production 

18 in spruce under drought. This finding suggested that drought was a more dominant stress than O3 for 

19 spruce (Nikolova et al., 2010, 626810). Andersen et al. (2010, 628559) labeled the canopies of 

20 European beech and Norway spruce with CO2 depleted in 13C at the same site. They did not observe 

21 any significant changes in soil respiration for either species. 

22 The nearly 10 year long studies at AspenFACE indicated that the response of soil respiration to 

23 O3 interacted with CO2 exposure and varied temporally (Table 9-7) (King et al., 2001, 

24 041751)(Pregitzer et al., 2006, 191676)(Pregitzer et al., 2008, 191677). Ozone treatment alone 

March 2011	 9-80 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

5 

 

10 

 

 

15 

 

 

20 

 

 

 

25 

 

 

30 

 

 

1 generally had the lowest mean soil respiration rates, although those differences between control and 

2 elevated O3 were usually not significant. However, soil respiration rates were different with O3 alone 

3 and when acting in combination with elevated CO2. In the first five years (1998-2002), soil 

4 respiration under +CO2+O3 treatment was similar to that under control and lower than that under 

+CO2 treatment (King et al., 2001, 041751)(Pregitzer et al., 2006, 191676). Since 2003, +CO2+O3 

6 treatment started to show the greatest impact on soil respiration. Compared to elevated CO2, soil 

7 respiration rate under +CO2+O3 treatment was 15-25% higher from 2003-2004, and 5-10% higher 

8 from 2005-2007 (Pregitzer et al., 2006, 191676)(Pregitzer et al., 2008, 191677). Soil respiration was 

9 highly correlated with the biomass of roots with diameters of <2 mm and <1 mm, across plant 

community and atmospheric treatments. The authors suggested that the increase in soil respiration 

11 rate may be due to +CO2+O3 increased fine root (<1.0 mm) biomass production (Pregitzer et al., 

12 2008, 191677). 

Soil Carbon Formation 

13 Ozone-induced reductions in plant growth can result in reduced C input to soil and therefore 

14 soil C content (Andersen, 2003, 041673). The simulations of most ecosystem models support this 

prediction (Felzer et al., 2004, 186927)(Zhang et al., 2007, 196983)(Ren et al., 2007, 191366). 

16 However, very few studies have directly measured soil C dynamics under elevated O3. After the first 

17 four years of fumigation (from 1998 to 2001) at the Aspen FACE site, Loya et al. (2003, 074380) 

18 found that forest stands exposed to both elevated O3 and CO2 accumulated 51% less total soil C, and 

19 48% less acid-insoluble soil C compared to stands exposed only to elevated CO2. Soil organic carbon 

(SOC) was continuously monitored at the Aspen FACE site, and the later data showed that the initial 

21 reduction in new C formation (soil C derived from plant litter since the start of the experiment) by 

22 O3 under elevated CO2 is only a temporary effect (Table 9-7) (Talhelm et al., 2009, 596189). The 

23 amount of new soil C in the elevated CO2 and the combined elevated CO2 and O3 treatments has 

24 converged since 2002. There was no significant effect of O3 on the new C formed under elevated 

CO2 over the last four years of the study (2004-2008). Talhelm et al. (2009, 596189) suggested the 

26 observed reduction in the early years of the experiment might be driven by a suppression of 

27 C allocated to fine root biomass. During the early exposure years, O3 had no significant impact on 

28 fine root production (King et al., 2001, 041751). However, the effect of O3 on fine root biomass was 

29 observed later in the experiment. Ozone increased fine root production and the highest fine root 

biomass was observed under the combined elevated CO2 and O3 treatment in the late exposure years 

31 (Table 9-7) (Pregitzer et al., 2006, 191676). This increase in fine root production was due to changes 

32 in community composition, such as better survival of O3-tolerant aspen genotype, birch and maple, 

33 rather than changes in C allocation at the individual tree level (Pregitzer et al., 2008, 191677)(Zak et 

34 al., 2007, 191239). 

March 2011 9-81 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

5

 

 

 

 

10 

 

 

15 

 

 

 

20 

 

 

 

25 

 

 

30

 

 

35 

9.6.4.4. Nutrient cycling 

1 Nutrient cycling is important for sustaining ecosystem production. Ozone can affect nutrient 

2 cycling by changing nutrient release from litter and uptake by plants. Nitrogen is the limiting 

3 nutrient for most ecosystems, and several studies examined N dynamics under elevated O3. Holmes 

4 et al. (2006, 191372) found that elevated O3 decreased gross N mineralization at the Aspen FACE 

site, indicating that O3 may reduce N availability. However, other N cycling processes, such as NH4

+ 

6 immobilization, gross nitrification, microbial biomass N and soil organic N, were not affected by 

7 elevated O3 (Holmes et al., 2006, 191372). Similarly, Kanerva et al. (2006, 191747) found total N, 

8 NO3-, microbial biomass N, potential nitrification and denitrification in their meadow mesocosms 

9 were not affected by elevated O3 (40-50 ppb). Ozone also showed small impact on other micro and 

macro nutrients. Liu et al. (2007, 093286) assessed N, P, K, S, Ca, Mg, Mn, B, Zn and Cu release 

11 dynamics at Aspen FACE, and they found that O3 had no effects on most nutrients, except to 

12 decrease N and Ca release from litter. 

13 Using the SImple NItrogen Cycle model (SINIC), Hong et al. (2006, 186989) evaluated the 

14 impacts of O3 exposure on soil N dynamics and streamflow nitrate flux. The detrimental effect of O3 

on plant growth was found to reduce plant uptake of N and therefore increase nitrate leaching. Their 

16 model simulation indicated that ambient O3 exposure increased the mean annual stream flow nitrate 

17 export by 12% (0.042 g N/m2/year) at the Hubbard Brook Experimental Watershed from 1964-1994 

18 (Hong et al., 2006, 186989). 

9.6.4.5. Dis s olve d Orga nic Ca rb on a nd Bio genic Trac e Gas es Emis s io n 

19	 The O3-induced changes in plant growth, C and N fluxes to soil and microbial metabolism can 

alter other biogeochemical cycling processes, such as soil dissolved organic carbon (DOC) turnover 

21 and trace gases emission. 

22 Jones et al. (2009, 199881) collected fen cores from two peatlands in North Wales, UK and 

23 exposed them to one of four levels of O3 (AOT40 of 0, 3.69, 5.87 and 13.80 ppm-h for 41 days). 

24 They found the concentration of porewater DOC in fen cores was significantly decreased by 

increased O3 exposure. A reduction of the low molecular weight fraction of DOC was concurrent 

26 with the observed decrease in DOC concentration. Their results suggested that O3 damage to 

27 overlying vegetation may decrease utilizable C flux to soil. Microbes, therefore, have to use labile C 

28 in the soil to maintain their metabolism, which, the authors hypothesized, leads to a decreased DOC 

29 concentration with a shift of the DOC composition to more aromatic, higher molecular weight 

organic compounds. 

31 Several studies since the 2006 O3 AQCD have examined the impacts of O3 on nitrous oxide 

32 (N2O) and methane (CH4) emission. Kanerva et al. (2007, 191405) measured the fluxes of N2O and 

33 CH4 in meadow mesocosms, which were exposed to elevated CO2 and O3 in OTCs in south-western 

34 Finland. They found that the daily N2O fluxes were decreased in the NF+O3 (non-filtered air + 

elevated O3, 40-50 ppb) after three seasons of exposure. Elevated O3 alone or combined with CO2 

March 2011	 9-82 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

5 

 

 

10 

 

 

 

15 

20 

 

25

 

 

30

 

 

 

 

1 did not have any significant effect on the daily fluxes of CH4 (Kanerva et al., 2007, 191405). In 

2 another study conducted in central Finland, the 4 year open air O3 fumigation (AOT40 of 20.8­

3 35.5 ppm-h for growing season) also did not affect the rate of potential CH4 production, but it 

4 slightly increased potential CH4 oxidation by 15% in the peatland microcosms (Morsky et al., 2008, 

191507). However, O3 has no overall effects on CH4 emissions, which is the net result of the 

6 potential CH4 production and oxidation (Morsky et al., 2008, 191507). 

9.6.5. Competition and biodivers ity 

7 The effects of O3 on species competition (AX9.3.3.4) and community composition (AX9.6.4) 

8 were summarized in the 2006 O3 AQCD (U.S. EPA, 2006, 088089). Plant species differ in their 

9 sensitivity to O3. Fast growing plants with high stomatal conductance and high specific leaf area 

(SLA) were more likely to be sensitive to O3 exposure. Further, different genotypes of a given 

11 species also vary in their sensitivity. This differential sensitivity could change the competitive 

12 interactions that lead to loss in O3 sensitive species or genotypes. A shift in community composition 

13 in forest and grassland ecosystems noted in the 2006 O3 AQCD (U.S. EPA, 2006, 088089) has 

14 continued to be observed from experimental and gradient studies. Additionally, research since the 

last review has shown that O3 can alter community composition and diversity of soil microbial 

16 communities. 

9.6.5.1. Fores t 

17 In the San Bernardino Mountains in southern California, O3 pollution caused a significant 

18 decline in ponderosa pine (Pinus ponderosa ) and Jeffrey pine (Pinus jeffreyi) (U.S. EPA, 2006, 

19 088089). Pine trees in the young mature age class group exhibited higher mortality rates compared 

with mature trees at a site with severe O3 visible foliar injury. The vulnerability of young mature 

21 pines was most likely caused by the fact that trees in this age class were emerging into the canopy, 

22 where higher O3 concentrations were encountered (McBride and Laven, 1999, 053050). Because of 

23 the loss of O3-sensitive pines, mixed forests of ponderosa pine, Jeffery Pine and white fir (Abies 

24 concolor) shifted to predominantly white fir (Miller, 1973, 039165). Ozone may have indirectly 

caused the decline in understory diversity in coniferous forests in the San Bernardino Mountains 

26 through an increase in pine litterfall. This increase in litterfall from O3 exposure results in an 

27 understory layer that may prohibit the establishment of native herbs, but not exotic annual Galium 

28 aparine (Allen et al., 2007, 196876). 

29 Ozone damage to conifer forests has also been observed in several other regions. In the Valley 

of Mexico, a widespread mortality of sacred fir (Abies religiosa) was observed in the heavily 

31 polluted area of the Desierto de los Leones National Park in the early 1980s (de Lourdes de Bauer 

32 and Hernandez-Tejeda, 2007, 196891)(Fenn et al., 2002, 626806). Ozone damage was widely 

33 believed to be an important causal factor in the dramatic decline of sacred fir. In alpine regions of 

34 southern France and the Carpathians Mountains, O3 was also considered as the major cause of the 

March 2011 9-83 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

5

 

 

 

10 

 

15

 

 

20 

 

 

25

 

 

30 

 

 

 

35

1 observed decline in cembran pine (Pinus cembra)(Wieser et al., 2006, 191391). However, for those 

2 pollution gradient studies, several possible factors, such as drought, insect outbreak and forest 

3 management, may also contribute to the mortality of trees (de Lourdes de Bauer and Hernandez4 Tejeda, 2007, 196891)(Wieser et al., 2006, 191391). 

New evidence from long-term free O3 fumigation experiments provided additional support for 

6 the potential impacts of O3 on species competition and community composition changes in forest 

7 ecosystems. At the Aspen FACE site, community composition at both the genetic and species levels 

8 was altered after seven years of fumigation with O3 (Kubiske et al., 2007, 191336). In the pure aspen 

9 community, O3 fumigation reduced growth and increased mortality of sensitive clone 259, while the 

O3 tolerant clone 8L emerged as the dominant clone. Growth of clone 8L was even greater under 

11 elevated O3 compared to controls, probably due to O3 alleviated competitive pressure on clone 8L by 

12 reducing growth of other clones. In the mixed aspen-birch and aspen-maple communities, O3 

13 reduced the competitive capacity of aspen compared to birch and maple (Kubiske et al., 2007, 

14 191336). In a phytotron study, O3 fumigation reduced growth of beech but not spruce in mixed 

culture, suggesting a higher susceptibility of beech to O3 under interspecific competition (Kozovits 

16 et al., 2005, 191282). 

9.6.5.2. Gra s s la nd a nd Ag ric ultura l La n d 

17 The response of managed pasture, often cultivated as a mixture of grasses and clover, to O3 

18 pollution has been studied for many years. The tendency for O3-exposure to shift the biomass of 

19 grass-legume mixtures in favor of grass species, reported in the previous O3 AQCD (U.S. EPA, 2006, 

088089) has been generally confirmed by recent studies. In a mesocosm study, Trifolium repens and 

21 Lolium perenne mixtures were exposed to an episodic rural O3 regime within solardomes for 

22 12 weeks. T. repens showed significant changes in biomass but not L. perenne, and the proportion of 

23 T. repens decreased in O3-exposed mixtures compared to the control (Hayes et al., 2009, 191360). 

24 After 5-year O3 fumigation (AOT40 of 13.3-59.5 ppm-h) at the Le Mouret FACE experiment, 

Switzerland, legumes in fumigated plots declined from their initial over-representation (128%) to a 

26 mere 59% in control plots(Volk et al., 2006, 191434). However, Stampfli and Fuhrer (2010, 102180) 

27 re-analyzed the species and soil data and suggested that Volk et al. (2006, 191434) overestimated the 

28 O3 effect. Stampfli and Fuhrer (2010, 102180) found that the difference in the species dynamics 

29 between control and O3 treatment was more caused by heterogeneous initial conditions than O3 

exposure. Several studies also suggested the mature/species-rich ecosystems were more resilient to 

31 O3 exposure. At another FACE experiment, located at Alp Flix, Switzerland, O3 fumigation (AOT40 

32 of 15.2-64.9 ppm-h) showed no significant impact on community composition of this species-rich 

33 pasture (Bassin et al., 2007, 191534). Pfleeger et al. (2010, 644281) collected seed bank soil from an 

34 agricultural field and examined how the plant community responded over several generations to 

elevated O3 exposures. Sixty plant species from 22 families emerged in the chambers over their 

36 four year study. Overall, they found that O3 appeared to have small impacts on seed germination and 

37 only a minor effect on species richness of pioneer plant communities. 

March 2011 9-84 DRAFT – DO NOT CITE OR QUOTE 
 

 

5

 

10 

 

 

 

15

 

 

 

20 

 

 

25

 

 

 

30 

 

35

 

1 Several review papers have discussed the physiological and ecological characteristics of O3­

2	 sensitive herbaceous plants. Hayes et al. (2007, 196911) assessed species traits associated with O3 

3	 sensitivity by the changes in biomass caused by O3 exposure. Plants of the therophyte (e.g., annual) 

4	 life form were particularly sensitive to O3. Species with higher mature leaf N concentration tended to 

be more sensitive than those with lower leaf N concentration. Plants growing under high oxidative 

6	 stress environments, such as high light or high saline, were more sensitive to O3. Using the same 

7	 dataset from Hayes et al. (2007, 196911), Mills et al. (2007, 196934) identified the O3 sensitive 

8	 communities. They found that the largest number of these O3 sensitive communities were associated 

9	 with grassland ecosystems. Among grassland ecosystems, alpine grassland, sub-alpine grassland,
 

woodland fringe, and dry grassland were identified as the most sensitive communities.
 

9.6.5.3. Mic ro be s 

11 Several methods have been used to study microbial composition changes associated with 

12 elevated O3. Phospholipid fatty acid (PLFA) analysis is widely used to determine whether O3 elicits 

13 an overall effect on microbial community composition. However, since PLFA markers cover a broad 

14 range of different fungi, resolution of this method may be not fine enough to detect small changes in 

the composition of fungal communities. Methods, such as microscopic analyses and polymerase 

16 chain reaction–denaturing gradient gel electrophoresis (PCR–DGGE), have better resolution to 

17 specifically analyze the fungal community composition. The resolution differences among those 

18 methods needs to be considered when assessing the O3 impact on microbial community composition. 

19 Kanerva et al. (2008, 191264) found that elevated O3 (40-50 ppb) decreased total, bacterial, 

actinobacterial and fungal PLFA biomass values as well as fungal:bacterial PLFA biomass ratio in 

21 their meadow mesocosms in south-western Finland. The relative proportions of individual PLFAs 

22 between the control and elevated O3 treatments were significantly different, suggesting that O3 

23 modified the structure of the microbial community. Morsky et al. (2008, 191507) exposed boreal 

24 peatland microcosms to elevated O3, with growing season AOT40 of 20.8-35.3 ppm-h, in an open-air 

O3 exposure field in Central Finland. They also found that microbial composition was altered after 

26 three growing seasons with O3 fumigation, as measured by PLFA. Ozone tended to increase the 

27 presence of Gram-positive bacteria and the biomass of fungi in the peatland microcosms. Ozone also 

28 resulted in higher microbial biomass, which co-occurred with the increases in concentrations of 

29 organic acids and leaf density of sedges (Morsky et al., 2008, 191507). In a lysimeter study in 

Germany, O3 was found to alter the PLFA profiles in the upper 0-20 cm rhizosphere soil of European 

31 beech. Elevated O3 reduced bacterial abundance but had no detectable effect on fungal abundance 

32 (Pritsch et al., 2009, 626808). Using microscopic analyses, Kasurinen et al. (2005, 191245) found 

33 that elevated O3, with 5 or 6 months of AOT40 of 20.6-30.9 ppb-h, decreased the proportions of 

34 black and liver-brown mycorrhizas and increased that of light brown/orange mycorrhizas. In an 

herbaceous plant study, SSCP (single-strand conformation polymorphism) profiles indicated that O3 

36 stress (about 75 ppb) had a very small effect on the structural diversity of the bacterial community in 

37 rhizospheres (Dohrmann and Tebbe, 2005, 191320). At the Aspen FACE site, O3 had no significant 

March 2011	 9-85 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

1 effect on fungal relative abundance, as indicated by PLFA profile. However, elevated O3 altered 

2 fungal community composition, according to the identification of 39 fungal taxonomic units from 

3 soil using polymerase chain reaction–denaturing gradient gel electrophoresis (PCR-DGGE) (Chung 

4 et al., 2006, 191729). Ozone was found to change microbial community composition in an 

5 agricultural system. Chen et al. (2010, 644257) found elevated O3 (100-150 ppb) had significant 

6 effects on soil microbial composition expressed as PLFA percentage in a rice paddy in China. 

9.6.6. Ins ec ts an d Wildlife 

9.6.6.1. Ins e c ts 

7 Insects may respond indirectly to changes to plants (i.e., increased reactive oxygen species, 

8 altered phytochemistry, altered nutrient content) that occur under elevated O3 conditions, or O3 can 

9 have a direct effect on insect performance (Menendez et al., 2009, 191430). Effects of O3 on insects 

10 occur at the species level (i.e., growth, survival, reproduction, development, feeding behavior) and at 

11 the population and community-level (i.e., population growth rate, community composition). In 

12 general, effects of O3 on insects are highly context- and species-specific (Bidart-Bouzat and Imeh13 Nathaniel, 2008, 191431; Lindroth, 2010, 596479). Furthermore, plant responses to O3 exposure and 

14 herbivore attack have been demonstrated to share signaling pathways, complicating characterization 

15 of these stressors (Lindroth, 2010, 596479; Menendez et al., 2009, 191430; Menendez et al., 2010, 

16 384046) 

Species-Level Responses 

17 In considering insect growth, survival and reproduction in elevated O3 conditions, several 

18 studies have indicated an effect while others have found no correlation. The performance of five 

19 herbivore species (three moths and two weevils) was assessed in an OTC experiment at 2 × ambient 

20 concentration (Peltonen et al., 2010, 596482). Growth of larvae of the Autumnal moth, Epirrita 

21 autumna, was significantly decreased in the O3 treatment while no effects were observed in the other 

22 species. In an aphid oviposition preference study using birch buds grown in a three year OTC 

23 experiment, O3 had neither a stimulatory or deterring effect on egg-laying (Peltonen et al., 2006, 

24 196946). Furthermore, changes in birch bud phenolic compounds associated with the doubled 

25 ambient concentrations of O3 did not correlate with changes in aphid oviposition (Peltonen et al., 

26 2006, 196946). Reproduction in Popillia japonica, that were fed soybeans and grown under elevated 

27 O3,appeared to be unaffected (O'Neill et al., 2008, 195653). In a meta-analysis of effects of elevated 

28 O3 on 22 species of trees and 10 species of insects, the rates of survival, reproduction and food 

29 consumption were typically unaffected while development times were reduced and pupal masses 

30 were increased (Valkama et al., 2007, 191348). 

31 At the Aspen FACE site insect performance under elevated (50-60 ppb) O3 conditions 

32 (approximately 1.5 × background ambient levels of 30-40 ppb O3) have been considered for several 

March 2011 9-86 DRAFT – DO NOT CITE OR QUOTE 
 

5 

 

 

 

10

 

 

15 

 

20 

 

 

 

 

25

 

30

 

 

 

35

 

 

1 species. Cumulative fecundity of aphids (Cepegillettea betulaefoliae), that were reared on O3­

2 exposed paper birch (Betula papyrifera) trees, was lower than aphids from control plots (Awmack et 

3 al., 2004, 052926). No effects on growth, development, adult weight, embryo number and birth 

4 weight of newborn nymphs were observed. In a study conducted using three aspen genotypes, 

performance of the aspen beetle (Chrysomela crochi) decreased across all parameters measured 

6 (development time, adult mass and survivorship) under elevated O3 (Vigue and Lindroth, 2010, 

7 644185). There was an increase in the development time of male and female aspen beetle larvae 

8 although the percentages varied across genotypes. Decreased beetle adult mass and survivorship was 

9 observed across all genotypes under elevated O3 conditions. Another study from the Aspen FACE 

site, did not find any significant effects of elevated O3 on performance (longevity, fecundity, 

11 abundance) of the invasive weevil (Polydrusus sericeus) (Hillstrom et al., 2010, 644679). 

12 Since the 2006 O3 AQCD, several studies have considered the effect of elevated O3 on feeding 

13 behavior of insects. In a feeding preference study, the common leaf weevil (Phyllobius pyri) 

14 consumed significantly more leaf discs from one aspen clone when compared to a second clone 

under ambient air conditions (Freiwald et al., 2008, 196903). In a moderately elevated O3 

16 environment (1.5 × ambient), this preference for a certain aspen clone was less evident, however, 

17 leaves from O3-exposed trees were significantly preferred to leaves grown under ambient conditions. 

18 Other plant-herbivore interactions have shown no effects of elevated O3 on feeding. Feeding 

19 behavior of Japanese beetles (P. japonica) appeared to be unchanged when beetles were fed soybean 

leaves grown under elevated O3 conditions (O'Neill et al., 2008, 195653). At the Aspen FACE site, 

21 feeding by the invasive weevil (Polydrusus sericeus), as measured by leaf area consumption, was not 

22 significantly different between foliage that was grown under elevated O3 versus ambient conditions 

23 (Hillstrom et al., 2010, 644679). 

Population-Level and Community-Level Responses 

24 Recent data on insects provide evidence of population-level and community-level responses to 

O3. Elevated levels of O3 can affect plant phytochemistry and nutrient content which in turn can alter 

26 population density and structure of the associated herbivorous insect communities and impact 

27 ecosystem processes (Lindroth, 2010, 596479). In a long-term study of elevated O3 on herbivore 

28 performance at the Aspen FACE site, individual performance and population-level effects of the 

29 aphid C. betulaefoliae were assessed. Elevated O3 levels had a strong positive effect on the 

population growth rates of the aphids; although effects were not detected by measuring growth, 

31 development, adult weight, embryo number or birth weight of newborn nymphs (Awmack et al., 

32 2004, 052926). Conversely, a lower rate of population growth was observed in aphids previously 

33 exposed to O3 in an OTC (Menendez et al., 2010, 384046). No direct effects of O3 were observed; 

34 however, nymphs born from adults exposed to and feeding on O3 exposed plants were less capable of 

infesting new plants when compared to nymphs in the control plots (Menendez et al., 2010, 384046). 

36 Elevated O3 reduced arthropod abundance by 17% at Aspen FACE, largely as a result of the negative 

37 effects on parasitoids, although phloem-feeding insects may benefit (Hillstrom and Lindroth, 2008, 

March 2011 9-87 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

1 191367). Herbivore communities effected by O3 and N were sampled along an air pollution gradient 

2 in the Los Angeles basin (Jones and Paine, 2006, 191301). Abundance, diversity, and richness of 

3 herbivores were not affected. However, a shift in community structure, from phloem-feeding to 

4 chewing dominated communities, was observed along the gradient. No consistent effect of elevated 

5 O3 on herbivory or insect population size was detected at SoyFACE (Dermody et al., 2008, 191850). 

6 Evidence of modification of insect populations and communities in response to elevated O3 

7 includes genotypic and phenotypic changes. In a study conducted at the Aspen FACE site, elevated 

8 O3 altered the genotype frequencies of the pea aphid (Acyrthosiphon pisum) grown on red clover 

9 (Trifolium pratense) over multiple generations (Mondor et al., 2005, 191217). Aphid color was used 

10 to distinguish between the two genotypes. Ozone increased the genotypic frequencies of 

11 pink-morph:green-morph aphids from 2:1 to 9:1, and depressed wing-induction responses more 

12 strongly in the pink than the green genotype (Mondor et al., 2005, 191217). Growth and 

13 development of individual green and pink aphids reared as a single genotype or mixed genotypes 

14 were unaffected by elevated O3 (Mondor et al., 2010, 644271). Furthermore, growth of pea aphid 

15 populations is not readily predictable using individual growth rates. 

9.6.6.2. Wildlife 

Herpatofauna 

16 Since the 2006 O3 AQCD, direct effects of O3 exposure including physiological changes and 

17 alterations of ecologically important behaviors such as feeding and thermoregulation have been 

18 observed in wildlife. These studies have been conducted in limited laboratory exposures, and the 

19 levels of O3 treatment (e.g. 0.2-0.8 ppm) were often unrealistically higher than the ambient levels. 

20 Amphibians may be especially vulnerable to airborne oxidants due to the significant gas exchange 

21 that occurs across the skin (Andrews et al., 2008, 645771)(Dohm et al., 2008, 604584). Exposure to 

22 0.2 ppm to 0.8 ppm O3 for 4 h resulted in a decrease of oxygen consumption and depressed lung 

23 ventilation in the California tree frog Pseudacris cadaverina (Mautz and Dohm, 2004, 644188). 

24 Following a single 4-h exposure to O3, reduced pulmonary macrophage phagocytosis was observed 

25 at 1 and 24 hours postexposure in the marine toad (Bufo marinus) indicating an effect on immune 

26 system function (Dohm et al., 2005, 180452). There was no difference in macrophage function at 

27 48 hours postexposure in exposed and control individuals. 

28 Behavioral effects of O3 observed in amphibians include responses to minimize the surface 

29 area of the body exposed to the air and a decrease in feeding rates (Dohm et al., 2008, 604584; 

30 Mautz and Dohm, 2004, 644188). The adoption of a low-profile “water conservation posture” during 

31 O3 exposure was observed in experiments with the California tree frog (Mautz and Dohm, 2004, 

32 644188). Toads, Bufo marinus, exposed to 0.06 μL/L O3 for 4 hours ate significantly fewer 

33 mealworms at 1 hour and 48 hours postexposure than control toads (Dohm et al., 2008, 604584). In 

34 the same study, escape/exploratory behavior as measured by total distance moved was not adversely 

35 affected in the O3-exposed individuals as compared to the controls (Dohm et al., 2008, 604584). 

March 2011 9-88 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

5 

 

 

 

10 

 

 

 

15 

 

 

 

20

 

25

 

30

 

1 Water balance and thermal preference in herpatofauna are altered with elevated O3. Toads 

2 exposed to 0.8 ppm O3 for 4 hours exhibited behavioral hypothermia when preferred body 

3 temperatures were measured at 1, 24 and 48 hours postexposure (Dohm et al., 2001, 017086). 

4 Ozone-exposed individuals lost almost 5g more body mass on average than controls due to 

evaporative water loss. At 24 hours after exposure, the individuals that had lost significant body 

6 mass had lower preferred body temperatures(Dohm et al., 2001, 017086). Behavioral hypothermia 

7 was also observed in reptiles following 4-h exposures to 0.6 ppm O3. Exposure of the Western Fence 

8 Lizard (Sceloporus occidentalis) at 25oC induced behavioral hypothermia that recovered to control 

9 temperatures by 24 hours (Mautz and Dohm, 2004, 644188). The behavioral hypothermic response 

persisted in lizards exposed to O3 at 35o

C at 24 hours postexposure resulting in a mean body 

11 temperature 3.3oC over controls. 

Soil Fauna Communities 

12 Ozone has also been shown to alter soil fauna communities (Kasurinen et al., 2007, 199826; 

13 Loranger et al., 2004, 072712; Meehan et al., 2010, 644184). Abundance of Acari (mites and ticks) 

14 decreased by 47% under elevated O3 at AspenFACE site, probably due to the higher secondary 

metabolites and lower N concentrations in litter and foliage under elevated O3 (Loranger et al., 2004, 

16 072712). In another study from the AspenFACE site, leaf litter collected from aspen grown under 

17 elevated O3 conditions were higher in fiber and lignin concentrations than trees grown under ambient 

18 conditions. These chemical characteristics of the leaves were associated with an increased springtail 

19 population growth following 10 weeks in a laboratory microcosm (Meehan et al., 2010, 644184). 

Consumption rates of earthworms fed on leaf litter for 6 weeks from trees grown under elevated O3 

21 conditions and ambient air did not vary significantly between treatments (Meehan et al., 2010, 

22 644184). In another study on juvenile earthworms Lumbricus terrestris, individual growth was 

23 reduced when worms were fed high-O3 birch litter from trees exposed for three years to elevated O3 

24 in an OTC system (Kasurinen et al., 2007, 199826). In the same study no significant growth or 

mortality effects were observed in isopods. 

9.6.6.3. Indire c t Effe c ts on Wildlife 

26 In addition to the direct effects of O3 exposure on physiological and behavioral endpoints 

27 observed in the laboratory, there are indirect effects to wildlife. These effects include changes in 

28 biomass and nutritive quality of O3-exposed plants (reviewed in Section 9.5) that are consumed by 

29 wildlife. Reduced digestibility of O3-exposed plants may alter dietary intake and foraging strategies 

in herbivores. In a study using native highbush blackberry (Rubus argutus) relative feed value of the 

31 plants decreased in bushes exposed to double ambient concentrations of O3 (Ditchkoff et al., 2009, 

32 192230). Indirect effects of elevated O3 on wildlife include changes in chemical signaling important 

33 in ecological interactions reviewed below. 

March 2011 9-89 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

5

 

 

10 

 

 

15

 

 

20 

 

 

 

25 

 

30 

 

 

Chemical Signaling in Ecological Interactions 

1 Ozone has been shown to degrade or alter biogenic VOC signals important to ecological 

2 interactions including; (1) attraction of pollinators and seed dispersers; (2) defense against herbivory; 

3 and (3) predator-prey interactions (McFrederick et al., 2009, 191329; Pinto et al., 2007, 196949; 

4 Pinto et al., 2007, 196950; Pinto et al., 2010, 596490; Yuan et al., 2009, 199779). Each signal 

released by emitters has an atmospheric lifetime and a unique chemical signature comprised of 

6 different ratios of individual hydrocarbons that is susceptible to atmospheric oxidants such as O3 

7 (Wright et al., 2005, 626463; Yuan et al., 2009, 199779). Under elevated O3 conditions, these 

8 olfactory cues may travel shorter distances before losing their specificity (McFrederick et al., 2008, 

9 196931; McFrederick et al., 2009, 191329). Additional non-phytogenic VOC-mediated 

interrelationships with the potential to be modified by O3 include territorial marking, pheromones for 

11 attraction of mates and various social interactions including scent trails, nestmate recognition and 

12 signals involved in aggregation behaviors (McFrederick et al., 2009, 191329). In general, effects of 

13 O3 on scent-mediated ecological interactions are highly context- and species-specific (Bidart-Bouzat 

14 and Imeh-Nathaniel, 2008, 191431; Lindroth, 2010, 596479). 

Pollination and Seed Dispersal 

Phytogenic VOC’s attract pollinators and seed dispersers to flowers and fruits (Dudareva et al., 

16 2006, 626458; Theis and Raguso, 2005, 626461). These floral scent trails in plant-insect interactions 

17 may be destroyed or transformed by O3 (McFrederick et al., 2008, 196931). Using a Lagrangian 

18 model, the rate of destruction of phytogenic VOC’s was estimated in air parcels at increasing 

19 distance from a source in response to increased regional levels of O3, hydroxyl and nitrate radicals 

(McFrederick et al., 2008, 196931). Based on the model, the ability of pollinators to locate highly 

21 reactive VOCs from emitting flowers may have decreased from kilometers during pre-industrial 

22 times to <200 m at current ambient conditions (McFrederick et al., 2008, 196931). Scents that travel 

23 shorter distances (0-10 m) are less susceptible to air pollutants, while highly reactive scents that 

24 travel longer distances (10 to 100’s of meters), are at a higher risk for degradation (McFrederick et 

al., 2009, 191329). For example, male euglossine bees can detect bait stations from a distance of at 

26 least one kilometer (Dobson, 1994, 626466). The alcohols, ketones and aldehydes comprising sex 

27 pheromones in moths could be especially vulnerable to degradation by O3, since some males travel 

28 >100 m to find mates (Carde and Haynes, 2004, 626467). 

Defense Against Herbivory 

29 Ozone can alter the chemical signature of VOCs emitted by plants and these VOCs are 

subsequently detected by herbivores (Blande et al., 2010, 643928; Cannon, 1990, 626460; Iriti and 

31 Faoro, 2009, 199313; Jackson et al., 1999, 026877; Pinto et al., 2007, 196949; Vuorinen et al., 2004, 

32 626462). These modifications can make the plant either more attractive or repellant to phytophagous 

33 insects (Pinto et al., 2010, 596490). For example, under elevated O3, the host plant preference by 

34 forest tent caterpillars increased for birch compared to aspen (Agrell et al., 2005, 074324). OzoneMarch 2011 9-90 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

5 

 

 

 

10 

 

 

 

 

15

20 

 

 

 

25

 

30

 

1 induced emissions from red spruce needles were found to repel spruce budworm larvae (Cannon, 

2 1990, 626460). Transcriptional profiles of field grown soybean (Glycine max) grown in elevated O3 

3 conditions were altered due to herbivory by Japanese beetles. The herbivory resulted in a higher 

4 number of transcripts in the leaves of O3-exposed plants and up-regulation of antioxidant metabolism 

associated with plant defense (Casteel et al., 2008, 191696). 

6 Ozone may modify signals involved in plant-to-plant interactions and plant defense against 

7 pathogens (Blande et al., 2010, 643928; McFrederick et al., 2009, 191329; Pinto et al., 2010, 

8 596490; Yuan et al., 2009, 199779). In a recent study with lima beans, 80 ppb O3 degraded several 

9 herbivore-induced VOC’s, reducing the distance over which plant-to-plant signaling occurred 

(Blande et al., 2010, 643928). 

Predator-Prey Interactions 

11 Elevated O3 conditions are associated with disruption of pheromone-mediated interactions at 

12 higher trophic levels (e.g., predators and parasitoids of herbivores). In a study from the Aspen FACE 

13 site, predator escape behaviors of the aphid (Chatophorus stevensis) were enhanced on O3-fumigated 

14 aspen trees although the mechanism of this response remains unknown (Mondor et al., 2004, 

074334). The predatory mite Phytoseiulus persimilis can distinguish between the VOC signature of 

16 ozonated lima bean plants and ozonated plants simultaneously damaged by T. urticae (Vuorinen et 

17 al., 2004, 626462) however, other tritrophic interactions have shown no effect (Pinto et al., 2007, 

18 196950). 

19 There are few studies that consider host location behaviors of parasites under elevated O3. In 

closed chambers fumigated with O3, the searching efficiency and proportion of the host larval fruit 

21 flies parasitized by Asobara tabida, declined when compared to filtered air controls (Gate et al., 

22 1995, 026655). The host location behavior and rate of parasitism of the wasp (Coesia plutellae) on 

23 Plutella xylostella-infested potted cabbage plants was tested under ambient and doubled O3 

24 conditions in an open-air fumigation system (Pinto et al., 2008, 616554). The number of wasps 

found in the field and the percentages of parasitized larvae were not significantly different from 

26 controls under elevated O3. 

27 Elevated O3 has the potential to perturb specialized food-web communication in transgenic 

28 crops. In insect-resistant oilseed rape Brassica napus grown under 100 ppb O3 in a growth chamber, 

29 reduced feeding damage by Putella xylostella led to deceased attraction of the endoparasitoid 

(Costesia vestalis) (Himanen et al., 2009, 191338). Under chronic O3-exposure, the insect resistance 

31 trait BT cry1Ac in transgenic B. napus was higher than the control (Himanen et al., 2009, 191369). 

32 There was a negative relative growth rate of the Bt target herbivore, P. xylostella, in all O3 

33 treatments. 

March 2011 9-91 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

9.7.	 Effects-Ba s e d Air Qu a lity Exp o s u re In d ices and Dos e 

Mo d e lin g 

9.7.1.	 Introd uc tion 

1	 Exposure indices are metrics that quantify exposure as it relates to measured plant damage 

2 (i.e., reduced growth). They are summary measures of monitored ambient O3 concentrations over 

3 time, intended to provide a consistent metric for reviewing and comparing exposure-response effects 

4 obtained from various studies. Such indices may also provide a basis for developing a biologically5 relevant air quality standard for protecting vegetation and ecosystems. Effects on plant growth and/or 

6 yield have been a major focus of the characterization of O3 impacts on plants for purposes of the air 

7 quality standard setting process (U.S. EPA, 1986, 017607)(U.S. EPA, 1996, 039046)(U.S. EPA, 

8 2007, 090207). The quantitative characterization of the relationship of O3 and plant responses has 

9 been referred to as “dose-response” and “exposure-response” alternatively. The distinction is in how 

10 the pollutant concentration is expressed: “dose” is the pollutant concentration absorbed by the leaf 

11 over some time period, and is very difficult to measure directly, whereas “exposure” is the ambient 

12 air concentration measured near the plant over some time period, and summarized for that period 

13 using an index. Exposure indices have been most useful in considering the form of secondary O3 

14 NAAQS standard, in large part because they only require ambient air quality data rather than more 

15 complex indirect calculations of dose to the plant. The attributes of exposure indices that most 

16 require consideration are the weighting of O3 concentrations, and the daily and seasonal time-periods 

17 that are most relevant to plant damage. Several different types of exposure indices are discussed in 

18 Section 9.7.2. 

19 Theoretically, a measure of plant O3 uptake or dose from ambient air (either rate of uptake or 

20 cumulative seasonal uptake) might be a more ideal predictor of O3 damage to plants than an 

21 exposure index and may be more useful in improving risk assessment. An uptake measure would 

22 have to integrate all those environmental factors that influence stomatal conductance, including but 

23 not limited to temperature, humidity, and soil water status (Section 9.7.4). Even when integrating 

24 those environmental factors, a direct measure of the internal leaf concentration of O3, however, is 

25 technically difficult. Therefore, uptake values are generally obtained with simulation models that 

26 require knowledge of species- and site-specific values for the variables mentioned. In addition, it has 

27 also been recognized that O3 detoxification processes and the temporal dynamics of detoxification 

28 must be taken into account in dose modeling (Heath et al., 2009, 196783) (Section 9.7.4). Because of 

29 this, research has focused historically on predictors of O3 damage to plants based only on exposure 

30 as a summary measure of monitored ambient pollutant concentration over some integral of time, 

31 rather than dose (Lee et al., 1988, 042136)(Lefohn and Benedict, 1982, 039395)(O'Gara, 1922, 

32 015018)(U.S. EPA, 1986, 017607)(U.S. EPA, 1992, 042599)(U.S. EPA, 1996, 080828). 

March 2011	 9-92 DRAFT – DO NOT CITE OR QUOTE 
 

1

2 

3 

4

5 

6

7 

8 

9 

10 

11 

12 

13

14 

15

16

9.7.2. De s cription o f Exp o s u re Indice s Availa b le in the Lite ratu re 

Mathematical approaches for summarizing ambient air quality information in biologically 

meaningful forms for O3 vegetation effects assessment purposes have been explored for more than 

80 years (O'Gara, 1922, 015018)(U.S. EPA, 1996, 080827). In the context of broad-scale national 

NAAQS standards that protect for "known or anticipated" effects on many plant species in a variety 

of habitats, exposure indices provide a numerical summary of very large numbers of ambient 

observations of concentration over extended periods. Like any summary statistic, exposure indices 

retain information on some characteristics of the original observations, at the exclusion of others. It 

is expected that indices that incorporate the characteristics that are most relevant to plant response 

will lead to more reliable protection. 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, 042128). In 

the 1996 O3 AQCD (U.S. EPA, 1996, 080828), 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, 042137). Figure 9-8 illustrates how 

several of the indices weight concentration and accumulate exposure. 

March 2011 9-93 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

 

 

 

 

 

1 

Source: Used with permission from Air and Waste Management Association, Tingey et al. (1991, 042623) 

Figure 9-8. 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 to 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. This is similar to the W126

function. The lower portion of the graphic illustrates how concentration is

accumulated over the exposure period. (c) second HDM and M-7: the upper graphic

illustrates an episodic exposure profile. The lower portion of the graphic illustrates

that the second HDM considers only a single exposure peak, while the M-7 (average 

of 7-h daily means) applies a constant exposure value over the exposure period. 

2 Various components of the exposure-response relationship, including concentration, time of 

3 day, respite time, frequency of peak occurrence, plant phenology, predisposition, etc., were weighted 

4 with various functions and evaluated on their ability in ordering the regression of exposure versus 

5 growth or yield response (Lee et al., 1989, 042137). The statistical evaluations for each of these 

6 indices were completed using growth or yield response data from many earlier exposure studies 

7 (e.g., NCLAN). This retrospective approach was necessary because there were no studies 

8 specifically designed to test the goodness of fit of the various indices. The goodness of fit of a set of 

9 linear and nonlinear models for exposure-response was ranked as various proposed indices were 

10 used in turn to quantify exposure. This approach provided evidence for the best indices. 

March 2011 9-94 DRAFT – DO NOT CITE OR QUOTE 
 

1 

2 

3 

4

5 

6

7 

8 

9

10 

11 

12

13 

14 

15

16 

17 

18 

19 

20 

21 

22 

23

24

25

26 

27 

28 

29 

30 

31

32

33 

34 

35

36 

37 

38

39 

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, 042135)(Lee et al., 

1988, 042136)(Lefohn et al., 1988, 042138)(Musselman et al., 1988, 042144)(U.S. EPA, 1992, 

042599)(U.S. EPA, 1986, 017607). These studies were previously reviewed by the EPA (U.S. EPA, 

1992, 042599)(U.S. EPA, 1996, 080828) and were in general agreement that the best fit of the data 

were cumulative concentration-weighted exposure indices. Lee et al. (1987, 042135) suggested that 

exposure indices that included all the 24-h data performed better than those that used only 7 hours of 

data; this was consistent with the conclusions of Heagle et al. (1987, 042124) 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 National Crop Loss Assessment Network (NCLAN) data, Lee et al. 

(1988, 042138) 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 provided the best statistical fit when used in the models under 

consideration, but it required data on 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, and AOT40). Amongst this group 

no index lead to consistently better fits across all studies and species (Heagle et al., 1994, 

042656)(Lefohn et al., 1988, 042138)(Musselman et al., 1988, 042144). 

Other factors, including predisposition time (Hogsett et al., 1988, 042128)(McCool et al., 

1988, 041877) and crop development stage (Heagle et al., 1991, 042530)(Tingey et al., 2002, 

040896) 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; therefore, 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, 043747) or higher (Finnan et al., 1996, 

042545)(Finnan et al., 1997, 083315) cutoff value in indices with a threshold may provide a better 

statistical fit to the experimental data. Finnan et al. (1997, 083315) 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 ranking 

of these indices differed depending on the exposure-response model used. Weighting those 

concentrations associated with sunshine hours in an attempt to incorporate an element of plant 

uptake did not improve the index performance (Finnan et al., 1997, 083315). 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 

March 2011 9-95 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

1 the SUM06 (Skarby et al., 2004, 080375). All the indices performed relatively well in regressing 

2 biomass and exposure index, with the AOT20 and AOT30 doing slightly better than others (r2 = 

3 0.46-0.47). In another comparative study of four independent data sets of potato yield and different 

4 cumulative uptake indices with different cutoff values, a similarly narrow range of r2 was observed 

5 (r2 = 0.3-0.4) (Pleijel et al., 2004, 036662). 

6 In both the U.S. and Europe, the adequacy of these numerical summaries of exposure in 

7 relating biomass and yield changes have, for the most part, all been evaluated using data from 

8 studies not necessarily designed to compare one index to another (Lefohn et al., 1988, 042138)(Lee 

9 et al., 1989, 042137)(Skarby et al., 2004, 080375). Very few studies in the U.S. have addressed this 

10 issue since the 2006 O3 AQCD. McLaughlin et al. (2007, 090348) reported that the cumulative 

11 exposure index of AOT60 related well to reductions in growth rates at forest sites in the southern 

12 Appalachian Mountains. However, the authors did not report an analysis to compare multiple 

13 indices. Overall, given the available data from previous O3 AQCDs and the few recent studies, the 

14 cumulative, concentration-weighted indices perform better than the peak or mean indices. It is still 

15 not possible, however, to distinguish the differences in performance among the cumulative, 

16 concentration-weighted indices. 

17 The main conclusions from the 1996 and 2006 O3 AQCDs (U.S. EPA, 1996, 

18 080828)(U.S. EPA, 2006, 088089) regarding an index based on ambient exposure are still valid. No 

19 information has come forth since the 2006 O3 AQCD to alter those conclusions significantly. These 

20 key conclusions can be restated as follows: 

21  O3 effects in plants are cumulative; 

22  higher O3 concentrations appear to be more important than lower concentrations in 

23 eliciting a response; 

24  plant sensitivity to O3 varies with time of day and plant development stage; and 

25  exposure indices that accumulate the O3 hourly concentrations and preferentially weight 

26 the higher concentrations have better statistical fits to growth/yield response than do the 

27 mean and peak indices. 

28 Following the 2006 criteria review process (U.S. EPA, 2006, 088089), the EPA proposed an 

29 alternative form of the secondary NAAQS for O3 using a cumulative, concentration-weighted 

30 exposure index to protect vegetation from damage (72 FR 37818 (2007, 684055), 75 FR 2938 (2010, 

31 684211), p. 3003). The EPA considered two specific concentration-weighted indices: the cutoff 

32 concentration weighted SUM06 and the sigmoid-weighted W126 exposure index (U.S. EPA, 2007, 

33 090207). These two indices performed equally well in predicting the exposure-response relationships 

34 observed in the crop and tree seedlings studies (Lee et al., 1989, 042137). At a workshop convened 

March 2011 9-96 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

1 to consider the science supporting these indices (Heck and Cowling, 1997, 084484) the participants 

2 agreed that these cumulative concentration-weighted indices being considered were equally capable 

3 of predicting plant response. A short-term daily 8-h avg was also proposed in the review that ended 

4 in 2008 (73 FR 16436, (2008, 684051)), but at the time there were no scientific studies supporting 

5 the utility of this index for vegetation. Below are the definitions of the three index forms considered 

6 in the previous staff paper review (U.S. EPA, 2007, 090207): 

7  8-h average form: 4th-highest daily max 8-h avg over the O3 season. 

8  SUM06: Sum of all hourly O3 concentrations greater than or equal to 0.06 ppm observed 

9 during a specified daily and seasonal time window (Figure 9-8a). 

10  W126: Sigmoidally weighted sum of all hourly O3 concentrations observed during a 

11 specified daily and seasonal time window (Similar to Figure 9-8b). The sigmoidal 

12 weighting of hourly O3 concentration is given in the equation below, where C is the 

13 hourly O3 concentration in ppm: 

c C e 

w 126 44031 

1 − + =

Equation 9-1 

14 The SUM06 and W126 indices have a variety of relevant time windows that may be applied and are 

15 discussed in Section 9.7.3. 

16 Other exposure indices are discussed in the literature and are currently used outside the U.S. In 

17 Europe, the cutoff concentration-weighted index AOT40 was selected in developing exposure18 response relationships based on OTC studies of a limited number of crops and trees (Grunhage and 

19 Jager, 2003, 052972). The United Nations Economic Commission for Europe (United Nations 

20 Economic Commission for Europe UNECE, 1988, 055354) adopted the critical levels approach for 

21 assessment of O3 risk to vegetation across Europe. As used by the UNECE, the critical levels are not 

22 like the air quality regulatory standards used in the U.S., but rather planning targets for reductions in 

23 pollutant emissions to protect ecological resources. Critical levels for O3 are intended to prevent 

24 long-term deleterious effects on the most sensitive plant species under the most sensitive 

25 environmental conditions, but not intended to quantify O3 effects. A critical level was defined as “the 

26 concentration of pollutant in the atmosphere above which direct adverse effects on receptors, such as 

27 plants, ecosystems, or materials may occur according to present knowledge” (United Nations 

28 Economic Commission for Europe UNECE, 1988, 055354). The nature of the “adverse effects” was 

29 not specified in the original definition, which provided for different levels for different types of 

30 harmful effect (e.g., visible injury or loss of crop yield). There are also different critical levels for 

31 crops, forests, and semi-natural vegetation. The caveat, “according to present knowledge” is 

March 2011 9-97 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

5

 

10 

 

 

15 

 

 

20

 

 

 

 

25

 

 

 

30

 

 

1 important because critical levels are not rigid; they are revised periodically as new scientific 

2 information becomes available. For example, the original critical level for O3 specified 

3 concentrations for three averaging times, but further research and debate led to the current critical 

4 level being stated as the cumulative exposure (concentration × hours) over a cutoff concentration of 

40 ppb (AOT40) (Fuhrer et al., 1997, 030380). 

6 In Europe, a decision was made to work towards a flux-based approach for the critical levels 

7 (“Level II”), with the goal of modeling O3 flux-effect relationships for three vegetation types: crops, 

8 forests, and semi-natural vegetation (Grunhage and Jager, 2003, 052972). Progress has been made in 

9 modeling flux (see section AX9.4.5; U.S. EPA, 2006, 088089) and the Mapping Manual is being 

revised (Ashmore et al., 2004, 056623)(Ashmore MEmberson et al., 2004, 056624)(Grennfelt, 2004, 

11 056625)(Karlsson et al., 2003, 055331). The revisions may include a flux-based approach for three 

12 crops: wheat, potatoes, and cotton. However, because of a lack of flux-response data, a cumulative, 

13 cutoff concentration-based (AOTx) exposure index will remain in use for the near future for most 

14 crops and for forests and semi-natural herbaceous vegetation (Ashmore et al., 2004, 056623). 

9.7.3. Impo rta n t Co mp o ne nts o f Expo s ure In dice s 

The efficacy of exposure indices in predicting biological responses requires that researchers 

16 identify a relationship between measured growth and/or yield effects and important components of 

17 exposure indices. In the previous O3 AQCDs it was established that higher hourly concentrations 

18 have greater effects on vegetation than lower concentrations (U.S. EPA, 1996, 080828)(U.S. EPA, 

19 2006, 088089). Further, it was determined that the diurnal and seasonal duration of exposure is 

important for plant response. Weighting of hourly concentrations and the diurnal and seasonal time 

21 window of exposure are the most important variables in a cumulative exposure index and will be 

22 discussed below. However, these variables must be taken in the context of plant phenology, diurnal 

23 conductance rates, plant canopy structure, and detoxification mechanisms of vegetation as well as 

24 the climate and meteorology, all of which are determinants of plant response. These more specific 

factors will be discussed in the uptake and dose modeling section (Section 9.7.4). 

9.7.3.1. Role of Co nc e n tra tio n 

26 The significant role of higher O3 concentrations was established based on several experimental 

27 studies (U.S. EPA, 1996, 080828). Several studies (Nussbaum et al., 1995, 030141)(Oksanen and 

28 Holopainen, 2001, 019538)(Yun and Laurence, 1999, 044165) have added support for the important 

29 role that peak concentrations, as well as the pattern of occurrence, plays in plant response to O3. 

Oksanen and Holopainen (2001, 019538) found that the peak concentrations and the shape of the O3 

31 exposure (i.e., duration of the event) were important determinants of foliar injury in European white 

32 birch saplings, but growth reductions were found to be more related to total cumulative exposure. 

33 Based on air quality data from 10 U.S. cities, three 4-week exposure treatments having the same 

34 SUM06 value were constructed by Yun and Laurence (1999, 044165). The authors used different 

March 2011 9-98 DRAFT – DO NOT CITE OR QUOTE 
 

1 

2 

3 

4

5 

6 

7 

8 

9 

10 

11

12 

13 

14

15 

16

17 

18 

19

20

21

22 

23 

24

25 

26 

27 

28

29

30

31 

32

33 

34

35

36

37 

38 

39 

exposure 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. (1995, 030141) 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 air 

quality analyses for the southern California area (Lee et al., 2003, 053031)(Tingey et al., 2004, 

042385). In the late 1960s and 1970s, extremely high O3 concentrations had impacted the San 

Bernardino National Forest. However, over the past 20+ years, significant reductions in O3 exposure 

have occurred (Davidson, 1993, 043366)(Lee et al., 2003, 053031)(Lefohn and Shadwick, 2000, 

040483)(Bytnerowicz et al., 2008, 196881). An illustration of this improvement in air quality is 

shown by the 37-year history of O3 air quality at the Crestline site in the San Bernardino Mountains 

(Figure 9-9) (Lee et al., 2003, 053031). Ozone 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 greater than or equal to 

95 ppb. The number of days having hourly averaged O3 concentrations greater than or equal to 

95 ppb declined significantly from 163 days in 1978 to 103 days in 1997. The changes in ambient O3 

air quality for the Crestline site were reflected in the changes in frequency and magnitude of the peak 

hourly concentration and the duration of exposure (Figure 9-9). 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, 053031). 

The inference for a role of higher concentrations comes both from results of measures of tree 

conditions on established plots and from results of model simulations. Across a broad area of the San 

Bernardino National Forest, 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, 

040702). A more recent analysis of forest changes in the San Bernardino National Forest 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, 052925). 

A model simulation of ponderosa pine growth over the 40-year period in the San Bernardino 

National Forest 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, 083540). To account for this potential interaction, the model simulations were 

conducted under conditions of unlimited soil N. The actual interactions are not known. The 

March 2011 9-99 DRAFT – DO NOT CITE OR QUOTE 
 

1 

2

3 

4 

5

6 

7 

8 

9 

10 

11 

12 

improvement in growth over the years 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, 042385). A summary of air quality data from 1980 to 2000 

for the San Bernardino National Forest area of the number of “mid-range” hourly concentrations 

indicated no dramatic changes over this 20-year period, ranging from about 1,500 to 2,000 hours per 

year (Figure 9-10). 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 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. 

March 2011 9-100 DRAFT – DO NOT CITE OR QUOTE 
 

 

Source: Used with permission from Elsevier Science Ltd., (Lee et al., 2003, 053031). 

Figure 9-9. Trends in May to September 12-h SUM06, peak 1-h ozone concentration and number

of daily exceedances of 95 ppb for the Crestline site 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 (NO X) for San Bernardino County. ﷒Annual ROG 

and NOX emissions data for San Bernardino County were obtained from Alexis et al. 

(2001, 079886) and the California Air Resource Board’s emission inventory available

at http://www.arb.ca.gov/aqd/aqdpage.htm (Cal EPA, 2009, 677487). 

March 2011 9-101 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

Figure 9-10.The number of hourly average concentrations between 50 and 89 ppb for the period 

1980-2000 for the Crestline, San Bernardino County, CA, monitoring site. 

9.7.3.2. Diurna l a n d Se a s o na l Exp os ure 

Diurnal Exposure 

1 The diurnal patterns of maximal leaf/needle conductance and occurrence of higher ambient 

2 concentrations can help determine which hours during the day over a season should be cumulated. 

3 Stomatal conductance is species and phenology dependent and is linked to both diurnal and seasonal 

4 meteorological activity as well as to soil/site conditions (e.g., soil moisture). Daily patterns of 

5 leaf/needle conductance are often highest in midmorning, whereas higher ambient O3 concentrations 

6 generally occurred in early to late afternoon when stomata were often partially closed and 

7 conductances were lower. Total O3 flux depends on atmospheric and boundary layer resistances, both 

8 of which exhibit variability throughout the day. Experimental studies with tree species demonstrated 

9 the decoupling of ambient O3 exposure, peak occurrence, and gas exchange, particularly in areas of 

10 drought (Panek, 2004, 079202). Several studies have suggested that ponderosa pine trees in the 

11 southern and northern Sierra Nevada Mountains may not be as susceptible to high O3 concentrations 

12 as to lower concentrations, due to reduced needle conductance and O3 uptake during the period when 

13 the highest concentrations occur (Arbaugh et al., 1998, 040297)(Bauer et al., 2000, 040315)(Panek et 

14 al., 2002, 040712)(Panek and Goldstein, 2001, 030190). Panek et al. (2002, 040712) compared direct 

15 O3 flux measurements into a canopy of ponderosa pine and demonstrated a lack of correlation of 

16 daily patterns of conductance and O3 occurrence, especially in the late season drought period; the 

March 2011 9-102 DRAFT – DO NOT CITE OR QUOTE 
 

1

2 

3

4

5 

6

7

8 

9 

10 

11

12 

13

14 

15

16 

17 

18 

19 

20

21 

22 

23

24

25 

26 

27 

28

29

30 

31

32 

33 

34 

35

36 

37 

38 

authors 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. (2002, 035283) 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, 040702) 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. Another consideration for the 

effect of O3 uptake is the diurnal pattern of detoxification capacity of the plant. The detoxification 

capacity may not follow the same pattern as stomatal conductance (Heath et al., 2009, 196783). 

A 12-h (8:00 a.m. to 8:00 p.m.) daylight period for a W126 cumulating exposure was proposed 

in 2007 (72 FR 37818, (2007, 684055)) and 2010 (75 FR 2938, (2010, 684211), p. 3003) following 

the release of the 2006 O3 AQCD, based primarily on evidence that the conditions for uptake of O3 

into the plant occur mainly during the daytime hours. In general, plants have the highest stomatal 

conductance during the daytime and in many areas atmospheric turbulent mixing is greatest during 

the day as well (U.S. EPA, 2006, 088089)(Uddling et al., 2010, 387073). However, notable 

exceptions to maximum daytime conductance are cacti and other plants with crassulacean acid 

metabolism (CAM photosynthesis) which only open their stomata at night. This section will focus on 

plants with C3 and C4 photosynthesis, which generally have maximum stomatal conductance during 

the daytime. 

Recent reviews of the literature reported that a large number of species had varying degrees of 

nocturnal stomatal conductance (Caird et al., 2007, 199337)(Dawson et al., 2007, 

670381)(Musselman and Minnick, 2000, 011612). This night-time conductance can also be 

enhanced by O3 damage during the day that could result in loss of stomatal control, and less 

complete closure of stomata, than under low O3 conditions (Grulke et al., 2007, 199376). In general, 

the rate of stomatal conductance at night is much lower than during the day (Caird et al., 2007, 

199337). Atmospheric turbulence at night is also often low, which results in stable boundary layers 

and unfavorable conditions for O3 uptake into vegetation (Finkelstein et al., 2000, 024029). 

Nevertheless, nocturnal turbulence does intermittently occur and may result in nonnegligible O3 flux 

into the plants. In addition, plants might be more susceptible to O3 exposure at night than during the 

daytime, because of potentially lower plant defenses (Musselman and Minnick, 2000, 

011612)(Musselman et al., 2006, 121678)(Loreto and Fares, 2007, 180259)(Heath et al., 2009, 

196783). For significant nocturnal stomatal flux and O3 effects to occur, specific conditions must 

exist. A susceptible plant with nocturnal stomatal conductance and low defense must be growing in 

an area with relatively high night-time O3 concentrations and appreciable nocturnal atmospheric 

turbulence. It is unclear how many areas there are in the U.S. where these conditions occur. It may 

be possible that these conditions exist in mountainous areas of southern California, front-range of 

Colorado (Turnipseed et al., 2009, 588752) and the Great Smoky Mountains of North Carolina and 

March 2011 9-103 DRAFT – DO NOT CITE OR QUOTE 
 

1 

2 

3

4 

5 

6 

7 

8

9 

10 

11 

12 

13

14

15

16 

17

18 

19

20 

21 

Tennessee. More information is needed in these locations to assess the local O3 patterns, 

micrometeorology and responses of potentially vulnerable plant species. 

Several field studies have attempted to quantify night-time O3 uptake with a variety of 

methods. However, many of these studies have not linked the night-time flux to measured effects on 

plants. Grulke et al. (2004, 042646) showed that the stomatal conductance at night for ponderosa 

pine in the San Bernardino National Forest (CA) ranged from one tenth to one fourth that of 

maximum daytime stomatal conductance. In June, at a high-elevation site, it was calculated that 11% 

of the total daily O3 uptake of pole-sized trees occurred at night. In late summer, however, O3 uptake 

at night was negligible. However, this study did not consider the turbulent conditions at night. 

Finklestein et al. (2000, 024029) investigated O3 deposition velocity to forest canopies at three 

different sites. The authors found the total flux (stomatal and non-stomatal) to the canopy to be very 

low during night-time hours as compared to day-time hours. However, the authors did note that 

higher nocturnal deposition velocities at conifer sites may be due to some degree of stomatal opening 

at night (Finkelstein et al., 2000, 024029). Work by Mereu et al. (2009, 102168) in Italy on 

mediterranean species indicated that nocturnal uptake was from 10 to 18% of total daily uptake 

during a weak drought and up to 24% as the drought became more pronounced. The proportion of 

night-time uptake was greater during the drought due to decreases in daytime stomatal conductance 

(Mereu et al., 2009, 102168). In a recent study at the AspenFACE site in Wisconsin, calculated leaflevel stomatal O3 flux was near zero from the night-time hours of 8:00 p.m. to 5:00 a.m. (Uddling et 

al., 2010, 387073). This was likely due to low horizontal wind speed (>1 m/s) and low O3 

concentrations (<25 ppb) during those same night-time hours (Figure 9-11). 

March 2011 9-104 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

 

 

 

 

Source: Used with permission from Elsevier Ltd., Uddling et al. (2010, 387073). 

Figure 9-11. Mean diurnal. 

(a) conductance through boundary layer and stomata (gbs), (b) Ozone concentration, and leaflevel stomatal ozone flux without flux cut-off threshold (Fst0l) in control plots from

mid-June through August in (c) 2004 and (d) 2005 in the Aspen FACE experiment.

Subscripts “max” and “min” refer to stomatal fluxes calculated neglecting and

accounting for potential non-stomatal ozone flux, respectively. 

1 A few studies have tested the biological effects of night-time O3 exposure on vegetation in 

2 controlled chambers. Biomass of ponderosa pine seedlings was significantly reduced when seedlings 

3 were exposed to either daytime or nighttime episodic profiles (Lee and Hogsett, 1999, 040451). 

4 However, the biomass reductions were much greater with daytime peak concentrations than with 

5 nighttime peak concentrations. Similarly, birch cuttings grown in field chambers that were exposed 

6 to O3 at night only, daytime only, and 24 hours showed similar reductions in biomass in night only 

7 and day only treatments. Birch seedling showed greater reductions in growth in 24-h exposures than 

8 those exposed to O3 at night or day only (Matyssek et al., 1995, 040700). Field mustard (Brassica 

9 rapa) plants exposed to O3 during the day or night showed little significant difference in the amounts 

10 of injury or reduced growth response to O3 treatment, although the stomatal conductance was 

11 70-80% lower at night (Winner et al., 1989, 043403). These studies show that effects can be seen 

March 2011 9-105 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

5

 

 

 

10 

 

15 

 

 

 

20 

 

 

 

25 

 

30

 

 

 

35

 

1 with night-time exposures to O3 but if atmospheric conditions are stable at night, it is uncertain how 

2 these exposures may affect plants and trees with complex canopies in the field. 

Seasonal Exposure 

3 Vegetation across the U.S. has widely varying periods of physiological activity during the year 

4 due to variability in climate and phenology. In order for a particular plant to be vulnerable to O3 

pollution, it must have foliage and be physiologically active. Annual crops are typically grown for 

6 periods of two to three months. In contrast, perennial species may be photosynthetically active 

7 longer (up to 12 months each year for some species) depending on the species and where it is grown. 

8 In general, the period of maximum physiological activity and thus, potential O3 uptake for vegetation 

9 coincides with some or all of the intra-annual period defined as the O3 season, which varies on a 

state-by-state basis (Figure 3-18). This is because the high temperature and high light conditions that 

11 typically promote the formation of tropospheric O3 also promote physiological activity in vegetation. 

12 There are very limited exceptions to this pattern where O3 can form in the winter in areas in the 

13 western U.S. with intense natural gas exploration (Pinto, 2009, 187038), but this is typically when 

14 plants are dormant and there is little chance of O3 uptake. The selection of any single window of 

time for a national standard to consider hourly O3 concentrations represents a compromise, given the 

16 significant variability in growth patterns and lengths of growing season among the wide range of 

17 vegetation species that may experience adverse effects associated with O3 exposure. 

18 Various intra-annual averaging and accumulation time periods have been considered for the 

19 protection of vegetation. The 2010 proposal for secondary O3 standard (75 FR 2938, (2010, 684211), 

p. 3003) proposed to use the maximum consecutive 3-month period within the O3 season. The U.S. 

21 Forest Service and federal land managers have used a 24-h W126 accumulated for 6 months from 

22 April through September (see FLAG report; Federal land managers' air quality related values 

23 workgroup (FLAG) phase I report, 2000, 088923). However, some monitors in the U.S. are 

24 operational for as little as four months and would not have enough data for a 6-month seasonal 

window. The exposure period in the vast majority of O3 exposure studies conducted in the U.S. has 

26 been much shorter than 6 months. Most of the crop studies done through NCLAN had exposures less 

27 than three months with an average of 77 days. Open-top chamber studies of tree seedlings, compiled 

28 by the EPA, had an average exposure of just over three months or 99 days. In more recent FACE 

29 experiments, SoyFACE exposed soybeans for an average of approximately 120 days per year and the 

Aspen FACE experiment exposed trees to an average of approximately 145 days per year of elevated 

31 O3, which included the entire growing season at those particular sites. Despite the possibility that 

32 plants may be exposed to ambient O3 longer than 3 months in some locations, there is a lack of 

33 exposure experiments conducted for longer than 3 months. 

34 In an analysis of the 3- and 6-month maximum W126 values calculated for over 1,200 AQS 

(Air Quality System) and CASTNET (Clean Air Status and Trend Network) EPA monitoring sites for 

36 the years 2008-2009, it was found that these 2 accumulation periods resulted in highly correlated 

37 metrics (Figure 9-12). The two cumulation periods were centered on the yearly maximum for each 

March 2011 9-106 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

1

2

3

4

5

6

7

monitoring site, and it is possible that this correlation would be weaker if the two periods were not 

temporally aligned. In the U.S., W126 cumulated over 3 months, and W126 cumulated over 6 

months are proxies of one another, as long as the period in which daily W126 is accumulated 

corresponds to the seasonal maximum. Therefore, it is expected that either statistic will predict 

vegetation response equally well. In other words, the strength of the correlation between maximum 

3-month W126 and maximum 6-month W126 is such that there is no material difference in their 

predictive value for vegetation response. 

March 2011 9-107 DRAFT – DO NOT CITE OR QUOTE 
Highest 6 month W126 Highest 6 month W126 

110 

100 

90 

80 

70 

60 

50 

40 

30 

20 

10 

0 

0 10 20 30 40 50 60 70

Highest 3 month W126 

A 

110 

100 

90 

80 

70 

60 

50 

40 

30 

20 

10 

0 

0 10 20 30 40 50 60 70

Highest 3 month W126 

B 

Figure 9-12.Maximum 3-month, 12-h W126 plotted against maximum 6-month, 12-h W126. ﷒Data 

are from the AQS and CASTNET monitors for the years 2008 and 2009. (A) W126, 

3 month versus 6 month, 2008 (Pearson correlation = 0.99); (B) W126, 3 month

versus 6 month, 2009 (Pearson correlation = 0.99).

9.7.4. Ozon e Upta ke/Do s e Mo d eling for Ve g eta tio n

1 Another approach for improving risk assessment of vegetation response to ambient O3 is based 

2 on estimating the O3 concentration from the atmosphere that enters the leaf (i.e., flux or deposition). 

March 2011 9-108 DRAFT – DO NOT CITE OR QUOTE 
 

1

2 

3 

4

5

6 

7

8 

9 

10 

11 

12 

13

14

15

16 

17

18

19

20 

21 

22 

23

24 

25 

26

27 

28

29 

30 

31

32 

33

34 

35 

36 

Interest has been increasing in recent years, particularly in Europe, in using mathematically tractable 

flux models for O3 assessments at the regional, national, and European scale (Matyssek et al., 2008, 

191262)(Paoletti and Manning, 2007, 180174)(Emberson et al., 2000, 040350)(Emberson et al., 

2000, 042537)(ICP M&M, 2004, 677471). Some researchers have claimed that using flux models 

can be used to better predict vegetation responses to O3 than exposure-based approaches (Matyssek 

et al., 2008, 191262). However, other research has suggested that flux models do not predict 

vegetation responses to O3 better than exposure-based models, such as AOT40 (Gonzalez-Fernandez 

et al., 2010, 381357). While some efforts have been made in the U.S. to calculate O3 flux into leaves 

and canopies (Grantz et al., 1997, 026664)(Grantz et al., 1995, 026659)(Grulke et al., 2004, 

042646)(Turnipseed et al., 2009, 588752)(Uddling et al., 2009, 596219)(Bergweiler et al., 2008, 

191656)(Hogg et al., 2007, 199349), little information has been published relating these fluxes to 

effects on vegetation. The lack of flux data in the U.S. and the lack of understanding of 

detoxification processes have made this technique less viable for vulnerability and risk assessments 

in the U.S. 

Flux calculations are data intensive and must be carefully implemented. 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, 025018). 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, 041857). 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, 

041857). 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, 

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 (Uddling et al., 2009, 596219). 

Determining this O3 uptake via canopy and stomatal conductance by necessity relies on models to 

predict flux and ultimately the “effective” flux (Grunhage et al., 2004, 056621)(Massman et al., 

2000, 011616)(Massman, 2004, 055350). “Effective flux” has been defined as the balance between 

O3 flux and detoxification processes (Dammgen et al., 1993, 055312)(Grunhage and Haenel, 1997, 

040392)(Musselman and Massman, 1999, 040706)(Heath et al., 2009, 196783). 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 below. 

The flux-based index is the goal for the “Level II” critical level for assessment of O3 risk to 

vegetation and ecosystems across Europe (Ashmore MEmberson et al., 2004, 056624). 

March 2011 9-109 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

5 

 

 

 

10

 

 

 

15

 

 

 

20 

 

 

25 

 

 

30

 

 

9.7.4.1. Canopy Structure 

1 A factor important in both O3 exposure and uptake is how canopy structure affects O3 

2 concentration in and under forest canopies. There have been several investigations of O3 

3 concentrations under tree canopies (Enders, 1992, 040355)(Fontan et al., 1992, 

4 040370)(Fredericksen et al., 1995, 038898)(Joss and Graber, 1996, 040408)(Kolb et al., 1997, 

052597)(Lorenzini and Nali, 1995, 040691)(Neufeld et al., 1992, 038971)(Samuelson and Kelly, 

6 1997, 040832). In general, they indicated a reduction in O3 of ~20 to 40% in the area below the 

7 canopy but above the shrub/herb layers. An essential component in the determination of the AOT40 

8 as a critical level was the height at which the O3 concentration was measured. The measurement 

9 heights are related to the O3 concentration measured at the top of the canopy, i.e., upper surface 

boundary of the (quasi-) laminar layer (Grunhage and Jager, 2003, 052972). This location is 

11 presumably more closely related to stomatal uptake. Weighting the O3 concentration at this location 

12 takes into account stomatal opening and, if weighted with the Jarvis-Steward factors for radiation, 

13 temperature, and soil moisture, the “toxicologically” effective AOT40 is obtained (Grunhage and 

14 Jager, 2003, 052972). 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., 

16 low soil moisture, high VPD), high concentrations of O3 can occur at the top of the canopy with 

17 minimal risk. 

9.7.4.2. Site and Climate Factors 

18 Soil moisture is a critical factor in controlling O3 uptake through its effect on plant water status 

19 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 

21 literature-based seedling exposure-response data from regions within the southern Appalachian 

22 Mountains that might have experienced O3 exposures sufficient to inhibit growth (Lefohn et al., 

23 1997, 082871). In a small number of areas within the region, O3 exposures and soil moisture 

24 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 

26 exposures in many of the areas and because the hydrologic index used might not reflect actual water 

27 stress. 

9.7.4.3. Pla nt De fe ns e Me c ha nis m – De toxification 

28 The non-stomatal component of plant defenses are the most difficult to quantify, but some 

29 studies are available (Barnes et al., 2002, 040313)(Chen et al., 1998, 040317)(Massman and Grantz, 

1995, 040698)(Plochl et al., 2000, 040802)(Heath et al., 2009, 196783). Massman et al. (2000, 

31 011616) developed a conceptual model of a dose-based index to determine how plant injury response 

32 to O3 relates to the traditional exposure-based parameters. The index used time-varying-weighted 

33 fluxes to account for the fact that flux was not necessarily correlated with plant injury or damage. 

March 2011 9-110 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

1 The model applied only to plant foliar injury and suggested that application of flux-based models for 

2 determining plant damage (yield or biomass) would require a better understanding and quantification 

3 of the relationship between injury and damage. 

9.8. Ozone Exposure -Plant Res pons e Relations hips 

9.8.1. Introd uc tion 

4 The adequate characterization of the effects of O3 on plants for the purpose of setting air 

5 quality standards is contingent not only on the choice of the index used (i.e. SUM06, W126) to 

6 summarize O3 concentrations (Section 9.7), but also on quantifying the response of the plant 

7 variables of interest at specific values of the selected index. The many factors that determine the 

8 response of plants to O3 exposure have been discussed in previous sections. They include species, 

9 genotype and other genetic characteristics (Section 9.4), biochemical and physiological status 

10 (Section 9.4), previous and current exposure to other stressors (Section 9.5), and characteristics of 

11 the exposure itself (Section 9.7). Establishing a secondary air quality standard requires the capability 

12 to generalize those observations, in order to obtain predictions that are reliable enough under a broad 

13 variety of scenarios, taking into account these factors. This section reviews results that have related 

14 specific quantitative observations of O3 exposure with quantitative observations of plant responses, 

15 and the predictions of responses that have been derived from those observations through empirical 

16 models. 

17 For four decades, exposure to O3 at ambient concentrations found in many areas of the U.S. 

18 has been known to cause detrimental effects in plants (U.S. EPA, 2006, 088089)(U.S. EPA, 1996, 

19 080827)(U.S. EPA, 1984, 029711)(U.S. EPA, 1978, 040586). Results published after the 2006 O3 

20 AQCD continue to support this finding, and the following sections deal with the quantitative 

21 characterizations of the relationship, and what new insights may have appeared since 2006. 

22 Detrimental effects on plants include visible injury, decreases in the rate of photosynthesis, reduced 

23 growth, and reduced yield of marketable plant parts. Most published exposure-response data have 

24 been reported O3 effects on the yield of crops and the growth of tree seedlings, and those two 

25 variables have been the focus of the characterization of ecological impacts of O3 for the purpose of 

26 setting secondary air quality standards. In order to support quantitative modeling of exposure27 response relationships, data should preferably include more than three levels of exposure, and some 

28 control of potential confounding or interacting factors should be present in order to model the 

29 relationship with sufficient accuracy. Letting potential confounders, such as other stressors, vary 

30 freely when generating O3 exposure-response data might improve the ‘realism’ of the data, but it also 

31 greatly increases the amount of data necessary to extract a clear quantitative description of the 

32 relationship. Conversely however, experimental settings should not be so exhaustively restrictive as 

33 to make generalization outside of them problematic. During the last four decades, many of the 

34 studies of the effects of O3 on growth and yield of plants have not included enough levels of O3 to 

March 2011 9-111 DRAFT – DO NOT CITE OR QUOTE 
 

1 

2

3 

4 

5 

6

7 

8

9

10 

11

12 

13 

14

15 

16

17 

18

19

20

21

22 

23 

24 

25 

26

27

28 

29 

30

31

32 

33 

34

35 

36

37 

38 

39 

parameterize more than the simplest linear model. The majority of these studies have only contrasted 

two levels, ambient and elevated, or sometimes three by adding carbon filtration in OTC studies, 

with little or no consideration of quantitatively relating specific values of exposure to specific values 

of growth or yield. This is not to say that studies that did not include more than two or three levels of 

O3 exposure, or studies that were conducted in uncontrolled environments, do not provide exposureresponse information that is highly relevant to reviewing air quality standards. In fact, they can be 

essential in verifying the agreement between predictions obtained through the empirical models 

derived from experiments such as NCLAN, and observations. The consensus of model predictions 

and observations from a variety of studies conducted in other locations, at other times, and using 

different exposure methods, greatly increases confidence in the reliability of both. Furthermore, if 

they are considered in the aggregate, studies with few levels of exposure or high unaccounted 

variability can provide additional independent estimates of decrements in plant growth and yield, at 

least within a few broad categories of exposure. 

Extensive exposure-response information on a wide variety of plant species has been produced 

by two long-term projects that were designed with the explicit aim of obtaining quantitative 

characterizations of the response of such an assortment of crop plants and tree seedlings to O3 under 

North American conditions: the NCLAN project for crops, and the EPA National Health and 

Environmental Effects Research Laboratory, Western Ecology Division tree seedling project 

(NHEERL/WED). 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 U.S. (Heck et al., 1982, 039525)(Heck et al., 1991, 042621). The cultural 

conditions used in the NCLAN studies approximated typical agronomic practices, and 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 (3) to advance understanding of cause-and-effect relationships that determine 

crop responses to pollutant exposures. 

NCLAN experiments yielded 54 exposure-response curves for 12 crop species, some of which 

were represented by multiple cultivars at several of 6 locations throughout the U.S. The 

NHEERL/WED project was initiated by EPA in 1988 with the same objectives for tree species, and 

yielded 49 exposure-responses curves for multiple genotypes of 11 tree species grown for up to three 

years in Oregon, Michigan, and the Great Smoky Mountain National Park. Both projects used OTCs 

to expose plants to three to five levels of O3. Eight of the 54 crop datasets were from plants grown 

under a combination of O3 exposure and experimental drought conditions. Figure 9-13 through 9-16 

summarize some of the NCLAN and NHEERL/WED results. 

It should be noted that data from FACE experiments might also be used for modeling 

exposure-response. They only use two levels of O3 (ambient concentration at the site and a multiple 

of it), but given that the value of both levels of exposure changes every year, and that they are 

typically run for many consecutive years, aggregating data over time produces twice as many levels 

March 2011 9-112 DRAFT – DO NOT CITE OR QUOTE 
 

1 

2 

3

4

5

6

7 

8 

9

10 

11 

12 

13

14 

15 

16 

17 

18 

19 

20 

21 

22 

23

24 

25 

26 

27

28 

29

30 

31

32

33

34 

35 

36

37 

38 

39

of O3 as there are years. As described in Section 9.3.4, FACE experiments seek to impose fewer 

constraints on the growth environment then OTCs. As a consequence, FACE studies have to contend 

with larger variability, especially year-to-year variability, but the difference in experimental 

conditions between the two methodologies makes comparisons between their results especially 

useful. 

Growth and yield of at least one crop (soybean) has been investigated in yearly experiments 

since 2001 at a FACE facility in Illinois (Morgan et al., 2006, 079186)(University of Illinois, 2010, 

670286), however almost all analyses of SoyFACE published so far have been based on subsets of 

one or two years, and have only contrasted ambient versus elevated O3 as categorical variables. They 

have not modeled the response of growth and yield to O3 exposure continuously over the range of 

exposure values that have occurred over time. The only exception is a study by Betzelberger et al. 

(2010, 644183), who used a linear regression model on data pooled over 2 years. Likewise, trees of 

three species (trembling aspen, paper birch, and sugar maple) were grown between 1998 and 2009 in 

a FACE experiment located in Rhinelander, Wisconsin (Dickson et al., 2000, 628220)(Pregitzer et 

al., 2008, 191677). The Aspen FACE experiment has provided extensive data on responses of trees 

beyond the seedling stage under long-term exposure, and also on ecosystem-level responses (Section 

9.6), but the only attempt to use those data in a continuous model of the response of tree growth to 

O3 exposure (Percy et al., 2007, 093287) suffered severe methodological problems, some of which 

are discussed in Section 9.8.3. Finally, one experiment was able to exploit a naturally occurring 

gradient of O3 concentrations to fit a linear regression model to the growth of cottonwood (Gregg et 

al., 2003, 046996; Gregg et al., 2006, 186961). Factors such as genotype, soil type and soil moisture 

were under experimental control, and the authors were able to partition out the effects of potential 

confounders such as temperature, atmospheric N deposition, and ambient CO2. 

A serious difficulty in assessing results of exposure-response research is the multiplicity of O3 

metrics that have been used in reporting. As described in Section 9.7, metrics that entail either 

weighting or thresholding of hourly values cannot be converted into one another, or into unweighted 

metrics such as hourly average. When using weighted or thresholded metrics, which include W126, 

AOTx or SUMx metrics; O3 exposure at every exposure-response data point must be computed 

separately for each metric, starting with the hourly data. Comparisons of exposure-response models 

can only be made between studies that used the same metric, and the value of exposure at which a 

given plant response is expected on one scale of exposure cannot be exactly converted to another 

scale. Determining the exposure value at which an effect would be observed in a different metric can 

only be accomplished by first computing the experimental exposures in this metric from the hourly 

data, then estimating (fitting) model coefficients again. This problem is irremediable, although useful 

comparisons might be made using categorical exposures such as ‘current ambient exposure’ or ‘2050 

projected exposure’, which can serve as a common reference for quantitative values expressed in 

various metrics. Studies that contained growth or yield exposure-response data at few levels of 

exposure, and/or using metrics other than W126 are summarized in Tables 9-16 and 9-17. 

March 2011 9-113 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

9.8.2.	 1996 and 2006 Ozo ne AQCDs Es tima te s Of Crop Yield Lo s s And 

Tre e S ee dlin g Bio mas s Lo s s 

1	 The 1996 and 2006 O3 AQCDs relied extensively on analyses of NCLAN and NHEERL/WED 

2 by Lee et al. (1987, 042135; 1988, 042136; 1989, 042137; 1994, 043268), Hogsett et al. (1997,
 

3 040402), Lee and Hogsett (1999, 040451), Heck et al. (1984, 039380), Rawlings and Cure (1985,
 

4 039419), Lesser et al. (1990, 043015), and Gumpertz and Rawlings (1992, 043259). Those analyses
 

5 concluded that a three-parameter Weibull model –
 

β  W126  − 

 



 η 

6 Y =α e Equation 9-2﷒

6 is the most appropriate model for the response of absolute yield and growth to O3 exposure, because 

7 of the interpretability of its parameters, its flexibility (given the small number of parameters), and its 

8 tractability for estimation. In addition, removing the intercept α results in a model of relative yield 

9 (yield relative to [yield at exposure=0]) without any further reparameterization. Formulating the 

10 model in terms of relative yield or relative yield loss (yield loss=[1 − relative yield]) is essential in 

11 comparing exposure-response across species, genotypes, or experiments for which absolute values of 

12 the response may vary greatly. In the 1996 and 2006 O3 AQCDs, the two-parameter model of relative 

13 yield was used in deriving common models for multiple species, multiple genotypes within species, 

14 and multiple locations. 

15 Given the disparate species, genotypes, and locations that were included in the NCLAN and 

16 NHEERL/WED projects, and in the absence of plausible distributional assumptions with respect to 

17 those variables, a three step process using robust methods was used to obtain parameter estimates 

18 that could be generalized. The models that were derived for each species or group of species were 

19 referred to as median composite functions. In the first step, the three parameters of the Weibull 

20 model were estimated (fitted) for absolute yield or biomass data from each NCLAN and 

21 NHEERL/WED experiment (54 crop datasets and 49 tree seedling datasets), using nonlinear 

22 regression. When data were only available for three levels of exposure because of experimental 

23 problems, the shape parameter β was constrained to 1, reducing the model to an exponential decay 

24 model. In the second step, α was dropped, and predicted values of relative yield or biomass were 

25 then computed for 12-hr W126 exposures between 0 and 60 ppm-h. At each of these W126 exposure 

26 values, the 25th, 50th, and 75th percentiles of the response were identified among the predicted 

27 curves of relative response. For example, for the 34 NCLAN studies of 12 crop species grown under 

28 non-droughted conditions for a complete cropping cycle (Figure 9-13), the 3 quartiles of the 

29 response were identified at every integer value of W126 between 0 and 60. The third step fitted a 

30 two-parameter Weibull model to those percentiles, yielding the median composite function for the 

31 relative yield or biomass response to O3 exposure for each grouping of interest (e.g., all crops, all 

32 trees, all datasets for one species), as well as composite functions for the other quartiles. In the 1996 

33 and 2006 O3 AQCDs, this modeling of crop yield loss and tree seedling biomass loss was conducted 

March 2011	 9-114 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

1 using the SUM06 metric for exposure. This section updates those results by using the 12-hr W126 as 

2 proposed in 2007 (72 FR 37818 (2007, 684055)) and 2010 (75 FR 2938 (2010, 684211), p. 3003). 

3 Figures 9-13 through 9-16 present quantiles of predicted relative yield or biomass loss at seven 

4 values of the 12-h W126 for some representative groupings of NCLAN and NHEERL/WED results. 

5 Tables 9-8 through 9-10 give the 90-day 12-h W126 O3 exposure values at which 10 and 20% yield 

6 or biomass losses are predicted in 50 and 75% of crop or tree species using the composite functions. 

100 

Percent Yield Loss (%) 

90 

80 

70 

60 

50 

40 

30 

20 

10 

0 

10th Pctile 

25th Pctile 

50th Pctile 

75th Pctile 

34 crop datasets 90th Pctile 

0 10 20 30 40 50 60 

12 hr W126 (ppm-hr) 

Source of Weibull parameters: Lee and Hogsett (1996, 670278). 

Figure 9-13. Quantiles of predicted relative yield loss for 34 NCLAN crop experiments. ﷒Quantiles 

of the predicted relative yield loss at 7 values of 12-hour W126 for 34 Weibull curves

estimated using nonlinear regression on data from 34 studies of 12 crop species 

grown under well-watered conditions for the full duration of 1 cropping cycle. 

March 2011 9-115 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

100 

90 

80 

70 

60 

50 

40 

30 

20 

10 

0 

25th Pctile 

50th Pctile 

75th Pctile 

11 Soybean datasets 

100 

90 

80 

70 

60 

50 

40 

30 

20 

10 

0 

25th Pctile 

50th Pctile 

75th Pctile 7 Wheat datasets 

Percent Yield Loss (%) Percent Yield Loss (%) 100 

90 

80 

70 

60 

50 

40 

30 

20 

10 

0 

0 10 20 30 40 50 60 0 10 20 30 40 50 60 

100
 

5 Cotton datasets 90
 2 Corn datasets 

80 

70 

60 

75th Pctile 

50 

40 

30 

20 50th Pctile 

10 

0 

0 10 20 30 40 50 60 0 10 20 30 40 50 60 

12 hr W126 (ppm-hr 12 hr W126 (ppm-hr) 

Source of Weibull parameters: Lee and Hogsett (1996, 670278). 

25th Pctile 

50th Pctile 

Figure 9-14. Quantiles of predicted relative yield loss for 4 crop species in NCLAN experiments.﷒

Quantiles of the predicted relative yield loss at 7 values of 12-h W126 for Weibull

curves estimated using nonlinear regression for 4 species grown under wellwatered conditions for the full duration of 1 cropping cycle. The number of studies

available for each species is indicated on each plot. 

March 2011 9-116 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

10th Pctile 

25th Pctile 

50th Pctile 

75th Pctile 

90th Pctile 

0 

10 

20 

30 

40 

50 

60 

70 

80 

90 

100 

Percent Biomass Loss (%) 

49 tree seedlings datasets 

0 10 20 30 40 50 60 

90 day 12 hr W126 (ppm-hr) 

Source of Weibull parameters: Lee and Hogsett (1996, 670278). 

Figure 9-15. Quantiles of predicted relative biomass loss for 49 tree species in NHEERL/WED 

experiments. ﷒Quantiles of the predicted relative above-ground biomass loss at 7

values of 12-h W126 for 49 Weibull curves estimated using nonlinear regression on 

data from 49 studies of 11 tree species grown under well-watered conditions for 1 or

2 years. Curves were standardized to 90-day W126. 

March 2011 9-117 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

40 

50 

80 

90 

100 

0 

100 

90 11 Ponderosa pine datasets 

80 

25th Pctile 

50th Pctile 

75th Pctile 

14 Aspen datasets 

0 10 20 30 40 50 60 

Percent Biomass Loss (%) Percent Biomass Loss (%) 

70 70

60 60

50 

40 

25th Pctile 

50th Pctile 

75th Pctile 

0 10 20 30 40 50 60 

30 30

20 20

10 10

0 0

100 100 

50th Pctile 

7 Douglas f ir datasets 

0 10 20 30 40 50 60 

50th Pctile 

3 Tulip poplar datasets 

0 10 20 30 40 50 60 

90 day 12 hr W126 (ppm-hr) 

90 

80 

70 

60 

50 

40 

30 

20 

10 

90 

80 

70 

60 

50 

40 

30 

20 

10 

0 

90 day 12 hr W126 (ppm-hr) 

Source of Weibull parameters: Lee and Hogsett (1996, 670278). 

Figure 9-16. Quantiles of predicted relative biomass loss for 4 tree species in NHEERL/WED

experiments. ﷒Quantiles of the predicted relative above-ground biomass loss at 7

exposure values of 12-h W126 for Weibull curves estimated using nonlinear

regression on data for 4 tree species grown under well-watered conditions for 1 or 2 

year. Curves were standardized to 90-day W126. The number of studies available for

each species is indicated on each plot. 

March 2011 9-118 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

 

Table 9-8. Ozone exposures at which 10 and 20% yield loss is predicted for 50 and 75% of crop 

species, based on composite functions for the 50th and 75th percentiles of 34 Weibull

curves for relative yield loss data from 34 non-droughted NCLAN studies of 12 crop species;

curves were standardized to 90-day W126 

90-day 12-h W126 for 10% yield loss (ppm-h) 90-day 12-h W126 for 20% yield loss (ppm-h) 

Model for the 50th Percentile of 34 curves 

Relative yield=exp(-(W126/104.82)**1.424) 22 37 

Model for the 75th Percentile of 34 curves 

Relative yield=exp(-(W126/78.12)**1.415) 16 27 

Source of parameters for the 34 curves: Lee and Hogsett (1996, 670278) 

Table 9-9. Ozone exposures at which 10 and 20% yield loss is predicted for 50 and 75% of crop 

species under drought conditions and adequate moisture, based on composite functions

for the 50th and 75th percentiles of 16 Weibull curves for relative yield loss data from 8

NCLAN studies that paired droughted and watered conditions for the same genotype;

curves were standardized to 90-day W126 

90 day 12-h W126 for 10% yield loss 

(ppm-h) 

90 day 12-h W126 for 20% yield loss 

(ppm-h) 

Model for the 50th Percentile of 2×8 curves 

Watered Relative yield=exp(-(W126/132.86)**1.170) 19 37 

Droughted Relative yield=exp(-(W126/179.84)**1.713) 48 75 

Model for the 75th Percentile of 2×8 curves 

Watered Relative yield=exp(-(W126/90.43)**1.310) 16 29 

Droughted Relative yield=exp(-(W126/105.16)**1.833) 31 46 

Source of parameters for the 16 curves: Lee and Hogsett (1996, 670278) 

March 2011 9-119 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

Table 9-10. Ozone exposures at which 10 and 20% biomass loss is predicted for 50 and 75 % of tree

species, based on composite functions for the 50th and 75th percentiles of 49 Weibull

curves for relative above-ground biomass loss data from 49 studies of 11 tree species 

grown under well-watered conditions for 1 or 2 year; curves were standardized to 90-day

W126 

90 day 12 h W126 for 10% yield loss 90 day 12 h W126 for 20% yield loss 

(ppm-h) (ppm-h) 

Model for the 50th Percentile of 49 curves 

Relative yield=exp(-(W126/131.57)**1.242) 21	 39 

Model for the 75th Percentile of 49 curves 

Relative yield=exp(-(W126/65.49)**1.500) 15	 24 

Source of parameters for the 49 curves: Lee and Hogsett (1996, 670278) 

9.8.3.	 Valida tio n o f 1996 a nd 2006 Ozo n e AQCD Mo d els an d Meth od olo gy

Us ing th e 90 day 12-h W126 a nd Cu rre n t FACE Data. 

1 Since the completion of the NCLAN and NHEERL/WED projects, almost no studies have 

2 been published that could provide a basis for estimates of exposure-response that can be compared to 

3 those of the 1996 and 2006 O3 AQCDs. Most experiments, regardless of exposure methodology, 

4 include only two levels of exposure. In addition, very few studies have included measurements of 

5 exposure using the W126 metric, or the hourly O3 concentration data that would allow computing 

6 exposure using the W126. Two FACE projects, however, were conducted over multiple years, and by 

7 adding to the number of exposure levels over time, may support independent model estimation and 

8 prediction using the same model and the same robust process as summarized in Section 9.8.2. 

9 Hourly O3 data were available from both FACE projects. 

10 The SoyFACE project is situated near Champaign, IL, and comprises 32 octagonal rings (20m11 diameter), 4 of which in a given year are exposed to ambient conditions, and 4 of which are exposed 

12 to elevated O3 as a fixed proportion of the instantaneous ambient concentration (Betzelberger et al., 

13 2010, 644183; Morgan et al., 2004, 072764; Morgan et al., 2006, 079186)(University of Illinois, 

14 2010, 670286). Since 2002, yield data have been collected for up to 8 genotypes of soybean grown 

15 in subplots within each ring. The Aspen FACE project is situated in Rhinelander, WI, and comprises 

16 12 rings (30m-diameter), 3 of which are exposed to ambient conditions, and 3 of which are exposed 

17 to O3 as a fixed proportion of the instantaneous ambient concentration (Dickson et al., 2000, 628220; 

18 Karnosky et al., 2005, 095556; Pregitzer et al., 2008, 191677). In the summer of 1997, half the area 

19 of each ring was planted with small (five to seven leaf sized) clonally propagated plants of five 

20 genotypes of trembling aspen, which were left to grow in those environments until 2009. Biomass 

21 data are currently available for the years 1997-2005 (King et al., 2005, 191701). Ozone exposure in 

22 these two FACE projects can be viewed as a categorical variable with two levels: ambient, and 

23 elevated. However, this overlooks the facts that yearly ambient and elevated exposure both vary with 

24 every year, and that the proportionality between them also changes. This change has two sources: 

March 2011	 9-120 DRAFT – DO NOT CITE OR QUOTE 
 

1

2

3

4

5 

6 

7 

8

9 

10 

11 

12 

13

14

15 

16 

17 

18

19 

20 

21

22 

23

24

25 

26 

27 

28 

29 

30

31 

32 

33 

34

35 

36

37

first, the dispensing of O3 into the elevated exposure rings varies from the proportionality set point to 

some extent, and for SoyFACE, the set point changed between years. Second, the proportionality 

does not propagate predictably from the hourly data to the yearly value when using thresholded or 

concentration-weighted cumulative metrics (such as AOT40, SUM06 or W126). Hourly average 

elevated exposures that are, for example, 1.5 times ambient do not result in AOT40, SUM06 or 

W126 values that are some constant multiple of the ambient values of those indices. The greater the 

fraction of elevated hourly values that are above the threshold or heavily weighted, compared to the 

fraction of hourly ambient values that are, the greater the difference between ambient and elevated 

yearly exposure, as measured using weighted cumulative indices. When elevated exposure is a 

multiple of ambient hourly intervals, the number of hours for which elevated exposure meets the 

threshold for inclusion can vary widely, even though the hourly mean for the year retains the 

proportionality. As a consequence, the number of exposure levels in multi-year experiments is twice 

the number of years. In the case of SoyFACE for the period between 2002 and 2008, ambient 

exposure in the highest year was approximately equal to elevated exposure in the lowest year, with 

14 levels of O3 exposure evenly distributed from lowest to highest. The particular conditions of the 

Aspen FACE experiment resulted in 12 exposure levels between 1998 and 2003, but they were not as 

evenly distributed between minimum and maximum over the 6-year period. 

There are necessary differences in the modeling of exposure-response in annual plants such as 

soybean, and in perennial plants such as aspen trees, when exposure takes place over multiple years. 

In annual plants, responses recorded at the end of the life cycle, i.e., yearly, are analyzed in 

relationship to that year's exposure. Yield of soybeans is affected by exposure during the year the 

crop was growing, and a new crop is planted every year. Thus an exposure-response relationship can 

be modeled from yearly responses matched to yearly exposures, with those exposure-response data 

points having been generated in separate years. For perennial organisms, which are not harvested 

yearly and continue to grow from year to year, such pairing of exposure and response cannot be done 

without accounting for time. Not only does the size of the organism at the beginning of each year of 

exposure increase, but size is also dependent on the exposure from previous years. Therefore the 

relationship of response and exposure must be analyzed either one year at a time, or by standardizing 

the response as a yearly increment relative to size at the beginning of each year. Furthermore, the 

relevant measurement of exposure is cumulative, or cumulative yearly average exposure, starting in 

the year exposure was initiated, up to the end of the year of interest. When analyzing the growth of 

trees over several years, it would be evidently incorrect to pair the exposure level in every discrete 

year with absolute size of the trees that year, and posit a direct relationship between them. In the 

Aspen FACE experiment, for example, one could not establish an exposure-response relationship by 

matching 12 yearly exposures and 12 yearly tree sizes, as if size did not also depend on time. This is 

the basis of the 2007 study of Aspen FACE data by Percy et al. (2007, 093287), and that study was 

therefore not informative. 

March 2011 9-121 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

1

2

3

4

5

6

7

8

9

10

11

12

13

14

15

16

17

18

19

20

21

22

23

24

25

26

27

28

29

30

31

32

33

34

35

36

37

38

9.8.3.1. Co mpa ris on of NCLAN-Ba s e d P re dic tion a nd SoyFACE Da ta . 

For this ISA, EPA conducted a comparison between yield as predicted by the composite 

function three-step process (Section 9.8.2) using NCLAN data for soybean yield as observed in 

SoyFACE. The median composite function for relative yield was derived for the 11 NCLAN soybean 

Weibull functions for non-droughted studies, and several comparisons between prediction and 

SoyFACE observations were conducted as follows. 

For the years 2007 and 2008, SoyFACE yield data were available for 7 and 6 genotypes, 

respectively. The EPA used those data to compare the change in relative yield observed in a given 

year between ambient O3 and elevated O3 in SoyFACE, versus the change in relative yield predicted 

by the NCLAN-based median composite function between those same two values of O3 exposure. 

The two parameter median composite function for relative yield of soybean was used to predict yield 

response at the two observed values of exposure in each year, and the change between yield under 

ambient and elevated was compared to the change observed in SoyFACE for the relevant year (Table 

9-11). This approach results in a direct comparison of predicted versus observed change in yield. 

Because the value of relative response between any two values of O3 exposure is independent of the 

intercept α, this comparison does not require prediction of the absolute values of the responses. 

Since comparisons of absolute values might be of interest, the predictive functions were also 

scaled to the observed data, using two distinct methods. In the first method, the intercept α was 

calculated algebraically by entering the observed W126 value at ambient exposure and the 

corresponding value of the response into the three parameter model with the shape and scale 

parameters (β and η) set to their value for the NCLAN predictive model. This method provides a 

comparison between the response observed under elevated exposure, and the response that would 

have been predicted with only the knowledge of what the response was under ambient exposure 

(Table 9-12; Method 1). In the second method, the intercept for the NCLAN predictive model was 

estimated by regression using both ambient and elevated data. This method gives a comparison of 

prediction and observation that takes all the observed information into account to provide the best 

possible estimate of the intercept, and thus the best possible scaling (Table 9-12, Method 2 and 

Figure 9-17). It should be noted that the similarity to each other of the predictions obtained by these 

two scaling methods is a reflection of the accuracy of the predictions: the distance between the 

predictions from the two methods increases as the distance between prediction and observation 

increases. 

For the comparison of NCLAN and SoyFACE, this validation was possible for 2007 and 2008, 

where data for 7 and 6 soybean genotypes, respectively, were available. The median composite 

function for relative yield was derived for the 11 NCLAN soybean Weibull functions for 

nondroughted studies, and the values of median yield under ambient exposure at SoyFACE in 2007 

and 2008 were used to obtain an estimate of the intercept α for the NCLAN median function in each 

of the two years 

Table 9-11 presents the results of ambient/elevated relative yield comparisons between the 

NCLAN-derived predictions and SoyFACE observations. Table 9-12 presents the results of 

March 2011 9-122 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

 

 

1 comparisons between NCLAN-derived predictions and SoyFACE observations of yield, using two 

2 methods for scaling the predictive function. Figure 9-17 presents yield observed in two years in 

3 SoyFACE, and predicted by the median composite function derived from NCLAN data using the 

4 second scaling method as described, with the intercept estimated using 14 observations in 2007, and 

5 12 in 2008. 

6 Finally, data were also available for one additional genotype from 2003 to 2007. By 

7 aggregating data for each genotype over the years (5 years for one genotype, 2 years for the 6 

8 others), a composite function for the 25th, 50th, and 75th percentiles was then developed for 

9 SoyFACE, and compared to the corresponding NCLAN-based function. NCLAN functions were 

10 obtained using 12-hr W126 standardized to 90 days. SoyFACE 12-hr W126 was cumulated over the 

11 highest 90 days of the growing season. 

12 

Table 9-11. Comparison between relative yield observed in the SoyFACE experiment, and relative yield

predicted at the same values of ozone by the median composite function for NCLAN (twoparameter relative yield model) 

Year 90-day 12-h W126 (ppm-h) Yield, 

Elevated Relative to Ambient 

Ambient Elevated Observed in SoyFACE Predicted by NCLAN 

2007 4.39 46.23 0.76 0.75 

2008 3.23 28.79 0.88 0.85 

Table 9-12. Comparison between yield observed in the SoyFACE experiment and yield predicted at the 

same values of ozone by the median composite function for NCLAN (three-parameter 

absolute yield model), using two scaling methods to calculate the intercept 

Year 90-day 12-h W126 (ppm-h) 

Ambient Elevated 

Yield observed in 

SoyFACE ambient

(g/m2) 

Yield observed in 

SoyFACE elevated

(g/m2) 

Yield predicted by

NCLAN in elevated, 

Method 1 (g/m2) 

Yield predicted by

NCLAN in ambient, 

Method 2 (g/m2) 

Yield predicted by

NCLAN in elevated, 

Method 2 (g/m2) 

2007 4.39 46.23 305.2 230.6 227.6 309.2 230.6 

2008 3.23 28.79 344.8 304.4 293.5 350.3 298.2 

March 2011 9-123 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

 

 

400 

350 

300 

2007, 7 genotypes 

400 

350 

2008, 6 genotypes 

300 

250 250 

Yield (g/m2) 

Yield (g/m2) 

200 

150 

200 

150 

100 100

50 50

0 0

0 10 20 30 40 50 60 70 0 10 20 30 40 50 60 70 

90 day 12 hr W126 (ppm-hr) 90 day 12 hr W126 (ppm-hr) 

Source of data: Betzelberger et al. (2010, 644183);Morgan et al. (2006, 079186); Lee and Hogsett (1996, 670278). 

Note: Black dots are median of 7 or 6 soybean genotypes in SoyFACE (2007, 2008); bars are IQR for genotypes; dashed line is median 

composite model for 11 studies in NCLAN. 

Figure 9-17. Comparison of yield observed in SoyFACE experiment in a given year with yield 

predicted by the median composite function based on NCLAN. 

1 Finally, a composite function for the 25th, 50th, and 75th percentiles was developed from 

2 SoyFACE annual yield data by pooling one genotype from 2003 to 2007, and six genotypes in 2007 

3 and 2008. NCLAN functions were obtained using 12-h W126 standardized to 90 days. SoyFACE 

4 12-h W126 was cumulated over the highest 90 days of the growing season. The correlation between 

5 W126 cumulated for the entire season in SoyFACE and W126 for the highest 90 days was greater 

6 than 0.99 in all years. The same process was used for SoyFACE: first, the three parameter Weibull 

7 model described in Section 9.8.2 was estimated using nonlinear regression on exposure-yield data 

8 for each genotype separately, over the years for which data were available, totaling seven curves. 

9 The 25th, 50th, and 75th percentiles of the predicted values for the two parameter relative yield 

10 curves were then identified at every integer of W126 between 0 and 60, and a two-parameter Weibull 

11 model estimated by regression for the three quartiles. The comparison between these composite 

12 functions for the quartiles of relative yield loss in SoyFACE and the corresponding composite 

13 functions for NCLAN is presented in Figure 9-18. 

March 2011 9-124 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

 

 

 

 

 

FACE 75th Pctile 

FACE median 

FACE 25th Pctile 

NCLAN 75th Pctile 

NCLAN median 

NCLAN 25th Pctile 

0 

10 

20 

30 

40 

50 

60 

70 

80 

90 

100 

Percent Yield Loss (%) 

0 10 20 30 40 50 60 70 

90 day 12 hr W126 (ppm-hr) 

Source of data: Betzelberger et al. (2010, 644183);Morgan et al. (2006, 079186); Lee and Hogsett (1996, 670278). 

Figure 9-18. Comparison of composite functions for the quartiles of 7 curves for 7 genotypes of

soybean grown in the SoyFACE experiment, and for the quartiles of 11 curves for 5 

genotypes of soybean grown in the NCLAN project. 

1 As seen in Tables 9-11 and 9-12, and in Figure 9-17, the agreement between predictions based 

2 on NCLAN data and SoyFACE observations was notably close in single-year comparisons. Together 

3 with the very high agreement between median composite models for NCLAN and SoyFACE, it 

4 provides very strong mutual confirmation of those two projects' results with respect to the response 

5 of yield of soybeans to O3 exposure. It is readily apparent from these results that the methodology 

6 described in Section 9.8.2 for obtaining predictions of yield or yield loss from NCLAN data is 

7 strongly validated by SoyFACE results. As described in Section 9.3, the exposure technologies used 

8 in the two projects were in sharp contrast, specifically with respect to the balance each achieved 

9 between control of potential interacting factors or confounders, and fidelity to real world conditions. 

10 The comparisons that EPA conducted therefore demonstrate that the methodology used in developing 

11 the composite functions is resistant to the influence of nuisance variables, and that predictions are 

12 reliable. They may also suggest that the aspects in which the two exposure technologies differ have 

13 less influence on exposure-response than initially supposed. 

9.8.3.2.	 Co mpa ris on of NHEERL/WED-Ba s e d Pre dic tion of Tre e Bioma s s Re s p ons e

a nd As pe n FACE Da ta 

14 EPA also conducted two comparisons between prediction of above-ground biomass loss based 

15 on NHEERL/WED results and observations from Aspen FACE. The median composite function was 

16 developed from NHEERL/WED data for 11 studies that used wild-type seedlings of aspen as well as 

March 2011	 9-125 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

 

 

 

 

1 four clonally propagated genotypes. All plants were grown in OTCs for one growing season before 

2 being destructively harvested. Aspen FACE data were from clonally propagated trees of five 

3 genotypes grown from 1998 to 2003, with above-ground biomass calculated using allometric 

4 equations derived from data for trees harvested destructively in 2000 and 2002 (King et al., 2005, 

5 191701). 

6 The two parameter median composite function for relative biomass was used to predict 

7 biomass response under the observed elevated exposure, relative to its value under observed ambient 

8 exposure. EPA first tested the accuracy of the prediction of biomass at elevated exposure relative to 

9 biomass at ambient exposure, for each separate year of Aspen FACE. Comparisons between 

10 observed and predicted biomass values were then conducted for each year by scaling the predictive 

11 function to yearly Aspen FACE data using the two scaling methods described in Section 9.8.3.1. 

12 Yearly 90 day 12-hour W126 values for Aspen FACE were computed as the cumulative average 

13 from the year of planting up to the year of interest. A comparison of composite functions between 

14 NHEERL/WED and Aspen FACE, similar to the one performed for NCLAN and SoyFACE, was not 

15 possible: as discussed in the introduction to Section 9.8, the pairing of 12 exposure values from 

16 separate years and 12 values of biomass cannot be the basis for a model of exposure-response, 

17 because the trees continued growing for the six-year period of exposure. Table 9-13 presents the 

18 results of ambient/elevated relative biomass comparisons between the NHEERL/WED-derived 

19 predictions and Aspen FACE observations. Table 9-14 presents the results of comparisons between 

20 NHEERL/WED-derived predictions and Aspen FACE observations of biomass, using two methods 

21 for scaling the predictive function. Figure 9-19 presents biomass observed in six years at Aspen 

22 FACE, and predicted by the median composite function derived from NHEERL/WED data using the 

23 second scaling method as described, with the intercept estimated using 2 observations in each year. 

Table 9-13. Comparison between above-ground biomass observed under elevated ozone in Aspen 

FACE experiment in 6 year, relative to above-ground biomass observed under ambient

ozone and relative above-ground biomass above-ground biomass at the same values of

ozone predicted by the median composite function for NHEERL/WED (two-parameter

relative biomass model) 

90-day 12-h W126 (ppm-h) Above-Ground Biomass, Year Cumulative Average Elevated Relative To Ambient 

Ambient Elevated Observed in Aspen FACE Predicted by NHEERL/WED 

1998 3.19 30.08 0.75 0.74 

1999 2.61 33.85 0.70 0.70 

2000 2.43 30.16 0.71 0.74 

2001 2.55 31.00 0.71 0.73 

2002 2.51 30.27 0.69 0.74 

2003 2.86 29.12 0.71 0.75 

March 2011 9-126 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

Table 9-14. Comparison between above-ground biomass observed in Aspen FACE experiment in 

6 year and biomass predicted by the median composite function based on NHEERL/WED

(three-parameter absolute biomass model), using 2 scaling methods to calculate the

intercept 

Year 90 day 12-h W126 (ppm-h)

Cumulative Average 

Biomass Observed in Aspen

FACE (g/m2

) 

Biomass Predicted by 

NHEERL/WED,

Method 1 (g/m2) 

Biomass Predicted by NHEERL/WED,

Method 2 (g/m2) 

Ambient Elevated Ambient Elevated Elevated Ambient Elevated 

1998 3.19 30.08 274.7 204.9 202.3 276.0 203.2 

1999 2.61 33.85 955.3 673.3 665.9 958.7 668.3 

2000 2.43 30.16 1400.3 998.6 1036.0 1382.4 1022.8 

2001 2.55 31.00 1620.7 1154.9 1183.7 1607.0 1173.7 

2002 2.51 30.27 2125.9 1468.41 1566.7 2079.0 1532.1 

2003 2.86 29.12 2695.2 1907.8 2022.5 2640.1 1981.2 

1998 

1999 

2000 

2001 

2002 

2003 

0 

500 

1000 

1500 

2000 

2500 

3000 

Biomass (g/m2 ) 

0 10 20 30 40 50 60 70 

90 day 12 hr W126 (yearly cumulative average, ppm-hr) 

Source of data: King et al. (2005, 191701), Lee and Hogsett (1996, 670278). 

Note: Black dots are aspen biomass/m2 for 3 FACE rings filled with an assemblage of 5 clonal genotypes of aspen at Aspen FACE; bars 

are SE for 3 rings; dashed line is median composite model for 4 clonal genotypes and wild-type seedlings in 11 NHEERL/WED 1-year 

OTC studies. 

Figure 9-19. Comparison between above-ground biomass observed in Aspen FACE experiment

in 6 year and biomass predicted by the median composite function based on 

NHEERL/WED. 

March 2011 9-127 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

5 

 

 

 

10 

 

 

15

 

20 

 

 

25 

 

30

 

 

35 

 

 

1 As in the comparisons between NCLAN and SoyFACE, the agreement between predictions 

2 based on NHEERL/WED data and Aspen FACE observations was exceptionally close. The results of 

3 the two projects strongly reinforce each other with respect to the response of aspen biomass to O3 

4 exposure. The methodology used for obtaining the median composite function is shown to be 

capable of deriving a predictive model despite potential confounders, and despite the added 

6 measurement error that is expected from calculating biomass using allometric equations. In addition, 

7 the function based on one year of growth was shown to be applicable to subsequent years. 

8 The results of experiments that used different exposure methodologies, different genotypes, 

9 locations, and durations converged to the same values of response to O3 exposure for each of two 

very dissimilar plant species, and predictions based on the earlier experiments were validated by the 

11 data from current ones. However, in these comparisons, the process used in establishing predictive 

12 functions involved aggregating data over variables such as time, locations, and genotypes, and the 

13 use of a robust statistic (quartiles) for that aggregation. The validating data, from SoyFACE and 

14 Aspen FACE, were in turn aggregated over the same variables. The accuracy of predictions is not 

expected to be conserved for individual values of those variables over which aggregation occurred. 

16 For example, the predicted values for soybean, based on data for five genotypes, are not expected to 

17 be valid for each genotype separately. As shown in the validation, however, aggregation that 

18 occurred over different values of the same variable did not affect accuracy: composite functions 

19 based on one set of genotypes were predictive for another set, as long as medians were used for both 

sets. A study of cottonwood (Populus deltoides) conducted using a naturally occurring gradient of O3 

21 exposure (Gregg et al., 2003, 046996; Gregg et al., 2006, 186961) may provide an illustration of the 

22 response of an individual species whose response is far from the median response for an aggregation 

23 of species. 

9.8.3.3. Expos ure -Res pons e in a Gra dient Study 

24	 Gregg et al. (2003, 046996) grew saplings of one clonally propagated genotype of cottonwood 

(Populus deltoides) in seven locations within New York City and in the surrounding region between 

26 July and September in 1992, 1993 and 1994, and harvested them 72 days after planting. Owing to 

27 regional gradients of atmospheric O3 concentration, the experiment yielded eight levels of exposure 

28 (Figure 9-20), and the authors were able to rule out environmental variables other than O3 to account 

29 for the large differences in biomass observed after one season of growth. The deficit in growth 

increased substantially faster with increasing O3 exposure than has been observed in aspen, another 

31 species of the same genus (Populus tremuloides, Section 9.8.3.2). Using a three parameter Weibull 

32 model (Figure 9-20), the biomass of cottonwood at a W126 exposure of 15 ppm-h, relative to 

33 biomass at 5 ppm-h, is estimated to be 0.18 (18% of growth at 5 ppm-h). The relative biomass of 

34 trembling aspen within the same 5-15 ppm-h range of exposure is estimated to be 0.92, using the 

median composite model for aspen whose very close agreement with Aspen FACE data was shown 

36 in Section 9.8.3.2. Using a median composite function for all deciduous trees in the NHEERL/WED 

37 project (6 species in 21 studies) also gives predictions that are very distant from the cottonwood 

March 2011	 9-128 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

 

 

 

 

 

 

1 response observed in this experiment. For all deciduous tree species in NHEERL/WED, biomass at a 

2 W126 exposure of 15 ppm-h, relative to biomass at 5 ppm-h, was estimated to be 0.87. 

100 

90 

80 

70 

0 10 20 30 40 50 60 70 

Biomass (g) 

60 

50 

40 

30 

20 

10 

0 

72 day 12 hr W126 (ppm-hr) 

Source: Modified with permission from Nature Publishing Group, Gregg et al. (2003, 046996). 

Figure 9-20. Above-ground biomass for one genotype of cottonwood grown in seven locations

for one season in 3 years. ﷒Line represents the three-parameter Weibull model. 

3 These cottonwood data confirms that, as should be expected, some individual tree species are 

4 substantially more sensitive than the median of NHEERL/WED (Figure 9-15). As shown in 

5 Section 9.8.2, the median models available for trembling aspen and soybean have verifiable 

6 predictive ability for those particular species. This suggests that the corresponding NCLAN- and 

7 NHEERL/WED-based models for multiple crop and tree species can provide reliable estimates of 

8 losses for similar assortments of species. However, their predictive ability would likely be poor for 

9 individual species not tested. 

10 An alternative hypothesis for the difference between the response of cottonwood in this 

11 experiment and deciduous tree species in NHEERL/WED, or the difference between the response of 

12 cottonwood and aspen in NHEERL/WED and Aspen FACE, could be the presence of confounding 

13 factors in the environments where the experiment was conducted. However, variability in 

14 temperature, moisture, soil fertility, and atmospheric deposition of N were all ruled out by Gregg 

15 et al. (2003, 046996) as contributing to the observed response to O3. In addition, this hypothesis 

16 would imply that the unrecognized confounder(s) were either absent from both OTC and FACE 

17 studies, or had the same value in both. This is not impossible, but the hypothesis that cottonwood is 

18 very sensitive to O3 exposure is more parsimonious, and sufficient. 

March 2011 9-129 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

9.8.3.4. Meta -analys es of growth and yield s tudies 

1 Since the 2006 O3 AQCD, five studies have used meta-analytic methods to integrate results 

2 from experimental studies of crops or tree species relevant to the U.S. It is possible to obtain 

3 exposure-response data for growth and yield from those meta-analyses, but because all of them 

4 provided summary measurements of O3 exposure as hourly averages of various lengths of exposures, 

5 comparisons with exposure-response results where exposure is expressed as W126 are problematic. 

6 Table 9-15 summarizes the characteristics of the five meta-analyses. They all included studies 

7 conducted in the U.S. and other locations worldwide, and all of them expressed responses as 

8 comparative change between levels of exposure to O3, with carbon filtered air (CF) among those 

9 levels. Using hourly average concentration to summarize exposure, CF rarely equates absence of O3, 

10 although it almost always near zero when exposure is summarized as W126, SUM06, or AOT40. 

Table 9-15. Meta-analyses of growth or yield studies published since 2005 

Years of Number of articles Number of Duration of Study publication Crop, species or genera Response included O3 levels exposure surveyed 

Ainsworth (2008, 12 1980-2007 rice Yield 2 unreported 191646) 

Feng et al. (2008, 53 1980-2007 wheat Yield 5 > 10 days 191453) 

Feng and Kobayashi Potato, barley, wheat, rice, All crops together : 81 1980-2007 Yield 3 > 10 days (2009, 199223) bean, soybean 

34 herbaceous dicots Grantz et al. (2006, Relative 16 1992-2004 21 herbaceous monocots 2 2-24 weeks 191545) Growth Rate 5 tree species 

All responses:263 Wittig et al. (2009, 4 gymnosperm tree genera Articles that included 1970-2006 Total biomass 4 > 7 days 191631) 11 angiosperm tree genera biomass:unreported 

11 The only effect of O3 exposure on yield of rice reported in Ainsworth (2008, 191646) was a 

12 decrease of 14% with exposure increasing from CF to 62 ppb average concentration. Feng et al. 

13 (2008, 191453) were able to separate exposure of wheat into four classes with average 

14 concentrations of 42, 69, 97, and 153 ppb, in data where O3 was the only treatment. Mean responses 

15 relative to CF were yield decreases of 17, 25, 49, and 61% respectively. Feng et al. (2008, 191453) 

16 observed that wheat yield losses were smaller under conditions of drought, and that Spring wheat 

17 and Winter wheat appeared similarly affected. However, mean exposure in studies of Winter wheat 

18 was substantially higher than in studies of Spring wheat (86 versus 64 ppb), which suggests that the 

19 yield of Spring wheat was in fact more severely affected, since yield was approximately the same, 

20 even though Spring wheat was exposed to lower concentrations. Exposures of the six crops 

21 considered in Feng and Kobayashi (2009, 199223) were classified into two ranges, each compared to 

22 CF air. In the lower range of exposure (41-49 ppb), potato studies had the highest average exposure 

23 (45 ppb), and wheat and rice the lowest (41 ppb). In the higher range (51-75 ppb), wheat studies had 

24 the highest average exposure (65 ppb), and potato, barley and rice the lowest (63 ppb). In other 

25 words, across the studies included, all crops were exposed to very similar levels of O3. At 

March 2011 9-130 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

1 approximately 42 ppb, the yield of potato, barley, wheat, rice, bean, and soybean declined by 5.3, 

2 8.9, 9.7, 17.5, 19, and 7.7% respectively, relative to CF air. At approximately 64 ppb O3, declines 

3 were 11.9, 12.5, 21.1, 37.5, 41.4, and 21.6%. Grantz et al. (2006, 191545) reported Relative Growth 

4 Rate (RGR) rather than growth, and did not report O3 exposure values in a way that would allow 

5 calculation of mean exposure for each of the three categories of plants for which RGR changes are 

6 reported. All studies used only two levels of exposure, with CF air as the lower one, and most used 

7 elevated exposure in the range of 40 to 70 ppb. Decline in RGR was 8.2% for the 34 herbaceous 

8 dicots, 4.5% for the 21 herbaceous monocots, and 17.9 for the 5 tree species. Finally, Wittig et al. 

9 (2009, 191631) divided the studies analyzed into three classes of comparisons: CF versus ambient, 

10 CF versus elevated, and ambient versus elevated, but reported comparisons between three average 

11 levels of exposure besides CF: 40 ppb, 64 ppb, and 97 ppb. Corresponding decreases in total biomass 

12 relative to CF were 7, 17, and 17%. 

13 These meta-analyses provide very strong confirmation of EPA's conclusions from previous O3 

14 AQCDs: compared to lower levels of ambient O3, current levels in many locations are having a 

15 substantial detrimental effect on the growth and yield of a wide variety of crops and natural 

16 vegetation. They also confirm strongly that decreases in growth and yield continue at exposure levels 

17 higher than current ambient levels. However, direct comparisons with the predictions of exposure18 response models that use concentration-weighted cumulative metrics are difficult. 

9.8.3.5. Ad ditiona l e xpos ure -res pons e data 

19 The studies summarized in Tables 9-16 and 9-17 contain growth or yield exposure-response 

20 data at too few levels of exposure for exposure-response models, and/or used metrics other than 

21 W126. These tables update Tables AX9-16 through AX9-19 of the 2006 O3 AQCD. 

March 2011 9-131 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

 

 

 

 

 

 

Table 9-16. Summary of studies of effects of ozone exposure on growth and yield of agricultural crops 

Species

Facility

Location 

Exposure

Duration 

O3 Exposure

(Additional Treatment) Response Measured 

percent change from 

CF 

(percent change from

ambient) 

Reference 

Alfalfa (Medicago 

sativa) 

OTC; 0.27m3 pots 

Federico, Italy 

2 yr, 2005,

2006 

AOT40: CF 0 ppm-h 

13.9 ppm-h (2005), 10.1 ppm-h 

(2006) 

(NaCl: 0.29, 0.65, 0.83, 

1.06 deciSiemens/meter) 

Total shoot yield n.s. (N/A) Maggio et al. 

(2009, 191645) 

Bean (Phaseolus 

vulgaris I. cv Borlotto) 

OTC; ground-planted 

Curno, Italy 

3 months, 2006 

Seasonal AOT40: 

CF (0.5 ppm-h);

ambient (4.6 ppm-h) 

(N/A) 

# Seeds per plant; 

100-seed weight 

-33 (N/A) 

n.s. (N/A) 

Gerosa et al. 

(2009, 191403) 

Big Blue Stem 

(Andropogon gerardii) 

OTC 

Alabama 

4 months, 

2003 

12-h avg:

CF (14 ppb),

Ambient (29 ppb),

Elevated (71 ppb) 

(N/A) 

Final harvest biomass; 

RVF 

n.s. (n.s.) 

-7 (-7) 

Lewis et al. 

(2006, 191542) 

Brassica napus cv. 

Westar 

Growth chambers 

Finland 

17-26 days 

8-h avg:

CF (0 ppb), 100 ppb 

(Bt/non-Bt; herbivory) 

Shoot biomass -30.70 (N/A) Himanen et al. 

(2009, 191338) 

Corn (Zea mays cv. 

Chambord) 

OTC 

France 

33 days 

AOT40 ppm-h: 1.1; 1.3; 4.9; 7.2;

9.3; 12.8 

(N/A) 

Total above-ground

biomass 

N/A (Highest treatment

caused -26% change) 

Leitao et al. 

(2007, 191456) 

Cotton cv. Pima 

OTC; 9-L pots 

France 

8 wk 

12-h avg: 12.8 ± 0.6; 79.9 ± 6.3;

122.7 ± 9.7 

(N/A) 

Above-ground biomass -76 (n.s.) 

Grantz and 

Shrestha (2006, 

191702) 

Eastern Gamagrass

(Tripsacum

dactyloides) 

OTC 

Alabama 

4 months, 

2003 

12-h avg:

CF (14ppb),

Ambient (29 ppb),

Elevated (71 ppb) 

(N/A) 

Final harvest biomass; 

RVF 

+68 (+42); 

-17 (-12) 

Lewis et al. 

(2006, 191542) 

Grapevine (Vitis 

vinivera) 

OTC 

Austria 

3 yr, May-Oct 

AOT40 ppm-h: 

CF (0),

Ambient (7-20),

Elevated. 1 (20-30), Elevated. 2 

(38-48) 

Total fruit yield/

Sugar yield -20 to -80 in different yr (-20 to -90 in different yr) 

Soja et al. (2004, 

094397) 

Mustard (Brassica 

campestris) 

Chambers; 

7.5-cm pots 

10 days 

CF & 

67.8 ppb for 7 h 

(N/A) 

Seeds/plant n.s. (N/A) Black et al. 

(2007, 191558) 

Oilseed Rape 

(Brassica napus) 

OTC 

Yangtze Delta, China 

39 days 

Daily avg: 100 ppb, one with 

diurnal variation and one with 

constant concentration 

(N/A) 

Biomass and pods per

plant 

Diurnal variability

reduced both biomass 

and pod number more

than constant fumigation 

(N/A) 

Wang et al.

(2008, 191276) 

Peanut (Arachis 

hypogaea) 

OTC 3 yr 

12-h avg:

CF (22 ppb),

Ambient (46 ppb),

Elevated (75ppb) Yield (seed weight, g/m) -33 (-8) Burkey et al.

(2007, 191371) 

Raleigh, NC (CO2: 375 ppm; 548 ppm; 

730 ppm) 

March 2011 9-132 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

 

 

 

 

 

Species

Facility

Location 

Exposure

Duration 

O3 Exposure

(Additional Treatment) Response Measured 

percent change from 

CF 

(percent change from

ambient) 

Reference 

Poa pratensis 

OTC 

Braunschweig,

Germany 

2000-2002: 

4-5 wk in the 

Spring 

8-h avg:

CF+25 (21.7),

NF+50 (73.1) 

(Competition) 

Total biomass (g 

DW/pot) N/A (n.s.) Bender et al. (2006, 191437) 

Potato (Solanum 

tuberosum) 

OTC; CHIP 

6 northern European 

locations 

1988,1999. 

Emergence to

harvest 

AOT40:CF (0); 

Ambient (0.27-5.19); NF (0.002­

2.93)

NF+ (3.10-24.78 

(N/A) 

Tuber yield averaged 

across 5 field-sites; 

Tuber starch content 

regressed against [O3] 

report sig. ± slope with 

increasing [O3] 

N/A (-27 % -+27%, most 

comparisons n.s.) Linear 

regression slope = 

-0.0098) 

Vandermeiren 

et al. (2005,

179992) 

Rice (Oryza sativa) 

OTC 

Raleigh, NC 

1997-1998, 

JuneSeptember 

12-h mean ppb:

CF (27.5),

Elevated (74.8) 

(CO2) 

Total biomass; 

Seed yield 

-25(N/A) 

-13 to 20 (N/A) 

Reid, et al. 

(2008, 191561) 

Rice (Oryza sativa) 20 

Asian cultivars 

OTC 

Gunma Prefecture, 

Japan 

2008 growing 

season 

Daily avg (ppb):

CF (2),

0.8×ambient (23);

1 ×ambient (28);

1.5×ambient (42);

2×ambient (57) 

(Cultivar comparisons) 

Yield From n.s. to -30 across 

all cultivars 

Sawada and 

Kohno (2009,

199426) 

Seminatural grass 

FACE 5 yr 

Seasonal AOT40: Ambient 

(0.1-7.2 ppm-h);

Elevated. (1.8-24.1 ppm-h) Relative annual yield N/A (2×faster

decrease in yield/yr) 

Volk et al. (2006, 

191434) 

Le Mouret, Switzerland (N/A) 

Soybean 

OTC; CRA 

Bari, Italy 

2003-2005 

growing 

seasons 

Seasonal AOT40 ppm-h: CF (0),

Ambient (3.4), High (9.0) 

(Drought) 

Yield -46 (-9) Jaoude et al. 

(2008, 191223) 

Soybean (Glycine max 

cv. 93B15) 

SoyFACE 

Urbana, IL 

2002, 2003 

growing 

seasons 

8-h avg:

Ambient (62 & 50 ppb), Elevated 

(75 & 63 ppb) 

(N/A) 

Yield 

N/A 

(-15 in 2002;

-25 in 2003) 

Morgan et al. 

(2006, 079186) 

Soybean (Glycine max 

cv. Essex) 

Chambers; 21 L 

Raleigh, NC 

2×3 months 

12-h avg: CF (28),

Elevated (79), 

Elevated flux (112) 

(CO2: 365 & 700) 

Seed mass per plant -30 (N/A) 

Booker and 

Fiscus (2005,

191652) 

Soybean (Glycine max 

cv. Essex) 

OTCs; 21-L pots 

Raleigh, NC 

2×3 months 

12-h avg: CF (18);

Elevated (72) 

(CO2: 367 & 718) 

Seed mass per plant -34 (N/A) Booker et al. 

(2004, 079138) 

Soybean (Glycine max 

cv. Tracaja) 

Chambers; pots 20 days 

12-h avg: CF & 30 ppb 

(N/A) 

Biomass -18 (N/A) Bulbovas et al. 

(2007, 199411) 

Brazil 

Soybean (Glycine 

max) 10 cultivars 

SoyFACE 

Urbana, IL 

2007 & 2008 

8-h avg: Ambient (46.3 & 37.9), 

Elevated (82.5 & 61.3) 

(Cultivar comparisons) 

Yield N/A (-17.20) 

Betzelberger 

et al. (2010,

644183) 

March 2011 9-133 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

Species

Facility

Location 

Exposure

Duration 

O3 Exposure

(Additional Treatment) Response Measured 

percent change from 

CF 

(percent change from

ambient) 

Reference 

Spring Wheat (Triticum 

aestivum cv. Minaret; 

Satu; Drabant; Dragon) 

OTCs 

Belgium, Finland, & 

Sweden 

1990-2006 

Seasonal AOT40s ranged from 0 

to16 ppm-h 

(N/A) 

Seed protein content; 

1,000-seed weight

regressed across all 

experiments 

N/A (significant negative 

correlation) 

N/A (sig negative 

correlation) 

Piikki et al. 

(2008, 199812) 

Strawberry (Fragaria x 

ananassa Duch. Cv 

Korona & Elsanta) 

Growth chambers 

2 months 

8-h avg: CF (0 ppb) &

Elevated (78 ppb) 

(N/A) 

Fruit yield (weight/plant) -16 (N/A) Keutgen et al.

(2005, 191295) 

Bonn, Germany 

Sugarbeet (Beta 

vulgaris cv. Patriot) 

OTC 

2003, 2004; 

5 months 

8-h avg: Ambient (36 ppb); 

Elevated (62 ppb) 

(N/A) Sugar yield N/A (-9) 

De Temmerman 

et al. (2007,

191361) 

Belgium 

Sugarcane

(Saccharum spp) 

CSTR 

2007; 

11-13 wk. 

12-h avg: CF (4 ppb);

Ambient (58); 

Elevated (147) Total biomass (g/plant) -40 (-30) Grantz and Vu 

(2009, 195237) 

San Joaquin Valley, CA (N/A) 

Sweet Potato 

Growth chambers 4 wk 

8-h avg: CF (0 ppb),

Ambient (<40 ppb) Elevated (255 

ppb) Tuber weight -14 (-11.5) Keutgen et al.

(2008, 191690) 

Bonn, Germany (N/A) 

Tomato (Lycopersicon 

esculentum) 

OTC 

Valencia, Spain 

133 days in 

1998 

8-h mean ppb:

CF 16.3, NF 30.1, 

NF+ 83.2 

(Various cultivars; early & late 

harvest) 

Yield n.s (n.s.) Calvo et al. 

(2005, 191570) 

Trifolium 

Subterraneum 

OTC; 2.5-L pots 29 days 

12-h avg: CF (<7.9±6.3); Ambient

(34.4±10.8);

Elevated (56.4±22.3) Above-ground biomass -45 (-35) Sanz et al. (2005,

196963) 

Madrid, Spain (N: 5, 15 & 30 kg/ha) 

Watermelon (Citrullus 

lanatus) 

OTC 

2000, 2001. 

90 days 

AOT40: CF (0 ppm-h)

Ambient (5.7 ppm-h), Elevated 

(34.1 ppm-h) total fruit yield (kg) n.s. (54) Calatayud et al.

(2006, 191482) 

Valencia, Spain (N:0, 14.0 & 29.6 g/pot) 

Yellow Nutsedge 

OTC; 9-L pots 

8 wk 

12-h avg: 12.8 ± 0.6; 79.9 ± 6.3; 

122.7 ± 9.7 

(N/A) 

above-ground biomass n.s. (n.s.) 

Grantz and 

Shrestha (2006,

191702) 

In studies where variables other than O3 were included in the experimental design, response to O3 is only provided for the control level of those variables. 

March 2011 9-134 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

Table 9-17. Summary of studies of effects of ozone exposure on growth of natural vegetation 

Species

Facility

Location 

Exposure

Duration 

O3 Exposure

(Additional

Treatment) 

Response Measured Response Reference 

Yellow nutsedge (Cyperus 

esculentus) 

CSTR 53 days in 2008 

12-h mean ppb: CF

(4); CF+ (60);

CF2+ (115) 

Above-ground biomass;

tubers (g/plant) 

ns; CF(4.1) CF+(3.9)

CF2+(2.7) 

Grantz et al. (2010,

102161) 

Parlier, CA 

35 herbaceous species 

OTC 

Corvallis, OR 

1999-2002, 

May-August 

4-yr avg; yearly

W126 ppm-h:

CF (0),

CF+ (21),

CF 2+ (49.5) 

Total community aboveground biomass

(35 species) after 4 years 

CF (459 g/m2), CF+ 

(457 g/m2

), CF2+ (398 

g/m2

) 

Pfleeger et al. (2010,

644281) 

Highbush blackberry (Rubus 

argutus) 

OTC 

Auburn, AL 

2004, 

May-August 

12-h mean ppb:

CF (21.7),

Ambient (32.3),

Elevated (73.3) 

Vegetative regrowth after 

pruning 

CF (75.1 g/plant),

Ambient (76.4 g/plant),

Elevated (73.1 g/plant) 

Ditchkoff et al. (2009,

192230) 

Horseweed (Conyza 

canadensis) 

CSTR 

San Joaquin Valley, CA 

2005, 2 runs, 

28 days each

(July-Aug, Sept) 

W126 ppm-hr:

CF(0),

CF+ (11),

CF 2+ (30) 

(Glyphosate 

resistance) 

Total biomass (g/plant) 

Glyphosate sensitive: 

CF (0.354)

CF+ (0.197)

CF2+ (0.106) 

Glyphosate resistant: 

CF(0.510)

CF+ (0.313)

CF2+ (0.143) 

Grantz et al. ( (2008,

191312) 

AOT60: 

Red Oak (Quercus rubrum) 

Forest sites 

Look Rock & Twin Creeks 

Forests, TN 

2001-2003, 

April-October 

2001 (11.5),

2002 (24.0),

2003 (11.7) 

(Observational 

study with multiple 

environmental 

Annual circumference 

increment (change

relative to 2001 in year 

2002;2003) 

-42.8%; +1% McLaughlin et al. (2007, 

090348) 

variables) 

AOT60: 

Pine species 

Forest sites 

Look Rock Forest, TN 

2001-2003, 

April-October 

2001 (11.5),

2002 (24.0),

2003 (11.7) 

(Observational 

study with multiple 

environmental 

Annual circumference 

increment (change

relative to 2001 in year 

2002;2003) 

-62.5%; -2.9% McLaughlin et al. (2007, 

090348) 

variables) 

AOT60: 

Hickory species 

Forest sites 

Look Rock Forest, TN 

2001-2003, 

April-October 

2001 (11.5),

2002 (24.0),

2003 (11.7) 

(Observational 

study with multiple 

environmental 

Annual circumference 

increment (change

relative to 2001 in year 

2002;2003) 

-14%; +30% McLaughlin et al. (2007, 

090348) 

variables) 

AOT60: 

Chestnut Oak (Quercus prinus) 

Forest sites 

Look Rock Forest, TN 

2001-2003, 

April-October 

2001 (11.5),

2002 (24.0),

2003 (11.7) 

(Observational 

study with multiple 

environmental 

Annual circumference 

increment (change

relative to 2001 in year 

2002;2003) 

+44%; +55% McLaughlin et al. (2007, 

090348) 

variables) 

March 2011 9-135 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

Species

Facility

Location 

Exposure

Duration 

O3 Exposure

(Additional

Treatment) 

Response Measured Response Reference 

AOT60: 

Black Cherry (Prunus rigida) 

Forest sites 

Twin Creeks Forest, TN 

2002-2003, 

April-October 

2002 (24.0),

2003 (11.7) 

(Observational 

study with multiple 

environmental 

Annual circumference 

increment (change 

relative to 2003 in year 

2002) 

-75% McLaughlin et al. (2007, 

090348) 

variables) 

AOT60: 

Shortleaf pine (Pinus echinata) 

Forest sites 

Twin Creeks Forest, TN 

2002-2003, 

April-October 

2002 (24.0),

2003 (11.7) 

(Observational 

study with multiple 

environmental 

Annual circumference 

increment (change

relative to 2003 in year 

2002) 

-16.8% McLaughlin et al. (2007, 

090348) 

variables) 

AOT60: 

Hemlock (Tsuga canadensis) 

Forest sites 

Twin Creeks Forest, TN 

2002-2003, 

April-October 

2002 (24.0),

2003 (11.7) 

(Observational 

study with multiple 

environmental 

Annual circumference 

increment (change

relative to 2003 in year 

2002) 

-21.9% McLaughlin et al. (2007, 

090348) 

variables) 

AOT60: 

Red Maple (Acer rubrum) 

Forest sites 

Twin Creeks Forest, TN 

2002-2003, 

April-October 

2002 (24.0),

2003 (11.7) 

(Observational 

study with multiple 

environmental 

Annual circumference 

increment (change

relative to 2003 in year 

2002) 

-59.6% McLaughlin et al. (2007, 

090348) 

variables) 

Yellow Poplar (Liriodendron 

tulipifera) 

Forest sites 

Look Rock, Oak Ridge, & Twin 

Creeks Forest, TN 

2002-2003, 

April-October 

AOT60: 

2002 (24.0),

2003 (11.7) 

(Observational 

study with multiple 

environmental 

variables) 

Annual circumference 

increment (change

relative to 2001 in years 2002; 2003) 

-45.9%; -15.25% McLaughlin et al. (2007, 

090348) 

AOT60: 

Sugar Maple (Acer saccharum) 

Forest sites 

Twin Creeks Forest, TN 

2002-2003, 

April-October 

2002 (24.0),

2003 (11.7) 

(Observational 

study with multiple 

environmental 

Annual circumference 

increment (change

relative to 2003 in year 

2002) 

-63.8% McLaughlin et al. (2007, 

090348) 

variables) 

Trembling aspen (Populus 

tremuloides), 5 genotypes 

Aspen FACE 

1998-2004, 

May-October 

Cumulative avg 

90-day 12-h W126.

Ambient 3.1 ppm-h 

Elevated: 27.2 

ppm-h 

main stem volume after 7 

years 

Ambient: 6.22 dm3; 

Elevated: 4.73 dm3 Kubiske et al. (2006,

093284) 

Rhinelander, WI (Competition with 

birch, maple) 

Hybrid Poplar (Populus 

trichocarpa x Populus deltoides) 

OTC 2003, 3 months 

Daily mean (μg/g):

CF(<9),

Elevated (85-128) 

Total biomass CF to elevated: 

-12.9% 

Woo and Hinckley (2005, 

191359) 

Seattle, WA 

In studies where variables other than O3 were included in the experimental design, response to O3 is only provided for the control level of those variables. 

March 2011 9-136 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

References
 

A list of all references considered for inclusion in this chapter can be found at 

http://hero.epa.gov/index.cfm?action=search.do&submit=Search&group_id=476 

Agrell, J.; Kopper, B. J.; McDonald, E. P.; Lindroth, R. L. (2005). CO2 and O3 effects on host plant preferences of the 

forest tent caterpillar (Malacosoma disstria). Global Change Biol, 11: 588-599. http://dx.doi.org/10.1111/j.1365­

2486.2005.00924.x 074324 

Ahlfors, R.; Brosche, M.; Kollist, H.; Kangasjarvi, J. (2009). Nitric oxide modulates ozone-induced cell death, hormone 

biosynthesis and gene expression in Arabidopsis thaliana. Plant J, 58: 1-12. http://dx.doi.org/10.1111/j.1365­

313X.2008.03756.x 191533 

Ahsan, N.; Nanjo, Y.; Sawada, H.; Kohno, Y.; Komatsu, S. (2010). Ozone stress-induced proteomic changes in leaf total 

soluble and chloroplast proteins of soybean reveal that carbon allocation is involved in adaptation in the early 

developmental stage. Proteomics, 10: 2605-2619. http://dx.doi.org/10.1002/pmic.201000180 644189 

Ainsworth, E. A. (2008). Rice production in a changing climate: a meta-analysis of responses to elevated carbon dioxide 

and elevated ozone concentration. Global Change Biol, 14: 1642-1650. http://dx.doi.org/10.1111/j.1365­

2486.2008.01594.x 191646 

Ainsworth, E. A.; Long, S. P. (2005). What have we learned from 15 years of free-air CO2 enrichment (FACE)? A metaanalytic review of the responses of photosynthesis, canopy properties and plant production to rising CO2. New 

Phytol, 165: 351-371. 042647 

Ainsworth, E. A.; Rogers, A. (2007). The response of photosynthesis and stomatal conductance to rising [CO2]: 

mechanisms and environmental interactions. Plant Cell Environ, 30: 258-270. 092940 

Air quality criteria and control techniques, Section 108 of the Clean Air Act. 42 USC Â§ 7408. (1990). 

http://www.epa.gov/air/caa/title1.html#ia. 080701 

Alexis, A.; Garcia, A.; Nystrom, M.; Rosenkranz, K. (2001). The 2001 California almanac of emissions and air quality. 

Retrieved August 10, 2005 from http://www.arb.ca.gov/aqd/almanac/almanac01/almanac01.htm. 079886 

Allen, E. B.; Temple, P. J.; Bytnerowicz, A.; Arbaugh, M. J.; Sirulnik, A. G.; Rao, L. E. (2007). Patterns of understory 

diversity in mixed coniferous forests of southern California impacted by air pollution. ScientificWorldJournal, 7: 

247-263. http://dx.doi.org/10.1100/tsw.2007.72 196876 

Alonso, R.; Bermejo, V.; Sanz, J.; Valls, B.; Elvira, S.; Gimeno, B. S. (2007). Stomatal conductance of semi-natural 

Mediterranean grasslands: Implications for the development of ozone critical levels. Environ Pollut, 146: 692-698. 

http://dx.doi.org/10.1016/j.envpol.2006.06.009 199289 

Amthor, J. S. (1988). Growth and maintenance respiration in leaves of bean (Phaseolus vulgaris L) exposed to ozone in 

open-top chambers in the field. New Phytol, 110: 319-325. 041870 

Andersen, C. P. (2003). Source-sink balance and carbon allocation below ground in plants exposed to ozone. New Phytol, 

157: 213-228. 041673 

Andersen, C. P.; Ritter, W.; Gregg, J.; Matyssek, R.; Grams, T. E. E. (2010). Below-ground carbon allocation in mature 

beech and spruce trees following long-term, experimentally enhanced O3 exposure in Southern Germany. Environ 

Pollut, 158: 2604-2609. http://dx.doi.org/10.1016/j.envpol.2010.05.008 628559 

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. 052923 

Andrews, K. M.; Gibbons, J. W.; Jochimsen, D. M. (2008). Ecological effects of roads on amphibians and reptiles: A 

literature review. In JC Mitchell; RE Jung Brown; B Bartholomew (Eds.), Urban Herpetology (pp. 121-143). Salt 

Lake City: Society for the Study of Amphibians and Reptiles. http://www.uga.edu/srel/Reprint/3091.htm 645771 

Note: Hyperlinks to the reference citations throughout this document will take you to the NCEA HERO database (Health and 

Environmental Research Online) at http://epa.gov/hero. HERO is a database of scientific literature used by U.S. EPA in the process of 

developing science assessments such as the Integrated Science Assessments (ISAs) and the Integrated Risk Information System (IRIS). 

March 2011 9-137 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

Aneja, M. K.; Sharma, S.; Fleischmann, F.; Stich, S.; Heller, W.; Bahnweg, G.; Munch, J. C.; Schloter, M. (2007). Influence 

of ozone on litter quality and its subsequent effects on the initial structure of colonizing microbial communities. 

Microb Ecol, 54: 151-160. http://dx.doi.org/10.1007/s00248-006-9183-0 191472 

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. 

http://dx.doi.org/10.1016/S0269-7491(98)00027-X 040297 

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. 052925 

Ariyaphanphitak, W.; Chidthaisong, A.; Sarobol, E.; Bashkin, V. N.; Towprayoon, S. (2005). Effects of elevated ozone 

concentrations on Thai Jasmine rice cultivars (Oryza sativa L.). Water Air Soil Pollut, 167: 179-200. 

http://dx.doi.org/10.1007/s11270-005-8650-4 191349 

Ashmore MEmberson, L.; Karlsson, P. E.; Pleijel, H. (2004). Introduction for ozone deposition special issue. Atmos 

Environ, 38: 2211-2212. 056624 

Ashmore, M. R. (2002). Effects of oxidants at the whole plant and community level. In Air Pollution and Plants (pp. 89­

118). London: Wiley. 672967 

Ashmore, M. R.; Bell, J. N. B.; Mimmack, A. (1988). Crop growth along a gradient of ambient air pollution. Environ 

Pollut, 53: 99-121. 037038 

Ashmore, M.; Emberson, L.; Karlsson, P. E.; Pleijel, H. (2004). New directions: a new generation of ozone critical levels 

for the protection of vegetation in Europe (correspondence). Atmos Environ, 38: 2213-2214. 056623 

Ashmore, M.; Toet, S.; Emberson, L. (2006). Ozone - a significant threat to future world food production? New Phytol, 

170: 201-204. http://dx.doi.org/10.1111/j.1469-8137.2006.01709.x 191557 

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, Biol.10: 1414-1423. 052926 

Awmack, C. S.; Mondor, E. B.; Lindroth, R. L. (2007). Forest understory clover populations in enriched CO2 and O-3 

atmospheres: Interspecific, intraspecific, and indirect effects. Environ Exp Bot, 59: 340-346. 

http://dx.doi.org/10.1016/j.envexpbot.2006.04.003 191415 

Bagard, M.; Le Thiec, D.; Delacote, E.; Hasenfratz-Sauder, M. P.; Banvoy, J.; Gerard, J.; Dizengremel, P.; Jolivet, Y. 

(2008). Ozone-induced changes in photosynthesis and photorespiration of hybrid poplar in relation to the 

developmental stage of the leaves. Physiol Plant, 134: 559-574. http://dx.doi.org/10.1111/j.1399-3054.2008.01160.x 

191593 

Baier, M.; Kandlbinder, A.; Golldack, D.; Dietz, K. (2005). Oxidative stress and ozone: Perception; signalling and 

response. Plant Cell Environ, 28: 1012-1020. http://dx.doi.org/10.1111/j.1365-3040.2005.01326.x 186866 

Ball, G. R.; Palmer-Brown, D.; Fuhrer, J.; Skarby, 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 Modell, 129: 153-168. 026354 

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. 026370 

Bandeff, J. M.; Pregitzer, K. S.; Loya, W. M.; Holmes, W. E.; Zak, D. R. (2006). Overstory community composition and 

elevated atmospheric CO2 and O-3 modify understory biomass production and nitrogen acquisition. Plant Soil, 

282: 251-259. http://dx.doi.org/10.1007/s11104-005-5930-0 191733 

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 

(p. 235–252). Tokyo: Springer-Verlag. 040313 

Bassin, S.; Volk, M.; Fuhrer, J. (2007). Factors affecting the ozone sensitivity of temperate European grasslands: An 

overview. Environ Pollut, 146: 678-691. http://dx.doi.org/10.1016/j.envpol.2006.06.010 196879 

Bassin, S.; Volk, M.; Suter, M.; Buchmann, N.; Fuhrer, J. (2007). Nitrogen deposition but not ozone affects productivity 

and community composition of subalpine grassland after 3 yr of treatment. New Phytol, 175: 523-534. 

http://dx.doi.org/10.1111/j.1469-8137.2007.02140.x 191534 

March 2011 9-138 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

Bassin, S.; Werner, R. A.; Sorgel, K.; Volk, M.; Buchmann, N.; Fuhrer, J. (2009). Effects of combined ozone and nitrogen 

deposition on the in situ properties of eleven key plant species of a subalpine pasture. Oecologia, 158: 747-756. 

http://dx.doi.org/10.1007/s00442-008-1191-y 191333 

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, 105: 22,123-22,136. 

040315 

Bender, J.; Muntifering, R. B.; Lin, J. C.; Weigel, H. J. (2006). Growth and nutritive quality of Poa pratensis as influenced 

by ozone and competition. Environ Pollut, 142: 109-115. http://dx.doi.org/10.1016/j.envpol.2005.09.012 191437 

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. 

039778 

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. 036321 

Bergweiler, C.; Manning, W. J.; Chevone, B. I. (2008). Seasonal and diurnal gas exchange differences in ozone-sensitive 

common milkweed (Asclepias syriaca L.) in relation to ozone uptake. Environ Pollut, 152: 403-415. 

http://dx.doi.org/10.1016/j.envpol.2007.06.019 191656 

Bernacchi, C. J.; Leaky, A. D. B.; Heady, L. E.; Morgan, P. B.; Dohleman, F. G.; McGrath, J. M.; Gillespie, K. M.; Wittig, 

V. E.; Rogers, A.; Long, S. P.; Ort, D. R. (2006). Hourly and seasonal variation in photosynthesis and stomatal 

conductance of soybean grown at future CO2 and ozone concentrations for 3 years under fully open-air field 

conditions. Plant Cell Environ, 29: 2077-2090. http://dx.doi.org/10.1111/j.1365-3040.2006.01581.x 158001 

Bernacchi, C. J.; Morgan, P. B.; Ort, D. R.; Long, S. P. (2005). The growth of soybean under free air CO2 enrichment 

(FACE) stimulates photosynthesis while decreasing in vivo Rubisco capacity. Planta, 220: 434-446. 

http://dx.doi.org/10.1007/s00425-004-1320-8 679713 

Betzelberger, A. M.; Gillespie, K. M.; McGrath, J. M.; Koester, R. P.; Nelson, R. L.; Ainsworth, E. A. (2010). Effects of 

chronic elevated ozone concentration on antioxidant capacity, photosynthesis and seed yield of 10 soybean 

cultivars. Plant Cell Environ, 33: 1569-1581. http://dx.doi.org/10.1111/j.1365-3040.2010.02165.x 644183 

Bidart-Bouzat, M. G.; Imeh-Nathaniel, A. (2008). Global change effects on plant chemical defenses against insect 

herbivores. J Integr Plant Biol, 50: 1339-1354. http://dx.doi.org/10.1111/j.1744-7909.2008.00751.x 191431 

Billings, W. D. (1978). Plants and the ecosystem. Belmont, CA: Wadsworth Publishing Company, Inc. 034165 

Biswas, D. K.; Xu, H.; Li, Y. G.; Sun, J. Z.; Wang, X. Z.; Han, X. G.; Jiang, G. M. (2008). Genotypic differences in leaf 

biochemical, physiological and growth responses to ozone in 20 winter wheat cultivars released over the past 60 

years. Global Change Biol, 14: 46-59. http://dx.doi.org/10.1111/j.1365-2486.2007.01477.x 191428 

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. 036322 

Black, V. J.; Stewart, C. A.; Roberts, J. A.; Black, C. R. (2007). Ozone affects gas exchange, growth and reproductive 

development in Brassica campestris (Wisconsin Fast Plants). New Phytol, 176: 150-163. 

http://dx.doi.org/10.1111/j.1469-8137.2007.02163.x 191558 

Black, V. J.; Stewart, C. A.; Roberts, J. A.; Black, C. R. (2010). Direct effects of ozone on reproductive development in 

Plantago major L. populations differing in sensitivity. Environ Exp Bot, 69: 121-128. 

http://dx.doi.org/10.1016/j.envexpbot.2010.04.006 625575 

Blande, J. D.; Holopainen, J. K.; Li, T. (2010). Air pollution impedes plant-to-plant communication by volatiles. Ecol Lett, 

13: 1172-1181. http://dx.doi.org/10.1111/j.1461-0248.2010.01510.x 643928 

Bohler, S.; Bagard, M.; Oufir, M.; Planchon, S.; Hoffmann, L.; Jolivet, Y.; Hausman, J. F.; Dizengremel, P.; Renaut, J. 

(2007). A DIGE analysis of developing poplar leaves subjected to ozone reveals major changes in carbon 

metabolism. Proteomics, 7: 1584-1599. http://dx.doi.org/10.1002/pmic.200600822 199408 

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. 020581 

March 2011 9-139 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

Booker, F. L.; Burkey, K. O.; Pursley, W. A.; Heagle, A. S. (2007). Elevated carbon dioxide and ozone effects on peanut: I. 

Gas-exchange, biomass, and leaf chemistry. Crop Sci, 47: 1475-1487. 

http://dx.doi.org/10.2135/cropsci2006.08.0537 191370 

Booker, F. L.; Fiscus, E. L. (2005). The role of ozone flux and antioxidants in the suppression of ozone injury by elevated 

CO2 in soybean. J Exp Bot, 56: 2139-2151. http://dx.doi.org/10.1093/jxb/eri214 191652 

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. J Environ Manage, 1: S355-S362. 079138 

Booker, F. L.; Prior, S. A.; Torbert, H. A.; Fiscus, E. L.; Pursley, W. A.; Hu, S. (2005). Decomposition of soybean grown 

under elevated concentrations of CO2 and O3. Global Change Biol, 11: 685-698. http://dx.doi.org/10.1111/j.1365­

2486.2005.00939.x 079151 

Booker, F. L.; Reid, C. D.; Brunschon-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. 026425 

Booker, F.; Muntifering, R.; McGrath, M.; Burkey, K.; Decoteau, D.; Fiscus, E.; Manning, W.; Krupa, S.; Chappelka, A.; 

Grantz, D. (2009). The ozone component of global change: Potential effects on agricultural and horticultural plant 

yield, product quality and interactions with invasive species. J Integr Plant Biol, 51: 337-351. 

http://dx.doi.org/10.1111/j.1744-7909.2008.00805.x 191569 

Borowiak, K.; Rucinska-Sobkowiak, R.; Rymer, K.; Gwozdz, E. A.; Zbierska, J. (2009). Biochemical markers of 

tropospheric ozone: Experimentation with test-plants. Polish Journal of Ecology, 57: 3-14. 191247 

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. 021325 

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. 041857 

Bulbovas, P.; de Souza, S. R.; de Moraes, R. M.; Luizao, F.; Artaxo, P. (2007). Soybean 'Tracaja' seedlings exposed to 

ozone under controlled conditions. Pesquisa Agropecuaria Brasileira, 42: 641-646. http://dx.doi.org/10.1590/S0100­

204X2007000500005 199411 

Burkey, K. O.; Booker, F. L.; Pursley, W. A.; Heagle, A. S. (2007). Elevated carbon dioxide and ozone effects on peanut: II. 

Seed yield and quality. Crop Sci, 47: 1488-1497. http://dx.doi.org/10.2135/cropsci2006.08.0538 191371 

Burkey, K. O.; Eason, G.; Fiscus, E. L. (2003). Factors that affect leaf extracellular ascorbic acid content and redox status. 

Physiol Plant, 117: 51-57. http://dx.doi.org/10.1034/j.1399-3054.2003.1170106.x 630251 

Bytnerowicz, A.; Arbaugh, M.; Schilling, S.; Fraczek, W.; Alexander, D. (2008). Ozone distribution and phytotoxic 

potential in mixed conifer forests of the San Bernardino Mountains, Southern California. Environ Pollut, 155: 398­

408. http://dx.doi.org/10.1016/j.envpol.2008.01.046 196881 

Caird, M. A.; Richards, J. H.; Donovan, L. A. (2007). Nighttime stomatal conductance and transpiration in C-3 and C-4 

plants. Plant Physiol, 143: 4-10. http://dx.doi.org/10.1104/pp.106.092940 199337 

Calatayud, A.; Alvarado, J. W.; Barreno, E. (2002). Similar effects of ozone on four cultivars of lettuce in open top 

chambers during winter. Photosynthetica, 40: 195-200. http://dx.doi.org/10.1023/A:1021333305592 684222 

Calatayud, A.; Pomares, F.; Barreno, E. (2006). Interactions between nitrogen fertilization and ozone in watermelon 

cultivar Reina de Corazones in open-top chambers. Effects on chlorophyll alpha fluorescence, lipid peroxidation, 

and yield. Photosynthetica, 44: 93-101. http://dx.doi.org/10.1007/s11099-005-0163-2 191482 

Calatayud, V.; Cervero, J.; Sanz, M. J. (2007). Foliar, physiologial and growth responses of four maple species exposed to 

ozone. Water Air Soil Pollut, 185: 239-254. http://dx.doi.org/10.1007/s11270-007-9446-5 191411 

Calatayud, V.; Sanz, M. J.; Calvo, E.; Cervero, J.; Ansel, W.; Klumpp, A. (2007). Ozone biomonitoring with Bel-W3 

tobacco plants in the city of Valencia (Spain). Water Air Soil Pollut, 183: 283-291. 

http://dx.doi.org/10.1007/s11270-007-9376-2 191568 

Cal EPA (2009). Air quality data branch main page. Retrieved January 28, 2011 from 

http://www.arb.ca.gov/aqd/aqdpage.htm. 677487 

March 2011 9-140 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

Campbell, S. J.; Wanek, R.; Coulston, J. W. (2007). Ozone injury in west coast forests: 6 years of monitoring ­

Introduction. Portland, OR: U.S. Department of Agriculture. 602360 

Cannon, W. N. (1990). Olfactory response of eastern spruce budworm larvae to red spruce needles exposed to acid rain and 

elevated levels of ozone. J Chem Ecol, 16: 3255-3261. 626460 

Carde, R. T.; Haynes, K. F. (2004). Stucture of the pheromone communication channel in moths. In Advances in insect 

chemical ecology (pp. 283-332). Cambridge: Cambridge University Press. 626467 

Casteel, C. L.; O'Neill, B. F.; Zavala, J. A.; Bilgin, D. D.; Berenbaum, M. R.; DeLucia, E. H. (2008). Transcriptional 

profiling reveals elevated CO2 and elevated O-3 alter resistance of soybean (Glycine max) to Japanese beetles 

(Popillia japonica). Plant Cell Environ, 31: 419-434. http://dx.doi.org/10.1111/j.1365-3040.2008.01782.x 191696 

Chapman, J. A.; King, J. S.; Pregitzer, K. S.; Zak, D. R. (2005). Effects of elevated concentrations of atmospheric CO2 and 

tropospheric O-3 on decomposition of fine roots. Tree Physiol, 25: 1501-1510. 191345 

Chappelka, A. H. (2002). Reproductive development of blackberry (Rubus cuneifolius) as influenced by ozone. New 

Phytol, 155: 249-255. 052946 

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. 093687 

Chappelka, A. H.; Somers, G. L.; Renfro, J. R. (2007). Temporal patterns of foliar ozone symptoms on tall milkweed 

(Asclepias exaltata L) in Great Smokey Mountains National Park. Environ Pollut, 149: 358-365. 

http://dx.doi.org/10.1016/j.envpol.2007.05.015 093290 

Chappelka, A.; Skelly, J.; Somers, G.; Renfro, J.; Hildebrand, E. (1999). Mature black cherry used as a bioindicator of 

ozone injury. Water Air Soil Pollut, 116: 261-266. 052952 

Chappelka, A.; Somers, G.; Renfro, J. (1999). Visible ozone injury on forest trees in Great Smoky Mountains National 

Park, USA. Water Air Soil Pollut, 116: 255-260. 041860 

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. 040317 

Chen, Z.; Gallie, D. R. (2005). Increasing tolerance to ozone by elevating foliar ascorbic acid confers greater protection 

against ozone than increasing avoidance. Plant Physiol, 138: 1673-1689. http://dx.doi.org/10.1104/pp.105.062000 

191465 

Chen, Z.; Wang, X. K.; Feng, Z. Z.; Xiao, Q.; Duan, X. N. (2009). Impact of Elevated O-3 on Soil Microbial Community 

Function Under Wheat Crop. Water Air Soil Pollut, 198: 189-198. http://dx.doi.org/10.1007/s11270-008-9838-1 

191452 

Chen, Z.; Wang, X. K.; Yao, F. F.; Zheng, F. X.; Feng, Z. Z. (2010). Elevated ozone changed soil microbial community in a 

rice paddy. Soil Sci Soc Am J, 74: 829-837. http://dx.doi.org/10.2136/sssaj2009.0258 644257 

Cheng, F. Y.; Burkey, K. O.; Robinson, J. M.; Booker, F. L. (2007). Leaf extracellular ascorbate in relation to O-3 tolerance 

of two soybean cultivars. Environ Pollut, 150: 355-362. http://dx.doi.org/10.1016/j.envpol.2007.01.022 191499 

Cho, K.; Shibato, J.; Agrawal, G. K.; Jung, Y. H.; Kubo, A.; Jwa, N. S.; Tamogami, S.; Satoh, K.; Kikuchi, S.; Higashi, T.; 

Kimura, S.; Saji, H.; Tanaka, Y.; Iwahashi, H.; Masuo, Y.; Rakwal, R. (2008). Integrated transcriptomics, 

proteomics, and metabolomics analyses to survey ozone responses in the leaves of rice seedling. J Proteome Res, 7: 

2980-2998. http://dx.doi.org/10.1021/pr800128q 603254 

Chung, H. G.; Zak, D. R.; Lilleskov, E. A. (2006). Fungal community composition and metabolism under elevated CO2 and 

O-3. Oecologia, 147: 143-154. http://dx.doi.org/10.1007/s00442-005-0249-3 191729 

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. 026469 

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. 041871 

Crous, K. Y.; Vandermeiren, K.; Ceulemans, R. (2006). Physiological responses to cumulative ozone uptake in two white 

clover (Trifolium repens L. cv. Regal) clones with different ozone sensitivity. Environ Exp Bot, 58: 169-179. 

http://dx.doi.org/10.1016/j.envexpbot.2005.07.007 199321 

March 2011 9-141 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

D'Haese, D.; Horemans, N.; De Coen, W.; Guisez, Y. (2006). Identification of late O-3-responsive genes in Arabidopsis 

thaliana by cDNA microarray analysis. Physiol Plant, 128: 70-79. http://dx.doi.org/10.1111/j.1399­

3054.2006.00711.x 191448 

D'Haese, D.; Vandermeiren, K.; Asard, H.; Horemans, N. (2005). Other factors than apoplastic ascorbate contribute to the 

differential ozone tolerance of two clones of Trifolium repens L. Plant Cell Environ, 28: 623-632. 

http://dx.doi.org/10.1111/j.1365-3040.2005.01308.x 191551 

Dalstein, L.; Vas, N. (2005). Ozone concentrations and ozone-induced symptoms on coastal and alpine mediterranean pines 

in southern France. Water Air Soil Pollut, 160: 181-195. 191570 

Dammgen, U.; Grunhage, L.; Haenel, H.-D.; Jager, H.-J. (1993). Climate and stress in ecotoxicology A coherent system of 

definitions and terms. J Appl Bot Food Qual, 67: 157-162. 055312 

Darbah, J. N. T.; Kubiske, M. E.; Neilson, N.; Oksanen, E.; Vaapavuori, E.; Karnosky, D. F. (2007). Impacts of elevated 

atmospheric CO2 and O3 on paper birch (Betula papyrifera): reproductive fitness. ScientificWorldJournal, 7: 240­

246. 093288 

Darbah, J. N. T.; Kubiske, M. E.; Nelson, N.; Oksanen, E.; Vapaavuori, E.; Kamosky, D. F. (2008). Effects of decadal 

exposure to interacting elevated CO2 and/or O-3 on paper birch (Betula papyrifera) reproduction. Environ Pollut, 

155: 446-452. http://dx.doi.org/10.1016/j.envpol.2008.01.033 196890 

Davidson, A. (1993). Update of ozone trends in California's South Coast Air Basin. J Air Waste Manag Assoc, 43: 226-227. 

043366 

Davis, D. D. (2007). Ozone injury to plants within the Seney National Wildlife Refuge in northern Michigan. Northeast 

Nat, 14: 415-424. 093291 

Davis, D. D. (2007). Ozone-induced symptoms on vegetation within the Moosehorn National Wildlife Refuge in Maine. 

Northeast Nat, 14: 403-414. 093292 

Davis, D. D. (2009). Ozone-Induced Stipple on Plants in the Cape Romain National Wildlife Refuge, South Carolina. 

Southeastern Naturalist, 8: 471-478. 199287 

Davis, D. D.; Orendovici, T. (2006). Incidence of ozone symptoms on vegetation within a National Wildlife Refuge in New 

Jersey, USA. Environ Pollut, 143: 555-564. http://dx.doi.org/10.1016/j.envpol.2005.10.051 093293 

Dawson, T. E.; Burgess, S. S.; Tu, K. P.; Oliveira, R. S.; Santiago, L. S.; Fisher, J. B.; Simonin, K. A.; Ambrose, A. R. 

(2007). Nighttime transpiration in woody plants from contrasting ecosystems. Tree Physiol, 27: 561-575. 

http://dx.doi.org/10.1093/treephys/27.4.561 670381 

De Temmerman, L.; Legrand, G.; Vandermeiren, K. (2007). Effects of ozone on sugar beet grown in open-top chambers. 

Eur J Agron, 26: 1-9. http://dx.doi.org/10.1016/j.eja.2006.08.001 191361 

Degl'Innocenti, E.; Guidi, L.; Soldatini, G. F. (2007). Effects of elevated ozone on chlorophyll a fluorescence in 

symptomatic and asymptomatic leaves of two tomato genotypes. Biol Plantarum, 51: 313-321. 

http://dx.doi.org/10.1007/s10535-007-0061-5 191350 

Dermody, O.; O'Neill, B. F.; Zangerl, A. R.; Berenbaum, M. R.; DeLucia, E. H. (2008). Effects of elevated CO2 and O-3 on 

leaf damage and insect abundance in a soybean agroecosystem. Arthropod-Plant Interactions, 2: 125-135. 191850 

de Lourdes de Bauer, M.; Hernandez-Tejeda, T. (2007). A review of ozone-induced effects on the forests of central Mexico. 

Environ Pollut, 147: 446-453. 196891 

Dickson, R. E.; Lewin, K. F.; Isebrands, J. G.; Coleman, M. D.; Heilman, W. E.; Riemenschneider, D. E.; Sober, J.; Host, G. 

E.; Zak, D. R.; Hendrey, G. R.; Pregitzer, K. S.; Karnosky, D. F. (2000). Forest Atmosphere Carbon Transfer and 

Storage (FACTS-II) the Aspen Free-Air CO2 and O3 Enrichment (FACE) project : An overview (Report No. 

General Technical Report NC-214). St. Paul, MN: U.S. Dept. of Agriculture, Forest Service, North Central 

Research Station. http://nrs.fs.fed.us/pubs/278. 628220 

Ditchkoff, S. S.; Lewis, J. S.; Lin, J. C.; Muntifering, R. B.; Chappelka, A. H. (2009). Nutritive quality of highbush 

blackberry (Rubus argutus) exposed to tropospheric ozone. R E M, 62: 364-370. 192230 

Dizengremel, P.; Le Thiec, D.; Bagard, M.; Jolivet, Y. (2008). Ozone risk assessment for plants: Central role of 

metabolism-dependent changes in reducing power. Environ Pollut, 156: 11-15. 

http://dx.doi.org/10.1016/j.envpol.2007.12.024 191587 

March 2011 9-142 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

Dizengremel, P.; Le Thiec, D.; Hasenfratz-Sauder, M. P.; Vaultier, M. N.; Bagard, M.; Jolivet, Y. (2009). Metabolicdependent changes in plant cell redox power after ozone exposure. Plant Biol (Stuttg), 11: 35-42. 

http://dx.doi.org/10.1111/j.1438-8677.2009.00261.x 199424 

Dizengremel, P.; Sasek, T.; Brown, K.; Richardson, C. (1994). Ozone-induced changes in primary carbon metabolism 

enzymes of loblolly pine needles. J Plant Physiol, 144: 300-306. 187217 

Di Baccio, D.; Castagna, A.; Paoletti, E.; Sebastiani, L.; Ranieri, A. (2008). Could the differences in O(3) sensitivity 

between two poplar clones be related to a difference in antioxidant defense and secondary metabolic response to 

O(3) influx? Tree Physiol, 28: 1761-1772. 199850 

Dobson, H. E. M. (1994). Floral volatiles in insect biology. In Insect-Plant Interactions (pp. 47-82). Boca Raton, FL: 

Bernays, E. 626466 

Dohm, M. R.; Mautz, W. J.; Andrade, J. A.; Gellert, K. S.; Salas-Ferguson, L. J.; Nicolaisen, N.; Fujie, N. (2005). Effects of 

ozone exposure on nonspecific phagocytic capacity of pulmonary macrophages from an amphibian, Bufo marinus. 

Environ Toxicol Chem, 24: 205-210. 180452 

Dohm, M. R.; Mautz, W. J.; Doratt, R. E.; Stevens, J. R. (2008). Ozone exposure affects feeding and locomotor behavior of 

adult bufo marinus. Environ Toxicol Chem, 27: 1209-1216. 604584 

Dohm, M. R.; Mautz, W. J.; Looby, P. G.; Gellert, K. S.; Andrade, J. A. (2001). Effects of ozone on evaporative water loss 

and thermoregulatory behavior of marine toads (Bufo marinus). Environ Res, 86: 274-286. 017086 

Dohrmann, A. B.; Tebbe, C. C. (2005). Effect of elevated tropospheric ozone on the structure of bacterial communities 

inhabiting the rhizosphere of herbaceous plants native to Germany. Appl Environ Microbiol, 71: 7750-7758. 

http://dx.doi.org/10.1128/AEM.71.12.7750-7758.2005 191320 

Drogoudi, P. D.; Ashmore, M. (2001). 14C-allocation of flowering and deblossomed strawberry in response to elevated 

ozone. New Phytol, 152: 455-461. 052959 

Drogoudi, P. D.; Ashmore, M. R. (2000). Does elevated ozone have differing effects in flowering and deblossomed 

strawberry? New Phytol, 147: 561-569. 042517 

Dudareva, N.; Negre, F.; Nagegowda, D. A.; Orlova, I. (2006). Plant Volatiles: recent advances and future perspectives. 

Crit Rev Plant Sci, 25: 417-440. 626458 

Ederli, L.; Morettini, R.; Borgogni, A.; Wasternack, C.; Miersch, O.; Reale, L.; Ferranti, F.; Tosti, N.; Pasqualini, S. (2006). 

Interaction between nitric oxide and ethylene in the induction of alternative oxidase in ozone-treated tobacco plants. 

Plant Physiol, 142: 595-608. http://dx.doi.org/10.1104/pp.106.085472 191479 

Ellsworth, D. S.; Reich, P. B.; Naumburg, E. S.; Koch, G. W.; Kubiske, M. E.; Smith, S. D. (2004). Photosynthesis, 

carboxylation and leaf nitrogen responses of 16 species to elevated pCO2 across four free-air CO2 enrichment 

experiments in forest, grassland and desert. Global Change Biol, 10: 2121-2138. 080092 

Eltayeb, A. E.; Kawano, N.; Badawi, G. H.; Kaminaka, H.; Sanekata, T.; Morishima, I.; Shibahara, T.; Inanaga, S.; Tanaka, 

K. (2006). Enhanced tolerance to ozone and drought stresses in transgenic tobacco overexpressing 

dehydroascorbate reductase in cytosol. Physiol Plant, 127: 57-65. http://dx.doi.org/10.1111/j.1399­

3054.2005.00624.x 191377
 

Eltayeb, A. E.; Kawano, N.; Badawi, G. H.; Kaminaka, H.; Sanekata, T.; Shibahara, T.; Inanaga, S.; Tanaka, K. (2007). 

Overexpression of monodehydroascorbate reductase in transgenic tobacco confers enhanced tolerance to ozone, salt 

and polyethylene glycol stresses. Planta, 225: 1255-1264. http://dx.doi.org/10.1007/s00425-006-0417-7 191553 

Emberson, L. D.; Wieser, G.; Ashmore, M. R. (2000). Modelling of stomatal conductance and ozone flux of Norway 

spruce: comparison with field data. Environ Pollut, 109: 393-402. 042537 

Emberson, L.; Ashmore, M. R.; Cambridge, H. M.; Simpson, D.; Tuovinen, J.-P. (2000). Modelling stomatal ozone flux 

across Europe. Environ Pollut, 109: 403-413. 040350 

Enders, G. (1992). Deposition of ozone to a mature spruce forest: measurements and comparison to models. Environ Pollut, 

75: 61-67. 040355 

Esperschutz, J.; Pritsch, K.; Gattinger, A.; Welzl, G.; Haesler, F.; Buegger, F.; Winkler, J. B.; Munch, J. C.; Schloter, M. 

(2009). Influence of chronic ozone stress on carbon translocation pattern into rhizosphere microbial communities of 

beech trees (Fagus sylvatica L.) during a growing season. Plant Soil, 323: 85-95. http://dx.doi.org/10.1007/s11104­

009-0090-2 595669 

March 2011 9-143 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

Fares, S.; Barta, C.; Brilli, F.; Centritto, M.; Ederli, L.; Ferranti, F.; Pasqualini, S.; Reale, L.; Tricoli, D.; Loreto, F. (2006). 

Impact of high ozone on isoprene emission, photosynthesis and histology of developing Populus alba leaves 

directly or indirectly exposed to the pollutant. Physiol Plant, 128: 456-465. http://dx.doi.org/10.1111/j.1399­

3054.2006.00750.x 191455 

Fares, S.; Loreto, F.; Kleist, E.; Wildt, J. (2008). Stomatal uptake and stomatal deposition of ozone in isoprene and 

monoterpene emitting plants. Plant Biol (Stuttg), 10: 44-54. http://dx.doi.org/10.1055/s-2007-965257 191683 

Fares, S.; Oksanen, E.; Lannenpaa, M.; Julkunen-Tiitto, R.; Loreto F (2010). Volatile emissions and phenolic compound 

concentrations along a vertical profile of Populus nigra leaves exposed to realistic ozone concentrations. Photosynth 

Res, 104: 61-74. http://dx.doi.org/10.1007/s11120-010-9549-5 628522 

Federal land managers' air quality related values workgroup (FLAG) phase I report. (2000). Lakewood, CO: U.S. Forest 

Service. 088923 

Felzer, B. S.; Cronin, T. W.; Melillo, J. M.; Kicklighter, D. W.; Schlosser, C. A. (2009). Importance of carbon-nitrogen 

interactions and ozone on ecosystem hydrology during the 21st century. J Geophys Res, 114: G01020. 

http://dx.doi.org/10.1029/2008jg000826 191460 

Felzer, B.; Kicklighter, D.; Melillo, J.; Wang, C.; Xhuang, Q.; Prinn, R. (2004). Effects of ozone on net primary production 

and carbon sequestration in the conterminous United States using a biogeochemistry model. Tellus B Chem Phys 

Meteorol, 56B: 230-248. 186927 

Felzer, B.; Reilly, J.; Melillo, J.; Kicklighter, D.; Sarofim, M.; Wang, C.; Prinn, R.; Zhuang, Q. (2005). Future effects of 

ozone on carbon sequestration and climate change policy using a global biogeochemical model. Clim Change, 73: 

345-373. http://dx.doi.org/10.1007/s10584-005-6776-4 186928 

Feng, Y. W.; Komatsu, S.; Furukawa, T.; Koshiba, T.; Kohno, Y. (2008). Proteome analysis of proteins responsive to 

ambient and elevated ozone in rice seedlings. Agric Ecosyst Environ, 125: 255-265. 

http://dx.doi.org/10.1016/j.agee.2008.01.018 191626 

Feng, Z. Z.; Kobayashi, K. (2009). Assessing the impacts of current and future concentrations of surface ozone on crop 

yield with meta-analysis. Atmos Environ, 43: 1510-1519. http://dx.doi.org/10.1016/j.atmosenv.2008.11.033 199223 

Feng, Z. Z.; Kobayashi, K.; Ainsworth, E. A. (2008). Impact of elevated ozone concentration on growth, physiology, and 

yield of wheat (Triticum aestivum L.): a meta-analysis. Global Change Biol, 14: 2696-2708. 

http://dx.doi.org/10.1111/j.1365-2486.2008.01673.x 191453 

Feng, Z. Z.; Yao, F. F.; Chen, Z.; Wang, X. K.; Meng, Q. W.; Feng, Z. W. (2007). Response of gas exchange and yield 

components of field-grown Triticum aestivum L. to elevated ozone in China. Photosynthetica, 45: 441-446. 

http://dx.doi.org/10.1007/s11099-007-0073-6 191641 

Fenn, M. E.; de Bauer, L. I.; Hernández-Tejeda, T. (2002). Summary of air pollution impacts on forests in the Mexico City 

air basin. In Urban Air Pollution and Forests (pp. 337-355). New York, NY: Springer-Verlag New York, Inc. 626806 

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. 083540 

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. 036337 

Finkelstein, P. L.; Ellestad, T. G.; Clarke, J. F.; Meyers, T. P.; Schwede, D. B.; Hebert, E. O.; Neal, J. A. (2000). Ozone and 

sulfur dioxide dry deposition to forests: observations and model evaluation. J Geophys Res, 105: 15,365-15,377. 

024029 

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. 083315 

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. 042545 

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. 079155 

Fiscus, E. L.; Philbeck, R.; Britt, A. M.; Booker, F. L. (1999). Growth of Arabidopsis flavonoid mutants under solar 

radiation and UV filters. Environ Exp Bot, 41: 231-245. http://dx.doi.org/10.1016/S0098-8472(99)00011-8 672964 

March 2011 9-144 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

Fishman, J.; Bowman, K. W.; Burrows, J. P.; Richter, A.; Chance, K. V.; Edwards, D. P.; Martin, R. V.; Morris, G. A.; 

Pierce, R. B.; Ziemke, J. R.; Al-Saadi, J. A.; Creilson, J. K.; Schaack, T. K.; Thompson, A. M. (2008). Remote 

sensing of tropospheric pollution from space. Bull Am Meteorol Soc, 89: 805-821. 

http://dx.doi.org/10.1175/2008BAMS2526.1 193927 

Fishman, J.; Creilson, J. K.; Parker, P. A.; Ainsworth, E. A.; Vining, G. G.; Szarka, J.; Booker, F. L.; Xu, X. J. (2010). An 

investigation of widespread ozone damage to the soybean crop in the upper Midwest determined from groundbased and satellite measurements. Atmos Environ, 44: 2248-2256. 

http://dx.doi.org/10.1016/j.atmosenv.2010.01.015 644259 

Flagler, R. B. (1998). Recognition of air pollution injury to vegetation: A pictorial atlas. Pittsburgh, PA: Air & Waste 

Management Association. 025525 

Flowers, M. D.; Fiscus, E. L.; Burkey, K. O.; Booker, F. L.; Dubois, J. J. B. (2007). Photosynthesis, chlorophyll 

fluorescence, and yield of snap bean (Phaseolus vulgaris L.) genotypes differing in sensitivity to ozone. Environ 

Exp Bot, 61: 190-198. http://dx.doi.org/10.1016/j.envexpbot.2007.05.009 191852 

Fontan, J. A.; Minga, A.; Lopez, A.; Druilhet, A. (1992). Vertical ozone profiles in a pine forest. Atmos Environ, 26A: 863­

869. 040370 

Foyer, C. H.; Noctor, G. (2005). Oxidant and antioxidant signalling in plants: A re-evaluation of the concept of oxidative 

stress in a physiological context. Plant Cell Environ, 28: 1056-1071. 191555 

Foyer, C. H.; Noctor, G. (2005). Redox homeostasis and antioxidant signaling: A metabolic interface between stress 

perception and physiological responses. Plant Cell, 17: 1866-1875. http://dx.doi.org/10.1105/tpc.105.033589 

631198 

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. 038898 

Fredericksen, T. S.; Kolb, T. E.; Skelly, J. M.; Steiner, K. C.; Joyce, B. J.; Savage, J. E. (1996). Light environment alters 

ozone uptake per net photosynthetic rate in black cherry trees. Tree Physiol, 16: 485-490. 026653 

Freiwald, V.; Haikio, E.; Julkunen-Tiitto, R.; Holopainen, J. K.; Oksanen, E. (2008). Elevated ozone modifies the feeding 

behaviour of the common leaf weevil on hybrid aspen through shifts in developmental, chemical, and structural 

properties of leaves. Entomol Exp Appl, 128: 66-72. http://dx.doi.org/10.1111/j.1570-7458.2008.00677.x 196903 

Fuhrer, J. (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. 055549 

Fuhrer, J.; Skarby, L.; Ashmore, M. R. (1997). Critical levels for ozone effects on vegetation in Europe. Environ Pollut, 97: 

91-106. 030380 

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. 026655 

Gerosa, G.; Marzuoli, R.; Rossini, M.; Panigada, C.; Meroni, M.; Colombo, R.; Faoro, F.; Iriti, M. (2009). A flux-based 

assessment of the effects of ozone on foliar injury, photosynthesis, and yield of bean (Phaseolus vulgaris L. cv. 

Borlotto Nano Lingua di Fuoco) in open-top chambers. Environ Pollut, 157: 1727-1736. 

http://dx.doi.org/10.1016/j.envpol.2008.06.028 191403 

Gielen, B.; Vandermeiren, K.; Horemans, N.; D'Haese, D.; Serneels, R.; Valcke, R. (2006). Chlorophyll a fluorescence 

imaging of ozone-stressed Brassica napus L. plants differing in glucosinolate concentrations. Plant Biol (Stuttg), 8: 

698-705. http://dx.doi.org/10.1055/s-2006-924150 191271 

Gitay, H.; Brown, S.; Easterling, W.; Jallow, B. (2001). Ecosystems and their goods and services. In Climate change 2001: 

impacts, adaptation and vulnerability. Contribution of working group II to the third assessment report of the 

Intergovernmental Panel on Climate Change (pp. 237-342). Cambridge, United Kingdom: Cambridge University 

Press. 092761 

Gombert, S.; Asta, J.; Seaward, M. R. D. (2006). Lichens and tobacco plants as complementary biomonitors of air pollution 

in the Grenoble area (Isere, southeast France). Ecol Indicat, 6: 429-443. 089309 

March 2011 9-145 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

Gonzalez-Fernandez, I.; Kaminska, A.; Dodmani, M.; Goumenaki, E.; Quarrie, S.; Barnes, J. D. (2010). Establishing ozone 

flux-response relationships for winter wheat: Analysis of uncertainties based on data for UK and Polish genotypes. 

Atmos Environ, 44: 621-630. http://dx.doi.org/10.1016/j.atmosenv.2009.11.021 381357 

Grantz, D. A.; Gunn, S.; Vu, H. B. (2006). O-3 impacts on plant development: a meta-analysis of root/shoot allocation and 

growth. Plant Cell Environ, 29: 1193-1209. http://dx.doi.org/10.1111/j.1365-3040.2006.01521.x 191545 

Grantz, D. A.; Shrestha, A. (2006). Tropospheric ozone and interspecific competition between yellow nutsedge and pima 

cotton. Crop Sci, 46: 1879-1889. http://dx.doi.org/10.2135/cropsci2005.06.0167 191702 

Grantz, D. A.; Shrestha, A.; Vu, H. B. (2008). Early vigor and ozone response in horseweed (Conyza canadensis) biotypes 

differing in glyphosate resistance. Weed Sci, 56: 224-230. http://dx.doi.org/10.1614/ws-07-130.1 191312 

Grantz, D. A.; Shrestha, A.; Vu, H. B. (2010). Ozone impacts on assimilation and allocation to reproductive sinks in the 

vegetatively propagated C-4 weed, yellow nutsedge. Crop Sci, 50: 246-252. 

http://dx.doi.org/10.2135/cropsci2009.03.0127 102161 

Grantz, D. A.; Vu, H. B. (2009). O3 sensitivity in a potential C4 bioenergy crop: Sugarcane in California. Crop Sci, 49: 

643-650. 195237 

Grantz, D. A.; Vu, H. B.; Aguilar, C.; Rea MA (2010). No interaction between methyl jasmonate and ozone in Pima cotton: 

Growth and allocation respond independently to both. Plant Cell Environ, 33: 717-728. 

http://dx.doi.org/10.1111/j.1365-3040.2009.02096.x 625419 

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. 026659 

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. Agr 

Forest Meteorol, 85: 19-31. 026664 

Grebenc, T.; Kraigher, H. (2007). Changes in the community of ectomycorrhizal fungi and increased fine root number 

under adult beech trees chronically fumigated with double ambient ozone concentration. Plant Biol (Stuttg), 9: 279­

287. http://dx.doi.org/10.1055/s-2006-924489 191265 

Gregg, J. W.; Jones, C. G.; Dawson, T. E. (2003). Urbanization effects on tree growth in the vicinity of New York City 

[letter]. Nature, 424: 183-187. 046996 

Gregg, J.; Jones, C.; Dawson, T. (2006). Physiological and developmental effects of O3 on cottonwood growth in urban 

and rural sites. Ecol Appl, 16: 2368-2381. http://dx.doi.org/10.1890/1051-0761(2006)016[2368:PADEOO]2.0.CO;2 

186961 

Grennfelt, P. (2004). New directions: recent research findings may change ozone control policies. Atmos Environ, 38: 

2215-2216. 056625 

Groppa, M. D.; Benavides, M. P. (2008). Polyamines and abiotic stress: Recent advances. Amino Acids, 34: 35-45. 

http://dx.doi.org/10.1007/s00726-007-0501-8 191616 

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 New York, NY: 

Springer-Verlag. 052983 

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. 042646 

Grulke, N. E.; Dobrowolski, W.; Mingus, P.; Fenn, M. E. (2005). California black oak response to nitrogen amendment at a 

high O3, nitrogen-saturated site. Environ Pollut, 137: 536-545. http://dx.doi.org/10.1016/j.envpol.2005.01.039 

199433 

Grulke, N. E.; Johnson, R.; Esperanza, A.; Jones, D.; Nguyen, T.; Posch, S.; Tausz, M. (2003). Canopy transpiration of 

Jeffrey pine in mesic and xeric microsites: O3 uptake and injury response. Trees Struct Funct, 17: 292-298. 042637 

Grulke, N. E.; Johnson, R.; Monschein, S.; Nikolova, P.; Tausz, M. (2003). Variation in morphological and biochemical O3 

injury attributes of mature Jeffrey pine within canopies and between microsites. Tree Physiol, 23: 923-929. 052984 

Grulke, N. E.; Paoletti, E.; Heath, R. L. (2007). Chronic vs. short-term acute O3 exposure effects on nocturnal transpiration 

in two Californian oaks. ScientificWorldJournal, 7: 134-140. http://dx.doi.org/10.1100/tsw.2007.33 199376 

March 2011 9-146 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

Grulke, N. E.; Preisler, H. K.; Rose, C.; Kirsch, J.; Balduman, L. (2002). O3 uptake and drought stress effects on carbon 

acquisition of ponderosa pine in natural stands. New Phytol, 154: 621-631. 035283 

Grulke, N.; Neufeld, H.; Davison, A.; Roberts, M.; Chappelka, A. (2007). Stomatal behavior of ozone-sensitive and ­

insensitive coneflowers (Rudbeckia laciniata var. digitata) in Great Smoky Mountains National Park. New Phytol, 

173: 100-109. http://dx.doi.org/10.1111/j.1469-8137.2006.01872.x 186963 

Grunhage, 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. 040392 

Grunhage, L.; Jager, 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. 052972 

Grunhage, L.; Krupa, S. V.; Legge, A. H.; Jager, 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. 056621 

Guderian, R. (1985). Air pollution by photochemical oxidants: formation, transport, control, and effects on plants. New 

York: Springer-Verlag. 019325 

Guenther, A.; Karl, T.; Harley, P.; Wiedinmyer, C.; Palmer, P. I.; Geron, C. (2006). Estimates of global terrestrial isoprene 

emissions using MEGAN (Model of Emissions of Gases and Aerosols from Nature). Atmos Chem Phys, 6: 3181­

3210. http://dx.doi.org/10.5194/acp-6-3181-2006 607080 

Guidi, L.; Degl'Innocenti, E.; Martinelli, F.; Piras, M. (2009). Ozone effects on carbon metabolism in sensitive and 

insensitive Phaseolus cultivars. Environ Exp Bot, 66: 117-125. http://dx.doi.org/10.1016/j.envexpbot.2008.12.005 

199825 

Guidi, L.; Degl'lnnocenti, E. (2008). Ozone effects on high light-induced photoinhibition in Phaseolus vulgaris. Plant Sci, 

174: 590-596. http://dx.doi.org/10.1016/j.plantsci.2008.03.003 191571 

Gumpertz, M. L.; Rawlings, J. O. (1992). Nonlinear regression with variance components: modeling effects of ozone on 

crop yield. Crop Sci, 32: 219-224. 043259 

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. 080097 

Haberer, K.; Herbinger, K.; Alexou, M.; Rennenberg, H.; Tausz, M. (2008). Effects of drought and canopy ozone exposure 

on antioxidants in fine roots of mature European beech (Fagus sylvatica). Tree Physiol, 28: 713-719. 

http://dx.doi.org/10.1093/treephys/28.5.713 191334 

Hamel, L. P.; Miles, G. P.; Samuel, M. A.; Ellis, B. E.; Seguin, A.; Beaudoin, N. (2005). Activation of stress-responsive 

mitogen-activated protein kinase pathways in hybrid poplar (Populus trichocarpa x Populus deltoides). Tree 

Physiol, 25: 277-288. http://dx.doi.org/10.1093/treephys/25.3.277 191214 

Handley, T.; Grulke, N. E. (2008). Interactive effects of O-3 exposure on California black oak (Quercus kelloggii Newb.) 

seedlings with and without N amendment. Environ Pollut, 156: 53-60. 

http://dx.doi.org/10.1016/j.envpol.2008.01.002 191485 

Hanson, P. J.; Wullschleger, S. D.; Norby, R. J.; Tschaplinski, T. J.; Gunderson, C. A. (2005). Importance of changing CO2, 

temperature, precipitation, and ozone on carbon and water cycles of an upland-oak forest: incorporating 

experimental results into model simulations. Global Change Biol, 11: 1402-1423. http://dx.doi.org/10.1111/j.1365­

2486.2005.00991.x 191461 

Harward, M.; Treshow, M. (1975). Impact of ozone on the growth and reproduction of understorey plants in the Aspen 

zone of western USA. Environ Conserv, 2: 17-23. 038366 

Hassan, R.; Scholes, R.; Ash, N. (2005). Ecosystems and human well-being: current state and trends, volume 1. United 

Kingdom: Shearwater Books. 092759 

Hayes, F.; Jones, M. L. M.; Mills, G.; Ashmore, M. (2007). Meta-analysis of the relative sensitivity of semi-natural 

vegetation species to ozone. Environ Pollut, 146: 754-762. http://dx.doi.org/10.1016/j.envpol.2006.06.011 196911 

Hayes, F.; Mills, G.; Ashmore, M. (2009). Effects of ozone on inter- and intra-species competition and photosynthesis in 

mesocosms of Lolium perenne and Trifolium repens. Environ Pollut, 157: 208-214. 

http://dx.doi.org/10.1016/j.envpol.2008.07.002 191360 

March 2011 9-147 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

He, X. Y.; Fu, S. L.; Chen, W.; Zhao, T. H.; Xu, S.; Tuba, Z. (2007). Changes in effects of ozone exposure on growth, 

photosynthesis, and respiration of Ginkgo biloba in Shenyang urban area. Photosynthetica, 45: 555-561. 

http://dx.doi.org/10.1007/s11099-007-0095-0 199789 

He, X. Y.; Ruan, Y. N.; Chen, W.; Lu, T. (2006). Responses of anti-oxidative system in leaves of Ginkgo biloba to elevated 

ozone concentration in urban area. Botanical Studies, 47: 409-416. 199430 

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. 039329 

Heagle, A. S. (1989). Ozone and crop yield*. Annu Rev Phytopathol, 27: 397-423. 

http://dx.doi.org/10.1146/annurev.py.27.090189.002145 093985 

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. 038348 

Heagle, A. S.; Brandenburg, R. L.; Burns, J. C.; Miller, J. E. (1994). Ozone and carbon dioxide effects on spider mites in 

white clover and peanut. J Environ Qual, 23: 1168-1176. 026699 

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. 042124 

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 Assessment of crop loss from air pollutants 

(pp. 141-149). London: Elsevier Applied Science. 043559 

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. 039372 

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. 042530 

Heagle, A. S.; Miller, J. E.; Sherrill, D. E. (1994). A white clover system to estimate effects of tropospheric ozone on 

plants. J Environ Qual, 23: 613-621. 042656 

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. 042660 

Heath, R. L. (2008). Modification of the biochemical pathways of plants induced by ozone: What are the varied routes to 

change? Environ Pollut, 155: 453-463. http://dx.doi.org/10.1016/j.envpol.2008.03.010 195632 

Heath, R. L.; Lefohn, A. S.; Musselman, R. C. (2009). Temporal processes that contribute to nonlinearity in vegetation 

responses to ozone exposure and dose. Atmos Environ, 43: 2919-2928. 

http://www.sciencedirect.com/science/article/B6VH3-4VV2NR8-5/2/1f59c10d7c941154fd5cfa8d69f22dbc. 196783 

Heck, W. W.; Cowling, E. B. (1997). The need for a long term cumulative secondary ozone standard - an ecological 

perspective. EM, January: 23-33. 084484 

Heck, W. W.; Cure, W. W.; Rawlings, J. O.; Zaragoza, L. J.; Heagle, A. S.; Heggestad, H. E.; Kohut, R. J.; Kress, L. W.; 

Temple, P. J. (1984). Assessing impacts of ozone on agricultural crops: II. Crop yield functions and alternative 

exposure statistics. J Air Pollut Control Assoc, 34: 810-817. 039380 

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 (pp. 225-254). Pittsburgh, 

PA: Air & Waste Management Association. 042621 

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. 

Government Printing Office. 037673 

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 Waste Manag Assoc, 32: 353-361. 039525 

Heggestad, H. E. (1991). Origin of Bel-W3, Bel-C and Bel-B tobacco varieties and their use as indicators of ozone. 

Environ Pollut, 74: 263-291. 042533 

March 2011 9-148 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

Heidenreich, B.; Haberer, G.; Mayer, K.; Sandermann, H.; Ernst, D. (2005). CDNA array-analysis of mercury- and ozoneinduced genes in Arabidopsis thaliana. Acta Physiologiae Plantarum, 27: 45-51. http://dx.doi.org/10.1007/s11738­

005-0035-1 191260 

Hendrey, G. R.; Ellsworth, D. S.; Lewin, K. F.; Nagy, J. (1999). A free-air enrichment system for exposing tall forest 

vegetation to elevated atmospheric CO2. Global Change Biol, 5: 293-309. 042641 

Hendrey, G. R.; Kimball, B. A. (1994). The FACE program. Agr Forest Meteorol, 70: 3-14. 040397 

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. 042670 

Hillstrom, M. L.; Lindroth, R. L. (2008). Elevated atmospheric carbon dioxide and ozone alter forest insect abundance and 

community composition. Insect Conservation and Diversity, 1: 233-241. http://dx.doi.org/10.1111/j.1752­

4598.2008.00031.x 191367 

Hillstrom, M. L.; Vigue, L. M.; Coyle, D. R.; Raffa, K. F.; Lindroth, R. L. (2010). Performance of the invasive weevil 

Polydrusus sericeus is influenced by atmospheric CO2 and host species. Agricultural and Forest Entomology, 12: 

285-292. http://dx.doi.org/10.1111/j.1461-9563.2010.00474.x 644679 

Himanen, S. J.; Nerg, A. M.; Nissinen, A.; Pinto, D. M.; Stewart, C. N.; Poppy, G. M.; Holopainen, J. K. (2009). Effects of 

elevated carbon dioxide and ozone on volatile terpenoid emissions and multitrophic communication of transgenic 

insecticidal oilseed rape (Brassica napus). New Phytol, 181: 174-186. http://dx.doi.org/10.1111/j.1469­

8137.2008.02646.x 191338 

Himanen, S. J.; Nerg, A. M.; Nissinen, A.; Stewart, C. N.; Poppy, G. M.; Holopainen, J. K. (2009). Elevated atmospheric 

ozone increases concentration of insecticidal Bacillus thuringiensis (Bt) Cry1Ac protein in Bt Brassica napus and 

reduces feeding of a Bt target herbivore on the non-transgenic parent. Environ Pollut, 157: 181-185. 

http://dx.doi.org/10.1016/j.envpol.2008.07.006 191369 

Hogg, A.; Uddling, J.; Ellsworth, D.; Carroll, M. A.; Pressley, S.; Lamb, B.; Vogel, C. (2007). Stomatal and non-stomatal 

fluxes of ozone to a northern mixed hardwood forest. Tellus B Chem Phys Meteorol, 59: 514-525. 

http://dx.doi.org/10.1111/j.1600-0889.2007.00269.x 199349 

Hogsett, W. E.; Olszyk, D.; Ormrod, D. P.; Taylor, G. E. Jr; Tingey, D. T. (1987). Air pollution exposure systems and 

experimental protocols: volume I: A review and evaluation of performance. Washington, DC: U.S. Environmental 

Protection Agency. 043780 

Hogsett, W. E.; Olszyk, D.; Ormrod, D. P.; Taylor, G. E. Jr; Tingey, D. T. (1987). Air pollution exposure systems and 

experimental protocols: volume II: Description of facilities. Washington, DC: U.S. Environmental Protection 

Agency. 043465 

Hogsett, W. E.; Tingey, D. T.; Holman, S. R. (1985). A programmable exposure control system for determination of the 

effects of pollutant exposure regimes on plant growth. Atmos Environ, 19: 1135-1145. 039383 

Hogsett, W. E.; Tingey, D. T.; Lee, E. H. (1988). Ozone exposure indices: Concepts for development and evaluation of their 

use. In Assessment of crop loss from air pollutants: Proceedings of an international conference (pp. 107-138). New 

York: Elsevier Applied Science. 042128 

Hogsett, W. E.; Tingey, D. T.; Lee, E. H.; Beedlow, P. A.; Andersen, C. P. (2008). An approach for evaluating the 

effectiveness of various ozone Air Quality Standards for protecting trees. Environ Manage, 41: 937-948. 

http://dx.doi.org/10.1007/s00267-007-9057-3 191229 

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. J Environ Manage, 21: 105-120. 040402 

Holmes, W. E.; Zak, D. R.; Pregitzer, K. S.; King, J. S. (2006). Elevated CO2 and O-3 alter soil nitrogen transformations 

beneath trembling aspen, paper birch, and sugar maple. Ecosystems, 9: 1354-1363. 

http://dx.doi.org/10.1007/s10021-006-0163-5 191372 

Hong, B.; Weinstein, D.; Swaney, D. (2006). Assessment of ozone effects on nitrate export from Hubbard Brook Watershed 

6. Environ Pollut, 141: 8-21. http://dx.doi.org/10.1016/j.envpol.2005.08.030 186989 

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. 035288 

March 2011 9-149 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

ICP M&M (2004). Mapping critical levels for vegetation. In Manual on Methodologies and Criteria for Modelling and 

Mapping Critical Loads and Levels; and Air Pollution Effects, Risks and Trends International Cooperative 

Programme on Modelling and Mapping. Retrieved January 28, 2011, from: 

http://www.rivm.nl/en/themasites/icpmm/manual-and-downloads/manual-english/index.html. 677471 

Innes, J. L.; Skelly, J. M.; Schaub, M. (2001). Ozone and broadleaved species: A guide to the identification of ozoneinduced foliar injury. Bern; Stuttgart; Wien: Haupt. 048954 

Iriti, M.; Di Maro, A.; Bernasconi, S.; Burlini, N.; Simonetti, P.; Picchi, V.; Panigada, C.; Gerosa, G.; Parente, A.; Faoro, F. 

(2009). Nutritional traits of bean (Phaseolus vulgaris) seeds from plants chronically exposed to ozone pollution. J 

Agric Food Chem, 57: 201-208. http://dx.doi.org/10.1021/jf802819m 195635 

Iriti, M.; Faoro, F. (2009). Chemical diversity and defence metabolism: How plants cope with pathogens and ozone 

pollution. International Journal of Molecular Sciences, 10: 3371-3399. http://dx.doi.org/10.3390/ijms10083371 

199313 

IRRI (2002). Annual Report. Los Baños, Laguna in the Philippines: International Rice Research Institute. http://irri.org/. 

688833 

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. 044174 

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. 036345 

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. 026877 

Janzik, I.; Preiskowski, S.; Kneifel, H. (2005). Ozone has dramatic effects on the regulation of the prechorismate pathway 

in tobacco (Nicotiana tabacum L. cv. Bel W3). Planta, 223: 20-27. http://dx.doi.org/10.1007/s00425-005-0060-8 

191581 

Jaoude, M. B.; Katerji, N.; Mastrorilli, M.; Rana, G. (2008). Analysis of the ozone effect on soybean in the Mediterranean 

region II. The consequences on growth, yield and water use efficiency. Eur J Agron, 28: 519-525. 

http://dx.doi.org/10.1016/j.eja.2007.09.001 191223 

Jaude, M. B.; Katerji, N.; Mastrorilli, M.; Rana, G. (2008). Analysis of the effect of ozone on soybean in the Mediterranean 

region I: the consequences on crop-water status. Eur J Agron, 28: 508-518. 

http://dx.doi.org/10.1016/j.eja.2007.09.002 191222 

Johnson, R. M.; Pregitzer, K. S. (2007). Concentration of sugars, phenolic acids, and amino acids in forest soils exposed to 

elevated atmospheric CO2 and O3. Soil Biol Biochem, 39: 3159-3166. 

http://dx.doi.org/10.1016/j.soilbio.2007.07.010 191287 

Jones, M. E.; Paine, T. D. (2006). Detecting changes in insect herbivore communities along a pollution gradient. Environ 

Pollut, 143: 377-387. http://dx.doi.org/10.1016/j.envpol.2005.12.013 191301 

Jones, M. L. M.; Hodges, G.; Mills, G. (2010). Nitrogen mediates above-ground effects of ozone but not below-ground 

effects in a rhizomatous sedge. Environ Pollut, 158: 559-565. http://dx.doi.org/10.1016/j.envpol.2009.08.002 

567354 

Jones, T. G.; Freeman, C.; Lloyd, A.; Mills, G. (2009). Impacts of elevated atmospheric ozone on peatland below-ground 

doc characteristics. Ecol Eng, 35: 971-977. http://dx.doi.org/10.1016/j.ecoleng.2008.08.009 199881 

Joo, J. H.; Wang, S. Y.; Chen, J. G.; Jones, A. M.; Fedoroff, N. V. (2005). Different signaling and cell death roles of 

heterotrimeric G protein alpha and beta subunits in the arabidopsis oxidative stress response to ozone. Plant Cell, 

17: 957-970. http://dx.doi.org/10.1105/tpc.104.029603 191307 

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. 040408 

Kanerva, T.; Palojarvi, A.; Ramo, K.; Manninen, S. (2008). Changes in soil microbial community structure under elevated 

tropospheric O3 and CO2. Soil Biol Biochem, 40: 2502-2510. http://dx.doi.org/10.1016/j.soilbio.2008.06.007 

191264 

March 2011 9-150 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

Kanerva, T.; Palojarvi, A.; Ramo, K.; Ojanpera, K.; Esala, M.; Manninen, S. (2006). A 3-year exposure to CO2 and O3 

induced minor changes in soil N cycling in a meadow ecosystem. Plant Soil, 286: 61-73. 

http://dx.doi.org/10.1007/s11104-006-9026-2 191747 

Kanerva, T.; Regina, K.; Ramo, K.; Ojanpera, K.; Manninen, S. (2007). Fluxes of N2O, CH4 and CO2 in a meadow 

ecosystem exposed to elevated ozone and carbon dioxide for three years. Environ Pollut, 145: 818-828. 

http://dx.doi.org/10.1016/j.envpol.2006.03.055 191405 

Kangasjarvi, J.; Jaspers, P.; Kollist, H. (2005). Signalling and cell death in ozone-exposed plants. Plant Cell Environ, 28: 

1021-1036. 180341 

Karlsson, P. E.; Sellden, G.; Plaijel, H. (2003). Establishing ozone critical levels II UNECE workshop report. Gothenburg, 

Sweden: IVL Swedish Environmental Institute. 055331 

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. 036347 

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. 

035307 

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. http://dx.doi.org/10.1111/j.1365-3040.2005.01362.x 095556 

Kasurinen, A.; Keinanen, M. M.; Kaipainen, S.; Nilsson, L. O.; Vapaavuori, E.; Kontro, M. H.; Holopainen, T. (2005). 

Below-ground responses of silver birch trees exposed to elevated CO2 and O-3 levels during three growing 

seasons. Global Change Biol, 11: 1167-1179. http://dx.doi.org/10.1111/j.1365-2486.2005.00970.x 191245 

Kasurinen, A.; Peltonen, P. A.; Julkunen-Tiitto, R.; Vapaavuori, E.; Nuutinen, V.; Holopainen, T.; Holopainen, J. K. (2007). 

Effects of elevated co2 and o-3 on leaf litter phenolics and subsequent performance of litter-feeding soil 

macrofauna. Plant Soil, 292: 25-43. http://dx.doi.org/10.1007/s11104-007-9199-3 199826 

Kasurinen, A.; Riikonen, J.; Oksanen, E.; Vapaavuori, E.; Holopainen, T. (2006). Chemical composition and decomposition 

of silver birch leaf litter produced under elevated CO2 and O-3. Plant Soil, 282: 261-280. 

http://dx.doi.org/10.1007/s11104-005-6026-6 191269 

Kats, G.; Olszyk, D. M.; Thompson, C. R. (1985). Open top experimental chambers for trees. J Air Waste Manag Assoc, 35: 

1298-1301. 055511 

Kats, G.; Thompson, C. R.; Kuby, W. C. (1976). Improved ventilation of open top greenhouses. J Air Waste Manag Assoc, 

26: 1089-1090. 039799 

Keutgen, A. J.; Noga, G.; Pawelzik, E. (2005). Cultivar-specific impairment of strawberry growth, photosynthesis, 

carbohydrate and nitrogen accumulation by ozone. Environ Exp Bot, 53: 271-280. 

http://dx.doi.org/10.1016/j.envexpbot.2004.04.003 191295 

Keutgen, N.; Keutgen, A. J.; Janssens, M. J. J. (2008). Sweet potato [Ipomoea batatas (L.) Lam.] cultivated as tuber or 

leafy vegetable supplier as affected by elevated tropospheric ozone. J Agric Food Chem, 56: 6686-6690. 

http://dx.doi.org/10.1021/jf8006272 191690 

King, J. S.; Kubiske, M. E.; Pregitzer, K. S.; Hendrey, G. R.; McDonald, E. P.; Giardina, C. P.; Quinn, V. S.; Karnosky, D. 

F. (2005). Tropospheric O-3 compromises net primary production in young stands of trembling aspen, paper birch 

and sugar maple in response to elevated atmospheric CO2. New Phytol, 168: 623-635. 

http://dx.doi.org/10.1111/j.1469-8137.2005.01557.x 191701 

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. 041751 

Kitao, M.; Low, M.; Heerdt, C.; Grams, T. E. E.; Haberle, K. H.; Matyssek, R. (2009). Effects of chronic elevated ozone 

exposure on gas exchange responses of adult beech trees (Fagus sylvatica) as related to the within-canopy light 

gradient. Environ Pollut, 157: 537-544. http://dx.doi.org/10.1016/j.envpol.2008.09.016 191331 

March 2011 9-151 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

Kline, L. J.; Davis, D. D.; Skelly, J. M.; Decoteau, D. R. (2009). Variation in ozone sensitivity within Indian hemp and 

common milkweed selections from the Midwest. Northeast Nat, 16: 307-313. 

http://dx.doi.org/10.1656/045.016.0210 196918 

Kline, L. J.; Davis, D. D.; Skelly, J. M.; Savage, J. E.; Ferdinand, J. (2008). Ozone sensitivity of 28 plant selections 

exposed to ozone under controlled conditions. Northeast Nat, 15: 57-66. http://dx.doi.org/10.1656/1092­

6194(2008)15[57:OSOPSE]2.0.CO;2 191591 

Kohut, R. (2007). Assessing the risk of foliar injury from ozone on vegetation in parks in the US National Park Service's 

Vital Signs Network. Environ Pollut, 149: 348-357. http://dx.doi.org/10.1016/j.envpol.2007.04.022 093289 

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. 052597 

Kollist, T.; Moldau, H.; Rasulov, B.; Oja, V.; Ramma, H.; Huve, K.; Jaspers, P.; Kangasjarvi, J.; Kollist, H. (2007). A novel 

device detects a rapid ozone-induced transient stomatal closure in intact Arabidopsis and its absence in abi2 mutant. 

Physiol Plant, 129: 796-803. http://dx.doi.org/10.1111/j.1399-3054.2006.00851.x 191539 

Kostka-Rick, R.; Hahn, H. U. (2005). Biomonitoring using tobacco Bel W3 provides supplemental information for risk 

assessment of vegetation injury due to ozone. Gefahrstoffe Reinhaltung Der Luft, 65: 485-491. 199428 

Kozovits, A. R.; Matyssek, R.; Blaschke, H.; Gottlein, A.; Grams, T. E. E. (2005). Competition increasingly dominates the 

responsiveness of juvenile beech and spruce to elevated CO2 and/or O-3 concentrations throughout two subsequent 

growing seasons. Global Change Biol, 11: 1387-1401. http://dx.doi.org/10.1111/j.1365-2486.2005.00993.x 191282 

Krupa, S. V.; Grunhage, L.; Jager, 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. 038895 

Krupa, S. V.; Nosal, M.; Peterson, D. L. (2001). Use of passive ozone O3 samplers in vegetation effects assessment. 

Environ Pollut, 112: 303-309. 040430 

Kubiske, M. E.; Quinn, V. S.; Heilman, W. E.; McDonald, E. P.; Marquardt, P. E.; Teclaw, R. M.; Friend, A. L.; Karnoskey, 

D. F. (2006). Interannual climatic variation mediates elevated CO2 and O3 effects on forest growth. Global Change 

Biol, 12: 1054-1068. http://dx.doi.org/10.1111/j.1365-2486.2006.01152.x 093284 

Kubiske, M. E.; Quinn, V. S.; Marquardt, P. E.; Karnosky, D. F. (2007). Effects of elevated atmospheric CO2 and/or O-3 on 

intra- and interspecific competitive ability of aspen. Plant Biol (Stuttg), 9: 342-355. http://dx.doi.org/10.1055/s2006-924760 191336 

Laffray, X.; Rose, C.; Garrec, J. P. (2007). Estimation of ozone concentration in a valley of the alps mountains based on 

bel-w3 tobacco leaf injury. Water Air Soil Pollut, 186: 29-42. http://dx.doi.org/10.1007/s11270-007-9460-7 191384 

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. J Plant 

Physiol, 95: 882-889. http://www.plantphysiol.org/cgi/content/abstract/95/3/882. 043506 

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. 053015 

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 Leiden, The Netherlands: Backhuys Publishers. 

029015 

Leakey, A. D. B.; Bernacchi, C. J.; Ort, D. R.; Long, S. P. (2006). Long-term growth of soybean at elevated CO2 does not 

cause acclimation of stomatal conductance under fully open-air conditions. Plant Cell Environ, 29: 1794-1800. 

http://dx.doi.org/10.1111/j.1365-3040.2006.01556.x 679714 

Lee, E. H.; Hogsett, W. E. (1996). Methodology for calculating inputs for ozone secondary standard benefits anaylsis: Part 

II. Research Triangle Park, NC: U.S. Environmental Protection Agency. 670278 

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 Manag Assoc, 49: 669-681. 040451 

Lee, E. H.; Hogsett, W. E.; Tingey, D. T. (1994). Attainment and effects issues regarding alternative secondary ozone air 

quality standards. J Environ Qual, 23: 1129-1140. http://jeq.scijournals.org/cgi/content/abstract/23/6/1129. 043268 

March 2011 9-152 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

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. 042135 

Lee, E. H.; Tingey, D. T.; Hogsett, W. E. (1988). Evaluation of ozone exposure indices in exposure-response modeling. 

Environ Pollut, 53: 43-62. 042136 

Lee, E. H.; Tingey, D. T.; Hogsett, W. E. (1988). Evaluation of ozone-exposure indices for relating exposure to plant 

production and for estimating agricultural losses (Report No. EPA/600/3-88/039). Washington, DC: U.S. 

Environmental Protection Agency. 594572 

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 (Report No. EPA/600/3-89/047). Corvallis, 

OR: U.S. Environmental Protection Agency. 042137 

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. 053031 

Lee, E. H.; Tingey, D. T.; Waschmann, R. S.; Phillips, D. L.; Olszyk, D. M.; Johnson, M. G.; Hogsett, W. E. (2009). 

Seasonal and long-term effects of CO2 and O-3 on water loss in ponderosa pine and their interaction with climate 

and soil moisture. Tree Physiol, 29: 1381-1393. http://dx.doi.org/10.1093/treephys/tpp071 595904 

Lee, S.; Yun, S. C. (2006). The ozone stress transcriptome of pepper (Capsicum annuum L.). Molecules and Cells, 21: 197­

205. 191592 

Lefohn, A. S.; Benedict, H. M. (1982). Development of a mathematical index that describes ozone concentration, 

frequency and duration. Atmos Environ, 16: 2529-2532. 039395 

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. 082871 

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. 

http://dx.doi.org/10.1016/0004-6981(88)90353-8 042138 

Lefohn, A. S.; Shadwick, D. S. (2000). Differences in trending estimates in the United States using several ozone metrics. 

In Proceedings of the 93rd Air & Waste Management Association Annual Conference and Exhibition (pp. AS 1d645). Pittsburgh, PA: Air & Waste Management Association. 040483 

Legge, A. H.; Grunhage, L.; Nosal, M.; Jager, 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. J Appl Bot 

Food Qual, 69: 192-205. 040689 

Leitao, L.; Bethenod, O.; Biolley, J. P. (2007). The impact of ozone on juvenile maize (Zea mays L.) plant photosynthesis: 

Effects on vegetative biomass, pigmentation, and carboxylases (PEPc and Rubisco). Plant Biol (Stuttg), 9: 478-488. 

http://dx.doi.org/10.1055/s-2007-964942 191456 

Leitao, L.; Delacote, E.; Dizengremel, P.; Le Thiec, D.; Biolley, J. P. (2007). Assessment of the impact of increasing 

concentrations of ozone on photosynthetic components of maize (Zea mays L.), a C-4 plant. Environ Pollut, 146: 5­

8. http://dx.doi.org/10.1016/j.envpol.2006.05.019 199379 

Leitao, L.; Maoret, J. J.; Biolley, J. P. (2007). Changes in PEP carboxylase, rubisco and rubisco activase mRNA levels from 

maize (Zea mays) exposed to a chronic ozone stress. Biol Res, 40: 137-153. http://dx.doi.org/10.4067/S0716­

97602007000200005 191263 

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. 043015 

Levine, J. S.; Pinto, J. P. (1998). The production of CO by biomass burning. In MAK Khalil; JP Pinto; MJ Shearer (Eds.), 

Atmospheric carbon monoxide and its environmental effects: Proceedings of the international conference; 

December 1997; Portland, Oregon (pp. 251-256). Research Triangle Park, NC: U.S. Environmental Protection 

Agency, Office of Research and Development. 029599 

Lewis, J. S.; Ditchkoff, S. S.; Lin, J. C.; Muntifering, R. B.; Chappelka, A. H. (2006). Nutritive quality of big bluestem 

(Andropogon gerardii) and eastern gamagrass (Tripsacum dactyloides) exposed to tropospheric ozone. R E M, 59: 

267-274. 191542 

March 2011 9-153 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

Li, P. H.; Mane, S. P.; Sioson, A. A.; Robinet, C. V.; Heath, L. S.; Bohnert, H. J.; Grene, R. (2006). Effects of chronic ozone 

exposure on gene expression in Arabidopsis thaliana ecotypes and in Thellungielia halophila. Plant Cell Environ, 

29: 854-868. http://dx.doi.org/10.1111/j.1365-3040.2005.01465.x 191332 

Lin, J. C.; Nosal, M.; Muntifering, R. B.; Krupa, S. V. (2007). Alfalfa nutritive quality for ruminant livestock as influenced 

by ambient air quality in west-central Alberta. Environ Pollut, 149: 99-103. 

http://dx.doi.org/10.1016/j.envpol.2006.12.009 196925 

Lindroth, R. L. (2010). Impacts of elevated atmospheric CO2 and O3 on forests: Phytochemistry, trophic interactions, and 

ecosystem dynamics. J Chem Ecol, 36: 21-Feb. http://dx.doi.org/10.1007/s10886-009-9731-4 596479 

Liu, L. L.; King, J. S.; Giardina, C. P.; Booker, F. L. (2009). The influence of chemistry, production and community 

composition on leaf litter decomposition under elevated atmospheric CO2 and tropospheric O-3 in a northern 

hardwood ecosystem. Ecosystems, 12: 401-416. http://dx.doi.org/10.1007/s10021-009-9231-y 191470 

Liu, L.; King, J. S.; Giardina, C. P. (2007). Effects of elevated atmospheric CO2 and tropospheric O3 on nutrient dynamics: 

Decomposition of leaf litter in termbling aspen and paper birch communities. Plant Soil, 299: 65-82. 093286 

Liu, L.; King, J.; Giardina, C. (2005). Effects of elevated concentrations of atmospheric CO2 and tropospheric O3 on leaf 

litter production and chemistry in trembling aspen and paper birch communities. Tree Physiol, 25: 1511-1522. 

187005 

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. 029709 

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. 029710 

Long, S. P.; Naidu, S. L. (2002). Effects of oxidants at the biochemical, cell and physiological levels, with particular 

reference to ozone. In Bell, J. N. B.; Treshow, M. (Eds.), Air Pollution and Plant Life West Sussex, United 

Kingdom: John Wiley & Sons, Ltd. 038179 

Loranger, G. I.; Pregitzer, K. S.; King, J. S. (2004). Elevated CO2 and O3t concentrations differentially affect selected 

groups of the fauna in temperate forest soils. Soil Biol Biochem, 36: 1521-1524. 072712 

Lorenzini, G.; Nali, C. (1995). Analysis of vertical ozone and nitrogen oxides profiles in a Prunus cerasifera canopy. Int J 

Biometeorol, 39: 1-4. 040691 

Loreto, F.; Fares, S. (2007). Is ozone flux inside leaves only a damage indicator? Clues from volatile isoprenoid studies. 

Plant Physiol, 143: 1096-1100. http://dx.doi.org/10.1104/pp.106.091892 180259 

Loreto, F.; Velikova, V. (2001). Isoprene produced by leaves protects the photosynthetic apparatus against ozone damage, 

quenches ozone products, and reduces lipid peroxidation of cellular membranes. Plant Physiol, 127: 1781-1787. 

http://www.ncbi.nlm.nih.gov/pubmed/11743121. 657212 

Low, M.; Herbinger, K.; Nunn, A. J.; Haberle, K. H.; Leuchner, M.; Heerdt, C.; Werner, H.; Wipfler, P.; Pretzsch, H.; Tausz, 

M.; Matyssek, R. (2006). Extraordinary drought of 2003 overrules ozone impact on adult beech trees (Fagus 

sylvatica). Trees Struct Funct, 20: 539-548. http://dx.doi.org/10.1007/s00468-006-0069-z 191396 

Loya, W. M.; Pregitzer, K. S.; Karberg, N. J.; King, J. S.; Giardina, C. P. (2003). Reduction of soil carbon formation by 

tropospheric ozone under elevated carbon dioxide. Nature, 425: 705-707. 074380 

Ludwikow, A.; Gallois, P.; Sadowski, J. (2004). Ozone-induced oxidative stress response in Arabidopsis: Transcription 

profiling by microarray approach. Cell Mol Biol Lett, 9: 829-842. 595939 

Ludwikow, A.; Kierzek, D.; Gallois, P.; Zeef, L.; Sadowski, J. (2009). Gene expression profiling of ozone-treated 

Arabidopsis abi1td insertional mutant: Protein phosphatase 2C ABI1 modulates biosynthesis ratio of ABA and 

ethylene. Planta, 230: 1003-1017. http://dx.doi.org/10.1007/s00425-009-1001-8 199440 

Ludwikow, A.; Sadowski, J. (2008). Gene networks in plant ozone stress response and tolerance. J Integr Plant Biol, 50: 

1256-1267. http://dx.doi.org/10.1111/j.1744-7909.2008.00738.x 191426 

Luo, Y. (2001). Transient ecosystem response to free-air CO2 enrichment (FACE): experimental evidence and methods of 

analysis. New Phytol, 152: 3-8. 035318 

Luo, Y.; Reynolds, J. F. (1999). Validity of extrapolating field CO2 experiments to predict carbon sequestration in natural 

ecosystems. Ecology, 80: 1568-1583. 035319 

March 2011 9-154 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

Lyons, T. M.; Barnes, J. D. (1998). Influence of plant age on ozone resistance in Plantago major. New Phytol, 138: 83-89. 

036354 

Maggio, A.; Chiaranda, F. Q.; Cefariello, R.; Fagnano, M. (2009). Responses to ozone pollution of alfalfa exposed to 

increasing salinity levels. Environ Pollut, 157: 1445-1452. http://dx.doi.org/10.1016/j.envpol.2008.09.013 191645 

Mahalingam, R.; Jambunathan, N.; Gunjan, S. K.; Faustin, E.; Weng, H.; Ayoubi, P. (2006). Analysis of oxidative 

signalling induced by ozone in Arabidopsis thaliana. Plant Cell Environ, 29: 1357-1371. 

http://dx.doi.org/10.1111/j.1365-3040.2006.01516.x 191221 

Mahalingam, R.; Shah, N.; Scrymgeour, A.; Fedoroff, N. (2005). Temporal evolution of the Arabidopsis oxidative stress 

response. Plant Mol Biol, 57: 709-730. http://dx.doi.org/10.1007/s11103-005-2860-4 191693 

Maier-Maercker, U. (1998). Predisposition of trees to drought stress by ozone. Tree Physiol, 19: 71-78. 029961 

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. 043987 

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. 039967 

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. 053773 

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. 044155 

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. 043678 

Massman, W. J. (2004). Toward an ozone standard to protect vegetation based on effective dose: A review of deposition 

resistances and a possible metric. Atmos Environ, 38: 2323-2337. 055350 

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. 040698 

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. 011616 

Matyssek, R.; Gunthardt-Goerg, M. S.; Maurer, S.; Keller, T. (1995). Nighttime exposure to ozone reduces whole-plant 

production in Betula pendula. Tree Physiol, 15: 159-165. 040700 

Matyssek, R.; Le Thiec, D.; Low, M.; Dizengremel, P.; Nunn, A. J.; Haberle, K. H. (2006). Interactions between drought 

and O-3 stress in forest trees. Plant Biol (Stuttg), 8: 11-17. http://dx.doi.org/10.1055/s-2005-873025 191481 

Matyssek, R.; Sandermann, H.; Wieser, G.; Booker, F.; Cieslik, S.; Musselman, R.; Ernst, D. (2008). The challenge of 

making ozone risk assessment for forest trees more mechanistic. Environ Pollut, 156: 567-582. 

http://dx.doi.org/10.1016/j.envpol.2008.04.017 191262 

Matyssek, R.; Wieser, G.; Ceulemans, R.; Rennenberg, H.; Pretzsch, H.; Haberer, K.; Low, M.; Nunn, A. J.; Werner, H.; 

Wipfler, P.; Obwald, W.; Nikolova, P.; Hanke, D. E.; Kraigher, H.; Tausz, M.; Bahnweg, G.; Kitao, M.; Dieler, J.; 

Sandermann, H.; Herbinger, K.; Grebenc, T.; Blumenrother, M.; Deckmyn, G.; Grams, T. E. E.; Heerdt, C.; 

Leuchner, M.; Fabian, P.; Haberle, K. H. (2010). Enhanced ozone strongly reduces carbon sink strength of adult 

beech (Fagus sylvatica)- Resume from the free-air fumigation study at Kranzberg Forest. Environ Pollut, 158: 

2527-2532. http://dx.doi.org/10.1016/j.envpol.2010.05.009 628553 

Mautz, W. J.; Dohm, M. R. (2004). Respiratory and behavioral effects of ozone on a lizard and a frog. Comp Biochem 

Physiol A Mol Integr Physiol, 139: 371-377. 644188 

Maňkovská, B.; Percy, K. E.; Karnosky, D. F. (2005). Impacts of greenhouse gases on epicuticular waxes of Populus 

tremuloides Michx.: Results from an open-air exposure and a natural O3 gradient. Environ Pollut, 137: 580-586. 

http://dx.doi.org/10.1016/j.envpol.2005.01.043 672965 

March 2011 9-155 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

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 PR Miller; JR McBride (Eds.), Oxidant air pollution impacts in the Montane 

forests of southern California: a case study of the San Bernardino Mountains (pp. 338-352). New York, NY: 

Springer-Verlag New York, Inc. 053050 

McCarthy, H. R.; Oren, R.; Johnsen, K. H.; Gallet-Budynek, A.; Pritchard, S. G.; Cook, C. W.; LaDeau, S. L.; Jackson, R. 

B.; Finzi, A. C. (2010). Re-assessment of plant carbon dynamics at the Duke free-air CO2 enrichment site: 

Interactions of atmospheric CO2 with nitrogen and water availability over stand development. New Phytol, 185: 

514-528. http://dx.doi.org/10.1111/j.1469-8137.2009.03078.x 679715 

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. 041877 

McFrederick, Q. S.; Fuentes, J. D.; Roulston, T.; Kathilankal, J. C.; Lerdau, M. (2009). Effects of air pollution on biogenic 

volatiles and ecological interactions. Oecologia, 160: 411-420. http://dx.doi.org/10.1007/s00442-009-1318-9 

191329 

McFrederick, Q. S.; Kathilankal, J. C.; Fuentes, J. D. (2008). Air pollution modifies floral scent trails. Atmos Environ, 42: 

2336-2348. http://dx.doi.org/10.1016/j.atmosenv.2007.12.033 196931 

McLaughlin, S. B.; Nosal, M.; Wullschleger, S. D.; Sun, G. (2007). Interactive effects of ozone and climate on tree growth 

and water use in a southern Appalachian forest in the USA. New Phytol, 174: 109-124. 090348 

McLaughlin, S. B.; Wullschleger, S. D.; Sun, G.; Nosal, M. (2007). Interactive effects of ozone and climate on water use, 

soil moisture content and streamflow in a southern Appalachian forest in the USA. New Phytol, 174: 125-136. 

090347 

McLeod, A. R.; Long, S. P. (1999). Free-air carbon dioxide enrichment (FACE) in global change research: A review. Adv 

Ecol Res, 28: 1-56. http://dx.doi.org/10.1016/S0065-2504(08)60028-8 688834 

Medlyn, B. E.; Barton, C. V. M.; Broadmeadow, M. S. J.; Ceulemans, R.; De Angelis, P.; Forstreuter, M.; Freeman, M.; 

Jackson, S. B.; Kellomaki, S.; Laitat, E.; Rey, A.; Roberntz, P.; Sigurdsson, B. D.; Strassemeyer, J.; Wang, K.; 

Curtis, P. S.; Jarvis, P. G. (2001). Stomatal conductance of forest species after long-term exposure to elevated CO2 

concentration: A synthesis. New Phytol, 149: 247-264. 679716 

Meehan, T. D.; Crossley, M. S.; Lindroth, R. L. (2010). Impacts of elevated CO2 and O3 on aspen leaf litter chemistry and 

earthworm and springtail productivity. Soil Biol Biochem, 42: 1132-1137. 

http://dx.doi.org/10.1016/j.soilbio.2010.03.019 644184 

Menendez, A. I.; Romero, A. M.; Folcia, A. M.; Martinez-Ghersa, M. A. (2009). Getting the interactions right: Will higher 

O-3 levels interfere with induced defenses to aphid feeding? Basic Appl Ecol, 10: 255-264. 

http://dx.doi.org/10.1016/j.baae.2008.03.010 191430 

Menendez, A. I.; Romero, A. M.; Folcia, A. M.; Martinez-Ghersa, M. A. (2010). Aphid and episodic O sub(3) injury in 

arugula plants (Eruca sativa Mill) grown in open-top field chambers. Agric Ecosyst Environ, 135: 10-14. 

http://dx.doi.org/10.1016/j.agee.2009.08.005 384046 

Mereu, S.; Gerosa, G.; Finco, A.; Fusaro, L.; Muys, B.; Manes, F. (2009). Improved sapflow methodology reveals 

considerable night-time ozone uptake by Mediterranean species. Biogeosciences, 6: 3151-3162. 102168 

Miles, G. P.; Samuel, M. A.; Zhang, Y. L.; Ellis, B. E. (2005). RNA interference-based (RNAi) suppression of AtMPK6, an 

Arabidopsis mitogen-activated protein kinase, results in hypersensitivity to ozone and misregulation of AtMPK3. 

Environ Pollut, 138: 230-237. http://dx.doi.org/10.1016/j.envpol.2005.04.017 191648 

Miller, P. L. (1973). Oxidant-induced community change in a mixed conifer forest. In JA Naegele (Ed.), Air pollution 

damage to vegetation (pp. 101-117). Washington, DC: American Chemical Society. 039165 

Miller, P. R.; Elderman, M. J. (1977). Photochemical oxidant air pollutant effects on a mixed conifer forest ecosystem: A 

progress report, 1976. Corvallis, Oregon: U.S. Environmental Protection Agency. 038488 

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. 040702 

Mills, G. (2002). Modification of plant response by environmental conditions. In Bell, J. N. B.; Treshow, M. (Eds.), Air 

pollution and plant life Chichester, United Kingdom: John Wiley & Sons Ltd. 035322 

March 2011 9-156 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

Mills, G.; Ball, G.; Hayes, F.; Fuhrer, J.; Skarby, L.; Gimeno, B.; De Temmerman, L. (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. 

030098 

Mills, G.; Buse, A.; Gimeno, B.; Bermejo, V.; Holland, M.; Emberson, L.; Pleijel, H. (2007). A synthesis of AOT40-based 

response functions and critical levels of ozone for agricultural and horticultural crops. Atmos Environ, 41: 2630­

2643. http://dx.doi.org/10.1016/j.atmosenv.2006.11.016 180221 

Mills, G.; Hayes, F.; Jones, M. L. M.; Cinderby, S. (2007). Identifying ozone-sensitive communities of (semi-)natural 

vegetation suitable for mapping exceedance of critical levels. Environ Pollut, 146: 736-743. 

http://dx.doi.org/10.1016/j.envpol.2006.04.005 196934 

Mills, G.; Hayes, F.; Wilkinson, S.; Davies, W. J. (2009). Chronic exposure to increasing background ozone impairs 

stomatal functioning in grassland species. Global Change Biol, 15: 1522-1533. http://dx.doi.org/10.1111/j.1365­

2486.2008.01798.x 191272 

Mondor, E. B.; Awmack, C. S.; Lindroth, R. L. (2010). Individual growth rates do not predict aphid population densities 

under altered atmospheric conditions. Agricultural and Forest Entomology, 12: 293-299. 

http://dx.doi.org/10.1111/j.1461-9563.2010.00478.x 644271 

Mondor, E. B.; Tremblay, M. N.; Awmack, C. S.; Lindroth, R. L. (2004). Divergent pheromone-mediated insect behaviour 

under global atmospheric change. Global Change Biol, 10: 1820-1824. 074334 

Mondor, E. B.; Tremblay, M. N.; Awmack, C. S.; Lindroth, R. L. (2005). Altered genotypic and phenotypic frequencies of 

aphid populations under enriched CO2 and O-3 atmospheres. Global Change Biol, 11: 1990-1996. 

http://dx.doi.org/10.1111/j.1365-2486.2005.01054.x 191217 

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. 055527 

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. J Plant Physiol, 135: 2348-2357. 

072764 

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, 170: 333-343. 079186 

Morison, J. I. L.; Lawlor, D. W. (1999). Interactions between increasing CO2 concentration and temperature on plant 

growth. Plant Cell Environ, 22: 659-682. http://dx.doi.org/10.1046/j.1365-3040.1999.00443.x 094194 

Morsky, S. K.; Haapala, J. K.; Rinnan, R.; Tiiva, P.; Saarnio, S.; Silvola, J.; Holopainen, T.; Martikainen, P. J. (2008). 

Long-term ozone effects on vegetation, microbial community and methane dynamics of boreal peatland 

microcosms in open-field conditions. Global Change Biol, 14: 1891-1903. http://dx.doi.org/10.1111/j.1365­

2486.2008.01615.x 191507 

Mudd, J. B. (1996). Biochemical basis for the toxicity of ozone. In Yunus, M.; Iqbal, M. (Eds.), Plant response to air 

pollution (pp. 267-283). New York, NY: John Wiley & Sons. 052795 

Muntifering, R. B.; Chappelka, A. H.; Lin, J. C.; Karnosky, D. F.; Somers, G. L. (2006). Chemical composition and 

digestibility of Trifolium exposed to elevated ozone and carbon dioxide in a free-air (FACE) fumigation system. 

Funct Ecol, 20: 269-275. http://dx.doi.org/10.1111/j.1365-2435.2006.01093.x 191270 

Musselman, R. C.; Lefohn, A. S.; Massman, W. J.; Heath, R. L. (2006). A critical review and analysis of the use of 

exposure- and flux-based ozone indices for predicting vegetation effects. Atmos Environ, 40: 1869-1888. 121678 

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. 040706 

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. 042144 

Musselman, R. C.; Minnick, T. J. (2000). Nocturnal stomatal conductance and ambient air quality standards for ozone. 

Atmos Environ, 34: 719-733. 011612 

March 2011 9-157 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

Nali, C.; Balducci, E.; Frati, L.; Paoli, L.; Loppi, S.; Lorenzini, G. (2007). Integrated biomonitoring of air quality with 

plants and lichens: A case study on ambient ozone from central Italy. Chemosphere, 67: 2169-2176. 

http://dx.doi.org/10.1016/j.chemosphere.2006.12.036 191478 

National ambient air quality standards for ozone (final rule). 73 Fed. Reg. 16436-16514 (2008). 684051 

National Ambient Air Quality Standards for Ozone (Proposed Rule). 75 Fed. Reg. 2938-3052 (2010). 684211 

National ambient air quality standards for ozone - Proposed rule. 72 Fed. Reg. 37818 (2007). 684055 

National primary and secondary ambient air quality standards, Section 109 of the Clean Air Act. 42 USC Â§ 7409. (1990). 

http://www.epa.gov/air/caa/title1.html#ia. 037658 

Neufeld, H. S.; Renfro, J. R.; Hacker, W. D.; Silsbee, D. (1992). Ozone in Great Smoky Mountains National Park: 

dynamics and effects on plants. Abstract from Tropospheric Ozone and the Environment II, Pittsburgh, PA. 038971 

Nikolova, P. S.; Andersen, C. P.; Blaschke, H.; Matyssek, R.; Häberle, K. H. (2010). Belowground effects of enhanced 

tropospheric ozone and drought in a beech/spruce forest (Fagus sylvatica L./Picea abies [L.] Karst). Environ Pollut, 

158: 1071-1078. http://dx.doi.org/10.1016/j.envpol.2009.07.036 626810 

Noctor, G.; Foyer, C. H. (1998). Ascorbate and glutathione: Keeping active oxygen under control. Annu Rev Plant Physiol 

Plant Mol Biol , 49: 249-279. 657213 

Norby, R. J.; DeLucia, E. H.; Gielen, B.; Calfapietra, C.; Giardina, C. P.; King, J. S.; Ledford, J.; McCarthy, H. R.; Moore, 

D. J. P.; Ceulemans, R.; De Angelis, P.; Finzi, A. C.; Karnosky, D. F.; Kubiske, M. E.; Lukac, M.; Pregitzer, K. S.; 

Scarascia-Mugnozza, G. E.; Schlesinger, W. H.; Oren, R. (2005). Forest response to elevated CO2 is conserved 

across a broad range of productivity. PNAS, 102: 18052-18056. http://dx.doi.org/10.1073/pnas.0509478102 679717 

Novak, K.; Cherubini, P.; Saurer, M.; Fuhrer, J.; Skelly, J. M.; Kräuchi, N.; Schaub, M. (2007). Ozone air pollution effects 

on tree-ring growth, 13C, visible foliar injury and leaf gas exchange in three ozone-sensitive woody plant species. 

Tree Physiol, 27: 941-949. 194630 

NPS (2006). Ozone bioindicators. Retrieved January 28, 2011 from 

http://www.nature.nps.gov/air/Pubs/bioindicators/index.cfm. 677536 

NPS (2007). Ozone effects studies. Retrieved January 28, 2011 from http://www.nature.nps.gov/air/studies/ecoOzone.cfm. 

677537 

Nussbaum, S.; Geissmann, M.; Fuhrer, J. (1995). Ozone exposure-response relationships for mixtures of perennial ryegrass 

and white clover depend on ozone exposure patterns. Atmos Environ, 29: 989-995. 030141 

O'Gara, P. J. (1922). Sulfur dioxide and fume problems and their solution. Journal of Industrial and Engineering Chemistry, 

14: 744-745. 015018 

O'Neill, B. F.; Zangerl, A. R.; Delucia, E. H.; Berenbaum, M. R. (2008). Longevity and fecundity of Japanese beetle 

(Popillia japonica) on foliage grown under elevated carbon dioxide. Environ Entomol, 37: 601-607. 

www.life.illinois.edu/delucia/PUBLICATIONS/O'Neill%20Env%20Ent.pdf. 195653 

Ogawa, D.; Nakajima, N.; Sano, T.; Tamaoki, M.; Aono, M.; Kubo, A.; Kanna, M.; Ioki, M.; Kamada, H.; Saji, H. (2005). 

Salicylic acid accumulation under O-3 exposure is regulated by ethylene in tobacco plants. Plant Cell Physiol, 46: 

1062-1072. http://dx.doi.org/10.1093/pcp/pci118 191653 

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. 019538 

Olbrich, M.; Betz, G.; Gerstner, E.; Langebartels, C.; Sandermann, H.; Ernst, D. (2005). Transcriptome analysis of ozoneresponsive genes in leaves of European beech (Fagus sylvatica L.). Plant Biol (Stuttg), 7: 670-676. 

http://dx.doi.org/10.1055/s-2005-873001 191697 

Olbrich, M.; Gerstner, E.; Bahnweg, G.; Haberle, K. H.; Matyssek, R.; Welzl, G.; Heller, W.; Ernst, D. (2010). 

Transcriptional signatures in leaves of adult European beech trees (Fagus sylvatica L.) in an experimentally 

enhanced free air ozone setting. Environ Pollut, 158: 977-982. http://dx.doi.org/10.1016/j.envpol.2009.08.001 

625424 

Olbrich, M.; Gerstner, E.; Welzl, G.; Winkler, J. B.; Ernst, D. (2009). Transcript responses in leaves of ozone-treated beech 

saplings seasons at an outdoor free air model fumigation site over two growing seasons. Plant Soil, 323: 61-74. 

http://dx.doi.org/10.1007/s11104-009-0129-4 596020 

March 2011 9-158 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

Ollinger, S. V.; Aber, J. D.; Reich, P. B. (1997). Simulating ozone effects on forest productivity: interactions among leafand stand-level processes. Ecol Appl, 123: 351-358. 040707 

Ollinger, S. V.; Aber, J. D.; Reich, P. B.; Freuder, R. J. (2002). Interactive effects of nitrogen deposition, tropospheric 

ozone, elevated CO2 and land use history on the carbon dynamics of northern hardwood forests. Global Change 

Biol, 8: 545-562. 180189 

Olszyk, D. M.; Kats, G.; Dawson, P. J.; Bytnerowicz, A.; Wolf, J.; Thompson, C. R. (1986). Characteristics of air exclusion 

systems vs chambers for field air pollution studies. J Environ Qual, 15: 326-334. 055530 

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. 049080 

Orendovici-Best, T.; Skelly, J. M.; Davis, D. D.; Ferdinand, J. A.; Savage, J. E.; Stevenson, R. E. (2008). Ozone uptake 

(flux) as it relates to ozone-induced foliar symptoms of Prunus serotina and Populus maximowizii x trichocarpa. 

Environ Pollut, 151: 79-92. http://dx.doi.org/10.1016/j.envpol.2007.03.003 196940 

Oshima, R. J.; Braegelmann, P. K.; Baldwin, D. W.; Van Way, V.; Taylor, O. C. (1977). Reduction of tomato fruit size and 

yield by ozone. J Am Soc Hortic Sci, 102: 289-293. 038938 

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 Waste Manag Assoc, 26: 861-865. 

038475 

Overmyer, K.; Brosche, M.; Kangasjarvi, J. (2003). Reactive oxygen species and hormonal control of cell death. Trends 

Plant Sci, 8: 335-342. 053537 

Overmyer, K.; Brosche, M.; Pellinen, R.; Kuittinen, T.; Tuominen, H.; Ahlfors, R.; Keinanen, M.; Saarma, M.; Scheel, D.; 

Kangasjarvi, J. (2005). Ozone-induced programmed cell death in the Arabidopsis radical-induced cell death1 

mutant. Plant Physiol, 137: 1092-1104. http://dx.doi.org/10.1104/pp.104.055681 191596 

Overmyer, K.; Kollist, H.; Tuominen, H.; Betz, C.; Langebartels, C.; Wingsle, G.; Kangasjarvi, S.; Brader, G.; Mullineaux, 

P.; Kangasjarvi, J. (2008). Complex phenotypic profiles leading to ozone sensitivity in Arabidopsis thaliana 

mutants. Plant Cell Environ, 31: 1237-1249. http://dx.doi.org/10.1111/j.1365-3040.2008.01837.x 191283 

Overmyer, K.; Tuominen, H.; Kettunen, R.; Betz, C.; Langebartels, C.; Sandermann, H. Jr; Kangasjarvi, J. (2000). Ozonesensitive Arabidopsis rcd1 mutant reveals opposite roles for ethylene and jasmonate signaling pathways in 

regulating superoxide-dependent cell death. Plant Cell, 12: 1849-1862. 036361 

Pan, Y. D.; Birdsey, R.; Hom, J.; McCullough, K. (2009). Separating effects of changes in atmospheric composition, 

climate and land-use on carbon sequestration of US Mid-Atlantic temperate forests. For Ecol Manage, 259: 151­

164. http://dx.doi.org/10.1016/j.foreco.2009.09.049 596032 

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. 079202 

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. 030190 

Panek, J.; Kurpius, M. R.; Goldstein, A. H. (2002). An evaluation of ozone exposure metrics for a seasonally droughtstressed ponderosa pine ecosystem. Environ Pollut, 117: 93-100. 040712 

Paolacci, A. R.; Miraldi, C.; Tanzarella, O. A.; Badiani, M.; Porceddu, E.; Nali, C.; Lorenzini, G.; Ciaffi, M. (2007). Gene 

expression induced by chronic ozone in the Mediterranean shrub Phillyrea latifolia: Analysis by cDNA-AFLP. Tree 

Physiol, 27: 1541-1550. http://dx.doi.org/10.1093/treephys/27.11.1541 191422 

Paoletti, E.; Grulke, N. E. (2010). Ozone exposure and stomatal sluggishness in different plant physiognomic classes. 

Environ Pollut, 158: 2664-2671. http://dx.doi.org/10.1016/j.envpol.2010.04.024 628561 

Paoletti, E.; Manning, W. J. (2007). Toward a biologically significant and usable standard for ozone that will also protect 

plants. Environ Pollut, 150: 85-95. http://dx.doi.org/10.1016/j.envpol.2007.06.037 180174 

Paoletti, E.; Seufert, G.; Della Rocca, G.; Thomsen, H. (2007). Photosynthetic responses to elevated CO2 and O-3 in 

Quercus ilex leaves at a natural CO2 spring. Environ Pollut, 147: 516-524. 

http://dx.doi.org/10.1016/j.envpol.2006.08.039 199365 

March 2011 9-159 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

Parsons, W. F. J.; Bockheim, J. G.; Lindroth, R. L. (2008). Independent, interactive, and species-specific responses of leaf 

litter decomposition to elevated CO2 and O3 in a northern hardwood forest. Ecosystems, 11: 505-519. 191853 

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. 053074 

Peltonen, P. A.; Julkunen-Tiitto, R.; Vapaavuori, E.; Holopainen, J. K. (2006). Effects of elevated carbon dioxide and ozone 

on aphid oviposition preference and birch bud exudate phenolics. Global Change Biol, 12: 1670-1679. 

http://dx.doi.org/10.1111/j.1365-2486.2006.01226.x 196946 

Peltonen, P. A.; Vapaavuori, E.; Heinonen, J.; Julkunen-Tiitto, R.; Holopainen, J. K. (2010). Do elevated atmospheric CO2 

and O-3 affect food quality and performance of folivorous insects on silver birch? Global Change Biol, 16: 918­

935. http://dx.doi.org/10.1111/j.1365-2486.2009.02073.x 596482 

Penn State (1987). Diagnosing injury to eastern forest trees. University Park, PA: Penn State College of Agricultural 

sciences. http://pubs.cas.psu.edu/PubTitle.asp?varTitle=diagnosing 626971 

Percy, K. E.; Nosal, M.; Heilman, W.; Dann, T.; Sober, J.; Legge, A. H.; Karnosky, D. F. (2007). New exposure-based 

metric approach for evaluating O3 risk to North American aspen forests. Environ Pollut, 147: 554-566. 

http://dx.doi.org/10.1016/j.envpol.2006.10.009 093287 

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. J Air Waste Manag Assoc, 

37: 906-912. 042148 

Pfleeger, T. G.; Plocher, M.; Bichel, P. (2010). Response of pioneer plant communities to elevated ozone exposure. Agric 

Ecosyst Environ, 138: 116-126. http://dx.doi.org/10.1016/j.agee.2010.04.009 644281 

Phillips, D. L.; Johnson, M. G.; Tingey, D. T.; Storm, M. J. (2009). Elevated CO2 and O3 effects on fine-root survivorship 

in ponderosa pine mesocosms. Oecologia, 160: 827-837. http://dx.doi.org/10.1007/s00442-009-1339-4 199802 

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. 041768 

Piikki, K.; De Temmerman, L.; Hogy, P.; Pleijel, H. (2008). The open-top chamber impact on vapour pressure deficit and 

its consequences for stomatal ozone uptake. Atmos Environ, 42: 6513-6522. 

http://dx.doi.org/10.1016/j.atmosenv.2008.04.014 199265 

Piikki, K.; De Temmerman, L.; Ojanpera, K.; Danielsson, H.; Pleijel, H. (2008). The grain quality of spring wheat 

(Triticum aestivum L.) in relation to elevated ozone uptake and carbon dioxide exposure. Eur J Agron, 28: 245-254. 

http://dx.doi.org/10.1016/j.eja.2007.07.004 199812 

Piikki, K.; Vorne, V.; Ojanpera, K.; Pleijel, H. (2007). Impact of elevated O-3 and CO2 exposure on potato (Solanum 

tuberosum L. cv. Bintje) tuber macronutrients (N, P, K, Mg, Ca). Agric Ecosyst Environ, 118: 55-64. 

http://dx.doi.org/10.1016/j.agee.2006.04.012 191451 

Pinto, D. M.; Blande, J. D.; Nykanen, R.; Dong, W. X.; Nerg, A. M.; Holopainen, J. K. (2007). Ozone degrades common 

herbivore-induced plant volatiles: Does this affect herbivore prey location by predators and parasitoids? J Chem 

Ecol, 33: 683-694. http://dx.doi.org/10.1007/s10886-007-9255-8 196949 

Pinto, D. M.; Blande, J. D.; Souza, S. R.; Nerg, A. M.; Holopainen, J. K. (2010). Plant volatile organic compounds (VOCs) 

in ozone (O-3) polluted atmospheres: The ecological effects. J Chem Ecol, 36: 22-34. 

http://dx.doi.org/10.1007/s10886-009-9732-3 596490 

Pinto, D. M.; Himanen, S. J.; Nissinen, A.; Nerg, A. M.; Holopainen, J. K. (2008). Host location behavior of Cotesia 

plutellae Kurdjumov (Hymenoptera: Braconidae) in ambient and moderately elevated ozone in field conditions. 

Environ Pollut, 156: 227-231. http://dx.doi.org/10.1016/j.envpol.2007.12.009 616554 

Pinto, D. M.; Nerg, A. M.; Holopainen, J. K. (2007). The role of ozone-reactive compounds, terpenes, and green leaf 

volatiles (GLVs), in the orientation of Cotesia plutellae. J Chem Ecol, 33: 2218-2228. 

http://dx.doi.org/10.1007/s10886-007-9376-0 196950 

Pinto, J. (2009). Wyoming winter smog. Nat Geosci, 2: 88-90. http://dx.doi.org/10.1038/ngeo430 187038 

Pleijel, H.; Danielsson h; Ojanpera, K.; De Temmerman, L.; Hogy, P. (2004). Relationships between ozone exposure and 

yield loss in wheat and potato - Suggestions of critical levels for ozone effects on crops. Atmos Environ, 38: 2259­

2269. http://dx.doi.org/10.1016/j.atmosenv.2003.09.076 036662 

March 2011 9-160 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

Pleijel, H.; Danielsson, H.; Gelang, J.; Sild, E.; Sellden, G. (1998). Growth stage dependence of the grain yield response to 

ozone in spring wheat (Triticum aestivum L). Agric Ecosyst Environ, 70: 61-68. 053021 

Pleijel, H.; Danielsson, H.; Ojanpera, K.; De Temmerman, L.; Hogy, P.; Badiani, M.; Karlsson, P. E. (2004). Relationships 

between ozone exposure and yield loss in European wheat and potato--a comparison of concentration- and fluxbased exposure indices. Atmos Environ, 38: 2259-2269. 056608 

Pleijel, H.; Ojanpera, 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 B Soil Plant Sci, 47: 20-25. 043747 

Plessl, M.; Elstner, E. F.; Rennenberg, H.; Habermeyer, J.; Heiser, I. (2007). Influence of elevated CO2 and ozone 

concentrations on late blight resistance and growth of potato plants. Environ Exp Bot, 60: 447-457. 

http://dx.doi.org/10.1016/j.envexpbot.2007.01.003 196952 

Plochl, 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. 040802 

Pollastrini, M.; Desotgiu, R.; Cascio, C.; Bussotti, F.; Cherubini, P.; Saurer, M.; Gerosa, G.; Marzuoli, R. (2010). Growth 

and physiological responses to ozone and mild drought stress of tree species with different ecological requirements. 

Trees Struct Funct, 24: 695-704. http://dx.doi.org/10.1007/s00468-010-0439-4 644392 

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 (pp. 193-213). San Diego, CA: Academic Press. 093689 

Pregitzer, K. S.; Burton, A. J.; King, J. S.; Zak, D. R. (2008). Soil respiration, root biomass, and root turnover following 

long-term exposure of northern forests to elevated atmospheric Co-2 and tropospheric O-3. New Phytol, 180: 153­

161. http://dx.doi.org/10.1111/j.1469-8137.2008.02564.x 191677 

Pregitzer, K.; Loya, W.; Kubiske, M.; Zak, D. (2006). Soil respiration in northern forests exposed to elevated atmospheric 

carbon dioxide and ozone. Oecologia, 148: 503-516. http://dx.doi.org/10.1007/s00442-006-0381-8 191676 

Pretzsch, H.; Dieler, J.; Matyssek, R.; Wipfler, P. (2010). Tree and stand growth of mature Norway spruce and European 

beech under long-term ozone fumigation. Environ Pollut, 158: 1061-1070. 

http://dx.doi.org/10.1016/j.envpol.2009.07.035 580435 

Pritsch, K.; Esperschuetz, J.; Haesler, F.; Raidl, S.; Winkler, B.; Schloter, M. (2009). Structure and activities of 

ectomycorrhizal and microbial communities in the rhizosphere of Fagus sylvatica under ozone and pathogen stress 

in a lysimeter study. Plant Soil, 323: 97-109. http://dx.doi.org/10.1007/s11104-009-9972-6 626808 

Puckette, M. C.; Tang, Y. H.; Mahalingam, R. (2008). Transcriptomic changes induced by acute ozone in resistant and 

sensitive Medicago truncatula accessions. BMC Plant Biol, 8: 46. http://dx.doi.org/10.1186/1471-2229-8-46 

191698 

Ramo, K.; Kanerva, T.; Ojanpera, K.; Manninen, S. (2007). Growth onset, senescence, and reproductive development of 

meadow species in mesocosms exposed to elevated O3 and CO2. Environ Pollut, 145: 850-860. 

http://dx.doi.org/10.1016/j.envpol.2006.03.054 191441 

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. J Plant Physiol, 109: 421-432. 030221 

Rapport, D. J.; Whitford, W. G. (1999). How ecosystems respond to stress: Common properties of arid and aquatic systems. 

Bioscience, 49: 193-203. 004595 

Rawlings, J. O.; Cure, W. W. (1985). The Weibull function as a dose-response model to describe ozone effects on crop 

yields. Crop Sci, 25: 807-814. 039419 

Reich, P. B. (1987). Quantifying plant response to ozone: A unifying theory. Tree Physiol, 3: 63-91. 019314 

Reid, C. D.; Fiscus, E. L. (2008). Ozone and density affect the response of biomass and seed yield to elevated CO2 in rice. 

Global Change Biol, 14: 60-76. http://dx.doi.org/10.1111/j.1365-2486.2007.01472.x 191561 

Reiling, K.; Davison, A. W. (1992). Effects of a short ozone exposure given at different stages in the development of 

Plantago major L. New Phytol, 121: 643-647. 043591 

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. 035373 

March 2011 9-161 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

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. 030252 

Reinert, R. A.; Ho, M. C. (1995). Vegetative growth of soybean as affected by elevated carbon dioxide and ozone. Environ 

Pollut, 89: 89-96. 030247 

Ren, W.; Tian, H. Q.; Liu, M. L.; Zhang, C.; Chen, G. S.; Pan, S. F.; Felzer, B.; Xu, X. F. (2007). Effects of tropospheric 

ozone pollution on net primary productivity and carbon storage in terrestrial ecosystems of China. J Geophys Res, 

112: D22S09. http://dx.doi.org/10.1029/2007jd008521 191366 

Ren, W.; Tian, H.; Chen, G.; Liu, M.; Zhang, C.; Chappelka, A. H.; Pan, S. (2007). Influence of ozone pollution and climate 

variability on net primary productivity and carbon storage in China's grassland ecosystems from 1961 to 2000. 

Environ Pollut, 149: 327-335. http://dx.doi.org/10.1016/j.envpol.2007.05.029 581541 

Rhea, L.; King, J.; Kubiske, M.; Saliendra, N.; Teclaw, R. (2010). Effects of elevated atmospheric CO2 and tropospheric O3 on tree branch growth and implications for hydrologic budgeting. Environ Pollut, 158: 1079-1087. 

http://dx.doi.org/10.1016/j.envpol.2009.08.038 647084 

Riikonen, J.; Kets, K.; Darbah, J.; Oksanen, E.; Sober, A.; Vapaavuori, E.; Kubiske, M. E.; Nelson, N.; Karnosky, D. F. 

(2008). Carbon gain and bud physiology in Populus tremuloides and Betula papyrifera grown under long-term 

exposure to elevated concentrations of CO2 and O3. Tree Physiol, 28: 243-254. 

http://dx.doi.org/10.1093/treephys/28.2.243 191258 

Riikonen, J.; Maenpaa, M.; Alavillamo, M.; Silfver, T.; Oksanen, E. (2009). Interactive effect of elevated temperature and 

O3 on antioxidant capacity and gas exchange in Betula pendula saplings. Planta, 230: 419-427. 

http://dx.doi.org/10.1007/s00425-009-0957-8 195664 

Rizzo, M.; Bernardi, R.; Salvini, M.; Nali, C.; Lorenzini, G.; Durante, M. (2007). Identification of differentially expressed 

genes induced by ozone stress in sensitive and tolerant poplar hybrids. J Plant Physiol, 164: 945-949. 

http://dx.doi.org/10.1016/j.jplph.2006.07.012 191447 

Rodenkirchen, H.; Gottlein, A.; Kozovits, A. R.; Matyssek, R.; Grams, T. E. E. (2009). Nutrient contents and efficiencies of 

beech and spruce saplings as influenced by competition and O3/CO2 regime. European Journal of Forest Research, 

128: 117-128. http://dx.doi.org/10.1007/s10342-008-0221-y 191540 

Rogers, A.; Allen, D. J.; Davey, P. A.; Morgan, P. B.; Ainsworth, E. A.; Bernacchi, C. J.; Cornic, G.; Dermody, O. C.; 

Dohleman, F. G.; 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, 27: 449-458. 079201 

Rowland-Bamford, A. J. (2000). Plant responses to changing carbon dioxide and temperature. In Singh, S. N. (Ed.), Trace 

gas emissions and plants Dordecht, The Netherlands: Kluwer Academic Publishers. 030257 

Ryan, A.; Cojocariu, C.; Possell, M.; Davies, W. J.; Hewitt, C. N. (2009). Defining hybrid poplar (Populus deltoides x 

Populus trichocarpa) tolerance to ozone: identifying key parameters. Plant Cell Environ, 32: 31-45. 

http://dx.doi.org/10.1111/j.1365-3040.2008.01897.x 191299 

Ryang, S. Z.; Woo, S. Y.; Kwon, S. Y.; Kim, S. H.; Lee, S. H.; Kim, K. N.; Lee, D. K. (2009). Changes of net 

photosynthesis, antioxidant enzyme activities, and antioxidant contents of Liriodendron tulipifera under elevated 

ozone. Photosynthetica, 47: 19-25. http://dx.doi.org/10.1007/s11099-009-0005-8 191267 

Samuel, M. A.; Ellis, B. E. (2002). Double jeopardy: Both overexpression and suppression of a redox-activated plant 

mitogen-activated protein kinase render tobacco plants ozone sensitive. Plant Cell, 14: 2059-2069. 

http://dx.doi.org/10.1105/tpc.002337 625703 

Samuel, M. A.; Miles, G. P.; Ellis, B. E. (2000). Ozone treatment rapidly activates MAP kinase signalling in plants. Plant J, 

22: 367-376. http://dx.doi.org/10.1046/j.1365-313x.2000.00741.x 625706 

Samuel, M. A.; Walia, A.; Mansfield, S. D.; Ellis, B. E. (2005). Overexpression of SIPK in tobacco enhances ozoneinduced ethylene formation and blocks ozone-induced SA accumulation. J Exp Bot, 56: 2195-2201. 

http://dx.doi.org/10.1093/jxb/eri219 199316 

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. 040832 

March 2011 9-162 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

Sánchez, M. J. S.; Peña, G. S.; Lorente, V. C.; Gallego, T. M.; Albert, J. C. (2001). La contaminación atmosférica en los 

bosques: Guía para la identificacíon de daños visibles causados por Ozono. Madrid, Spain: Ministerio de Medio 

Ambiente. http://www.ceam.es/VentaLibros/guia_O3/index.htm 626980 

Sanz, J.; Bermejo, V.; Gimeno, B. S.; Elvira, S.; Alonso, R. (2007). Ozone sensitivity of the Mediterranean terophyte 

Trifolium striatum is modulated by soil nitrogen content. Atmos Environ, 41: 8952-8962. 

http://dx.doi.org/10.1016/j.atmosenv.2007.08.016 199245 

Sanz, J.; Muntifering, R. B.; Bermejo, V.; Gimeno, B. S.; Elvira, S. (2005). Ozone and increased nitrogen supply effects on 

the yield and nutritive quality of Trifolium subterraneum. Atmos Environ, 39: 5899-5907. 

http://dx.doi.org/10.1016/j.atmosenv.2005.06.022 196963 

Sarkar, A.; Rakwal, R.; Agrawal, S. B.; Shibato, J.; Ogawa, Y.; Yoshida, Y.; Agrawal, G. K.; Agrawal, M. (2010). 

Investigating the impact of elevated levels of ozone on tropical wheat using integrated phenotypical, physiological, 

biochemical, and proteomics approaches. J Proteome Res, 9: 4565-4584. http://dx.doi.org/10.1021/Pr1002824 

657214 

Saviranta, N. M. M.; Julkunen-Tiitto, R.; Oksanen, E.; Karjalainen, R. O. (2010). Leaf phenolic compounds in red clover 

(Trifolium pratense L.) induced by exposure to moderately elevated ozone. Environ Pollut, 158: 440-446. 

http://dx.doi.org/10.1016/j.envpol.2009.08.029 102177 

Sawada, H.; Kohno, Y. (2009). Differential ozone sensitivity of rice cultivars as indicated by visible injury and grain yield. 

Plant Biol (Stuttg), 11: 70-75. http://dx.doi.org/10.1111/j.1438-8677.2009.00233.x 199426 

Scebba, F.; Giuntini, D.; Castagna, A.; Soldatini, G.; Ranieri, A. (2006). Analysing the impact of ozone on biochemical and 

physiological variables in plant species belonging to natural ecosystems. Environ Exp Bot, 57: 89-97. 

http://dx.doi.org/10.1016/j.envexpbot.2005.04.005 191219 

Schaub, M.; Skelly, J. M.; Zhang, J. W.; Ferdinand, J. A.; Savage, J. E.; Stevenson, R. E.; Davis, D. D.; Steiner, K. C. 

(2005). Physiological and foliar symptom response in the crowns of Prunus serotina, Fraxinus americana and Acer 

rubrum canopy trees to ambient ozone under forest conditions. Environ Pollut, 133: 553-567. 

http://dx.doi.org/10.1016/j.envpol.2004.06.012 191608 

Schraudner, M.; Moeder, W.; Wiese, C.; Van Camp, W.; Inze, D.; Langebartels, C.; Sandermann, H. Jr (1998). Ozoneinduced oxidative burst in the ozone biomonitor plant, tobacco Bel W3. Plant J, 16: 235-245. 056358 

Severino, J. F.; Stich, K.; Soja, G. (2007). Ozone stress and antioxidant substances in Trifolium repens and Centaurea jacea 

leaves. Environ Pollut, 146: 707-714. http://dx.doi.org/10.1016/j.envpol.2006.04.006 199293 

Sharkey, T. D.; Wiberley, A. E.; Donohue, A. R. (2008). Isoprene emission from plants: Why and how. Ann Bot, 101: 5-18. 

http://dx.doi.org/10.1093/aob/mcm240 191492 

Singh, E.; Tiwari, S.; Agrawal, M. (2009). Effects of elevated ozone on photosynthesis and stomatal conductance of two 

soybean varieties: A case study to assess impacts of one component of predicted global climate change. Plant Biol 

(Stuttg), 11: 101-108. http://dx.doi.org/10.1111/j.1438-8677.2009.00263.x 199427 

Singh, E.; Tiwari, S.; Agrawal, M. (2010). Variability in antioxidant and metabolite levels, growth and yield of two soybean 

varieties: An assessment of anticipated yield losses under projected elevation of ozone. Agric Ecosyst Environ, 135: 

168-177. http://dx.doi.org/10.1016/j.agee.2009.09.004 386275 

Sitch, S.; Cox, P. M.; Collins, W. J.; Huntingford, C. (2007). Indirect radiative forcing of climate change through ozone 

effects on the land-carbon sink. Nature, 448: 791-794. http://dx.doi.org/10.1038/nature06059 093294 

Skarby, L.; Ottosson, S.; Karlsson, P. E.; Wallina, G.; Sellden, 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. 

080375 

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. 044183 

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. 030388 

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. 094397 

March 2011 9-163 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

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. 044010 

Souza, L.; Neufeld, H. S.; Chappelka, A. H.; Burkey, K. O.; Davison, A. W. (2006). Seasonal development of ozoneinduced foliar injury on tall milkweed (Asclepias exaltata) in Great Smoky Mountains National Park. Environ 

Pollut, 141: 175-183. http://dx.doi.org/10.1016/j.envpol.2005.07.022 191658 

Stampfli, A.; Fuhrer, J. (2010). Spatial heterogeneity confounded ozone-exposure experiment in semi-natural grassland. 

Oecologia, 162: 515-522. http://dx.doi.org/10.1007/s00442-009-1462-2 102180 

Stewart, C. A.. (1998). Impact of ozone on the reproductive biology of Brassica campestris L and Plantago major L. [Ph. D] 

(Doctoral dissertation). Loughborough University, England. 094233 

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. 036384 

Talhelm, A. F.; Pregitzer, K. S.; Zak, D. R. (2009). Species-specific responses to atmospheric carbon dioxide and 

tropospheric ozone mediate changes in soil carbon. Ecol Lett, 12: 1219-1228. http://dx.doi.org/10.1111/j.1461­

0248.2009.01380.x 596189 

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. http://dx.doi.org/10.1023/B:PLAN.0000019064.55734.52 080053 

Temple, P. J.; Kupper, R. S.; Lennox, R. W.; Rohr, K. (1988). Injury and yield responses of differentially irrigated cotton to 

ozone. Agron Sustain Dev, 80: 751-755. 043237 

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. 043349 

Theis, N.; Raguso, R. A. (2005). The effect of pollination on floral fragrance in thistles. J Chem Ecol, 31: 2581-2600. 

626461 

Thomas, V. F. D.; Braun, S.; Fluckiger, W. (2005). Effects of simultaneous ozone exposure and nitrogen loads on 

carbohydrate concentrations, biomass, and growth of young spruce trees (Picea abies). Environ Pollut, 137: 507­

516. http://dx.doi.org/10.1016/j.envpol.2005.02.002 075930 

Thomas, V. F. D.; Braun, S.; Fluckiger, W. (2006). Effects of simultaneous ozone exposure and nitrogen loads on 

carbohydrate concentrations, biomass, growth, and nutrient concentrations of young beech trees (Fagus sylvatica). 

Environ Pollut, 143: 341-354. http://dx.doi.org/10.1016/j.envpol.2005.11.036 191362 

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 RL Berglund; DR Lawson; DJ McKee (Eds.), 

Tropospheric Ozone and the Environment (pp. 272-288). Pittsburgh, PA: Air & Waste Management Association. 

042623 

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. J Environ Manage, 34: 397-405. 042385 

Tingey, D. T.; Johnson, M. G.; Lee, E. H.; Wise, C.; Waschmann, R.; Olszyk, D. M.; Watrud, L. S.; Donegan, K. K. (2006). 

Effects of elevated CO2 and O-3 on soil respiration under ponderosa pine. Soil Biol Biochem, 38: 1764-1778. 

http://dx.doi.org/10.1016/j.soilbio.2005.12.003 191341 

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. 

055536 

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. 040896 

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. 029011 

Tissue, D. T.; Thomas, R. B.; Strain, B. R. (1997). Atmospheric CO2 enrichment increases growth and photosynthesis of 

Pinus taeda: A 4 year experiment in the field. Plant Cell Environ, 20: 1123-1134. http://dx.doi.org/10.1046/j.1365­

3040.1997.d01-140.x 679718 

March 2011 9-164 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

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. 035394 

Tong, D. Q.; Mauzerall, D. L. (2008). Summertime State-Level Source-Receptor Relationships between Nitrogen Oxides 

Emissions and Surface Ozone Concentrations over the Continental United States. Environ Sci Technol, 42: 7976­

7984. http://dx.doi.org/10.1021/es7027636 621169 

Tong, D.; Mathur, R.; Schere, K.; Kang, D.; Yu, S. (2007). The use of air quality forecasts to assess impacts of air pollution 

on crops: Methodology and case study. Atmos Environ, 41: 8772-8784. 

http://dx.doi.org/10.1016/j.atmosenv.2007.07.060 107431 

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. 035398 

Tosti, N.; Pasqualini, S.; Borgogni, A.; Ederli, L.; Falistocco, E.; Crispi, S.; Paolocci, F. (2006). Gene expression profiles of 

O-3-treated Arabidopsis plants. Plant Cell Environ, 29: 1686-1702. http://dx.doi.org/10.1111/j.1365­

3040.2006.01542.x 191425 

Turnipseed, A. A.; Burns, S. P.; Moore, D. J. P.; Hu, J.; Guenther, A. B.; Monson, R. K. (2009). Controls over ozone 

deposition to a high elevation subalpine forest. Agr Forest Meteorol, 149: 1447-1459. 

http://dx.doi.org/10.1016/j.agrformet.2009.04.001 588752 

U.S. EPA (1978). Air quality criteria for ozone and other photochemical oxidants (Report No. EPA/600/8-78/004). 

Washington, DC: U.S. Environmental Protection Agency, Office of Research and Development. 040586 

U.S. EPA (1984). Air quality criteria for ozone and other photochemical oxidants, v 3 (Report No. EPA/600/8-84/020A). 

Research Triangle Park, NC: U.S. Environmental Protection Agency, Office of Research and Development. 

http://nepis.epa.gov/Exe/ZyPURL.cgi?Dockey=2000AVEV.txt. 029711 

U.S. EPA (1986). Air quality criteria for ozone and other photochemical oxidants (Report No. EPA-600/8-84-020aF - EPA600/8-84-020eF). Research Triangle Park, NC: U.S. Environmental Protection Agency. 017607 

U.S. EPA (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 (Report No. EPA/600/8-88/105F). Research Triangle 

Park, NC: U.S. Environmental Protection Agency, Office of Health and Environmental Assessment. 

http://nepis.epa.gov/Exe/ZyPURL.cgi?Dockey=30001HVP.txt. 042599 

U.S. EPA (1996). Air quality criteria for ozone and related photochemical oxidants Volume II of III (Report No. 

EPA/600/P-93/004BF). Research Triangle Park, NC: U.S. Environmental Protection Agency, Office of Research 

and Development. 080827 

U.S. EPA (1996). Air quality criteria for ozone and related photochemical oxidants Volume III of III (Report No. 

EPA/600/P-93/004cF). Research Triangle Park, NC: U.S. Environmental Protection Agency. (NTIS No. PB96­

185608) 080828 

U.S. EPA (1996). Review of national ambient air quality standards for ozone: assessment of scientific and technical 

information OAQPS staff paper (Report No. EPA/452/R-96/007). Research Triangle Park, NC: U.S. Environmental 

Protection Agency. (NTIS No. PB96-203435) 039046 

U.S. EPA (2006). Air quality criteria for ozone and related photochemical oxidants (Report No. EPA/600/R-05/004AF). 

Research Triangle Park, NC: U.S. Environmental Protection Agency, Office of Research and Development. 

http://cfpub.epa.gov/ncea/cfm/recordisplay.cfm?deid=149923. 088089 

U.S. EPA (2007). Review of the national ambient air quality standards for ozone: policy assessment of scientific and 

technical information OAQPS staff paper (Report No. EPA/452/R-07/003). Research Triangle Park, NC: U.S. 

Environmental Protection Agency, Office of Air Quality Planning and Standards. 090207 

Uddling, J.; Hogg, A. J.; Teclaw, R. M.; Carroll, M. A.; Ellsworth, D. S. (2010). Stomatal uptake of O3 in aspen and aspenbirch forests under free-air CO2 and O3 enrichment. Environ Pollut, 158: 2023-2031. 

http://dx.doi.org/10.1016/j.envpol.2009.12.001 387073 

Uddling, J.; Teclaw, R. M.; Kubiske, M. E.; Pregitzer, K. S.; Ellsworth, D. S. (2008). Sap flux in pure aspen and mixed 

aspen-birch forests exposed to elevated concentrations of carbon dioxide and ozone. Tree Physiol, 28: 1231-1243. 

191655 

March 2011 9-165 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

Uddling, J.; Teclaw, R. M.; Pregitzer, K. S.; Ellsworth, D. S. (2009). Leaf and canopy conductance in aspen and aspenbirch forests under free-air enrichment of carbon dioxide and ozone. Tree Physiol, 29: 1367-1380. 

http://dx.doi.org/10.1093/treephys/tpp070 596219 

United Nations Economic Commission for Europe UNECE (1988). ECE critical levels workshop; March; Bad Harzburg, 

Germany [final report]. In [Book title missing] [City missing]: [Publisher missing] 055354 

University of Illinois (2010). SoyFACE. Retrieved December 08, 2010 from http://soyface.illinois.edu/. 670286 

Unsworth, M. H.; Heagle, A. S.; Heck, W. W. (1984). Gas exchange in open-top field chambers - I measurement and 

analysis of atmospheric resistances to gas exchange. Atmos Environ, 18: 373-380. 041239 

Unsworth, M. H.; Heagle, A. S.; Heck, W. W. (1984). Gas exchange in open-top field chambers - II resistances to ozone 

uptake by soybeans. Atmos Environ, 18: 381-385. 040024 

USDA (2011). Ozone Biomonitoring Program. Retrieved January 28, 2011 from http://www.nrs.fs.fed.us/fia/topics/ozone/. 

677550 

Valkama, E.; Koricheva, J.; Oksanen, E. (2007). Effects of elevated O3, alone and in combination with elevated CO2, on 

tree leaf chemistry and insect herbivore performance: A meta-analysis. Global Change Biol, 13: 184-201. 

http://dx.doi.org/10.1111/j.1365-2486.01284.x 191348 

Vandermeiren, K.; Black, C.; Pleijel, H.; de Temmerman, L. (2005). Impact of rising tropospheric ozone on potato: effects 

on photosynthesis, growth, productivity and yield quality. Plant Cell Environ, 28: 982-996. 

http://dx.doi.org/10.1111/j.1365-3040.2005.01316.x 179992 

van Buuren, M. L.; Guidi, L.; Fornale, S.; Ghetti, F.; Franceschetti, M.; Soldatini, G. F.; Bagni, N. (2002). Ozone-response 

mechanisms in tobacco: Implications of polyamine metabolism. New Phytol, 156: 389-398. 

http://dx.doi.org/10.1046/j.1469-8137.2002.00539.x 631199 

Van Dingenen, R.; Dentener, F. J.; Raes, F.; Krol, M. C.; Emberson, L.; Cofala, J. (2009). The global impact of ozone on 

agricultural crop yields under current and future air quality legislation. Atmos Environ, 43: 604-618. 

http://dx.doi.org/10.1016/j.atmosenv.2008.10.033 199765 

Velikova, V.; Pinelli, P.; Pasqualini, S.; Reale, L.; Ferranti, F.; Loreto, F. (2005). Isoprene decreases the concentration of 

nitric oxide in leaves exposed to elevated ozone. New Phytol, 166: 419-426. http://dx.doi.org/10.1111/j.1469­

8137.2005.01409.x 199410 

Vickers, C. E.; Possell, M.; Cojocariu, C. I.; Velikova, V. B.; Laothawornkitkul, J.; Ryan, A.; Mullineaux, P. M.; Hewitt, C. 

N. (2009). Isoprene synthesis protects transgenic tobacco plants from oxidative stress. Plant Cell Environ, 32: 520­

531. http://dx.doi.org/10.1111/j.1365-3040.2009.01946.x 191497 

Vigue, L. M.; Lindroth, R. L. (2010). Effects of genotype, elevated CO2 and elevated O3 on aspen phytochemistry and 

aspen leaf beetle Chrysomela crotchi performance. Agricultural and Forest Entomology, 12: 267-276. 

http://dx.doi.org/10.1111/j.1461-9563.2010.00475.x 644185 

Volk, M.; Bungener, P.; Contat, F.; Montani, M.; Fuhrer, J. (2006). Grassland yield declined by a quarter in 5 years of freeair ozone fumigation. Global Change Biol, 12: 74-83. http://dx.doi.org/10.1111/j.1365-2486.2005.01083.x 191434 

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. 055568 

Vollenweider, P.; Woodcock, H.; Kelty, M. J.; Hofer, R.-M. (2003). Reduction of stem growth and site dependency of leaf 

injury in Massachusetts black cherries exhibiting ozone symptoms. Environ Pollut, 125: 467-480. 051157 

Vollsnes, A. V.; Kruse, O. M. O.; Eriksen, A. B.; Oxaal, U.; Futsaether, C. M. (2010). In vivo root growth dynamics of 

ozone exposed Trifolium subterraneum. Environ Exp Bot, 69: 183-188. 

http://dx.doi.org/10.1016/j.envexpbot.2010.03.007 625576 

Vuorinen, T.; Nerg, A. M.; Holopainen, J. K. (2004). Ozone exposure triggers the emission of herbivore-induced plant 

volatiles, but does not disturb tritrophic signalling. Environ Pollut, 131: 305-311. 626462 

Wang, L. L.; He, X. Y.; Chen, W. (2009). Effects of elevated ozone on photosynthetic CO2 exchange and chlorophyll a 

fluorescence in leaves of Quercus mongolica grown in urban area. Bull Environ Contam Toxicol, 82: 478-481. 

http://dx.doi.org/10.1007/s00128-008-9606-3 199303 

March 2011 9-166 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

Wang, X. K.; Zheng, Q. W.; Feng, Z. Z.; Xie, J. Q.; Feng, Z. W.; Ouyang, Z.; Manning, W. J. (2008). Comparison of a 

diurnal vs steady-state ozone exposure profile on growth and yield of oilseed rape (Brassica napus L.) in open-top 

chambers in the Yangtze Delta, China. Environ Pollut, 156: 449-453. 

http://dx.doi.org/10.1016/j.envpol.2008.01.027 191276 

Wang, X.; Mauzerall, D. L. (2004). Characterizing distributions of surface ozone and its impact on grain production in 

China, Japan and South Korea: 1990 and 2020. Atmos Environ, 38: 4383-4402. 

http://dx.doi.org/10.1016/j.atmosenv.2004.03.067 179978 

Wang, X.; Taub, D. R. (2010). Interactive effects of elevated carbon dioxide and environmental stresses on root mass 

fraction in plants: A meta-analytical synthesis using pairwise techniques. Oecologia, 163: 1-11. 

http://dx.doi.org/10.1007/s00442-010-1572-x 387478 

Watanabe, M.; Yamaguchi, M.; Tabe, C.; Iwasaki, M.; Yamashita, R.; Funada, R.; Fukami, M.; Matsumura, H.; Kohno, Y.; 

Izuta, T. (2007). Influences of nitrogen load on the growth and photosynthetic responses of Quercus serrata 

seedlings to O3. Trees Struct Funct, 21: 421-432. http://dx.doi.org/10.1007/s00468-007-0134-2 191474 

Weinstein, D. A.; Beloin, R. M.; Yanai, R. D. (1991). Modeling changes in red spruce carbon balance and allocation in 

response to interacting ozone and nutrient stresses. Tree Physiol, 9: 127-146. 043993 

Weinstein, D. A.; Laurence, J. A.; Retzlaff, W. A.; Kern, J. S.; Lee, E. H.; Hogsett, W. E.; Weber, J. (2005). Predicting the 

effects of tropospheric ozone on regional productivity of ponderosa pine and white fir. For Ecol Manage, 205: 73­

89. http://dx.doi.org/10.1016/j.foreco.2004.10.007 179965 

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 Int, 9: 117-121. 053040 

Wesely, M. L.; Hicks, B. B. (2000). A review of the current status of knowledge on dry deposition. Atmos Environ, 34: 

2261-2282. 025018 

Whitfield, C. P.; Davison, A. W.; Ashenden, T. W. (1996). Interactive effects of ozone and soil volume on Plantago major. 

New Phytol, 134: 287-294. 055544 

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. 036397 

Wieser, G.; Manning, W. J.; Tausz, M.; Bytnerowicz, A. (2006). Evidence for potential impacts of ozone on Pinus cembra 

L. at mountain sites in Europe: An overview. Environ Pollut, 139: 53-58.
 

http://dx.doi.org/10.1016/j.envpol.2005.04.037 191391
 

Wilkinson, S.; Davies, W. J. (2009). Ozone suppresses soil drying- and abscisic acid (ABA)-induced stomatal closure via 

an ethylene-dependent mechanism. Plant Cell Environ, 32: 949-959. http://dx.doi.org/10.1111/j.1365­

3040.2009.01970.x 199758 

Wilkinson, S.; Davies, W. J. (2010). Drought, ozone, ABA and ethylene: New insights from cell to plant to community. 

Plant Cell Environ, 33: 510-525. http://dx.doi.org/10.1111/j.1365-3040.2009.02052.x 598245 

Will, R. E.; Ceulemans, R. (1997). Effects of elevated CO2 concentration on photosynthesis, respiration and carbohydrate 

status of coppice Populus hybrids. Physiol Plant, 100: 933-939. http://dx.doi.org/10.1111/j.1399­

3054.1997.tb00020.x 679719 

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. PNAS, 86: 8828-8832. 

043403 

Wittig, V. E.; Ainsworth, E. A.; Long, S. P. (2007). To what extent do current and projected increases in surface ozone 

affect photosynthesis and stomatal conductance of trees? A meta-analytic review of the last 3 decades of 

experiments. Plant Cell Environ, 30: 1150-1162. 191695 

Wittig, V. E.; Ainsworth, E. A.; Naidu, S. L.; Karnosky, D. F.; Long, S. P. (2009). Quantifying the impact of current and 

future tropospheric ozone on tree biomass, growth, physiology and biochemistry: a quantitative meta-analysis. 

Global Change Biol, 15: 396-424. http://dx.doi.org/10.1111/j.1365-2486.2008.01774.x 191631 

Woo, S. Y.; Hinckley, T. M. (2005). The effects of ozone on growth and stomatal response in the F-2 generation of hybrid 

poplar (Populus trichocarpa x Populus deltoides). Biol Plantarum, 49: 395-404. http://dx.doi.org/10.1007/s10535­

005-0014-9 191359 

March 2011 9-167 DRAFT – DO NOT CITE OR QUOTE 
 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

Wright, G. A.; Lutmerding, A.; Dudareva, N.; Smith, B. H. (2005). Intensity and the ratios of compounds in the scent of 

snapdragon flowers affect scent discrimination by honeybees (Apis mellifera). J Comp Physiol A Neuroethol Sens 

Neural Behav Physiol, 191: 105-114. 626463 

Wrzaczek, M.; Brosché, M.; Salojärvi, J.; Kangasjärvi, S.; Idänheimo, N.; Mersmann, S.; Robatzek, S.; Karpiński, S.; 

Karpińska, B.; Kangasjärvi, J. (2010). Transcriptional regulation of the CRK/DUF26 group of receptor-like protein 

kinases by ozone and plant hormones in Arabidopsis. BMC Plant Biol, 10: 95. 644190 

Yamaguchi, M.; Watanabe, M.; Iwasaki, M.; Tabe, C.; Matsumura, H.; Kohno, Y.; Izuta, T. (2007). Growth and 

photosynthetic responses of Fagus crenata seedlings to O-3 under different nitrogen loads. Trees Struct Funct, 21: 

707-718. http://dx.doi.org/10.1007/s00468-007-0163-x 191438 

Yan, K.; Chen, W.; He, X. Y.; Zhang, G. Y.; Xu, S.; Wang, L. L. (2010). Responses of photosynthesis, lipid peroxidation 

and antioxidant system in leaves of Quercus mongolica to elevated O-3. Environ Exp Bot, 69: 198-204. 

http://dx.doi.org/10.1016/j.envexpbot.2010.03.008 628514 

Yoshida, S.; Tamaoki, M.; Ioki, M.; Ogawa, D.; Sato, Y.; Aono, M.; Kubo, A.; Saji, S.; Saji, H.; Satoh, S.; Nakajima, N. 

(2009). Ethylene and salicylic acid control glutathione biosynthesis in ozone-exposed Arabidopsis thaliana. Physiol 

Plant, 136: 284-298. http://dx.doi.org/10.1111/j.1399-3054.2009.01220.x 191385 

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. 044162 

Yuan, J. S.; Himanen, S. J.; Holopainen, J. K.; Chen, F.; Stewart, C. N. Jr (2009). Smelling global climate change: 

Mitigation of function for plant volatile organic compounds. Trends Ecol Evol, 24: 323-331. 

http://dx.doi.org/10.1016/j.tree.2009.01.012 199779 

Yun, S.-C.; Laurence, J. A. (1999). The response of sensitive and tolerant clones of Populus tremuloides to dynamic ozone 

exposure under controlled environmental conditions. New Phytol, 143: 305-313. 044165 

Zak, D. R.; Holmes, W. E.; Pregitzer, K. S. (2007). Atmospheric CO2 and O-3 alter the flow of N-15 in developing forest 

ecosystems. Ecology, 88: 2630-2639. 191239 

Zhang, C.; Tian, H. Q.; Chappelka, A. H.; Ren, W.; Chen, H.; Pan, S. F.; Liu, M. L.; Styers, D. M.; Chen, G. S.; Wang, Y. H. 

(2007). Impacts of climatic and atmospheric changes on carbon dynamics in the Great Smoky Mountains National 

Park. Environ Pollut, 149: 336-347. http://dx.doi.org/10.1016/j.envpol.2007.05.028 196983 

Zhang, J.; Schaub, M.; Ferdinand, J. A.; Skelly, J. M.; Steiner, K. C.; Savage, J. E. (2010). Leaf age affects the responses of 

foliar injury and gas exchange to tropospheric ozone in Prunus serotina seedlings. Environ Pollut, 158: 2627-2634. 

http://dx.doi.org/10.1016/j.envpol.2010.05.003 628555 

March 2011 9-168 DRAFT – DO NOT CITE OR QUOTE