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9 ENVIRONMENTAL EFFECTS: OZONE EFFECTS 

ON VEGETATION AND ECOSYSTEMS

9.1 Introduction

1 This chapter synthesizes and evaluates the relevant science to help form the scientific 

2 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) 

4 definition of welfare effects includes, but is not limited to, effects on soils, water, 

5 wildlife, vegetation, visibility, weather, and climate, as well as effects on materials, 

6 economic values, and personal comfort and well-being. The effects of O3 as a greenhouse 

7 gas and its direct effects on climate are discussed in 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” (42 U.S.C.7408

10 and 42 U.S.C.7409 (1999). This chapter of the ISA includes scientific research from 

11 biogeochemistry, soil science, plant physiology, and ecology conducted at multiple scales 

12 (e.g., organ, organism, population, community, ecosystem). Key information and 

13 judgments formerly found in the AQCDs regarding O3 effects on vegetation and 

14 ecosystems are found in this chapter. This chapter of the O3 ISA serves to update and 

15 revise Chapter 9 and AX9 of the 2006 O3 AQCD (U.S. EPA, 2006b).

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 

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

19 demonstrated plausible mechanisms for these effects. Ozone effect studies published 

20 from 2005 to July 2011 are reviewed in this document in the context of the previous O3

21 AQCDs. From 2005 to 2011, some areas have had very little new research published and 

22 the reader is referred back to sections of the 2006 O3 AQCD for a more comprehensive 

23 discussion of those subjects. This chapter is focused on studies of vegetation and 

24 ecosystems that occur in the U.S. and that report endpoints or processes most relevant to 

25 the review of the secondary standard. Many studies have been published about vegetation 

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

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

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

29 species and ecosystems. For example, studies outside North America are discussed that 

30 consider physiological and biochemical processes that contribute to the understanding of 

31 effects of O3 across species. Also, ecosystem studies outside of North America that 
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1 contribute to the understanding of O3 effects on general ecosystem processes are 

2 discussed in the chapter. 

3 Sections of this chapter first discuss exposure methods, followed by effects on vegetation 

4 and ecosystems at various spatial scales and ends with policy-relevant discussions of 

5 exposure indices and exposure-response. Figure 9-1 is a simplified illustrative diagram of 

6 the major pathway through which O3 enters plants and the major endpoints O3 may 

7 affect. First, Section 9.2 presents a brief overview of various methodologies that have 

8 been, and continue to be, central to quantifying O3 effects on vegetation (AX9.1 of the 

9 2006 O3 AQCD for more detailed discussion) (U.S. EPA, 2006b). Sections 9.3 through 

10 9.4 begin with a discussion of effects at the cellular and subcellular level followed by 

11 consideration of the O3 effects on plant and ecosystem processes (Figure 9-1). In Section 

12 9.3, research is reviewed from the molecular to the biochemical and physiological levels 

13 in impacted plants, offering insight into the mode of action of O3. Section 9.4 provides a 

14 review of the effects of O3 exposure on major endpoints at the whole plant scale 

15 including growth, reproduction, visible foliar injury and leaf gas exchange in woody and 

16 herbaceous plants in the U.S., as well as a brief discussion of O3 effects on agricultural 

17 crop yield and quality. Section 9.4 also integrates the effects of O3 on individual plants in 

18 a discussion of available research for assessing the effect of O3 on ecosystems, along 

19 with available studies that could inform assessments of various ecosystem services (See 

20 section 9.4.1.2). The development of indices of O3 exposure and dose modeling is 

21 discussed in Section 9.5. Finally, exposure-response relationships for a number of tree 

22 species, native vegetation, and crop species and cultivars are reviewed, tabulated, and 

23 compared in Section 9.6 to form the basis for an assessment of the potential risk to 

24 vegetation from current ambient levels of O3.
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Figure 9-1 An illustrative diagram of the major pathway through which O3

enters plants and the major endpoints that O3 may affect in plants 

and ecosystems.

9.2 Experimental Exposure Methodologies

9.2.1 Introduction

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

2 over the last several decades. Methodological advancements since 2006 have not 

3 fundamentally altered our understanding of O3 effects on plants or ecosystems. The 

4 majority of methodologies currently used have been discussed in detail in the 1996 O3

5 AQCD and 2006 O3 AQCD. This section will serve as a short overview of the 

6 methodologies and the reader is referred to the previous O3 AQCDs for more in-depth 

7 discussion. 

O3 exposure

O3 uptake & physiology (Fig 9-2) •Antioxidant metabolism up-regulated

•Decreased photosynthesis

•Decreased stomatal conductance 

or sluggish stomatal response

Effects on leaves

•Visible leaf injury

•Altered leaf production 

•Altered leaf chemical composition

Plant growth (Fig 9.8) •Decreased biomass accumulation

•Altered reproduction

•Altered carbon allocation

•Altered crop quality

Belowground processes (Fig 9.8) •Altered litter production and decomposition

•Altered soil carbon and nutrient cycling

•Altered soil fauna and microbial communities

Differential O3 sensitivity

Affected ecosystem services •Decreased productivity 

•Decreased C sequestration

•Altered water cycling (Fig 9-7)

•Altered community composition 

(i.e., plant, insect & microbe)
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9.2.2 “Indoor,” Controlled Environment, and Greenhouse Chambers

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

2 glass or plastic-covered chambers, often located within greenhouses, into which a flow of 

3 O3-enriched air or oxygen could be passed to provide the exposure. The types, shapes, 

4 styles, materials of construction, and locations of these chambers have been numerous. 

5 Hogsett et al. (1987a) have summarized the construction and performance of more 

6 elaborate and better instrumented chambers since the 1960s, including those installed in 

7 greenhouses (with or without some control of temperature and light intensity).

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

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

10 stirred tank reactors (CSTRs) first described by Heck et al. (1978). Although originally 

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

12 widened to include short-term physiological and growth studies of O3 × CO2 interactions 

13 (Loats and Rebbeck, 1999; Reinert et al., 1997; Rao et al., 1995; Reinert and Ho, 1995; 

14 Heagle et al., 1994a), and validation of visible foliar injury on a variety of plant species 

15 (Kline et al., 2009; Orendovici et al., 2003). In many cases, supplementary lighting and 

16 temperature control of the surrounding structure have been used to control or modify the 

17 environmental conditions (Heagle et al., 1994a).

18 Many investigations have utilized commercially available controlled environment 

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

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

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

22 modes of action. For example, some of the studies of the O3 responses of common 

23 plantain (Plantago major) populations have been conducted in controlled environment 

24 chambers (Whitfield et al., 1996; Reiling and Davison, 1994).

25 More recently, some researchers have been interested in attempting to investigate direct 

26 O3 effects on reproductive processes, separate from the effects on vegetative processes 

27 (Black et al., 2010). For this purpose, controlled exposure systems have been employed 

28 to expose the reproductive structures of annual plants to gaseous pollutants independently 

29 of the vegetative component (Black et al., 2010; Stewart et al., 1996).

9.2.3 Field Chambers

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

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

32 (1973). The OTC method continues to be a widely used technique in the U.S. and Europe 
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1 for exposing plants to varying levels of O3. Most of the new information confirms earlier 

2 conclusions and provides additional support for OTC use in assessing plant species and in 

3 developing exposure-response relationships. Chambers are generally ~3 m in diameter 

4 with 2.5-m-high walls. Hogsett et al. (1987b) described in detail many of the various 

5 modifications to the original OTC designs that appeared subsequently, e.g., the use of 

6 larger chambers for exposing small trees (Kats et al., 1985) or grapevines (Mandl et al., 

7 1989), the addition of a conical baffle at the top to improve ventilation (Kats et al., 1976), 

8 a frustum at the top to reduce ambient air incursions, and a plastic rain-cap to exclude 

9 precipitation (Hogsett et al., 1985). All versions of OTCs included the discharge of air via 

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

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

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

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

14 several above ambient concentrations for O3 experiments. Depending on experimental 

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

16 experimental designs. The OTC can provide a consistent, definable exposure because of 

17 the constant wind speed and delivery systems. Statistically robust concentration-response 

18 (C-R) functions can be developed using such systems for evaluating the implications of 

19 various alternative air quality scenarios on vegetation response. Nonetheless, there are 

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

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

22 subjected to constant air flow turbulence, which, by lowering the boundary layer 

23 resistance to diffusion, may result in increased uptake. This may lead to an 

24 overestimation of effects relative to areas with less turbulence (Krupa et al., 1995; Legge 

25 et al., 1995). However, other research has found that OTC’s may slightly change vapor 

26 pressure deficit (VPD) in a way that may decrease the uptake of O3 into leaves (Piikki et 

27 al., 2008b). As with all methods that expose vegetation to modified O3 concentrations in 

28 chambers, OTCs create internal environments that differ from ambient air. This so-called 

29 “chamber effect” refers to the modification of microclimatic variables, including reduced 

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

31 and increased air temperatures (Fuhrer, 1994; Manning and Krupa, 1992). However, in at 

32 least one case where canopy resistance was quantified in OTCs and in the field, it was 

33 determined that gaseous pollutant exposure to crops in OTCs was similar to that which 

34 would have occurred at the same concentration in the field (Unsworth et al., 1984a, b). 

35 Because of the standardized methodology and protocols used in National Crop Loss 

36 Assessment Network (NCLAN) and other programs, the database can be assumed to be 

37 internally consistent.
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1 While it is clear that OTCs can alter some aspects of the microenvironment and plant 

2 growth, it is important to establish whether or not these differences affect the relative 

3 response of a plant to O3. As noted in the 1996 O3 AQCD, evidence from a number of 

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

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

6 significantly affect response. For example, a study of chamber effects examined the 

7 responses of tolerant and sensitive white clover clones (Trifolium repens) to ambient O3

8 in greenhouse, open top, and ambient plots (Heagle et al., 1996). The response found in 

9 OTCs was the same as in ambient plots.

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

11 recent studies (Lee et al., 2009a). Concern over the need to establish realistic plant-litter12 soil relationships as a prerequisite to studies of the effects of O3 and CO2 enrichment on 

13 ponderosa pine (Pinus ponderosa) seedlings led Tingey et al. (1996) to develop closed, 

14 partially environmentally controlled, sun-lit chambers (“terracosms”) incorporating 1-m15 deep lysimeters containing forest soil in which the appropriate horizon structure was 

16 retained.

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

18 called recirculating Outdoor Plant Environment Chambers (OPECs) (Flowers et al., 

19 2007). These closed chambers are approximately 2.44 m×1.52 m with a growth volume 

of approximately 3.7 m3 20 in each chamber. These chambers admit 90% of full sunlight and 

21 control temperature, humidity and vapor pressure (Fiscus et al., 1999).

9.2.4 Plume and FACE-Type Systems

22 Plume systems are chamberless exposure facilities in which the atmosphere surrounding 

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

24 around them from multiple orifices spaced to permit diffusion and turbulence, so as to 

25 establish relatively homogeneous conditions as the individual plumes disperse and mix 

26 with the ambient air. They can only be used to increase the O3 levels in the ambient air.

27 The most common plume system used in the U.S. is a modification of the free-air carbon28 dioxide/ozone enrichment (FACE) system (Hendrey et al., 1999; Hendrey and Kimball, 

29 1994). Although originally designed to provide chamberless field facilities for studying 

30 the CO2 effects of climate change, FACE systems have been adapted to include the 

31 dispensing of O3 (Karnosky et al., 1999). This method has been employed in Illinois 

32 (SoyFACE) to study soybeans (Morgan et al., 2004; Rogers et al., 2004) and in 

33 Wisconsin (Aspen FACE) to study trembling aspen (Populus tremuloides), birch (Betula 

34 papyrifera) and maple (Acer saccharum) (Karnosky et al., 1999). Volk et al. (2003) also 
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1 described a similar system for exposing grasslands that uses 7-m diameter plots. FACE 

2 systems discharge the pollutant gas (O3 and/or CO2) through orifices spaced along an 

3 annular ring (or torus) or at different heights on a ring of vertical pipes. Computer4 controlled feedback from the monitoring of gas concentration regulates the feed rate of 

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

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

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

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

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

10 throughputs tends to limit the number of enrichment treatments, although Hendrey et al. 

11 (1999) argued that the cost on an enriched volume basis is comparable to that of chamber 

12 systems.

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

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

15 the appropriate numbers, sizes, and orientations of the discharge orifices to avoid “hot16 spots” resulting from the direct impingement of jets of pollutant-enriched air on plant 

17 foliage (Werner and Fabian, 2002). Because mixing is unassisted and completely 

18 dependent on wind turbulence and diffusion, local gradients are inevitable especially in 

19 large-scale systems. FACE systems have provisions for shutting down under low wind 

20 speed or calm conditions and for an experimental area that is usually defined within a 

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

22 the treatment area. They are also dependent upon continuous computer-controlled 

23 feedback of the O3 concentrations in the mixed treated air and of the meteorological 

24 conditions. Plume and FACE systems also are unable to reduce O3 levels below ambient 

25 in areas where O3 concentrations are phytotoxic.

9.2.5 Ambient Gradients

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

27 responses over multiple levels of exposure. However, few such gradients can be found 

28 that meet the rigorous statistical requirements for comparable site characteristics such as 

29 soil type, temperature, rainfall, radiation, and aspect (Manning and Krupa, 1992); 

30 although with small plants, soil variability can be avoided by the use of plants in large 

31 pots. The use of soil monoliths transported to various locations along natural O3 gradients 

32 is another possible approach to overcome differences in soils; however, this approach is 

33 also limited to small plants.
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1 Studies in the 1970s used the natural gradients occurring in southern California to assess 

2 yield losses of alfalfa and tomato (Oshima et al., 1977; Oshima et al., 1976). A transect 

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

4 confounded by differences in the concurrent gradients of SO2 and NO2 pollution 

5 (Ashmore et al., 1988). Studies of forest tree species in national parks in the eastern U.S. 

6 (Winner et al., 1989) revealed increasing gradients of O3 and visible foliar injury with 

7 increased elevation.

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

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

10 and Jeffrey pine (Jones and Paine, 2006; Arbaugh et al., 2003; Grulke, 1999; Miller and 

11 Elderman, 1977). However, it is difficult to separate the effects of N and O3 in some 

12 instances in these studies (Arbaugh et al., 2003). An O3 gradient in Wisconsin has been 

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

14 differing in O3 sensitivity (Maňkovská et al., 2005; Karnosky et al., 1999). 

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

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

17 saplings grown in an urban to rural gradient of O3 in the New York City area. The 

18 secondary nature of the reactions of O3 formation and NOX titration reactions within the 

19 city center resulted in significantly higher cumulative O3 exposures in the rural sites. The 

20 results of this gradient study were similar to those of a parallel OTC study. Also, the U.S. 

21 forest service Forest Inventory and Analysis (FIA) program uses large-scale O3 exposure 

22 patterns across the continental U.S. to study occurrences of foliar injury due to O3

23 exposure (Smith et al., 2003) (Section 9.4.2). Finally, McLaughlin et al. (2007a; 2007b)

24 used spatial and temporal O3 gradients to study forest growth and water use in the 

25 southern Appalachians. These studies found varying O3 exposures between years and 

26 between sites.

9.2.6 Comparative Studies

27 All experimental approaches used to expose plants to O3 have strengths and weaknesses. 

28 One potential weakness of laboratory, greenhouse, or field chamber studies is the 

29 potential effect of the chamber on micrometeorology. In contrast, plume, FACE and 

30 gradient systems are limited by the very small number of possible exposure levels 

31 (almost always no more than two), small replication and an inability to reduce O3 levels 

32 below ambient. In general, experiments that aim at characterizing the effect of a single 

33 variable, e.g., exposure to O3, must not only manipulate the levels of that variable, but 

34 also control potentially interacting variables and confounders, or else account for them. 
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1 However, while increasing control of environmental variables makes it easier to discern 

2 the effect of the variable of interest, it must be balanced with the ability to extend 

3 conclusions to natural, non-experimental settings. More naturalistic exposure systems, on 

4 the other hand, let interacting factors vary freely, resulting in greater unexplainable 

5 variability. The various exposure methodologies used with O3 vary in the balance each 

6 strikes between control of environmental inputs, closeness to the natural environment, 

7 noisiness, and ability to make general inferences. 

8 Studies have examined the comparability of results obtained though the various exposure 

9 methodologies. As noted in the 1996 O3 AQCD, evidence from the comparative studies 

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

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

12 differences were found for fewer than 10% of the growth parameters measured, the 

13 responses were, in general, essentially the same regardless of exposure system used, and 

14 chamber effects did not significantly affect response. In 1988, Heagle et al. (1988)

15 concluded: “Although chamber effects on yield are common, there are no results showing 

16 that this will result in a changed yield response to O3.” A study of the effects of an 

17 enclosure examined the responses of tolerant and sensitive white clover clones (Trifolium 

18 repens) to ambient O3 in a greenhouse, open-top chamber, and ambient (no chamber) 

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

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

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

22 response of yield to O3 for plants grown in pots or in the ground, suggesting that even 

23 such a significant change in environment does not alter the proportional response to O3, 

24 providing that the plants are well watered (Heagle et al., 1983; Heagle, 1979).

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

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

27 FACE (Isebrands et al., 2001; Isebrands et al., 2000) showed that O3 symptom 

28 expression was generally similar in OTCs, FACE, and ambient O3 gradient sites, and 

29 supported the previously observed variation among trembling aspen clones using OTCs 

30 (Maňkovská et al., 2005; Karnosky et al., 1999). In the SoyFACE experiment in Illinois, 

31 soybean (Pioneer 93B15 cultivar) yield loss data from a two-year study was published 

32 (Morgan et al., 2006). This cultivar is a recent selection and, like most modern cultivars, 

33 has been selected under an already high current O3 exposure. It was found to have 

34 average sensitivity to O3 compared to 22 other cultivars tested at SoyFACE. In this 

35 experiment, ambient hourly O3 concentrations were increased by approximately 20% and 

36 measured yields were decreased by 15% in 2002 as a result of the increased O3 exposure 

37 (Morgan et al., 2006). To compare these results to chamber studies, Morgan et al. (2006)

38 calculated the expected yield loss from a linear relationship constructed from chamber 
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1 data using seven-hour seasonal averages (Ashmore, 2002). They calculated an 8% 

2 expected yield loss from the 2002 O3 exposure using that linear relationship. In another 

3 study, Gregg et al. (2006, 2003) found similar O3 effects on cottonwood sapling biomass 

4 growth along an ambient O3 gradient in the New York City area and a parallel OTC 

5 study.

6 Finally, EPA conducted comparisons of exposure-response model predictions based on 

7 OTC studies, and more recent FACE observations. These comparisons include yield of 

8 annual crops, and biomass growth of trees. They are presented in section 9.6.3 of this 

9 document.

9.3 Mechanisms Governing Vegetation Response to Ozone

9.3.1 Introduction

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

11 on the molecular, biochemical and physiological levels. First, the pathway of O3 uptake 

12 into the leaf and the initial chemical reactions occurring in the substomatal cavity and 

13 apoplast will be described (Section 9.3.2); additionally, direct effects of O3 on the 

14 stomatal apparatus will be discussed. Once O3 has entered the substomatal cavity and 

15 apoplast, it is thought that the cell must be able to sense the presence of O3 or its 

16 breakdown products in order to initiate the rapid changes in signaling pathways and gene 

17 expression that have been measured in O3-treated plants. While it remains unclear exactly 

18 how O3 and/or its breakdown products are sensed in the apoplast, much progress has been 

19 made in examining several different mechanisms that may contribute both to sensing the 

20 presence of O3 and its breakdown products, and also initiating a signal transduction 

21 cascade, which will be described in Section 9.3.3.1. The next section focuses on changes 

22 in gene and protein expression measured in plants exposed to O3, with particular 

23 emphasis on results from transcriptome (all RNA molecules produced in a cell) and 

24 proteome (all proteins produced in a cell) analyses (Section 9.3.3.2). Subsequently, the 

25 role of phytohormones such as salicylic acid (SA), ethylene (ET), jasmonic acid (JA), and 

26 abscisic acid (ABA) and their interactions in both signal transduction processes and in 

27 determining plant response to O3 is discussed in Section 9.3.3.3. After O3 uptake and 

28 sensing, some plants can respond to the oxidative stress with detoxification to minimize 

29 damage. These mechanisms of detoxification, with particular emphasis on antioxidant 

30 enzymes and metabolites, are reviewed in Section 9.3.4. The next section focuses on 

31 changes in primary and secondary metabolism in plants exposed to O3, looking at 

32 photosynthesis, respiration and several secondary metabolites, some of which may also 
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1 act as antioxidants and protect the plant from oxidative stress (Section 9.3.5). For many 

2 of these topics, information from the 2006 O3 AQCD has been summarized, as this 

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

4 genomics and proteomics, which have arisen due to the availability of new technologies, 

5 the information is based solely on new publications with no reference to the 2006 O3

6 AQCD.

7 As Section 9.3 focuses on mechanisms underlying effects of O3 on plants and their 

8 response to it, the conditions that are used to study these mechanisms do not always 

9 reflect conditions that a plant may be exposed to in an agricultural setting or natural 

10 ecosystem. The goal of many of these studies is to generate an O3 effect in a relatively 

11 short period of time and not always to simulate ambient O3 exposures. Therefore, plants 

12 are often exposed to unrealistically high O3 concentrations for several hours or days 

13 (acute exposure), and only in some cases to ambient or slightly elevated O3

14 concentrations for longer time periods (chronic exposure). Additionally, the plant species 

15 utilized in these studies are often not agriculturally important or commonly found as part 

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

17 as they are easy to work with, and mutants or transgenic plants are easy to develop or 

18 have already been developed. Furthermore, the Arabidopsis genome has been sequenced, 

19 and much is known about the molecular basis of many biochemical and cellular 

20 processes. 

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

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

23 gene expression) do not always translate into changes in protein quantity and/or activity, 

24 as there are many levels of post-transcriptional and post-translational modifications 

25 which impact protein quantity and activity. Many studies do not evaluate whether the 

26 observed changes in gene expression lead to changes at the protein level and, therefore, it 

27 is not always clear how relevant the changes in gene expression are in determining plant 

28 response to O3. However, with the advent of proteomics, some very recent studies have 

29 evaluated changes in protein expression for large numbers of proteins in O3 treated 

30 plants, and the findings from these studies support the previous results regarding changes 

31 in gene expression studies as a result of O3 exposure. The next step in the process is to 

32 determine the implications of the measured changes occurring at the cellular level to 

33 whole plants and ecosystems, which is an important topic of study which has not been 

34 widely addressed.

35 The most significant new body of research since the 2006 O3 AQCD is on the 

36 understanding of molecular mechanisms underlying how plants are affected by O3; a 

37 significant number of recent studies reviewed here focus on changes in gene expression 
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1 in plants exposed to elevated O3. Conclusions from the 2006 O3 AQCD have been 

2 supported by these new studies, and the advent of new technologies has allowed for a 

3 more comprehensive understanding of the mechanisms governing how plants are affected 

4 by O3. 

5 In summary, these new studies have increased knowledge of the molecular, biochemical 

6 and cellular mechanisms occurring in plants in response to O3 by often using artificial 

7 exposure conditions and model organisms. This information adds to the understanding of 

8 the basic biology of how plants are affected by oxidative stress in the absence of any 

9 other potential stressors. The results of these studies provide important insights, even 

10 though they may not always directly translate into effects observed in other plants under

11 more realistic exposure conditions. 

9.3.2 Ozone Uptake into the Leaf

12 Appendix AX9.2.3 of the 2006 O3 AQCD clearly described the process by which O3

13 enters plant leaves through open stomata (U.S. EPA, 2006b). This information continues 

14 to be valid and is only summarized here. 

15 Stomata provide the principal pathway for O3 to enter and affect plants (Massman and 

16 Grantz, 1995; Fuentes et al., 1992; Reich, 1987; Leuning et al., 1979). Ozone moves into 

17 the leaf interior by diffusing through open stomata, and environmental conditions which 

18 promote high rates of gas exchange will favor the uptake of the pollutant by the leaf. 

19 Factors that may limit uptake include boundary layer resistance and the size of the 

20 stomatal aperture (Figure 9-2) (U.S. EPA, 2006b). Once inside the substomatal cavity, O3

21 is thought to rapidly react with the aqueous apoplast to form breakdown products known 

22 as reactive oxygen species (ROS), such as hydrogen peroxide (H2O2), superoxide (O2 ), 

hydroxyl radicals (HO·) and peroxy radicals (HO2

· 23 ) (Figure 9-3). Hydrogen peroxide is 

24 not only a toxic breakdown product of O3, but has been shown to function as a signaling 

25 molecule, which is activated in response to both biotic and abiotic stressors. The role of 

26 H2O2 in signaling was described in detail in the 2006 O3 AQCD. Additional organic 

27 molecules present in the apoplast or cell wall, such as those containing double bonds or 

28 sulfhydryls that are sensitive to oxidation, could also be converted to oxygenated 

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

30 dependent, but are also influenced by the presence of other molecules in the apoplast 

31 (U.S. EPA, 2006b). The 2006 O3 AQCD provided a comprehensive summary of what is 

32 known about the possible interactions of O3 with other biomolecules (U.S. EPA, 2006b). 

33 It is in the apoplast that initial detoxification reactions by antioxidant metabolites and 

34 enzymes take place, and these initial reactions are critical to reduce concentrations of the 
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1 oxidative breakdown products of O3; these reactions are described in more detail in 

2 Section 9.3.4 of this document. 

9.3.2.1 Changes in Stomatal Function

3 The effects of O3 exposure on stomatal conductance have been reviewed in detail in

4 previous O3 AQCDs. Although the nature of these effects depends upon many different 

5 factors, including the plant species, concentration and duration of the O3 exposure, and 

6 prevailing meteorological conditions, stomatal conductance is often negatively affected 

7 by plant exposure to O3 (Wittig et al., 2007). Decreases in conductance have been shown 

8 to result from declines in photosynthetic carboxylation capacity, leading to a buildup of 

9 CO2 in the substomatal cavity and subsequent stomatal closure (Wittig et al., 2007). 

10 However, results from the use of Arabidopsis mutants and new technologies, which allow 

11 for analysis of guard cell function in whole plants rather than in isolated guard cells or 

12 epidermal peels, suggest that O3 may also have a direct impact on stomatal guard cells, 

13 leading to alterations in stomatal conductance. The use of a new simultaneous O3

14 exposure/gas exchange device has demonstrated that exposure of Arabidopsis ecotypes 

15 Col-0 and Ler to 150 ppb O3 resulted in a 60-70% decline in stomatal conductance within 

16 9-12 minutes of beginning the exposure. Twenty to thirty minutes later, stomatal 

17 conductance had returned to its initial value, even with continuing exposure to O3, 

18 indicating a rapid direct effect of O3 on stomatal function (Kollist et al., 2007). This 

19 transient decrease in stomatal conductance was not observed in the abscisic acid 

20 insensitive (ABI2) Arabidopsis mutant. As the ABI2 protein is thought to regulate the 

21 signal transduction process involved in stomatal response downstream of ROS 

22 production, the authors suggest that the transient decrease in stomatal conductance in the 

23 Col-0 and Ler ecotypes results from the biological action of ROS in transducing signals, 

24 rather than direct physical damage to guard cells by ROS (Kollist et al., 2007). This rapid 

25 transient decrease in stomatal conductance was also not observed when exposing the 

26 Arabidopsis mutant slac1 (slow anion channel-associated 1) to 200 ppb O3 (Vahisalu et 

27 al., 2008). The SLAC1 protein was shown to be essential for guard cell slow anion 

28 channel functioning and for stomatal closure in response to O3. Based on additional 

29 studies using a variety of Arabidopsis mutants impaired in various aspects of stomatal 

30 function, Vahisalu et al. (2010) suggest that the presence of ROS in the guard cell 

31 apoplast (formed either by O3 breakdown or through ROS production from NADPH 

32 oxidase activity) leads to the activation of a signaling pathway in the guard cells, which 

33 includes SLAC1, and results in stomatal closure.

34 A review by McAinsh et al. (2002) discusses the role of calcium as a part of the signal 

35 transduction pathway involved in regulating stomatal responses to pollutant stress. A
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1 number of studies in this review provide some evidence that exposure to O3 increases the 

cytosolic free calcium concentration ([Ca2+ 2 ]cyt) in guard cells, which may result in an 

inhibition of the plasma membrane inward-rectifying K+ 3 channels in guard cells, which 

allow for the K+ 4 uptake needed for stomatal opening (McAinsh et al., 2002; Torsethaugen 

5 et al., 1999). This would compromise the ability of the stomata to respond to various 

6 stimuli, including light, CO2 concentration and drought. Pei et al. (2000) reported that the 

presence of H2O2 activated Ca2+ 7 -permeable channels, which mediate increases in 

[Ca2+ 8 ]cyt in guard cell plasma membranes of Arabidopsis. They also determined that 

9 abscisic acid (ABA) induced H2O2 production in guard cells, leading to ABA-induced 

stomatal closure via activation of the membrane Ca2+ 10 channels. Therefore, it is possible 

that H2O2, a byproduct of O3 breakdown in the apoplast, could disrupt the Ca2+ 11 -ABA 

12 signaling pathway that is involved in regulating stomatal responses (McAinsh et al., 

13 2002). The studies described here provide some evidence to suggest that O3 and its 

14 breakdown products can directly affect stomatal functioning by impacting the signal 

15 transduction pathways which regulate guard cells. Stomatal sluggishness has been 

16 described as a delay in stomatal response to changing environmental conditions in 

17 sensitive species exposed to higher concentrations and/or longer-term O3 exposures 

18 (Paoletti and Grulke, 2010, 2005; McAinsh et al., 2002). It is possible that the signaling 

19 pathways described above could be involved in mediating this stomatal sluggishness in 

20 some plant species under certain O3 exposure conditions (Paoletti and Grulke, 2005; 

21 McAinsh et al., 2002).
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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.
Draft – Do Not Cite or Quote 9-16 September 2011

Source: Adapted from Mudd (1996).

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.
Draft – Do Not Cite or Quote 9-17 September 2011

9.3.3 Cellular to Systemic Responses

9.3.3.1 Ozone Sensing and Signal Transduction

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

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

3 with the perception and integration of responses to the stress. Many of these studies have 

4 been conducted using Arabidopsis or tobacco plants, for which a variety of mutants are 

5 available and/or which can be easily genetically modified to generate either loss-of6 function or over-expressing genotypes. Several comprehensive review articles provide an 

7 overview of what is known of O3-induced signal transduction processes and how they 

8 may help to explain differential sensitivity of plants to the pollutant (Ludwikow and 

9 Sadowski, 2008; Baier et al., 2005; Kangasjarvi et al., 2005). Additionally, analysis of 

10 several studies of transcriptome changes has also allowed for the compilation of these 

11 data to determine an initial time-course for O3-induced activation of various signaling 

12 compounds (Kangasjarvi et al., 2005).

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

14 there is still much that is not known about this process. Some of the earliest events that 

15 occur in plants exposed to O3 have been described in the guard cells of stomata. Reactive 

16 oxygen species were observed in the chloroplasts of guard cells in the O3 tolerant Col-0 

17 Arabidopsis thaliana ecotype plants within 5 minutes of plant exposure to 350 ppb O3

18 (Joo et al., 2005). Reactive oxygen species from the breakdown of O3 in the apoplast are 

19 believed to activate GTPases (G-proteins), which, in turn, activate several intracellular 

20 sources of ROS, including ROS derived from the chloroplasts. G-proteins are also 

21 believed to play a role in activating membrane-bound NADPH oxidases to produce ROS 

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

23 Therefore, G-proteins are recognized as important molecules involved in plant responses 

24 to O3 and may play a role in perceiving ROS from the breakdown of O3 in the apoplast 

25 (Kangasjarvi et al., 2005; Booker et al., 2004b). 

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

27 may represent a means of perception and signaling of oxidative stress in plants. 

28 Disulfide-thiol conversions in proteins and the redox state of the glutathione pool may be 

29 important components of redox sensing and signal transduction (Foyer and Noctor, 

30 2005a, b). 

Calcium (Ca2+ 31 ) has also been implicated in the transduction of signals to the nucleus in 

response to oxidative stress. The influx of Ca2+ 32 from the apoplast into the cell occurs 

33 early during plant exposure to O3, and it is thought to play a role in regulating the activity 
Draft – Do Not Cite or Quote 9-18 September 2011

1 of protein kinases, which are discussed below (Baier et al., 2005; Hamel et al., 2005). 

2 Calcium channel blockers inhibited O3-induced activation of protein kinases in tobacco 

3 suspension cells exposed to 500 ppb O3 for 10 minutes, indicating that the opening of 

Ca2+ 4 channels is an important upstream signaling event or that the as yet unknown 

upstream process has a requirement for Ca2+ 5 (Samuel et al., 2000). 

6 Further transmission of information regarding the presence of ROS to the nucleus t 

7 involves mitogen-activated protein kinases (MAPK), which phosphorylate proteins and 

8 activate various cellular responses (Hamel et al., 2005). Mitogen-activated protein 

9 kinases are induced in several different plant species in response to O3 exposure, 

10 including tobacco (Samuel et al., 2005), Arabidopsis (Ludwikow et al., 2004), the shrub 

11 Phillyrea latifolia (Paolacci et al., 2007) and poplar (Hamel et al., 2005). Disruption of 

12 these signal transduction pathways by over-expressing or suppressing MAP kinase 

13 activity in different Arabidopsis and tobacco lines resulted in increased plant sensitivity 

14 to O3 (Miles et al., 2005; Samuel and Ellis, 2002). Additionally, greater O3 tolerance of 

15 several Arabidopsis ecotypes was correlated with greater up-regulation of MAP kinase 

16 signaling pathways upon O3 exposure than in more sensitive Arabidopsis ecotypes (Li et 

17 al., 2006b; Mahalingam et al., 2006; Overmyer et al., 2005), indicating that determination 

18 of plant sensitivity and plant response to O3 may, in part, be determined not only by 

19 whether these pathways are turned on, but also by the magnitude of the signals moving 

20 through these communication channels.

21 In conclusion, experimental evidence suggests that there are likely several different 

22 mechanisms by which the plant senses the presence of O3 or its breakdown products. 

23 These mechanisms may vary by species or developmental stage of the plant, or may co24 exist and be activated by different exposure conditions. Calcium and protein kinases are 

25 likely involved in relaying information about the presence of the stressor to the nucleus 

26 and other cellular compartments as a first step in determining whether and how the plant 

27 will respond to the stress. 

9.3.3.2 Gene and Protein Expression Changes in Response to 

Ozone

28 The advent of DNA microarray technology has allowed for the study of gene expression 

29 in cells on a large scale. Rather than assessing changes in gene expression of individual 

30 genes, DNA microarrays facilitate the evaluation of entire transcriptomes, providing a 

31 comprehensive picture of simultaneous alterations in gene expression. In addition, these 

32 studies have provided more insight into the complex interactions between molecules, how 

33 those interactions lead to the communication of information in the cell (or between 
Draft – Do Not Cite or Quote 9-19 September 2011

1 neighboring cells) , and which role these interactions play in determining tolerance or 

2 sensitivity and how a plant may respond to stresses such as O3 (Ludwikow and 

3 Sadowski, 2008). Transcriptome analysis of O3-treated plants has been performed in 

4 several species, including Arabidopsis thaliana (Li et al., 2006b; Tosti et al., 2006; 

5 Heidenreich et al., 2005; Mahalingam et al., 2005; Tamaoki et al., 2003), pepper 

6 (Capsicum annuum) (Lee and Yun, 2006), clover (Medicago truncatula) (Puckette et al., 

7 2008), Phillyrea latifolia (Paolacci et al., 2007), poplar (Street et al., 2011), and European 

8 beech (Fagus sylvatica) (Olbrich et al., 2010; Olbrich et al., 2009; Olbrich et al., 2005). 

9 In some cases, researchers compared transcriptomes of two or more cultivars, ecotypes or 

10 mutants that differed in their sensitivity to O3 (Puckette et al., 2008; Rizzo et al., 2007; 

11 Lee and Yun, 2006; Li et al., 2006b; Tamaoki et al., 2003). Species, O3 exposure 

12 conditions (concentration, duration of exposure) and sampling times varied significantly 

13 in these studies. However, functional classification of the genes that were either up- or 

14 down-regulated by plant exposure to O3 exhibited common trends. Genes involved in 

15 plant defense, signaling and those associated with the synthesis of plant hormones and 

16 secondary metabolism were generally up-regulated, while those related to photosynthesis 

17 and general metabolism were typically down-regulated in O3-treated plants (Puckette et 

18 al., 2008; Lee and Yun, 2006; Li et al., 2006b; Tosti et al., 2006; Olbrich et al., 2005; 

19 Tamaoki et al., 2003). 

20 Analysis of the transcriptome has been used to evaluate differences in gene expression 

21 between O3 sensitive and tolerant plants. In pepper, 67% of the 180 genes studied that 

22 were affected by O3 were differentially regulated in the sensitive and tolerant cultivars. 

23 At both 0 hours and 48 hours after a 3-day exposure at 150 ppb, O3 responsive genes 

24 were either up- or down-regulated more markedly in the sensitive than in the tolerant 

25 cultivar (Lee and Yun, 2006). Transcriptome analysis also revealed differences in timing 

26 and magnitude of changes in gene expression between sensitive and tolerant clovers. 

27 Acute exposure (300 ppb O3 for 6 hours) led to the production of an oxidative burst in 

28 both clovers (Puckette et al., 2008). However, the sensitive Jemalong cultivar exhibited a 

29 sustained ROS burst and a concomitant down-regulation of defense response genes at 

30 12 hours after the onset of exposure, while the tolerant JE 154 accession showed much 

31 more rapid and large-scale transcriptome changes than the Jemalong cultivar (Puckette et 

32 al., 2008). 

33 Arabidopsis ecotypes WS and Col-0 were exposed to 1.2 × ambient O3 concentrations for 

34 8-12 days at the SoyFACE site (Li et al., 2006b). The sensitive WS ecotype showed a far 

35 greater number of changes in gene expression in response to this low-level O3 exposure 

36 than the tolerant Col-0 ecotype. In a different study, exposure of the WS ecotype to 

37 300 ppb O3 for 6 hours showed a rapid induction of genes leading to cell death, such as 
Draft – Do Not Cite or Quote 9-20 September 2011

1 proteases, and down-regulation or inactivation of cell signaling genes, demonstrating an 

2 ineffective defense response in this O3 sensitive ecotype (Mahalingam et al., 2006). 

3 The temporal response of plants to O3 exposure was evaluated in the Arabidopsis Col-0 

4 ecotype during a 6-h exposure at 350 ppb O3 and for 6 hours after the exposure was 

5 completed. Results of this study, shown in Figure 9-5, indicate that genes associated with 

6 signal transduction and regulation of transcription were in the class of early up-regulated 

7 genes, while genes associated with redox homeostasis and defense/stress response were 

8 in the class of late up-regulated genes (Mahalingam et al., 2005). 

9 A few studies have been conducted to evaluate transcriptome changes in response to 

10 longer term chronic O3 exposures in woody plant species. Longer term exposures 

11 resulted in the up-regulation of genes associated with secondary metabolites, including 

12 isoprenoids, polyamines and phenylpropanoids in 2-year-old seedlings of the 

13 Mediterranean shrub Phillyrea latifolia exposed to 110 ppb O3 for 90 days (Paolacci et 

14 al., 2007). In 3-year-old European beech saplings exposed to O3 for 20 months, with 

15 monthly average twice ambient O3 concentrations ranging from 11 to 80 ppb, O3-induced 

16 changes in gene transcription were similar to those observed for herbaceous species 

17 (Olbrich et al., 2009). Genes encoding proteins associated with plant stress response, 

18 including ethylene biosynthesis, pathogenesis-related proteins and enzymes detoxifying 

19 ROS, were up-regulated. Some genes associated with primary metabolism, cell structure, 

20 cell division and cell growth were reduced (Olbrich et al., 2009). In a similar study using 

21 adult European beech trees, it was determined that the magnitude of the transcriptional 

22 changes described above was far greater in the saplings than in the adult trees exposed to 

23 the same O3 concentrations for the same time period (Olbrich et al., 2010). 

24 The results from transcriptome studies described above have been substantiated by results 

25 from proteome analysis in rice, poplar, European beech, wheat, and soybean. Exposure of 

26 soybean to 120 ppb O3 for 12-h/day for 3 days in growth chambers resulted in decreases 

27 in the quantity of proteins associated with photosynthesis, while proteins involved with 

28 antioxidant defense and carbon metabolism increased (Ahsan et al., 2010). Young poplar 

29 plants exposed to 120 ppb O3 in a growth chamber for 35 days also showed significant 

30 changes in proteins involved in carbon metabolism (Bohler et al., 2007). Declines in 

31 enzymes associated with carbon fixation, the Calvin cycle and photosystem II were 

32 measured, while ascorbate peroxidase and enzymes associated with glucose catabolism 

33 increased in abundance. In another study to determine the impacts of O3 on both 

34 developing and fully expanded poplar leaves, young poplars were exposed to 120 ppb O3

35 for 13-h per day for up to 28 days (Bohler et al., 2010). Impacts on protein quantity only 

36 occurred after the plants had been exposed to O3 for 14 days, and at this point in time, 

37 several Calvin cycle enzymes were reduced in quantity, while the effects on the light 
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1 reactions appeared later, at 21 days after beginning treatment. Some of the antioxidant 

enzymes increased in abundance with O3 2 treatment, while others (ascorbate peroxidase) 

3 did not. In relationship to leaf expansion, it was shown that O3 did not affect protein 

4 quantity until leaves had reached full expansion, after about 7 days (Bohler et al., 2010). 

5 Two-week-old rice seedlings exposed to varying levels of O3 (4, 40, 80, 120 ppb) in a 

6 growth chamber for 9 days showed reductions in quantities of proteins associated with 

7 photosynthesis and energy metabolism, and increases in some antioxidant and defense 

8 related proteins (Feng et al., 2008a). A subsequent study of O3-treated rice seedlings 

9 (exposed to 200 ppb O3 for 24-h) focusing on the integration of transcriptomics and 

10 proteomics, supported and further enhanced these results (Cho et al., 2008). The authors 

11 found that of the 22,000 genes analyzed from the rice genome, 1,535 were differentially 

12 regulated by O3. Those differentially regulated genes were functionally categorized as 

13 transcription factors, MAPK cascades, those encoding for enzymes involved in the 

14 synthesis of jasmonic acid (JA),ethylene (ET), shikimate, tryptophan and lignin, and 

15 those involved in glycolysis, the citric acid cycle, oxidative respiration and 

16 photosynthesis. The authors determined that the proteome and metabolome (all small 

17 molecule metabolites in a cell) analysis supported the results of the transcriptome 

18 changes described above (Cho et al., 2008). This type of study, which ties together results 

19 from changes in gene expression, protein quantity and activity, and metabolite levels, 

20 provides the most complete picture of the molecular and biochemical changes occurring 

21 in plants exposed to a stressor such as O3. 

22 Sarkar et al. (2010) compared proteomes of two cultivars of wheat grown in OTCs at 

23 several O3 concentrations, including filtered air, ambient O3 (mean concentration 

24 47 ppb), ambient + 10 ppb and ambient + 20 ppb for 5-h/day for 50 days. Declines in the 

25 rate of photosynthesis and stomatal conductance were related to decreases in proteins 

26 involved in carbon fixation and electron transport and increased proteolysis of 

27 photosynthetic proteins such as the large subunit of ribulose-1,6-bisphosphate 

28 carboxylase/oxygenase (Rubisco). Enzymes that take part in energy metabolism, such as 

29 ATP synthesis, were also down-regulated, while defense/stress related proteins were up30 regulated in O3 treated plants. In comparing the two wheat cultivars, Sarkar et al. (2010)

31 found that while the qualitative changes in protein expression between the two cultivars 

32 were similar, the magnitude of these changes differed between the sensitive and tolerant 

33 wheat cultivars. Greater foliar injury and a smaller decline in stomatal conductance was 

34 observed in the sensitive cultivar as compared to the more tolerant cultivar, along with 

35 greater losses in photosynthetic enzymes and higher quantities of antioxidant enzymes. 

36 Results from a three year exposure of European beech saplings to elevated O3 (AOT 40 

37 value was 52.6 ul l-1-h for 2006 when trees were sampled) supported the results from the 

38 short-term exposure studies described above (Kerner et al., 2011). The O3 treatment of 
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1 the saplings resulted in reductions in enzymes associated with the Calvin cycle, which 

2 could lead to reduced carbon fixation. Enzymes associated with carbon 

3 metabolism/catabolism were increased, and quantities of starch and sucrose were reduced 

4 in response to the O3 treatment in these trees, indicating a potential impact of O3 on 

5 overall carbon metabolism in long-term exposure conditions (Kerner et al., 2011).
Draft – Do Not Cite or Quote 9-23 September 2011

Source: Used with permission from Springer (Mahalingam et al., 2005).

(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.

Figure 9-5 Composite diagram of major themes in the temporal evolution of 

1 the genetic response to ozone stress. All of these studies describe common 

2 trends for changes in gene and protein expression which occur in a variety of plant 

3 species exposed to O3. While genes associated with carbon assimilation and general 

4 metabolism are typically down-regulated, genes associated with signaling, catabolism, 

5 and defense are up-regulated. The magnitude of these changes in gene and protein 

6 expression appears to be related to plant species, age and their sensitivity or tolerance to 

7 O3. averaged (normalized) 

transcript profiles for gene groups
Draft – Do Not Cite or Quote 9-24 September 2011

9.3.3.3 Role of Phytohormones in Plant Response to Ozone

1 Many studies of O3 effects on plants have analyzed the importance of plant hormones 

2 such as SA, ET and JA in determining plant response to O3; some of the roles of these 

3 hormones were described in the 2006 O3 AQCD. Transcriptome analysis and the use of a 

4 variety of mutants have allowed for further elucidation of the complex interactions 

5 between SA, ET, JA and the role of abscisic acid (ABA) in mediating plant response to 

6 O3 (Ludwikow and Sadowski, 2008). In addition to their roles in signaling pathways, 

7 phytohormones also appear to regulate, and be regulated by, the MAPK signaling 

8 cascades described previously. Most evidence suggests that while ET and SA are needed 

9 to develop O3-induced leaf lesions, JA acts antagonistically to SA and ET to limit the 

10 lesions (Figure 9-6) (Kangasjarvi et al., 2005). 

11 The rapid production of ET in O3 treated plants has been described in many plant species 

12 and has been further characterized through the use of a variety of mutants that either 

13 over-produce or are insensitive to ET. Production of stress ET in O3-treated plants, which 

14 is thought to be part of a wounding response, was found to be correlated to the degree of 

15 injury development in leaves (U.S. EPA, 2006b). More recent studies have supported 

16 these conclusions and have also focused on the interactions occurring between several 

17 oxidative-stress induced phytohormones. Yoshida et al. (2009) determined that ET likely 

18 amplifies the oxidative signal generated by ROS, thereby promoting lesion formation. By 

19 analyzing the O3-induced transcriptome of several Arabidopsis mutants of the Col-0 

20 ecotype, Tamaoki et al. (2003) determined that at 12 hours after initiating the O3

21 exposure (200 ppb for 12 hours), the ET and JA signaling pathways were the main 

22 pathways used to activate plant defense responses, with a lesser role for SA. The authors 

23 also demonstrated that low levels of ET production could stimulate the expression of 

24 defense genes, rather than promoting cell death which occurs when ET production is 

25 high. Tosti et al. (2006) supported these findings by showing that plant exposure to O3

26 not only results in activation of the biosynthetic pathways of ET, JA and SA, but also 

27 increases the expression of genes related to the signal transduction pathways of these 

28 phytohormones in O3-treated Arabidopsis plants (300 ppb O3 for 6 hours). Conversely, in 

29 the O3 sensitive Ws ecotype, its sensitivity may, in part, be due to intrinsically high ET 

30 levels leading to SA accumulation, and the high ET and SA may act to repress JA31 associated genes, which would serve to inhibit the spread of lesions (Mahalingam et al., 

32 2006). Ogawa et al. (2005) found that increases in SA in O3-treated plants leads to the 

33 formation of leaf lesions in tobacco plants exposed to 200 ppb O3 for 6 hours. 

34 Furthermore, in transgenic tobacco plants with reduced levels of ET production in 

35 response to O3 exposure, several genes encoding for enzymes in the biosynthetic pathway 

36 of SA were suppressed, suggesting that SA levels are, in part, controlled by ET in the 

37 presence of O3.
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1 Exposure of the Arabidopsis mutant rcd1 to acute doses of O3 (250 ppb O3 for 8-h/day 

2 for 3 days) resulted in programmed cell death (PCD) and the formation of leaf lesions 

3 (Overmyer et al., 2000). They determined that the observed induction of ET synthesis 

4 promotes cell death, and that ET perception and signaling are required for the 

5 accumulation of superoxide, which leads to cell death and propagation of lesions. . 

6 Jasmonic acid, conversely, contains the spread of leaf lesions (Overmyer et al., 2000). 

7 Transcriptome analysis of several Arabidopsis mutants, which are insensitive to SA, ET 

8 and JA, exposed to 12-h of 200 ppb O3 showed that approximately 78 of the up-regulated 

9 genes measured in this study were controlled by ET and JA signaling pathways, while SA 

10 signaling pathways were suggested to antagonize ET and JA pathways (Tamaoki et al., 

11 2003). In a subsequent transcriptome study on the Col-0 ecotype exposed to 150 ppb O3

12 for 48-h, JA and ET synthesis was down-regulated, while SA was up-regulated in O3-

13 treated plants. In cotton plants exposed to a range of O3 concentrations (0-120 ppb) and 

14 methyl jasmonate (MeJA), Grantz et al. (2010a) determined that exogenous applications 

15 of MeJA did not protect plants from chronic O3 exposure. 

16 Abscisic acid has been investigated for its role in regulating stomatal aperture and also 

17 for its contribution to signaling pathways in the plant. The role of ABA and the 

18 interaction between ABA and H2O2 in O3-induced stomatal closure was described in the 

19 2006 O3 AQCD. More recently, it was determined that synthesis of ABA was induced in 

20 O3-treated Arabidopsis plants (250-350 ppb O3 for 6 hours), with a more pronounced 

21 induction in the O3 sensitive rcd3 mutant as compared to the wildtype Col-0 (Overmyer 

22 et al., 2008). The rcd3 mutant also exhibited a lack of O3-induced stomatal closure, and 

23 the RCD3 protein has been shown to be required for slow anion channels (Overmyer et 

24 al., 2008) (see Section 9.3.4.1). Ludwikow et al. (2009) used Arabidopsis ABI1td 

25 mutants, in which a key negative regulator of ABA action (abscisic acid insensitive1 

26 protein phosphatase 2C) has been knocked out, to examine O3 responsive genes in this 

27 mutant compared to the Arabidopsis Col-0. Results of this study indicate a role for ABI1 

28 in negatively regulating the synthesis of both ABA and ET in O3-treated plants (350 ppb 

29 O3 for 9 hours). Additionally, ABI1 may stimulate JA-related gene expression, providing 

30 evidence for an antagonistic interaction between ABA and JA signaling pathways 

31 (Ludwikow et al., 2009).

32 Nitric oxide (NO) has also been shown to play a role in regulating gene expression in 

33 plants in response to O3 exposure. However, little is known to date about NO and its role 

34 in the complex interactions of molecules in response to O3. Exposure of tobacco to O3

35 (150 ppb for 5 hours) stimulated NO and NO-dependent ET production, while NO 

36 production itself did not depend on the presence of ET (Ederli et al., 2006). Analysis of 

37 O3-treated Arabidopsis indicated the possibility of a dual role for NO in the initiation of 

38 cell death and later lesion containment (Ahlfors et al., 2009). 
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1 While much work remains to be done to better elucidate how plants sense O3, what 

2 determines their sensitivity to the pollutant and how they might respond to it, it is clear 

3 that the mechanism for O3 sensing and signal transduction is very complex. Many of the 

4 phytohormones and other signaling molecules thought to be involved in these processes 

5 are interactive and depend upon a variety of other factors, which could be either internal 

6 or external to the plant. This results in a highly dynamic and complex system, capable of 

7 resulting in a spectrum of plant sensitivity to oxidative stress and generating a variety of 

8 plant responses to that stress. 

Source: Used with permission from Blackwell Publishing Ltd. (Kangasjarvi et al., 2005).

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.

Figure 9-6 The oxidative cell death cycle. Detoxification

9.3.4.1 Overview of Ozone-Induced Defense Mechanisms

9 Plants are exposed to an oxidizing environment on a continual basis, and many reactions 

10 that are part of the basic metabolic processes, such as photosynthesis and respiration, 

11 generate ROS. As a result, there is an extensive and complex mechanism in place to 

12 detoxify these oxidizing radicals, including both enzymes and metabolites, which are 

13 located in several locations in the cell and also in the apoplast of the cell. As O3 enters 

14 the leaf through open stomata, the first point of contact of O3 with the plant is likely in 

the apoplast, where it breaks down to form oxidizing radicals such as H2O2, O2

- 15 , HO· and 
Draft – Do Not Cite or Quote 9-27 September 2011

1 HO2. Another source of oxidizing radicals is an oxidative burst, generated by a 

2 membrane-bound NADPH oxidase enzyme, which is recognized as an integral 

3 component of the plant’s defense system against pathogens (Schraudner et al., 1998). 

4 Antioxidant metabolites and enzymes located in the apoplast are thought to form a first 

5 line of defense by detoxifying O3 and/or the ROS that are formed as breakdown products 

6 of O3 (Section 9.3.2.). However, even with the presence of several antioxidants, 

7 including ascorbate, the redox buffering capacity of the apoplast is far less than that of 

8 the cytoplasm, as it lacks the regeneration systems necessary to retain a reduced pool of 

9 antioxidants (Foyer and Noctor, 2005b). 

10 Redox homeostasis is regulated by the presence of a pool of antioxidants, which are 

11 typically found in a reduced state and detoxify ROS produced by oxidases or electron 

12 transport components. As ROS increase due to environmental stress such as O3, it is 

13 unclear whether the antioxidant pool can maintain its reduced state (Foyer and Noctor, 

14 2005b). As such, not only the quantity and types of antioxidant enzymes and metabolites 

15 present, but also the cellular ability to regenerate those antioxidants are important 

16 considerations in mechanisms of plant tolerance to oxidative stress (Dizengremel et al., 

17 2008). Molecules such as glutathione (GSH), thioredoxins and NADPH play very 

18 important roles in this regeneration process; additionally, it has been hypothesized that 

19 alterations in carbon metabolism would be necessary to supply the needed reducing 

20 power for antioxidant regeneration (Dizengremel et al., 2008). 

9.3.4.2 Role of Antioxidants in Plant Defense Responses

21 Ascorbate has been the focus of many different studies as an antioxidant metabolite that 

22 protects plants from exposure to O3. It is found in several cellular locations, including the 

23 chloroplast, the cytosol and the apoplast (Noctor and Foyer, 1998). Ascorbate is 

24 synthesized in the cell and transported to the apoplast. Apoplastic ascorbate can be 

25 oxidized to dehydroascorbate (DHA) with exposure to O3 and is then transported back to 

26 the cytoplasm. Here, DHA is reduced to ascorbate by the enzyme dehydroascorbate 

27 reductase (DHAR) and reduced GSH, which is part of the ascorbate-glutathione cycle 

28 (Noctor and Foyer, 1998). Many studies have focused on evaluating whether ascorbate is 

29 the primary determining factor in differential sensitivity of plants to O3. An evaluation of 

30 several species of wildflowers in Great Smoky Mountains National Park showed a 

31 correlation between higher quantities of reduced apoplastic ascorbate and lower levels of 

32 foliar injury from O3 exposure in the field in tall milkweed plants (Asclepsias exaltata L.) 

33 (Burkey et al., 2006; Souza et al., 2006). Cheng et al. (2007) exposed two soybean 

34 cultivars to elevated O3 (77 ppb) and filtered air for 7-h/day for 6 days. The differences in 

35 sensitivity between the two cultivars could not be explained by differential O3 uptake or 
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1 by the fraction of reduced ascorbate present in the apoplast. However, total antioxidant 

2 capacity of the apoplast was twofold higher in the tolerant Essex cultivar as compared to 

3 the sensitive Forrest cultivar, indicating that there may be other compounds in the leaf 

4 apoplast that scavenge ROS. D’Haese et al. (2005) exposed the NC-S (sensitive) and NC5 R (resistant) clones of white clover (Trifolium repens) to 60 ppb O3 for 7-h/day for 

6 5 days in environmental chambers. Surprisingly, the NC-S clone had a higher constitutive 

7 concentration of apoplastic ascorbate with a higher redox status than the NC-R clone. 

8 However, the redox status of symplastic GSH was higher in NC-R, even though the 

9 concentration of GSH was not higher than in NC-S. In addition, total symplastic 

10 antioxidative capacity was not a determining factor in differential sensitivity between 

11 these two clones. Severino et al. (2007) also examined the role of antioxidants in the 

12 differential sensitivity of the two white clover clones by growing them in the field for a 

13 growing season and then exposing them to elevated O3 (100 ppb for 8-h/day for 10 days) 

14 in OTC at the end of the field season. The NC-R clone had greater quantities of total 

15 ascorbate and total antioxidants than the NC-S clone at the end of the experiment. In snap 

16 bean, plants of the O3 tolerant Provider cultivar had greater total ascorbate and more 

17 ascorbate in the apoplast than the sensitive S156 cultivar after exposure to 71 ppb O3 for 

18 10 days in OTC (Burkey et al., 2003). While most of the apoplastic ascorbate was in the 

19 oxidized form, the ratio of reduced ascorbate to total ascorbate was higher in Provider 

20 than S156, indicating that Provider is better able to maintain this ratio to maximize plant 

21 protection from oxidative stress. Exposure of two wheat varieties to ambient (7-h average 

22 44 ppb O3) and elevated (7-h average 56 ppb O3) for 60 days in open-air field conditions 

23 showed higher concentrations of reduced ascorbate in the apoplast in the tolerant Y16 

24 variety than the more sensitive Y2 variety, however no varietal differences were seen in 

25 the decrease in reduced ascorbate quantity in response to O3 exposure (Feng et al., 2010). 

26 There is much evidence that supports an important role for ascorbate, particularly 

27 apoplastic ascorbate, in protecting plants from oxidative stressors such as O3; however, it 

28 is also clear that there is much variation in the importance of ascorbate for different plant 

29 species and differing exposure conditions. Additionally, the work of several authors 

30 suggests that there may be other compounds in the apoplast which have the capacity to 

31 act as antioxidants.

32 While the quantities of antioxidant metabolites such as ascorbate are an important 

33 indicator of plant tolerance to O3, the ability of the plant to recycle oxidized ascorbate 

34 efficiently also plays a large role in determining the plant’s ability to effectively protect 

35 itself from sustained exposure to oxidative stress. Tobacco plants over-expressing DHAR 

36 were better protected from exposure to either chronic (100 ppb O3 4-h/day for 30 days) or 

37 acute (200 ppb O3 for 2 hours) conditions than control plants and those with reduced 

38 expression of DHAR (Chen and Gallie, 2005). The DHAR over-expressing plants 

39 exhibited an increase in guard cell ascorbic acid, leading to a decrease in stomatal 
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1 responsiveness to O3 and an increase in stomatal conductance and O3 uptake. Despite 

2 this, the presence of higher levels of ascorbic acid led to a lower oxidative load and a 

3 higher level of photosynthetic activity in the DHAR over-expressing plants (Chen and 

4 Gallie, 2005). A subsequent study with tobacco plants over-expressing DHAR confirmed 

5 some of these results. Levels of ascorbic acid were higher in the transgenic tobacco 

6 plants, and they exhibited greater tolerance to O3 exposure (200 ppb O3) as demonstrated 

7 by higher photosynthetic rates in the transgenic plants as compared to the control plants 

8 (Eltayeb et al., 2006). Over-expression of monodehydroascorbate reductase (MDAR) in 

9 tobacco plants also showed enhanced stress tolerance in response to O3 exposure 

10 (200 ppb O3), with higher rates of photosynthesis and higher levels of reduced ascorbic 

11 acid as compared to controls (Eltayeb et al., 2007). Results of these studies demonstrate 

12 the importance of ascorbic acid as a detoxification mechanism in some plant species, and 

13 also emphasize that the recycling of oxidized ascorbate to maintain a reduced pool of 

14 ascorbate is a factor in determining plant tolerance to oxidative stress.

15 The roles of other antioxidant metabolites and enzymes, including GSH, catalase (CAT), 

16 and superoxide dismutase (SOD), were comprehensively reviewed in the 2006 O3

17 AQCD. Additional studies have supported the findings reported in that document. 

18 Superoxide dismutase (SOD) and peroxidase (POD) activities were measured in both the 

19 tolerant Bel B and sensitive Bel W3 tobacco cultivars exposed to ambient O3

20 concentrations for 2 weeks 3 times throughout a growing season (Borowiak et al., 2009). 

21 In this study, SOD and POD activity, including that of several different isoforms, 

22 increased in both the sensitive and tolerant tobacco cultivars with exposure to O3, 

23 however the isoenzyme composition for POD differed between the sensitive and tolerant 

24 tobacco cultivars (Borowiak et al., 2009) Tulip poplar (Liriodendron tulipifera) trees 

25 exposed to increasing O3 concentrations (from 100 to 300 ppb O3 during a 2-week 

26 period) showed increases in activities of SOD, ascorbate peroxidase (APX), glutathione 

27 reductase (GR), MDAR, DHAR, CAT and POD in the 2-week period, although 

28 individual enzyme activities increased at different times during the 2-week period (Ryang 

29 et al., 2009).

30 Longer, chronic O3 exposures in trees revealed increases in SOD and APX activity in 

31 Quercus mongolica after 45 days of plant exposure to 80 ppb O3, which were followed 

32 by declines in the activities and quantities of these enzymes after 75 days of exposure 

33 (Yan et al., 2010). Similarly, activities of SOD, APX, DHAR, MDAR, and GR increased 

34 in Gingko biloba trees during the first 50 days of exposure to 80 ppb O3, followed by 

35 decreases in activity below control values after 50 days of exposure (He et al., 2006). 

36 Soybean plants exposed to 70 or 100 ppb O3 for 4-h/day over the course of a growing 

37 season showed elevated POD activity and a decrease in CAT activity at 40 and 60 days 

38 after germination (Singh et al., 2010a).
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1 Antioxidant enzymes and metabolites have been shown to play an important role in 

2 determining plant tolerance to O3 and mediating plant responses to O3. However, there is 

3 also some evidence to suggest that the direct reaction of ascorbate with O3 could lead to 

4 the formation of secondary toxicants, such as peroxy compounds, which may act upon 

5 signal transduction pathways and modulate plant response to O3 (Sandermann, 2008). 

6 Therefore, the role of ascorbate and other antioxidants and their interaction with other 

7 plant responses to O3, such as the activation of signal transduction pathways, is likely far 

8 more complex than is currently understood.

9.3.5 Effects on Primary and Secondary Metabolism

9.3.5.1 Light and Dark Reactions of Photosynthesis

9 Declines in the rate of photosynthesis and stomatal conductance in O3-treated plants have 

10 been documented for many different plant species (Booker et al., 2009; U.S. EPA, 2006b)

11 (Wittig et al., 2007). The 2006 O3 AQCD outlined what is known about the effects of O3

12 on carbon assimilation, and the more recent scientific literature confirms these findings. 

13 While several measures of the light reactions of photosynthesis are sensitive to exposure 

14 to O3 (see below), photosynthetic carbon assimilation is generally considered to be more 

15 affected by pollutant exposure, resulting in an overall decline in photosynthesis (Guidi

16 and Degl'lnnocenti, 2008; Heath, 2008; Fiscus et al., 2005). Loss of carbon assimilation 

17 capacity has been shown to result primarily from declines in the quantity of Rubisco 

18 (Singh et al., 2009; Calatayud et al., 2007a). Experimental evidence suggests that both 

19 decreases in Rubisco synthesis and enhanced degradation of the protein contribute to the 

20 measured reduction in its quantity (U.S. EPA, 2006b). Reduced carbon assimilation has 

21 been linked to reductions in biomass and yield (Wang et al., 2009b; He et al., 2007; 

22 Novak et al., 2007; Gregg et al., 2006; Keutgen et al., 2005). Recent studies evaluating 

23 O3 induced changes in the transcriptome and proteome of several different species 

24 confirm these findings. Levels of mRNA for the small subunit of Rubisco (rbcS) declined 

25 in European beech saplings exposed to 300 ppb O3 for 8-h/day for up to 26 days (Olbrich 

26 et al., 2005). Similar declines in rbcS mRNA were also measured in the beech saplings in 

27 a free air exposure system over a course of two growing seasons (Olbrich et al., 2009). 

28 Proteomics studies have also confirmed the effects of O3 on proteins involved in carbon 

29 assimilation. Reductions in quantities of the small and large subunit (rbcL) of Rubisco

30 and Rubisco activase were measured in soybean plants exposed to 120 ppb O3 for 3 days 

31 in growth chambers (Ahsan et al., 2010). Exposure of young poplar trees to 120 ppb O3

32 for 35 days in exposure chambers resulted in reductions of Rubisco, Rubisco activase, 

33 and up to 24 isoforms of Calvin cycle enzymes, most of which play a role in regenerating 
Draft – Do Not Cite or Quote 9-31 September 2011

1 the CO2 acceptor molecule, ribulose-1.5-bisphosphate (Bohler et al., 2007). Reductions 

2 in protein quantity of both the small and large subunit of Rubisco were seen in wheat 

3 plants exposed to ambient (average concentration 47.3 ppb O3) and elevated O3 (ambient 

4 + 10 or 20 ppb O3) in open-top chambers for 5-h/day for 50 days (Sarkar et al., 2010). 

5 Lettuce plants exposed to 100 ppb O3 in growth chambers for 8-h/day for 3 weeks also 

6 showed reductions in transcript and protein levels of the small and large subunits of 

7 Rubisco and Rubisco activase (Goumenaki et al., 2010). The reductions in carbon 

8 assimilation have been associated with declines in both the mRNA of the small and large 

9 subunits of Rubisco, and with reductions in Rubisco activase mRNA and protein. 

10 Additionally, the reduction in Rubisco quantity has also been associated with the O3-

11 induced oxidative modification of the enzyme, which is evidenced by the increases in 

12 carbonyl groups on the protein after plant exposure to O3. 

13 In addition to impacts on carbon assimilation, the deleterious effects of O3 on the 

14 photosynthetic light reactions have received more attention in recent years. Chlorophyll 

15 fluorescence provides a useful measure of changes to the photosynthetic process from 

16 exposure to oxidative stress. Decreases in the Fv/Fm ratio (a measure of the maximum 

17 efficiency of Photosystem II) in dark adapted leaves indicate a decline in the efficiency of 

18 the PSII photosystems and a concomitant increase in non-photochemical quenching 

19 (Guidi and Degl'lnnocenti, 2008; Scebba et al., 2006). Changes in these parameters have 

20 been correlated to differential sensitivity of plants to the pollutant. In a study to evaluate 

21 the response of 4 maple species to O3 (exposed to an 8-h avg of 51 ppb for ambient and 

22 79 ppb for elevated treatment in OTC), the 2 species which were most sensitive based on 

23 visible injury and declines in CO2 assimilation also showed the greatest decreases in 

24 Fv/Fm in symptomatic leaves. In asymptomatic leaves, CO2 assimilation decreased 

25 significantly but there was no significant decline in Fv/Fm (Calatayud et al., 2007a). Degl 

26 ‘Innocenti et al. (2007) measured significant decreases in Fv/Fm in young and 

27 symptomatic leaves of a resistant tomato genotype (line 93.1033/1) in response to O3

28 exposure (150 ppb O3 for 3 hours in a growth chamber), but only minor decreases in 

29 asymptomatic leaves with no associated changes in net photosynthetic rate. In the O3

30 sensitive tomato cultivar Cuor Di Bue, the Fv/Fm ratio did not change, while the 

31 photosynthetic rate declined significantly in asymptomatic leaves (Degl'Innocenti et al., 

32 2007). In two soybean cultivars, Fv/Fm also declined significantly with plant exposure to 

33 O3 (Singh et al., 2009). It appears that in asymptomatic leaves, photoinhibition, as 

34 indicated by a decrease in Fv/Fm, is not the main reason for a decline in photosynthesis.

35 An evaluation of photosynthetic parameters of two white clover (Trifolium repens cv. 

36 Regal) clones that differ in their O3 sensitivity revealed that O3 (40-110 ppb O3 for 7-

37 h/day for 5 days) increased the coefficient of non-photochemical quenching (qNP) in both 

38 the resistant (NC-R) and sensitive (NC-S) clones, however qNP was significantly lower 
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1 for the sensitive clone (Crous et al., 2006). Sensitive Acer clones had a lower coefficient 

2 of non-photochemical quenching, while exposure to O3 increased qNP in both sensitive 

3 and tolerant clones (Calatayud et al., 2007a). While exposure to O3 also increased qNP in 

4 tomato, there were no differences in the coefficient of photochemical quenching between 

5 cultivars thought to be differentially sensitive to O3. (Degl'Innocenti et al., 2007). Higher 

6 qNP as a result of exposure to O3 indicates a reduction in the proportion of absorbed light 

7 energy being used to drive photochemistry. A lower coefficient of non-photochemical 

8 quenching in O3 sensitive plants could indicate increased vulnerability to ROS generated 

9 during exposure to oxidative stress (Crous et al., 2006). 

10 Most of the research on O3 effects on photosynthesis has focused on C3 (Calvin cycle) 

11 plants because C4 (Hatch-Slack) plants have lower stomatal conductance and are, 

12 therefore, thought to be less sensitive to O3 stress. However, a few studies have been 

13 conducted to evaluate the effects of O3 on C4 photosynthesis. In older maize leaves, 

14 Leitao et al. (2007b; 2007c) found that the activity, quantity and transcript levels of both 

15 Rubisco and phosphoenolpyruvate carboxylase (PEPc) decreased as a function of rising 

16 O3 concentration. In younger maize leaves, the quantity, activity, and transcript levels of 

17 the carboxylases were either increased or unaffected in plants exposed to 40 ppb O3 for 

18 7- h/day for 28-33 days, but decreased at 80 ppb (Leitao et al., 2007a; Leitao et al., 

19 2007b). 

9.3.5.2 Respiration and Dark Respiration

20 While much research emphasis regarding O3 effects on plants has focused on the 

21 negative impacts on carbon assimilation, other studies have measured impacts on 

22 catabolic pathways such as shoot respiration and photorespiration. Generally, shoot 

23 respiration has been found to increase in plants exposed to O3. Bean plants exposed to 

24 ambient (average 12-h mean 43 ppb) and twice ambient (average 12-h mean 80 ppb) O3

25 showed increases in respiration. When mathematically partitioned, the maintenance 

26 coefficient of respiration was significantly increased in O3 treated plants, while the 

27 growth coefficient of respiration was not affected (Amthor, 1988). Loblolly pines were 

28 exposed to ambient (12-h daily mean was 45 ppb) and twice ambient (12 hours daily 

29 mean was 86 ppb) O3 for 12-h/day for approximately seven months per year for 3 and 

30 4 years. While photosynthetic activity declined with the age of the needles and increasing 

31 O3 concentration, enzymes associated with respiration showed higher levels of activity 

32 with increasing O3 concentration (Dizengremel et al., 1994). In their review on the role of 

33 metabolic changes in plant redox status after O3 exposure, Dizengremel et al. (2009)

34 summarized multiple studies in which several different tree species were exposed to O3

35 concentrations ranging from ambient to 200 ppb O3 for at least several weeks. In all 
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1 cases, the activity of enzymes, including phosphofructokinase, pyruvate kinase and 

2 fumarase, which are part of several catabolic pathways, were increased in O3 treated 

3 plants.

4 Photorespiration is a light-stimulated process which consumes O2 and releases CO2. 

5 While it has been regarded as a wasteful process, more recent evidence suggests that it 

6 may play a role in photoprotection during photosynthesis (Bagard et al., 2008). The few 

7 studies that have been conducted on O3 effects on photorespiration suggest that rates of 

8 photorespiration decline concomitantly with rates of photosynthesis. Soybean plants were 

9 exposed to ambient (daily averages 43-58 ppb) and 1.5 ambient O3 (daily averages 63-

10 83 ppb) O3 in OTCs for 12-h/day for 4 months. Rates of photosynthesis and 

11 photorespiration and photorespiratory enzyme activity declined only at the end of the 

12 growing season and did not appear to be very sensitive to O3 exposure (Booker et al., 

13 1997). Young hybrid poplars exposed to 120 ppb O3 for 13-h/day for 35 days in 

14 phytotron chambers showed that effects on photorespiration and photosynthesis were 

15 dependent upon the developmental stage of the leaf. While young leaves were not 

16 impacted, reductions in photosynthesis and photorespiration were measured in fully 

17 expanded leaves (Bagard et al., 2008).

9.3.5.3 Secondary Metabolism

18 Transcriptome analysis of Arabidopsis plants has revealed modulation of several genes 

19 involved in plant secondary metabolism (Ludwikow and Sadowski, 2008). Phenylalanine 

20 ammonia lyase (PAL) has been the focus of many studies involving plant exposure to O3

21 due to its importance in linking the phenylpropanoid pathway of plant secondary 

22 metabolism to primary metabolism in the form of the shikimate pathway. Genes encoding 

23 several enzymes of the phenylpropanoid pathway and lignin biosynthesis were up24 regulated in transcriptome analysis of Arabidopsis plants (Col-0) exposed to 350 ppb O3

25 for 6 hours, while 2 genes involved in flavonoid biosynthesis were down-regulated 

26 (Ludwikow et al., 2004). Exposure of Arabidopsis (Col-0) to lower O3 concentrations 

27 (150 ppb for 8-h/day for 2 days) resulted in the induction of 11 transcripts involved in 

28 flavonoid synthesis. In their exposure of 2-year-old Mediterranean shrub Phillyrea 

29 latifolia to 110 ppb O3 for 90 days, Paolacci et al. (2007) identified four clones that were 

30 up-regulated and corresponded to genes involved in the synthesis of secondary 

31 metabolites, such as isoprenoids, polyamines and phenylpropanoids. Up-regulation of 

32 genes involved in isoprene synthesis was also observed in Medicago trunculata exposed 

33 to 300 ppb O3 for 6 hours, while genes encoding enzymes of the flavonoid synthesis 

34 pathway were either up- or down-regulated (Puckette et al., 2008). Exposure of red clover 

35 to 1.5 × ambient O3 (average concentrations of 32.4 ppb) for up to 9 weeks in an open 
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1 field exposure system resulted in increases in leaf total phenolic content. However, the 

2 types of phenolics that were increased in response to O3 exposure differed depending 

3 upon the developmental stage of the plant. While almost all of the 31 different phenolic 

4 compounds measured increased in quantity initially during the exposure, after 3 weeks 

5 the quantity of isoflavones decreased while other phenolics increased (Saviranta et al., 

6 2010). Exposure of beech saplings to ambient and 2 × ambient O3 concentrations over 2 

7 growing seasons resulted in the induction of several enzymes which contribute to lignin 

8 formation, while enzymes involved in flavonoid biosynthesis were down-regulated 

9 (Olbrich et al., 2009). Exposure of tobacco Bel W3 to 160 ppb O3 for 5 hours showed up10 regulation of almost all genes encoding for enzymes which are part of the prechorismate 

11 pathway (Janzik et al., 2005). Isoprenoids can serve as antioxidant compounds in plants 

12 exposed to oxidative stress (Paolacci et al., 2007). 

13 The prechorismate pathway is the pathway leading to the formation of chorismate, a 

14 precursor to the formation of the aromatic amino acids tryptophan, tyrosine and 

15 phenylalanine. These amino acids are precursors for the formation of many secondary 

16 aromatic compounds, and, therefore, the prechorismate pathway represents a branch17 point in the regulation of metabolites into either primary or secondary metabolism (Janzik 

18 et al., 2005). Exposure of the O3 sensitive Bel W3 tobacco cultivar at 160 ppb for 5 hours 

19 showed an increase in transcript levels of most of the genes encoding enzymes of the 

20 prechorismate pathway. However, shikimate kinase (SK) did not show any change in 

21 transcript levels and only one of three isoforms of DAHPS (3-deoxy-D-arabino22 heptulosonat-7-phosphate synthase), the first enzyme in this pathway, was induced by O3

23 exposure (Janzik et al., 2005). Differential induction of DAHPS isoforms was also 

24 observed in European beech after 40 days of exposure to 150-190 ppb O3. At this time 

25 point in the beech experiment, transcript levels of shikimate pathway enzymes, including 

26 SK, were generally strongly induced after an only weak initial induction after the first 

27 40 days of exposure. Both soluble and cell-wall bound phenolic metabolites showed only 

28 minimal increases in response to O3 for the duration of the exposure period (Alonso et 

29 al., 2007). Total leaf phenolics decreased with leaf age in Populus nigra exposed to 

30 80 ppb O3 for 12-h/day for 14 days. Ozone increased the concentration of total leaf 

31 phenolics in newly expanded leaves, with the most significant increases occurring in 

32 compounds such as quercitin glycoside, which has a high antioxidant capacity (Fares et 

33 al., 2010b). While several phenylpropanoid pathway enzymes were induced in two poplar 

34 clones exposed to 60 ppb O3 for 5-h/day for 15 days, the degree of induction differed 

35 between the two clones. In the tolerant I-214 clone, PAL activity increased nine fold in 

36 O3-treated plants as compared to controls, while there was no significant difference in 

37 PAL activity in the sensitive Eridano clone (Di Baccio et al., 2008). 
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1 Polyamines such as putrescine, spermidine and spermine play a variety of roles in plants 

2 and have been implicated in plant defense responses to both abiotic and biotic stresses. 

3 They exist in both a free form and conjugated to hydroxycinnamic acids. Investigations 

4 on the role of polyamines have found that levels of putrescine increase in response to 

5 oxidative stress. This increase stems largely from the increase in the activity of arginine 

6 decarboxylase (ADC), a key enzyme in the synthesis of putrescine (Groppa and 

7 Benavides, 2008). Langebartels et al. (1991) described differences in putrescine 

8 accumulation in O3-treated tobacco plants exposed to several O3 concentrations, ranging 

9 from 0-400 ppb for 5-7 hours. A large and rapid increase in putrescine occurred in the 

10 tolerant Bel B cultivar and only a small increase in the sensitive Bel W3 cultivar, which 

11 occurred only after the formation of necrotic leaf lesions. Van Buuren et al. (2002)

12 further examined the role of polyamines in these two tobacco cultivars during an acute 

13 (130 ppb O3 for 7-h in a growth chamber) exposure. They found that while free 

14 putrescine accumulated in undamaged tissue of both cultivars, conjugated putrescine 

15 predominantly accumulated in tissues undergoing cell death after plant exposure to O3

16 (van Buuren et al., 2002). The authors suggest that while free putrescine may not play a 

17 role in conferring tolerance in the Bel B cultivar, conjugated putrescine may play a role in 

18 O3-induced programmed cell death in Bel W3 plants.

19 Isoprene is emitted by some plant species and represents the predominant biogenic source 

20 of hydrocarbon emissions in the atmosphere (Guenther et al., 2006). In the atmosphere, 

21 the oxidation of isoprene by hydroxyl radicals can enhance O3 formation in the presence 

22 of NOX, thereby impacting the O3 concentration that plants are exposed to. While 

23 isoprene emission varies widely between species, it has been proposed to stabilize 

24 membranes and provide those plant species that produce it with a mechanism of 

25 thermotolerance (Sharkey et al., 2008). It has also been suggested that isoprene may act 

26 as an antioxidant compound to scavenge O3 (Loreto and Velikova, 2001). Recent studies 

27 using a variety of plant species have shown conflicting results in trying to understand the 

28 effects of O3 on isoprene emission. Exposure to acute doses of O3 (300 ppb for 3-h) in 

29 detached leaves of Phragmites australis resulted in stimulation of isoprene emissions 

30 (Velikova et al., 2005). Similar increases in isoprene emissions were measured in 

31 Populus nigra after exposure to 100 ppb O3 for 5 days continuously (Fares et al., 2008). 

32 Isoprene emission in attached leaves of Populus alba, which were exposed to 150 ppb O3

33 for 11-h/day for 30 days inside cuvettes, was inhibited, while isoprene emission and 

34 transcript levels of isoprene synthase mRNA were increased in the leaves exposed to 

35 ambient O3 (40 ppb), which were located above the leaves enclosed in the exposure 

36 cuvettes (Fares et al., 2006). Exposure of 2 genotypes of hybrid poplar to 120 ppb O3 for 

37 6-h/day for 8 days resulted in a significant reduction in isoprene emission in the O3-

38 sensitive but not the tolerant genotype (Ryan et al., 2009). Similarly, O3 treatment 

39 (80 ppb 12-h/day for 14 days) of Populus nigra showed that isoprene emission was 
Draft – Do Not Cite or Quote 9-36 September 2011

1 reduced in the treated plants relative to the control plants (Fares et al., 2010b). Based on 

2 results of this and other studies, Fares et al. (2010b) concluded that the isoprenoid 

3 pathway may be induced in plants exposed to acute O3 doses, while at lower doses 

4 isoprene emission may be inhibited. Vickers et al. (2009) developed transgenic tobacco 

5 plants with the isoprene synthase gene from Populus alba and exposed them to 120 ppb 

6 O3 for 6-h/day for 2 days. They determined that the wildtype plants showed significantly 

7 more O3 damage, including the development of leaf lesions and a decline in 

8 photosynthetic rates, than the transgenic, isoprene-emitting plants. Transgenic plants also 

9 accumulated less H2O2 and had lower levels of lipid peroxidation following exposure to 

10 O3 than the wildtype plants (Vickers et al., 2009). These results indicate that isoprene 

11 may have a protective role for plants exposed to oxidative stress.

9.3.6 Summary

12 The results of recent studies on the effects of O3 stress on plants support and strengthen 

13 those reported in the 2006 O3 AQCD. The most significant new body of evidence since 

14 the 2006 O3 AQCD comes from research on molecular mechanisms of the biochemical 

15 and physiological changes observed in many plant species in response to O3 exposure. 

16 Recent studies have employed new techniques, such as those used in evaluating 

17 transcriptomes and proteomes to perform very comprehensive analyses of changes in 

18 gene transcription and protein expression in plants exposed to O3. These newer molecular 

19 studies not only provide very important information regarding the many mechanisms of 

20 plant responses to O3, they also allow for the analysis of interactions between various 

21 biochemical pathways which are induced in response to O3. However, many of these 

22 studies have been conducted in artificial conditions with model plants, which are 

23 typically exposed to very high, short doses of O3. Therefore, additional work remains to 

24 elucidate whether these plant responses are transferable to other plant species exposed to 

25 more realistic ambient conditions. 

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

27 O3 is thought to rapidly react with the aqueous layer surrounding the cell (apoplast) to 

form breakdown products such as hydrogen peroxide (H2O2), superoxide (O2 hydroxyl , )28 ־

radicals (HO·) and peroxy radicals (HO2

· 29 ). Plants could be sensing the presence of O3

30 and/or its breakdown products in a variety different ways, depending upon the plant 

31 species and the exposure parameters. Experimental evidence suggests that mitogen32 activated protein kinases and calcium are important components of the signal 

33 transduction pathways, which communicate signals to the nucleus and lead to changes in 

34 gene expression in response to O3. It is probable that there are multiple sensing 

35 mechanisms and signal transduction pathways, and their activation may depend upon the 
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1 plant species, its developmental stage and/or O3 exposure conditions. Initiation of signal 

2 transduction pathways in O3 treated plants has also been observed in stomatal guard cells. 

3 Reductions in stomatal conductance in have been described for many plant species 

4 exposed to O3, and new experimental evidence suggests that this reduction may be due 

5 not only to a decrease in carboxylation efficiency, but also to a direct impact of O3 on 

6 stomatal guard cell function, leading to a changes in stomatal conductance. 

7 Alterations in gene transcription that have been observed in O3-treated plants are now 

8 evaluated more comprehensively using DNA microarray studies, which measure changes 

9 in the entire transcriptome rather than measuring the transcript levels of individual genes. 

10 These studies have demonstrated very consistent trends, even though O3 exposure 

11 conditions (concentration, duration of exposure), plant species and sampling times vary 

12 significantly. Genes involved in plant defense, signaling, and those associated with the 

13 synthesis of plant hormones and secondary metabolism are generally up-regulated in 

14 plants exposed to O3, while those related to photosynthesis and general metabolism are 

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

16 concomitant increases or decreases in the proteins encoded by these genes. Transcriptome 

17 analysis has also illuminated the complex interactions that exist between several different 

18 phytohormones and how they modulate plant sensitivity and response to O3. 

19 Experimental evidence suggests that while ethylene and salicylic acid are needed to 

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

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

22 regulating stomatal aperture, may also act antagonistically to the jasmonic acid signaling 

23 pathway. Changes in the quantity and activity of these phytohormones and the 

24 interactions between them reveal some of the complexity of plant responses to an 

25 oxidative stressor such as O3. 

26 Another critical area of interest is to better understand and quantify the capacity of the 

27 plant to detoxify oxygen radicals using antioxidant metabolites, such as ascorbate and 

28 glutathione, and the enzymes that regenerate them. Ascorbate remains an important focus 

29 of research, and, due to its location in the apoplast in addition to other cellular 

30 compartments, it is regarded as a first line of defense against oxygen radicals formed in 

31 the apoplast. Most studies demonstrate that antioxidant metabolites and enzymes increase 

32 in quantity and activity in plants exposed to O3, indicating that they play an important 

33 role in protecting plants from oxidative stress. However, attempts to quantify the 

34 detoxification capacity of plants have remained unsuccessful, as high quantities of 

35 antioxidant metabolites and enzymes do not always translate into greater protection of the 

36 plant. Considerable variation exists between plant species, different developmental 

37 stages, and the environmental and O3 exposure conditions which plants are exposed to. 
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1 As indicated earlier, the described alterations in transcript levels of genes correlate with 

2 observed changes quantity and activity of the enzymes and metabolites involved in 

3 primary and secondary metabolism. In addition to the generalized up-regulation of the 

4 antioxidant defense system, photosynthesis typically declines in O3 treated plants. 

5 Declines in C fixation due to reductions in quantity and activity of Rubisco were 

6 extensively described in the 2006 O3 AQCD. More recent studies support these results 

7 and indicate that declines in Rubisco activity may also result from reductions in Rubisco 

8 activase enzyme quantity. Other studies, which have focused on the light reactions of 

9 photosynthesis, demonstrate that plant exposure to O3 results in declines in electron 

10 transport efficiency and a decreased capacity to quench oxidizing radicals. Therefore, the 

11 overall declines in photosynthesis observed in O3 -treated plants likely result from 

12 combined impacts on stomatal conductance, carbon fixation and the light reactions. 

13 While photosynthesis generally declines in plants exposed to O3, catabolic pathways such 

14 as respiration have been shown to increase. It has been hypothesized that increased 

15 respiration may result from greater energy needs for defense and repair. Secondary 

16 metabolism is generally up-regulated in a variety of species exposed to O3 as a part of a 

17 generalized plant defense mechanism. Some secondary metabolites, such as flavonoids 

18 and polyamines, are of particular interest as they are known to have antioxidant 

19 properties. The combination of decreases in C assimilation and increases in catabolism 

20 and the production of secondary metabolites would negatively impact plants by 

21 decreasing the energy available for growth and reproduction.

9.4 Nature of Effects on Vegetation and Ecosystems

9.4.1 Introduction

22 Ambient O3 concentrations have long been known to cause visible symptoms, decreases 

23 in photosynthetic rates, decreases in growth and yield of plants as well as many other 

24 effects on ecosystems (U.S. EPA, 2006b, 1996c, 1986, 1978a). Numerous studies have 

25 related O3 exposure to plant responses, with most effort focused on the yield of crops and 

26 the growth of tree seedlings. Many experiments exposed individual plants grown in pots 

27 or soil under controlled conditions to known concentrations of O3 for a segment of 

28 daylight hours for some portion of the plant’s life span. Information in this section also 

29 goes beyond individual plant scale responses to consider effects at the broader ecosystem 

30 scale, including effects related to ecosystem services. 

31 This section will focus mainly on studies published since the release of the 2006 O3

32 AQCD. However, because much O3 research was conducted prior to the 2006 O3 AQCD, 
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1 the present discussion of vegetation and ecosystem response to O3 exposure is largely 

2 based on the conclusions of the 1978, 1986, 1996, and 2006 O3 AQCDs.

9.4.1.1 Ecosystem Scale, Function, and Structure

3 Information presented in this section was collected at multiple spatial scales, ranging 

4 from the physiology of a given species to population, community, and ecosystem 

5 investigations. An ecological population is a group of individuals of the same species and 

6 a community is an assemblage of populations of different species interacting with one 

7 another that inhabit an area. For this assessment, “ecosystem” is defined as the interactive 

8 system formed from all living organisms and their abiotic (physical and chemical) 

9 environment within a given area (IPCC, 2007a). The boundaries of what could be called 

10 an ecosystem are somewhat arbitrary, depending on the focus of interest or study. Thus, 

11 the extent of an ecosystem may range from very small spatial scales to, ultimately, the 

12 entire Earth (IPCC, 2007a). All ecosystems, regardless of size or complexity, have 

13 interactions and physical exchanges between biota and abiotic factors, this includes both 

14 structural (e.g., soil type and food web trophic levels) and functional (e.g., energy flow, 

15 decomposition, nitrification) attributes. 

16 Ecosystems are most often defined by their structure based on the number and type of 

17 species present. Structure may refer to a variety of measurements including the species 

18 richness, abundance, community composition and biodiversity as well as landscape 

19 attributes. Competition among and within species and their tolerance to environmental 

20 stressors are key elements of survivorship. When environmental conditions are shifted, 

21 for example, by the presence of anthropogenic air pollution, these competitive 

22 relationships may change and tolerance to stress may be exceeded. Ecosystems may also 

23 be defined on a functional basis. “Function” refers to the suite of processes and 

24 interactions among the ecosystem components and their environment that involve 

25 nutrient and energy flow as well as other attributes including water dynamics and the flux 

26 of trace gases. Plant processes including photosynthesis, respiration, C allocation, 

27 nutrient uptake and evaporation, are directly related to functions of energy flow and C, 

28 nutrient and water cycling. The energy accumulated and stored by vegetation (via 

29 photosynthetic C capture) is available to other organisms. Energy moves from one 

30 organism to another through food webs, until it is ultimately released as heat. Nutrients 

31 and water can be recycled. Air pollution alters the function of ecosystems when elemental 

32 cycles or the energy flow are altered. This alteration can also be manifested in changes in 

33 the biotic composition of ecosystems. 
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1 There are at least three levels of ecosystem response to pollutants: (1) the individual 

2 organism and its environment; (2) the population and its environment; and (3) the 

3 biological community composed of many species and their environment (Billings, 1978). 

4 Individual organisms within a population vary in their ability to withstand the stress of 

5 environmental change. The response of individual organisms within a population is based 

6 on their genetic constitution, stage of growth at time of exposure to stress, and the 

7 microhabitat in which they are growing (Levine and Pinto, 1998). The stress range within 

8 which organisms can exist and function determines the ability of the population to 

9 survive. Those best able to cope with environmental stressors survive and reproduce. 

10 Competition among different species results in succession (community change over time) 

11 and, ultimately, sensitive species may be progressively replaced and communities shift to 

12 favor those species that may have the capability to tolerate stressors such as O3 (Rapport 

13 and Whitford, 1999; Guderian, 1985). 

9.4.1.2 Ecosystem Services

14 Ecosystem structure and function may be translated into ecosystem services. Ecosystem 

15 services are the benefits people obtain from ecosystems (UNEP, 2003). Ecosystems 

16 provide many goods and services that are of vital importance for the functioning of the 

17 biosphere and provide the basis for the delivery of tangible benefits to human society. 

18 Hassan et al. (2005) define these benefits to include supporting, provisioning, regulating, 

19 and cultural services: 

20  Supporting services are necessary for the production of all other ecosystem 

21 services. Some examples include biomass production, production of 

22 atmospheric O2, soil formation and retention, nutrient cycling, water cycling, 

23 and provisioning of habitat. Biodiversity is a supporting service that is 

24 increasingly recognized to sustain many of the goods and services that humans 

25 enjoy from ecosystems. These provide a basis for three higher-level categories 

26 of services. 

27  Provisioning services, such as products (Gitay et al., 2001), i.e., food 

28 (including game, roots, seeds, nuts and other fruit, spices, fodder), water, fiber 

29 (including wood, textiles), and medicinal and cosmetic products (such as 

30 aromatic plants, pigments). 

31  Regulating services that are of paramount importance for human society such 

32 as (1) C sequestration, (2) climate and water regulation, (3) protection from 

33 natural hazards such as floods, avalanches, or rock-fall, (4) water and air 

34 purification, and (5) disease and pest regulation. 
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1  Cultural services that satisfy human spiritual and aesthetic appreciation of 

2 ecosystems and their components including recreational and other nonmaterial 

3 benefits.

4 In the sections that follow, available information on individual, population and 

5 community response to O3 will be discussed. Effects of O3 on productivity and 

6 C sequestration, water cycling, below-ground processes, competition and biodiversity, 

7 and insects and wildlife are considered below and in the context of ecosystem services 

8 where appropriate. 

9.4.2 Visible Foliar Injury and Biomonitoring

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

10 documented over several decades on many tree, shrub, herbaceous, and crop species 

11 (U.S. EPA, 2006b, 1996b, 1984, 1978a). Visible foliar injury symptoms are considered 

12 diagnostic as they have been verified experimentally in exposure-response studies, using 

13 exposure methodologies such as CSTRs, OTCs, and free-air fumigation (see Section 9.2 

14 for more detail on exposure methodologies). Several pictorial atlases and guides have 

15 been published, providing details on diagnosis and identification of O3-induced visible 

16 foliar injury on many plant species throughout North America (Flagler, 1998; NAPAP, 

17 1987) and Europe (Innes et al., 2001; Sánchez et al., 2001). Typical visible injury 

18 symptoms on broad-leaved plants include: stippling, flecking, surface bleaching, bifacial 

19 necrosis, pigmentation (e.g., bronzing), chlorosis, and/or premature senescence. Typical 

20 visible injury symptoms for conifers include: chlorotic banding, tip burn, flecking, 

21 chlorotic mottling, and/or premature senescence of needles. Although common patterns 

22 of injury develop within a species, these foliar lesions can vary considerably between and 

23 within taxonomic groups. Furthermore, the degree and extent of visible foliar injury 

24 development varies from year to year and site to site (Orendovici-Best et al., 2008; 

25 Chappelka et al., 2007; Smith et al., 2003), even among co-members of a population 

26 exposed to similar O3 levels, due to the influence of co-occurring environmental and 

27 genetic factors. Nevertheless, Chappelka et al. (2007) reported that the average incidence 

28 of O3-induced foliar injury was 73% on milkweed in the Great Smoky Mountains 

29 National Park in the years 1992-1996. 

30 Although the majority of O3-induced visible foliar injury occurrence has been observed 

31 on seedlings and small plants, many studies have reported visible injury of mature 

32 coniferous trees, primarily in the western U.S. (Arbaugh et al., 1998) and to mature 

33 deciduous trees in eastern North America (Schaub et al., 2005; Vollenweider et al., 2003; 
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1 Chappelka et al., 1999a; Chappelka et al., 1999b; Somers et al., 1998; Hildebrand et al., 

2 1996). 

3 It is important to note that visible foliar injury occurs only when sensitive plants are 

4 exposed to elevated O3 concentrations in a predisposing environment. A major modifying 

5 factor for O3-induced visible foliar injury is the amount of soil moisture available to a 

6 plant during the year that the visible foliar injury is being assessed. This is because lack 

7 of soil moisture generally decreases stomatal conductance of plants and, therefore, limits 

8 the amount of O3 entering the leaf that can cause injury (Matyssek et al., 2006; Panek, 

9 2004; Grulke et al., 2003a; Panek and Goldstein, 2001; Temple et al., 1992; Temple et 

10 al., 1988). Consequently, many studies have shown that dry periods in local areas tend to 

11 decrease the incidence and severity of O3-induced visible foliar injury; therefore, the 

12 incidence of visible foliar injury is not always higher in years and areas with higher O3, 

13 especially with co-occurring drought (Smith et al., 2003). Other factors such as leaf age 

14 influence the severity of symptom expression with older leaves showing greater injury 

15 severity as a result of greater seasonal exposure (Zhang et al., 2010a).

16 Although visible injury is a valuable indicator of the presence of phytotoxic 

17 concentrations of O3 in ambient air, it is not always a reliable indicator of other negative 

18 effects on vegetation. The significance of O3 injury at the leaf and whole plant levels 

19 depends on how much of the total leaf area of the plant has been affected, as well as the 

20 plant’s age, size, developmental stage, and degree of functional redundancy among the 

21 existing leaf area. Previous O3 AQCDs have noted the difficulty in relating visible foliar 

22 injury symptoms to other vegetation effects such as individual plant growth, stand 

23 growth, or ecosystem characteristics (U.S. EPA, 2006b, 1996b). As a result, it is not 

24 presently possible to determine, with consistency across species and environments, what 

25 degree of injury at the leaf level has significance to the vigor of the whole plant. 

26 However, in some cases, visible foliar symptoms have been correlated with decreased 

27 vegetative growth (Somers et al., 1998; Karnosky et al., 1996; Peterson et al., 1987; 

28 Benoit et al., 1982) and with impaired reproductive function (Chappelka, 2002; Black et 

29 al., 2000). Conversely, the lack of visible injury does not always indicate a lack of 

30 phytotoxic concentrations of O3 or a lack of non-visible O3 effects (Gregg et al., 2006, 

31 2003). 

9.4.2.1 Biomonitoring

32 The use of biological indicators to detect phytotoxic levels of O3 is a longstanding and 

33 effective methodology (Chappelka and Samuelson, 1998; Manning and Krupa, 1992). A 

34 plant bioindicator can be defined as a vascular or nonvascular plant exhibiting a typical 
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1 and verifiable response when exposed to a plant stress such as an air pollutant (Manning, 

2 2003). To be considered a good indicator species, plants must (1) exhibit a distinct, 

3 verified response; (2) have few or no confounding disease or pest problems; and (3) 

4 exhibit genetic stability (U.S. EPA, 2006b). Such sensitive plants can be used to detect 

5 the presence of a specific air pollutant such as O3 in the ambient air at a specific location 

6 or region and, as a result of the magnitude of their response, provide unique information 

7 regarding specific ambient air quality. Bioindicators can be either introduced sentinels, 

8 such as the widely used tobacco (Nicotiana tabacum) variety Bel W3 (Calatayud et al., 

9 2007b; Laffray et al., 2007; Nali et al., 2007; Gombert et al., 2006; Kostka-Rick and 

10 Hahn, 2005; Heggestad, 1991) or detectors, which are sensitive native plant species 

11 (Chappelka et al., 2007; Souza et al., 2006). The approach is especially useful in areas 

12 where O3 monitors are not operated (Manning, 2003). For example, in remote wilderness 

13 areas where instrument monitoring is generally not available, the use of bioindicator 

14 surveys in conjunction with the use of passive samplers (Krupa et al., 2001) may be a 

15 useful methodology (Manning, 2003). However, it requires expertise in recognizing those 

16 signs and symptoms uniquely attributable to exposure to O3 as well as in their 

17 quantitative assessment.

18 Since the 2006 O3 AQCD, new sensitive plant species have been identified from field 

19 surveys and verified in controlled exposure studies (Kline et al., 2009; Kline et al., 2008). 

20 Several multiple-year field surveys have also been conducted at National Wildlife 

21 Refuges in Maine, Michigan, New Jersey, and South Carolina (Davis, 2009, 2007a, b; 

22 Davis and Orendovici, 2006). 

23 The USDA Forest Service through the Forest Health Monitoring Program (FHM) (1990 -

24 2001) and currently the Forest Inventory and Analysis (FIA) Program has been collecting 

25 data regarding the incidence and severity of visible foliar injury on a variety of O3

26 sensitive plant species throughout the U.S. (Coulston et al., 2003; Smith et al., 2003). The 

27 plots where these data are taken are known as biosites. These biosites are located 

28 throughout the country and analysis of visible foliar injury within these sites follows a set 

29 of established protocols. For more details, see http://www.nrs.fs.fed.us/fia/topics/ozone/

30 (USDA, 2011). The network has provided evidence of O3 concentrations high enough to 

31 induce visible symptoms on sensitive vegetation. From repeated observations and 

32 measurements made over a number of years, specific patterns of areas experiencing 

33 visible O3 injury symptoms can be identified. Coulston et al. (2003) used information 

34 gathered over a 6-year period (1994-1999) from the network to identify several species 

35 that were sensitive to O3 over a regional scale including sweetgum (Liquidambar 

36 styraciflua), loblolly pine (Pinus taeda), and black cherry (Prunus serotina). In a study of 

37 the west coast of the U.S, Campbell et al. (2007) reported O3 injury in 25-37% of biosites 

38 in California forested ecosystems from 2000-2005.
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1 A study by Kohut (2007) assessed the risk of O3-induced visible foliar injury on 

2 bioindicator plants (NPS, 2006) in 244 national parks in support of the National Park 

3 Service’s Vital Signs Monitoring Network (NPS, 2007). The risk assessment was based 

4 on a simple model relating response to the interaction of species, level of O3 exposure, 

5 and exposure environment. Kohut (2007) concluded that the risk of visible foliar injury 

6 was high in 65 parks (27%), moderate in 46 parks (19%), and low in 131 parks (54%). 

7 Some of the well-known parks with a high risk of O3-induced visible foliar injury include 

8 Gettysburg, Valley Forge, Delaware Water Gap, Cape Cod, Fire Island, Antietam, 

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

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

11 Lichens have also long been used as biomonitors of air pollution effects on forest health 

12 (Nash, 2008). It has been suspected, based on field surveys in the San Bernardino 

13 Mountains surrounding the Los Angeles air basin, that declines in lichen diversity and 

14 abundance were correlated with measured O3 gradients (Gül et al., 2011). Several recent 

15 studies in North America (Geiser and Neitlich, 2007; Gombert et al., 2006; Jovan and 

16 McCune, 2006) and Europe (Nali et al., 2007; Gombert et al., 2006) have used lichens as 

17 biomonitors of atmospheric deposition (e.g., N and S) and O3 exposure. Nali et al. (2007)

18 found that epiphytic lichen biodiversity was not related to O3 geographical distribution. 

19 In addition, a recent study by Riddell et al. (2010) found that lichen species, Ramalina 

20 menziesii, showed no decline in physiological response to low and moderate 

21 concentrations of O3 and may not be a good indicator for O3 pollution. Mosses have also 

22 been used as biomonitors of air pollution; however, there remains a knowledge gap in the 

23 understanding of the effects of ozone on mosses as there has been very little information 

24 available on this topic in recent years.

9.4.2.2 Summary

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

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

27 species (U.S. EPA, 2006b, 1996b, 1984, 1978a). Ozone-induced visible foliar injury 

28 symptoms on certain bioindicator plant species are considered diagnostic as they have 

29 been verified experimentally in exposure-response studies, using exposure methodologies 

30 such as continuous stirred tank reactors (CSTRs), OTCs, and free-air fumigation. 

31 Experimental evidence has clearly established a consistent association of visible injury 

32 with O3 exposure, with greater exposure often resulting in greater and more prevalent 

33 injury. Since the 2006 O3 AQCD, several multi-year field surveys of O3-induced visible 

34 foliar injury have been conducted at National Wildlife Refuges in Maine, Michigan, New 
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1 Jersey, and South Carolina. New sensitive species showing visible foliar injury continue 

2 to be identified from field surveys and verified in controlled exposure studies.

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

4 longstanding and effective methodology. The USDA Forest Service through the Forest 

5 Health Monitoring (FHM) Program (1990-2001) and currently the Forest Inventory and 

6 Analysis (FIA) Program has been collecting data regarding the incidence and severity of 

7 visible foliar injury on a variety of O3 sensitive plant species throughout the U.S. The 

8 network has provided evidence that O3 concentrations were high enough to induce visible 

9 symptoms on sensitive vegetation. From repeated observations and measurements made 

10 over a number of years, specific patterns of areas experiencing visible O3 injury 

11 symptoms can be identified. As noted in the preceding section, a study of 244 national 

12 parks indicated that the risk of visible foliar injury was high in 65 parks (27%), moderate 

13 in 46 parks (19%), and low in 131 parks (54%). 

14 Evidence is sufficient to conclude that there is a causal relationship between 

15 ambient O3 exposure and the occurrence of O3-induced visible foliar injury on 

16 sensitive vegetation across the U.S.

9.4.3 Growth, productivity and carbon storage in natural ecosystems

17 Ambient O3 concentrations have long been known to cause decreases in photosynthetic 

18 rates, decreases in growth, and decreases in yield (U.S. EPA, 2006b, 1996c, 1986, 

19 1978a). The O3-induced damages at the plant scale may translate to the ecosystem scale, 

20 and cause changes in productivity and C storage. This section focuses on the responses of 

21 C cycling to seasonal or multi-year exposures to O3 from the plant to ecosystem scale. 

22 Quantitative responses include changes in plant growth, plant biomass allocation, 

23 ecosystem production and ecosystem C sequestration. Because of the available 

24 information, most of discussion at the plant scale focuses on the response of individual 

25 plants, especially tree seedlings and crops, with limited discussion of mixtures of 

26 herbaceous species. Changes at the ecosystem scale are difficult to evaluate directly due 

27 to the complexity and the large spatial and temporal scale. The discussion of ecosystem 

28 effects focuses on the new studies at the large-scale FACE experiments and on ecological 

29 model simulations.

9.4.3.1 Plant growth and biomass allocation

30 The previous O3 AQCDs concluded that there is strong evidence that exposure to O3

31 decreases photosynthesis and growth in numerous plant species (U.S. EPA, 2006b, 
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1 1996b, 1984, 1978a). Studies published since the last review support those conclusions 

2 and are summarized below. 

3 In general, research conducted over several decades has indicated that exposure to O3 

4 alters stomatal conductance and reduces photosynthesis in a wide variety of plant species. 

5 In a review of more than 55 studies, Wittig et al. (2007) reported that current O3

6 concentrations in the northern hemisphere are decreasing stomatal conductance (13%) 

7 and photosynthesis (11%) across tree species. It was also found that younger trees (<4 

8 year) were affected less by O3 than older trees. Further, the authors also found that 

9 decreases in photosynthesis are consistent with the cumulative uptake of O3 into the leaf. 

10 In contrast, several studies reported that O3 exposure may result in loss of stomatal 

11 control, incomplete stomatal closure at night and a decoupling of photosynthesis and 

12 stomatal conductance, which may have implications for whole- plant water use (Section 

13 9.4.5).

14 In a recently published meta-analysis, Wittig et al. (2009) quantitatively compiled peer 

15 reviewed studies from the past 40 years on the effect of current and future O3 exposures 

16 on the physiology and growth of forest species. Wittig et al. (2009) reported that current 

17 ambient O3 concentrations as reported in those studies (~40 ppb) significantly decreased 

18 annual total biomass growth (7%) across 263 studies. However, this effect could be 

19 greater (11 to 17%) in areas that have higher O3 concentrations and as background O3

20 increases in the future (Wittig et al., 2009). This meta-analysis demonstrates the 

21 coherence of O3 effects across numerous studies and species that used a variety of 

22 experimental techniques, and these results support the conclusion of the previous AQCD.

23 In two companion papers, McLaughlin et al. (2007a; 2007b) investigated the effects of 

24 ambient O3 on tree growth and hydrology at forest sites in the southern Appalachian 

25 Mountains. The authors reported that the cumulative effects of ambient levels of O3

26 decreased seasonal stem growth by 30-50% for most trees species in a high O3 year in 

27 comparison to a low O3 year (McLaughlin et al., 2007a). The authors also report that 

28 high ambient O3 concentrations can disrupt whole-tree water use and in turn reduce late29 season streamflow (McLaughlin et al., 2007b); see Section 9.4.5 for more on water 

30 cycling. 

31 Since the 2006 O3 AQCD, several new studies based on the Aspen FACE “free air” O3

32 and CO2 exposure experiment in a forest in Wisconsin were published (Darbah et al., 

33 2008; Riikonen et al., 2008; Darbah et al., 2007; Kubiske et al., 2007; Kubiske et al., 

34 2006; King et al., 2005). Over the first seven years of stand development , Kubiske et al. 

35 (2006) observed that elevated O3 decreased tree heights, diameters, and main stem 

36 volumes in the aspen community by 11, 16, and 20%, respectively. In addition, Kubiske 

37 et al. (2007) reported that elevated O3 may change the intra- and inter-species 
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1 competition. For example, O3 treatments increased the rate of conversion from a mixed 

2 aspen-birch community to a birch dominated community. In a comparison presented in 

3 Section 9.6.3 of this document, EPA found that effects on biomass accumulation in aspen 

4 during the first seven years closely agreed with the exposure-response function based on 

5 data from earlier OTC experiments.

6 Several studies at the Aspen FACE site also considered other growth-related effects of 

7 elevated O3. Darbah et al. (2008; 2007) reported that O3 treatments decreased paper birch 

8 seed weight and seed germination and that this would likely lead to a negative impact of 

9 regeneration for that species. Riikonen et al. (2008) found that elevated O3 decreased the 

10 amount of starch in birch buds by 16%, and reduced aspen bud size, which may have 

11 been related to the observed delay in spring leaf development. The results suggest that 

12 elevated O3 concentrations have the potential to alter C metabolism of overwintering 

13 buds, which may have carry-over effects in the subsequent growing season (Riikonen et 

14 al., 2008).

15 Effects on growth of understory vegetation were also investigated at Aspen FACE. 

16 Bandeff et al. (2006) found that the effects of elevated CO2 and O3 on understory species 

composition, total and individual species biomass, N content, and 15 17 N recovery were a 

18 result of overstory community responses to those treatments; however, there were no 

19 apparent direct treatment effects due to high variability of the data. Total understory 

20 biomass increased with increasing light and was greatest under the open canopy of the 

21 aspen/maple community, as well as the more open canopy of the elevated O3 treatments 

22 (Bandeff et al., 2006). Similarly, data from a study by Awmack et al. (2007) suggest that 

23 elevated CO2 and O3 may have indirect growth effects on red (Trifolium pratense) and 

24 white (Trifolium repens) clover in the understory via overstory community effects; 

25 however, no direct effects of elevated O3 were observed. 

26 Overall, the studies at the Aspen FACE experiment are consistent with many of the OTC 

27 studies that were evaluated in previous O3 AQCDs. These results strengthen our 

28 understanding of O3 effects on forests and demonstrate the relevance of the knowledge 

29 gained from trees grown in open-top chamber studies.

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

31 O3 exposure is yield or growth, e.g., production of grain. For plants grown in mixtures 

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

33 species), endpoints other than production of biomass may be important. Such endpoints 

34 include biodiversity or species composition, and effects may result from competitive 

35 interactions among plants in mixed-species communities. Most of the available data on 

36 non-crop herbaceous species are for grasslands, with many of the recent studies 
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1 conducted in Europe. See Section 9.4.7 for a review of the recent literature on O3 effects 

2 on competition and biodiversity in grasslands.

Root Growth

3 Although O3 does not penetrate soil, it could alter root development by decreasing 

4 C assimilation via photosynthesis leading to less C allocation to the roots (Andersen, 

5 2003). The response of root development to O3 exposure depends on available

6 photosynthate within the plant and could vary over time. Many biotic and abiotic factors, 

7 such as community dynamics and drought stress, have been found to alter root 

8 development under elevated O3. An earlier study at the Aspen FACE experiment found 

9 that elevated O3 reduced coarse root and fine roots biomass in young stands of paper 

10 birch and trembling aspen (King et al., 2001). However, this reduction disappeared 

11 several years later. Ozone significantly increased fine-root (<1.0 mm) in the aspen 

12 community (Pregitzer et al., 2008). This increase in fine root production was due to 

13 changes in community composition, such as better survival of the O3-tolerant aspen 

14 genotype, birch, and maple, rather than changes in C allocation at the individual tree level 

15 (Pregitzer et al., 2008; Zak et al., 2007). In an adult European beech/Norway spruce 

16 forest in Germany, drought was found to nullify the O3-driven stimulation of fine root 

17 growth. Ozone stimulated fine-root production of beech during the humid year, but had 

18 no significant impact on fine root production in the dry year (Matyssek et al., 2010; 

19 Nikolova et al., 2010).

20 Using a non-destructive method, Vollsnes et al. (2010) studied the in vivo root 

21 development of subterranean clover (Trifolium subterraneum) before, during and after 

22 short-term O3 exposure. It was found that O3 reduced root tip formation, root elongation, 

23 the total root length, and the ratios between below- and above-ground growth within 

24 one week after exposure. Those effects persisted for up to three weeks; however, biomass 

25 and biomass ratios were not significantly altered at the harvest five weeks after exposure.

26 Several recent meta-analyses have generally indicated that O3 reduced C allocated to 

27 roots. In one meta-analysis, Grantz et al. (2006) estimated the effect of O3 on the 

28 root:shoot allometric coefficient (k), the ratio between the relative growth rate of the root 

29 and shoot. The results showed that O3 reduced the root:shoot allometric coefficient by 

30 5.6%, and the largest decline of the root:shoot allometric coefficient was observed in 

31 slow-growing plants. In another meta-analysis including 263 publications, Wittig et al. 

32 (2009) found that current O3 exposure had no significant impacts on root biomass and 

33 root:shoot ratio when compared to pre-industrial O3 exposure. However, if O3

34 concentrations rose to 81-101 ppb (projected O3 levels in 2100), both root biomass and 

35 root:shoot ratio were found to significantly decrease. Gymnosperms and angiosperms 
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1 differed in their responses, with gymnosperms being less sensitive to elevated O3. In two 

2 other meta-analyses, Wang et al. (2010) found elevated O3 reduced biomass allocation to 

3 roots by 8.3% at ambient CO2 and 6.0% at elevated CO2, and Morgan et al. (2003) found 

4 O3 reduced root dry weight of soybean. While there is clear evidence that O3 reduced C 

5 allocation to roots, results of recent individual studies have been mixed, showing negative 

6 (Jones et al., 2010), non-significant (Andersen et al., 2010; Phillips et al., 2009) and 

7 positive effects (Pregitzer et al., 2008; Grebenc and Kraigher, 2007) on root biomass and 

8 root: shoot ratio. 

9.4.3.2 Summary

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

10 ambient concentrations of O3 decrease photosynthesis and growth in numerous plant 

11 species across the U.S. Studies published since the last review continue to support that 

12 conclusion.

13 The meta-analysis by Wittig et al.(2007) and (2009) demonstrates the coherence of O3

14 effects on plant photosynthesis and growth across numerous studies and species using a 

15 variety of experimental techniques. Since the 2006 O3 AQCD, several studies were 

16 published based on the Aspen FACE experiment using “free air,” O3, and CO2 exposures 

17 in a forest in Wisconsin. Overall, the studies at the Aspen FACE experiment were 

18 consistent with many of the open-top chamber (OTC) studies that were the foundation of 

19 previous O3 NAAQS reviews. These results strengthen our understanding of O3 effects 

20 on forests and demonstrate the relevance of the knowledge gained from trees grown in 

21 open-top chamber studies.

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

23 effects on root biomass and root:shoot ratio. While the findings of individual studies were 

24 mixed, recent meta-analyses have generally indicated that O3 reduced C allocated to roots 

25 (Wittig et al., 2009; Grantz et al., 2006).

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

27 exposure and reduced growth of woody and herbaceous vegetation.

9.4.3.3 Reproduction

28 Studies during recent decades have demonstrated O3 effects on various stages of plant 

29 reproduction. The impacts of O3 on reproductive development, as reviewed by Black et 

30 al. (2000), can occur by influencing (1) age at which flowering occurs, particularly in 
Draft – Do Not Cite or Quote 9-50 September 2011

1 long-lived trees that often have long juvenile periods of early growth without flower and 

2 seed production; (2) flower bud initiation and development; (3) pollen germination and 

3 pollen tube growth; (4) seed, fruit, or cone yields; and (5) seed quality (Table 9-1) (U.S. 

4 EPA, 2006b). Several recent studies since the 2006 O3 AQCD further demonstrate the 

5 effects of O3 on reproductive processes in herbaceous and woody plant species. Although 

6 there have been documented effects of ozone on reproductive processes, a knowledge gap 

7 still exists pertaining to the exact mechanism of these responses. 

8 Rämö et al. (2007) exposed several meadow species to elevated O3 (40-50 ppb) and CO2

9 (+100 ppm), both individually and combined, over three growing seasons in ground10 planted mesocosms, using OTCs. Elevated O3 delayed flowering of Campanula 

11 rotundifolia and Vicia cracca. Ozone also reduced the overall number of produced 

12 flowers and decreased fresh weight of individual Fragaria vesca berries. 

13 Black et al. (2007) exposed Brassica campestris to 70 ppb for two days during late 

14 vegetative growth or ten days during most of the vegetative phase. The two-day exposure 

15 had no effect on growth or reproductive characteristics, while the 10 day exposure 

16 reduced vegetative growth and reproductive site number on the terminal raceme, 

17 emphasizing the importance of exposure duration and timing. Mature seed number and 

18 weight per pod were unaffected due to reduced seed abortion, suggesting that, although 

19 O3 affected reproductive processes, indeterminate species such as B. campestris possess 

20 enough compensatory flexibility to avoid reduced seed production (Black et al., 2007). 

21 In the determinate species, Plantago major, Black et al. (2010) found that O3 may have 

22 direct effects on reproductive development in populations of differing sensitivity. Only 

23 the first flowering spike was exposed to 120 ppb O3 for 7 hours per day on 9 successive 

24 days (corresponding to flower development) while the leaves and second spike were 

25 exposed to charcoal-filtered air. Exposure of the first spike to O3 affected seed number 

26 per capsule on both spikes even though spike two was not exposed. The combined seed 

27 weight of spikes one and two was increased by 19% in the two resistant populations, 

28 suggesting an overcompensation for injury; whereas, a decrease of 21% was observed in 

29 the most sensitive population (Black et al., 2010). The question remains as to whether 

30 these effects are true direct ozone-induced effects or compensatory responses.

31 A study by Darbah et al. (2008; 2007) of paper birch (Betula papyrifera) trees at the 

32 Aspen FACE site in Rhinelander, WI investigated the effects of elevated O3 and/or CO2

33 on reproductive fitness. Elevated O3 increased flowering, but decreased seed weight and 

34 germination success rate of seeds from the exposed trees. These results suggest that O3

35 can dramatically affect flowering, seed production, and seed quality of paper birch, 

36 ultimately affecting its reproductive fitness (Darbah et al., 2008; Darbah et al., 2007).
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Table 9-1 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)

Buddleia davidii Flowering time Findley et al. (1997)

Rubus cuneifolius Pollen germination Chappelka (2002)

Plantago major Pollen tube elongation Stewart (1998)

Fragaria × ananassa Fruit yield Drogoudi and Ashmore (2001); Drogoudi and 

Ashmore (2000)

Plantago major Seed yield Lyons and Barnes (1998); Pearson et al. (1996); 

Reiling and Davison (1992); Whitfield et al. (1997)

Understory herbs Seed yield Harward and Treshow (1975)

Source: Adapted from 2006 O3 AQCD 

9.4.3.4 Ecosystem Productivity and Carbon Sequestration

1 During the previous NAAQS review, there were limited studies that investigated the 

2 effect of O3 exposure on ecosystem productivity and C sequestration. Recent studies 

3 from long-term FACE experiments provide more evidence of the association of O3

4 exposure and changes in productivity at the ecosystem scale. In addition to experimental 

5 studies, model studies also assessed the impact of O3 exposure on productivity and 

6 C sequestration from stand to global scales. 

7 Two types of models are most often used to study the ecological consequences of O3

8 exposure: (1) single plant growth models such as TREGRO and PnET–II (Hogsett et al., 

9 2008; Martin et al., 2001; Ollinger et al., 1997b), and (2) process-based ecosystem 

10 models such as PnET-CN, Dynamic Land Ecosystem Model (DLEM), Terrestrial 

11 Ecosystem Model (TEM), or Met Office Surface Exchange Scheme - Top-down 

12 Representation of Interactive Foliage and Flora Including Dynamics (MOSES-TRIFFID)

13 (Felzer et al., 2009; Ren et al., 2007a; Sitch et al., 2007; Ollinger et al., 2002) (Table 9-2). 

14 In these models, carbon uptake is simulated through photosynthesis (TREGRO, PnET –

15 II, PnET- CN, DLEM and MOSES-TRIFFID) or gross primary production (TEM).

16 Photosynthesis rate at leaf level is modeled by a function of stomatal conductance and 

17 other parameters in TREGRO, PnET –II, PnET- CN, DLEM and MOSES-TRIFFID.

18 Photosynthesis at canopy level is calculated by summing either photosynthesis of 

19 different leaf types (TREGRO, DLEM, and MOSES-TRIFFID) or photosynthesis of 

20 different canopy layers (PnET –II, PnET- CN). The detrimental effect of O3 on plant 

21 growth is often simulated by multiplying photosynthesis rate by a coefficient that is 

22 dependent on stomatal conductance and cumulative O3 uptake (Table 9-2). Different 
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1 plant functional groups (PFTs, such as deciduous trees, coniferous trees or crops) show 

2 different responses to O3 exposure. PnET-II, PnET-CN, TEM, DLEM and MOSES3 TRIFFID estimate this difference by modifying net photosynthesis with coefficients that 

4 represent the O3 induced fractional reduction of photosynthesis for each functional group. 

5 The coefficients used in PnET-II, PnET-CN, TEM, DLEM are derived from the functions 

6 of O3 exposure (AOT40) versus photosynthesis reduction from Reich (1987) and 

7 Tjoelker et al. (1995). The coefficients used in MOSES-TRIFFID are derived from the 

8 O3 dose-photosynthesis response function from Pleijel et al. (2004a) and Karlsson et al. 

9 (2004), where O3 dose is estimated by a metric named CUOt (cumulative stomatal uptake 

of O3). The O3 threshold of CUOt is 1.6 nmol/m2/s for woody PFT and 5 nmol/m2 10 /s for 

11 grass PFT, and is different from AOT40, which has an O3 threshold level of 40 ppb for 

12 all PFTs. Experimental and model studies on ecosystem productivity and C sequestration 

13 at the forest stand scale as well as regional and global scales are reviewed in the 

14 following section.
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Table 9-2 Comparison of models used to simulate the ecological 

consequences of O3 exposure

Model Model 

feature Carbon uptake Ozone effect Reference

TREGRO Hourly or 

daily step, 

single plant 

model 

simulating 

vegetation 

growth 

process 

Leaf: leaf photosynthesis is a function of 

stomatal conductance, mesophyll 

conductance and the gradient of CO2 from 

atmosphere to the mesophyll cells

Canopy: Leaf is divided into different ages. 

The canopy photosynthesis rate is the sum 

the photosynthesis of all foliage groups 

The effect of O3 on photosynthesis is 

simulated by reducing mesophyll 

conductance, and increasing respiration. 

The degree of O3 damage is determined 

by ambient O3 exposure, and the 

threshold O3 concentration below which 

O3 does not affect mesophyll 

conductance and respiration

Hogsett et al. 

(2008); 

Weinstein et al. 

(2005); Tingey 

et al. (2004)

PnET-II 

and PnET 

-CN

PnET-II: 

monthly timestep, single 

plant model 

PnET –CN: 

monthly timestep, 

ecosystem 

mode

Leaf: Maximum photosynthesis rate is 

determined by a function of foliar N 

concentration, and stomatal conductance is 

determined by a function of the actual rate 

of the photosynthesis.

Canopy: canopy is divided into multiple, 

even-mass layers and photosynthesis is 

simulated by a multilayered canopy 

submodel

The effect of O3 on photosynthesis is 

simulated by an equation of stomatal 

conductance and O3 dose (AOT40). The 

model assumes that photosynthesis and 

stomatal conductance remain coupled 

under O3 exposure, with a reduction in 

photosynthesis for a given month causing 

a proportion reduction in stomatal 

conductance. 

Ollinger et al. 

(2002; 1997b); 

Pan et al. 

(2009)

TEM monthly timestep, 

ecosystem 

mode

Ecosystem: TEM is run at a 0.5*0.5 degree 

resolution. Each grid cell is classified by 

vegetation type and soil texture, and 

vegetations and detritus are assumed to 

distribute homogeneously within grid cells. 

Carbon flows into ecosystem via gross 

primary production, which is a function of 

maximum rate of assimilation, 

photosynthetically active radiation, the leaf 

area relative to the maximum annual leaf 

area, mean monthly air temperate, and 

nitrogen availability.

The direct O3 reduction on GPP is 

simulated by multiplying GPP by f(O3)t, 

where f(O3)t is determined by 

evapotranspiration, mean stomatal 

conductance, ambient AOT40, and 

empirically O3 response coefficient 

derived from previous publications. 

Felzer et al. 

(2005; 2004)

DLEM daily timestep 

ecosystem 

model

Leaf: photosynthesis is a function of 6 

parameters: photosynthetic photon flux 

density, stomatal conductance, daytime 

temperature, the atmospheric CO2

concentration, the leaf N content and the 

length of daytime.

Canopy: Photosynthetic rates for sunlit leaf 

and shaded leaf scale up to the canopy 

level by multiplying the estimated leaf area 

index

Ecosystem: GPP is the sum of gross C 

fixation of different plant function groups

The detrimental effect of O3 is simulated 

by multiplying the rate of photosynthesis 

by O3eff, where O3eff is a function of 

stomatal conductance, ambient AOT40, 

and O3 sensitive coefficient. Ozone’s 

indirect effect on stomatal conductance is 

also simulated, with a reduction in 

photosynthesis for a given month causing 

a reduction in stomatal conductance, and 

therefore canopy conductance.

Ren et al. 

(2007a; 

2007b); Zhang 

et al. (2007a)

MOSESTRIFFID30 minutes 

time-step, 

dynamic 

global 

vegetation 

model

Leaf: photosynthesis is a function of 

environmental and leaf parameters and 

stomatal conductance; Stomatal 

conductance is a function of the 

concentration of CO2 and H2O in air at the 

leaf surface and the current rate of 

photosynthesis of the leaf

Canopy: Photosynthetic rates scale up to 

the canopy level by multiplying a function of 

leaf area index and PAR extinction 

coefficient

Ecosystem: GPP is the sum of gross C 

fixation of different plant function groups

The effect of O3 is simulated by 

multiplying the rate of photosynthesis by 

F, where F depends upon stomatal 

conductance, O3 exposure, a critical 

threshold for O3 damage, and O3

sensitive coefficient (functional type 

dependent)

Sitch et al. 

(2007)
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Local Scale

1 The above- and below-ground biomass and net primary production (NPP) were measured 

2 at the Aspen FACE site after 7 years of O3 exposure. Elevated O3 caused 23, 13 and 14% 

3 reductions in total biomass relative to the control in the aspen, aspen–birch and aspen–

4 maple communities, respectively (King et al., 2005). At the Kranzberg Forest FACE 

experiment in Germany, O3 reduced annual volume growth by 9.5 m3 5 /ha in a mixed 

6 mature stand of Norway spruce and European beech (Pretzsch et al., 2010). At the 

7 grassland FACE experiment at Alp Flix, Switzerland, O3 reduced the seasonal mean rates 

8 of ecosystem respiration and GPP by 8%, but had no significant impacts on aboveground 

9 dry matter productivity or growing season net ecosystem production (NEP) (Volk et al., 

10 2011). Ozone also altered C accumulation and turnover in soil, as discussed in Section 

11 9.4.6.

12 Changes in forest stand productivity under elevated O3 were assessed by several model 

13 studies. TREGRO (Table 9-2) has been widely used to simulate the effects of O3 on the 

14 growth of several species in different regions in the U.S. Hogsett et al. (2008) used 

15 TREGRO to evaluate the effectiveness of various forms and levels of air quality 

16 standards for protecting tree growth in the San Bernardino Mountains of California. They 

17 found that O3 exposures at the Crestline site resulted in a mean 20.9% biomass reduction 

18 from 1980 to 1985 and 10.3% biomass reduction from 1995 to 2000, compared to the 

19 ‘‘background’’ O3 concentrations (O3 concentration in Crook County, Oregon). The 

20 level of vegetation protection projected was different depending on the air quality 

21 scenarios under consideration. Specifically, when air quality was simulated to just meet 

22 the California 8-h average maximum of 70 ppb and the maximum 3 months 12-h SUM06 

23 of 25 ppm-h, annual growth reductions were limited to 1% or less, while air quality that 

24 just met a previous NAAQS (the second highest 1-h max [125 ppb]) resulted in 6-7% 

25 annual reduction in growth, resulting in the least protection relative to background O3

26 (Hogsett et al., 2008).

27 ZELIG is a forest succession gap model, and has been used to evaluate the dynamics of 

28 natural stand succession. Combining TREGRO with ZELIG, Weinstein et al. (2005)

29 simulated the effects of different O3 levels ( 0.5, 1.5, 1.75, and 2 times ambient) on the 

30 growth and competitive interactions of white fir and ponderosa pine at three sites in 

31 California: Lassen National Park, Yosemite National Park, and Crestline. Their results 

32 suggested that O3 had little impact on white fir, but greatly reduced the growth of 

33 ponderosa pine. If current O3 concentrations continue over the next century, ambient O3

34 exposure (SUM06 of 110 ppm-h) at Crestline was predicted to decrease individual tree 

35 C budget by 10% and decrease ponderosa pine abundance by 16%. Effects at Lassen 
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1 National Park and Yosemite National Park sites were found to be smaller because of 

2 lower O3 exposure levels (Weinstein et al., 2005).

3 The effects of O3 on stand productivity and dynamics were also studied by other tree 

4 growth or stand models, such as ECOPHYS, INTRASTAND and LINKAGES. 

5 ECOPHYS is a functional-structural tree growth model. The model used the linear 

6 relationship between the maximum capacity of carboxylation and O3 dose to predict the 

7 relative effect of O3 on leaf photosynthesis (Martin et al., 2001). Simulations with 

8 ECOPHYS found that O3 decreased stem dry matter production, stem diameter and leaf 

9 dry matter production, induced earlier leaf abscission, and inhibited root growth (Martin 

10 et al., 2001). INTRASTAND is an hourly time step model for forest stand carbon and 

11 water budgets. LINKAGES is a monthly time step model simulating forest growth and 

12 community dynamics. Linking INTRASTAND with LINKAGES, Hanson et al. (2005)

13 found that a simulated increase in O3 concentration in 2100 (a mean 20-ppb increase over 

14 the current O3 concentration) yields a 35% loss of net ecosystem C exchange (NEE) with 

respect to the current conditions (174 g C/m2 15 /year). 

Regional and Global Scales

16 Since the publication of the 2006 O3 AQCD, there is additional evidence suggesting that 

17 O3 exposure alters ecosystem productivity and biogeochemical cycling at the regional 

18 and continental scale. Most of those studies were conducted by using process-based 

19 ecosystem models (Table 9-2) and are briefly reviewed in the following sections.

20 Ollinger et al. (1997a) simulated the effect of O3 on hardwood forest productivity of 64 

21 hardwood sites in the northeastern U.S. with PnET-II (Table 9-2). Their simulations 

22 indicated that O3 caused a 3-16% reduction in NPP from 1987 to 1992 (Table 9-3). The 

23 interactive effects of O3, N deposition, elevated CO2 and land use history on C dynamics 

24 were estimated by PnET-CN (Table 9-2) (Ollinger et al., 2002). The results indicated that 

25 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 26 /year) under 

27 timber harvest site history. PnET-CN was also used to assess changes in C sequestration 

28 of U.S. Mid-Atlantic temperate forest. Pan et al. (2009) designed a factorial modeling 

29 experiment to separate the effects of changes in atmospheric composition, historical 

30 climatic variability and land-disturbances on the C cycle. They found that O3 acted as a 

31 negative factor, partially offsetting the growth stimulation caused by elevated CO2 and N 

32 deposition in U.S. Mid-Atlantic temperate forest. Ozone decreased NPP of most forest 

33 types by 7-8%. Among all the forest types, spruce-fir forest was most resistant to O3

34 damage, and NPP decreased by only 1% (Pan et al., 2009). 
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1 Felzer et al. (2004) developed TEM 4.3 (Table 9-2) to simulate the effects of O3 on plant 

2 growth and estimated effects of O3 on NPP and C sequestration of deciduous trees, 

3 conifers and crops in the conterminous U.S. The results indicated that O3 reduced NPP 

4 and C sequestration in the U.S. (Table 9-3) with the largest decreases (over 13% in some 

5 locations) in NPP occurring in the Midwest agricultural lands during the mid-summer. 

6 TEM was also used to evaluate the magnitude of O3 damage at the global scale (Table 9-

7 3) (Felzer et al., 2005). Simulations for the period 1860 to 1995 show that the largest 

8 reductions in NPP and net C exchange occurred in the mid western U.S., eastern Europe, 

9 and eastern China (Felzer et al., 2005). DLEM (Table 9-2) was developed to simulate the 

10 detrimental effect of O3 on ecosystems, and has been used to examine the O3 damage on 

11 NPP and C sequestration in Great Smoky Mountains National Park (Zhang et al., 2007a), 

12 grassland ecosystems and terrestrial ecosystems in China (Ren et al., 2007a; Ren et al., 

13 2007b). Results of those simulations are listed in Table 9-3. 

14 Instead of using AOT40 as their O3 exposure metric as PnET, TEM and DLEM did, 

15 Sitch et al. (2007) incorporated a different O3 metric named CUOt (cumulative stomatal 

16 uptake of O3), derived from Pleijel et al. (2004a), into the MOSES-TRIFFID coupled 

17 model (Table 9-2). In the CUOt metric, the fractional reduction of plant production is 

18 dependent on O3 uptake by stomata over a critical threshold for damage with this 

19 threshold level varying by plant functional type. Consistent with previous studies, their 

20 model simulation indicated that O3 reduced global gross primary production (GPP), 

21 C exchange rate and C sequestration (Table 9-3). The largest reductions in GPP and land22 C storage were projected over North America, Europe, China and India. In the model, 

23 reduced ecosystem C uptake due to O3 damage, results in additional CO2 accumulation 

24 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 25 ) could have 

26 even greater impact on global warming than the direct radiative forcing of O3

(0.89 W/m2 27 ) (Sitch et al., 2007).

28 Results from the various model studies presented in Table 9-3 are difficult to compare 

29 because of the various spatial and temporal scales used in these studies. However, all the 

30 studies showed that O3 exposure decreased ecosystem productivity and C sequestration. 

31 These results are consistent and coherent with experimental results from the leaf, plant 

32 and ecosystem scales (Sitch et al., 2007; Felzer et al., 2005). Many of the models use the 

33 same underlying function to simulate the effect of O3 exposure to C uptake. For example 

34 the functions of O3 exposure (AOT40) versus photosynthesis reduction for PnET-II, 

35 PnET-CN, TEM, DLEM were all from Reich (1987) and Tjoelker et al. (1995). 

36 Therefore, it is not surprising that the results are similar. While these models can be 

37 improved and more evaluation with experimental data can be done, these models 
Draft – Do Not Cite or Quote 9-57 September 2011

1 represent the state of the science for estimating the effect of O3 exposure on productivity 

2 and C sequestration.

9.4.3.5 Summary

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

4 effect of O3 exposure on ecosystem productivity and C sequestration. Recent studies 

5 from long-term FACE experiments, such as Aspen FACE, SoyFACE and the Kranzberg 

6 Forest (Germany), provided evidence of the association of O3 exposure and reduced 

7 productivity at the ecosystem level. Studies at the leaf and plant scales showed that O3

8 reduced photosynthesis and plant growth, which provided coherence and biological 

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

10 models, such as TREGRO, PnET, TEM and DLEM, were consistent with the FACE 

11 experimental evidence, which showed that O3 reduced ecosystem productivity. 

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

13 response varies among plant communities. For example, O3 had little impact on white fir, 

14 but greatly reduced growth of ponderosa pine in southern California (Weinstein et al., 

15 2005). Ozone decreased net primary production (NPP) of most forest types in Mid16 Atlantic region, but had small impacts on spruce-fir forest (Pan et al., 2009). 

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

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

19 simulations consistently found that O3 exposure caused negative impacts on those 

20 indicators, but the severity of these impacts was influenced by multiple interactions of 

21 biological and environmental factors. The suppression of ecosystem C sinks results in 

22 more CO2 accumulation in the atmosphere. Globally, the indirect radiative forcing caused 

23 by O3 exposure through lowering ecosystem C sink could have an even greater impact on 

24 global warming than the direct radiative forcing of O3 (Sitch et al., 2007). Ozone could 

25 also affect regional C budgets through interacting with multiple factors, such as N 

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

27 partially offset the growth stimulation caused by elevated CO2 and N deposition in both 

28 Northeast- and Mid-Atlantic-region forest ecosystems of the U.S. (Pan et al., 2009; 

29 Ollinger et al., 2002).

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

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

32 exposure and reduced carbon sequestration in terrestrial ecosystems.
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Table 9-3 Modeled effects of ozone on primary production, C exchange, and 

C sequestration 

Scale Model Index Ozone Impacts Reference

GPP Global MOSESTRIFFID CUOta Decreased by 14-23% over the period 1901-2100 Sitch et al. 

(2007)

NPP

Global TEM AOT40

Decreased by 0.8% without agricultural 

management and a decrease of 2.9% with optimal 

agricultural management

Felzer et al. 

(2005)

U.S. TEM AOT40 Reduced by 2.3% without optimal N fertilization and 

7.2% with optimal N fertilization from 1983-1993

Felzer et al. 

(2005)

U.S. TEM AOT40 Reduced by 2.6–6.8% during the late 1980s-early 

1990s. 

Felzer et al. 

(2004)

Northeastern 

U.S. PnET AOT40 A reduction of 3-16% from 1987-1992 Ollinger et al. (1997a)

U.S. MidAtlantic PnET AOT40 Decreased NPP of most forest types by 7-8% Pan et al. 

(2009)

China DLEM AOT40 Reduced NPP of grassland in China by 8.5 Tg C 

from 1960s to 1990s

Ren et al. 

(2007b)

C exchange

Global TEM AOT40

Reduced net C exchange (1950–1995) by 0.1 Pg 

C/yr without agricultural management and 0.3 Pg 

C/yr with optimal agricultural management

Felzer et al. 

(2005)

Global MOSESTRIFFID CUOt

Decreased global mean land–atmosphere C fluxes 

by 1.3 Pg C/yr and 1.7 Pg C/yr for the ‘high’ and 

‘low’ plant O3 sensitivity models, respectively

Sitch et al. 

(2007)

C sequestration

Global MOSESTRIFFID CUOt Reduced land-C storage accumulation by between 

143 Pg C/yr and 263 Pg C/yr from 1900–2100

Sitch et al. 

(2007)

U.S. TEM AOT40 Reduced C sequestration by 18–38 Tg C/yr from 

1950 to 1995

Felzer et al. 

(2004)

GSM National 

Park DLEM AOT40

Decreased the ecosystem C storage of deciduous 

forests by 2.5% and pine forest by 1.4% from 1971 

to 2001

Zhang et al. 

(2007a)

China DLEM AOT40 Reduced total C storage by 0.06% in 1960s and

1.6% in 1990s in China’s terrestrial ecosystems

Ren et al. 

(2007a)

China DLEM AOT40 O3 exposure reduced the net C sink of China’s 

terrestrial ecosystem by 7% from 1961 to 2005

Tian et al. 

(2011)

China DLEM AOT40

Ozone induced net carbon exchange reduction 

ranged from 0.4-43.1% , depending on different 

forest type

Ren et al. 

(2011)

a

CUOt is defined as the cumulative stomatal uptake of O3, using a constant O3-uptake rate threshold of t nmol/m2/s. d

Pg equals 1 × 1015 grams.

1

9.4.4 Crop yield and quality in agricultural systems

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

3 a large body of research has stemmed from that recognition. Previous O3 AQCDs have 

4 extensively reviewed this body of literature. Table 9-4 summarizes recent experimental 

5 studies of O3 effects on agricultural crops, exclusive of growth and yield. Growth and 

6 yield results are summarized in Table 9-17. 
Draft – Do Not Cite or Quote 9-59 September 2011

1 The actual concentration and duration threshold for O3 damage varies from species to 

2 species and sometimes even among genotypes of the same species (Guidi et al., 2009; 

3 Sawada and Kohno, 2009; Biswas et al., 2008; Ariyaphanphitak et al., 2005; Dalstein and 

4 Vas, 2005; Keutgen et al., 2005). A number of comprehensive reviews and meta-analyses 

5 have recently been published discussing both the current understanding of the 

6 quantitative effects of O3 concentration on a variety of crop species and the potential 

7 focus areas for biotechnological improvement to a future growing environment that will 

8 include higher O3 concentrations (Bender and Weigel, 2011; Booker et al., 2009; 

9 Van Dingenen et al., 2009; Ainsworth, 2008; Feng et al., 2008b; Hayes et al., 2007; Mills 

10 et al., 2007b; Grantz et al., 2006; Morgan et al., 2003). Since the 2006 O3 AQCD(U.S. 

11 EPA, 2006b), exposure-response indices for a variety of crops have been suggested 

12 (Mills et al., 2007b) and many reports have investigated the effects of O3 concentration 

13 on seed or fruit quality to extend the knowledge base beyond yield quantity. This section 

14 will outline the key findings from these papers as well as highlight some of the recent 

15 research addressing the endpoints such as yields and crop quality. 

16 This section will also highlight recent literature that focuses on O3 damage to crops as 

17 influenced by other environmental factors. Genetic variability is not the only factor that 

18 determines crop response to O3 damage. Ozone concentration throughout a growing19 season is not homogeneous and other environmental conditions such as elevated CO2

20 concentrations, drought, cold or nutrient availability may alleviate or exacerbate the 

21 oxidative stress response to a given O3 concentration. 

9.4.4.1 Yield

22 It is well known that yield is negatively impacted in many crop species in response to 

23 high O3 concentration. However, the concentrations at which damage is observed vary 

24 from species to species. Numerous analyses of experiments conducted in OTCs and with 

25 naturally occurring gradients demonstrate that the effects of O3 exposure also vary 

26 depending on the growth stage of the plant; plants grown for seed or grain are often most 

27 sensitive to exposure during the seed or grain-filling period (Soja et al., 2000; Pleijel et 

28 al., 1998; Younglove et al., 1994; Lee et al., 1988a). AX9.5.4.1 of the 2006 O3 AQCD 

29 summarized many previous studies on crop yield. 

Field studies and meta-analyses

30 The effect of O3 exposure on U.S. crops remains an important area of research and

31 several studies have been published on this topic since the 2006 O3 AQCD (U.S. EPA, 

32 2006b) (Table 9-4 and 9-17). For example, one study with cotton in a crop-weed 
Draft – Do Not Cite or Quote 9-60 September 2011

1 interaction study (Grantz and Shrestha, 2006) utilizing OTCs suggests that 12-hour 

2 average O3 concentrations of 79.9 ppb decreased cotton biomass by 25% and 12-hour 

3 average O3 concentration of 122.7 ppb decreased cotton biomass by 75% compared to 

4 charcoal filtered control (12-h avg: 12.8 ppb). Further, this study suggests that the weed, 

5 yellow nutsedge, was less sensitive to increasing O3 concentration, which would increase 

6 weed competition (Grantz and Shrestha, 2006). In a study of peanuts in North Carolina, 

7 near ambient and elevated exposures of O3 reduced photosynthesis and yield compared to 

8 very low O3 conditions (Booker et al., 2007; Burkey et al., 2007). In another study, 

9 Grantz and Vu (2009) reported that sugarcane biomass growth significantly declined 

10 under O3 exposure. 

11 The average yield loss reported across a number of meta-analytic studies have been 

12 published recently for soybean (Morgan et al., 2003), wheat (Feng et al., 2008b), rice 

13 (Ainsworth, 2008), semi-natural vegetation (Hayes et al., 2007), potato, bean and barley 

14 (Feng and Kobayashi, 2009). Meta-analysis allows for the objective development of a 

15 quantitative consensus of the effects of a treatment across a wide body of literature. 

16 Further, this technique allows for a compilation of data across a range of O3 fumigation 

17 techniques, durations and concentrations in order to assemble the existing literature in a 

18 meaningful manner.

19 Morgan et al. (2003) reported an average seed yield loss for soybean of 24% compared to 

20 charcoal filtered air across all O3 concentrations used in the 53 compiled studies. The 

21 decrease in seed yield appeared to be the product of nearly equal decreases (7-12%) in 

22 seed weight, seed number and pod number. As would be expected, the lowest O3

23 concentration (30-59 ppb) resulted in the smallest yield losses, approximately 8%, while 

24 the highest O3 concentration (80-120 ppb ) resulted in the largest yield losses, 

25 approximately 35% (Morgan et al., 2003). Further, the oil/protein ratio within the 

26 soybean seed was altered due to growth at elevated O3 concentrations, with a decrease in 

27 oil content. The studies included in this meta-analysis all used enclosed fumigation 

28 systems or growth chambers which may have altered the coupling of the atmosphere to 

29 the lower plant canopy (McLeod and Long, 1999) , although the results of Morgan et al. 

30 (2006), Betzelberger et al. (2010), and the comparisons presented in Section 9.6.3 

31 strongly suggest that decreases in yield between ambient and elevated exposures are not 

32 affected by exposure method. Utilizing the Soybean Free Air gas Concentration 

33 Enrichment Facility (SoyFACE; www.soyface.illinois.edu), Morgan et al. (2006) report a 

34 20% seed yield loss due to a 23% increase in average daytime O3 concentration 

35 (56-69 ppb) within a single soybean cultivar across two growing seasons in Illinois, 

36 which lies within the range predicted by the meta-analysis. A further breakdown of the 

37 effects of current O3 concentrations (AOT40 of 4.7 ppm-h) on bean seed quality 

38 (Phaseolus vulgaris) has identified that growth at current O3 concentrations compared to 
Draft – Do Not Cite or Quote 9-61 September 2011

1 charcoal-filtered air raised total lipids, total crude protein and dietary fiber content (Iriti et 

2 al., 2009). An increase in total phenolics was also observed, however the individual 

3 phenolics compounds responded differently, with significant decreases in anthocyanin 

4 content. The seeds from ambient O3 exposed plants also displayed increased total 

5 antioxidant capacity compared to charcoal-filtered air controls (Iriti et al., 2009). 

6 Betzelberger et al. (2010) has recently utilized the SoyFACE facility to compare the 

7 impact of elevated O3 concentrations across 10 soybean cultivars to investigate 

8 intraspecific variability of the O3 response to find physiological or biochemical markers 

9 for eventual O3 tolerance breeding efforts (Betzelberger et al., 2010). They report an 

10 average 17% decrease in yield across all 10 cultivars across two growing seasons due to a 

11 doubling of ambient O3 concentrations, with the individual cultivar responses ranging 

12 from -7% to -36%. The exposure-response functions derived for these 10 current 

13 cultivars were similar to the response functions derived from the NCLAN studies 

14 conducted in the 1980s (Heagle, 1989) suggesting there has not been any selection for 

15 increased tolerance to O3 in more recent cultivars. More complete comparisons between 

16 yield predictions based on data from cultivars used in NCLAN studies, and yield data for 

17 modern cultivars from SoyFACE are reported in Section 9.6.3 of this document. They 

18 confirm that the response of soybean yield to O3 exposure has not changed in current 

19 cultivars. 

20 A meta-analysis has also been performed on studies investigating the effects of O3

21 concentrations on wheat (Feng et al., 2008b). Across 23 studies included, elevated O3

22 concentrations (ranging from a 7-h daily average of 31-200 ppb) decreased grain yield by 

23 29%. Winter wheat and spring wheat did not differ in their responses; however the 

24 response in both varieties to increasing O3 concentrations resulted in successively larger 

25 decreases in yield, from a 20% decrease in 42 ppb to 60% in 153 ppb O3. These yield 

26 losses were mainly caused by a combination of decreases in individual grain weight (-

27 18%), ear number per plant (-16%), and grain number per ear (-11%). Further, the grain 

28 starch concentration decreased by 8% and the grain protein yield decreased by 18% due 

29 to growth at elevated O3 concentrations as well. However, increases in grain calcium and 

30 potassium levels were reported (Feng et al., 2008b).

31 A recent meta-analysis found that growth at elevated O3 concentrations negatively 

32 impacts nearly every aspect of rice performance as well (Ainsworth, 2008). While rice is 

33 not a major crop in the U.S., it provides a staple food for over half of the global 

34 population (IRRI, 2002) and the effects of rising O3 concentrations on rice yields merits 

35 consideration. On average, rice yields decreased 14% in 62 ppb O3 compared to charcoal36 filtered air. This yield loss was largely driven by a 20% decrease in grain number 

37 (Ainsworth, 2008). 
Draft – Do Not Cite or Quote 9-62 September 2011

1 Feng and Kobayashi (2009) have recently compiled yield data for six major crop species, 

2 potato, barley, wheat, rice, bean and soybean and grouped the O3 treatments used in those 

3 studies into three categories: baseline O3 concentrations (<26 ppb), current ambient 7- or 

4 12-h daily O3 concentrations (31-50 ppb), and future ambient 7- or 12-h daily O3

5 concentrations (51-75 ppb). Using these categories, they have effectively characterized 

6 the effects of current O3 concentrations and the effects of future O3 concentrations 

7 compared to baseline O3 concentrations. At current O3 concentrations, which ranged 

8 from 41-49 ppb in the studies included, soybean (-7.7%), bean (-19.0%), barley (-8.9%), 

9 wheat (-9.7%), rice (-17.5%) and potato (-5.3%) all had yield losses compared to the 

10 baseline O3 concentrations (<26 ppb). At future O3 concentrations, averaging 63 ppb, 

11 soybean (-21.6%), bean (-41.4%), barley (-14%), wheat (-28%), rice (-17.5%) and potato 

12 (-11.9%) all had significantly larger yield losses compared to the losses at current O3

13 concentrations (<26 ppb) (Feng and Kobayashi, 2009). 

14 A review of OTC studies has determined the AOT40 critical level that causes a 5% yield 

15 reduction across a variety of agricultural and horticultural species (Mills et al., 2007b). 

16 The authors classify the species studied into three groups: sensitive, moderate and 

17 tolerant. The sensitive crops, including watermelon, beans, cotton, wheat, turnip, onion, 

18 soybean, lettuce, and tomato, respond with a 5% reduction in yield under a 3-month 

19 AOT40 of 6 ppm-h. Watermelon was the most sensitive with a critical level of 

20 1.6 ppm-h. The moderately sensitive crops, including sugar beet, oilseed rape, potato, 

21 tobacco, rice, maize, grape and broccoli, responded with a 5% reduction in yield between 

22 8.6 and 20 ppm-h. The crops classified as tolerant, including strawberry, plum and barley, 

23 responded with a 5% yield reduction between 62-83.3 ppm-h (Mills et al., 2007b).

24 Feng and Kobayashi (2009) compared their exposure-response results to those published 

25 by Mills et al. (2007b) and found the ranges of yield loss to be similar for soybean, rice 

26 and bean. However, Feng and Kobayasi (2009) reported smaller yield losses for potato 

27 and wheat and larger yield losses for barley compared to the dose-response functions 

28 published by Mills et al. (2007b), which they attributed to their more lenient criteria for 

29 literature inclusion.

30 While the studies investigating the impact of various O3 concentrations on yield are 

31 important and aid in determining the vulnerability of various crops to a variety of O3

32 concentrations, there is still uncertainty as to how these crops respond under field 

33 conditions with interacting environmental factors such as temperature, soil moisture, CO2

34 concentration, and soil fertility (Booker et al., 2009). Further, there appears to be a 

35 distinct developmental and genotype dependent influence on plant sensitivity to O3 that 

36 has yet to be fully investigated across O3 concentrations in a field setting. The potentially 

37 mitigating effect of breeding selection for O3 resistance has received very little attention 
Draft – Do Not Cite or Quote 9-63 September 2011

1 in the published scientific literature. Anecdotal reports suggest that such selection may 

2 have occurred in recent decades for some crops in areas of the country with high ambient 

3 exposures. However, the only published literature available is on soybean and these 

4 studies indicate that sensitivity has not changed in cultivars of soybean between the 

5 1980s and the 2000s (Betzelberger et al., 2010). This conclusion for soybeans is 

6 confirmed by comparisons presented in Section 9.6.3 of this document.

Yield loss at regional and global scales 

7 Because O3 is heterogeneous in both time and space and O3 monitoring stations are 

8 predominantly near urban areas, the impacts of O3 on current crop yields at large spatial 

9 scales are difficult to estimate. Fishman et al. (2010) have used satellite observations to 

10 estimate O3 concentrations in the contiguous tri-state area of Iowa, Illinois and Indiana 

11 and have combined that information with other measured environmental variables to 

12 model the historical impact of O3 concentrations on soybean yield across the 2002-2006 

13 growing seasons. When soybean yield across Iowa, Indiana and Illinois was modeled as a 

14 function of seasonal temperature, soil moisture and O3 concentrations, O3 had the largest 

15 contribution to the variability in yield for the southern-most latitudes included in the 

16 dataset. Fishman et al. (2010) determined that O3 concentrations significantly reduced 

17 soybean yield by 0.38 to 1.63% for every additional ppb of exposure across the 5 years. 

18 This value is consistent with previous chamber studies (Heagle, 1989) and results from 

19 SoyFACE (Morgan et al., 2006). Satellite estimates of tropospheric O3 concentrations 

20 exist globally (Fishman et al., 2008), therefore utilizing this historical modeling approach 

21 is feasible across a wider geographical area, longer time-span and perhaps for more crop 

22 species.

23 The detrimental effects of O3 on crop production at regional or global scales were also 

24 assessed by several model studies. Two large scale field studies were conducted in the 

25 U.S. (NCLAN) and in Europe (European Open Top Chamber Programme, EOTCP) to 

26 assess the impact of O3 on crop production. Ozone exposure-response regression models 

27 derived from the two programs have been widely used to estimate crop yield loss 

28 (Avnery et al., 2011a, b; Van Dingenen et al., 2009; Tong and Mauzerall, 2008; Wang 

29 and Mauzerall, 2004). Those studies found that O3 generally reduced crop yield and that 

30 different crops showed different sensitivity to O3 pollution (Table 9-5). Ozone was 

31 calculated to induce a possible 45-82 million metric tons loss for wheat globally. 

32 Production losses for rice, maize and soybean were on the order of 17-23 million metric 

33 tons globally (Van Dingenen et al., 2009). The largest yield losses occur in high34 production areas exposed to high O3 concentrations, such the Midwest and the 

35 Mississippi Valley regions in the U.S., Europe, China and India (Van Dingenen et al., 

36 2009; Tong et al., 2007).
Draft – Do Not Cite or Quote 9-64 September 2011

9.4.4.2 Crop Quality

1 In general, it appears that increasing O3 concentrations above current ambient 

2 concentrations can cause species-dependent biomass losses, decreases in root biomass 

3 and nutritive quality, accelerated senescence and shifts in biodiversity. A study conducted 

4 with highbush blackberry has demonstrated decreased nutritive quality with increasing 

5 O3 concentration despite no change in biomass between charcoal-filtered control, 

6 ambient O3 and 2 × ambient O3 exposures (Ditchkoff et al., 2009). A study conducted 

7 with sedge using control (30 ppb), low (55 ppb), medium (80 ppb) and high (105 ppb) O3

8 treatments has demonstrated decreased root biomass and accelerated senescence in the 

9 medium and high O3 treatments (Jones et al., 2010). Alfalfa showed no biomass changes 

10 across two years of double ambient O3 concentrations (AOT40 of 13.9 ppm-h) using 

11 FACE fumigation (Maggio et al., 2009). However a modeling study has demonstrated 

12 that 84% of the variability in the relative feed value in high-yielding alfalfa was due to 

13 the variability in mean O3 concentration from 1998-2002 (Lin et al., 2007). Further, in a 

14 managed grassland FACE system, the reduction in total biomass harvest over five years 

15 decreased twice as fast in the elevated treatment (AOT40 of 13-59 ppm-h) compared to 

16 ambient (AOT40 of 1-20.7 ppm-h). Compared with the ambient control, loss in annual 

17 dry matter yield was 23% after 5 year. Further, functional groups were differentially 

18 affected, with legumes showing the strongest negative response (Volk et al., 2006). 

19 However, a later study by Stampfli and Fuhrer (2010) at the same site suggested that 

20 Volk et al.(2006) was likely overestimated the effects of O3 on yield reduction because 

21 the overlapping effects of species dynamics caused by heterogeneous initial conditions 

22 and a change in management were not considered in Volk et al. (2006). An OTC study 

23 conducted with Trifolium subterraneum exposed to filtered ( <15 ppb), ambient, and 

24 40 ppb above ambient O3 demonstrates decreases in biomass in the highest O3 treatment 

25 as well as 10-20% decreased nutritive quality which was mainly attributed to accelerated 

26 senescence (Sanz et al., 2005). A study conducted with Eastern gamagrass and big 

27 bluestem in OTCs suggested that big bluestem is not sensitive to O3, but gamagrass 

28 displayed decreased nutritive quality in the 2 × ambient O3 treatment, due to higher 

29 lignin content and decreased N, (Lewis et al., 2006).

9.4.4.3 Summary

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

31 and a large body of research has subsequently stemmed from those initial findings. 

32 Previous O3 AQCDs have extensively reviewed this body of literature (U.S. EPA, 

33 2006b). Current O3 concentrations across the U.S. are high enough to cause yield loss for 

34 a variety of agricultural crops including, but not limited to, soybean, wheat, potato, 
Draft – Do Not Cite or Quote 9-65 September 2011

1 watermelon, beans, turnip, onion, lettuce, and tomato. Continued increases in O3

2 concentration may further decrease yield in these sensitive crops. Despite the well3 documented yield losses due to increasing O3 concentration, there is still a knowledge 

4 gap pertaining to the exact mechanisms of O3-induced yield loss. Research has linked 

5 increasing O3 concentration to decreased photosynthetic rates and accelerated 

6 senescence, which are related to yield.

7 New research is beginning to consider the mechanism of damage caused by prolonged, 

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

9 concentration (so-called acute exposure). Both types of O3 exposure cause damage to 

10 agricultural crops, but through very different mechanisms. Historically, most research on 

11 the mechanism of O3 damage used acute exposure studies. During the last decade, it has 

12 become clear that the cellular and biochemical processes involved in the response to 

13 acute O3 exposure are not involved in response to chronic O3 exposure, even though both 

14 cause yield loss in agriculturally important crops.

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

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

17 concentrations in fruits and vegetable crops, and decreases cotton fiber quality. These 

18 areas of research require further investigation to determine mechanisms and exposure19 response relationships.

20 During the previous NAAQS reviews, there were very few studies that estimated O3

21 impacts on crop yields at large spatial scales. Recent modeling studies found that O3

22 generally reduced crop yield, but the impacts varied across regions and crop species. For 

23 example, the largest O3-induced crop yield losses occurred in high-production areas 

24 exposed to high O3 concentrations, such the Midwest and the Mississippi Valley regions 

25 of the U.S. (Van Dingenen et al., 2009). Among crop species, the estimated yield loss for 

26 wheat and soybean were higher than for rice and maize (Van Dingenen et al., 2009). 

27 Using satellite air-column observations with direct air-sampling O3 data, Fishman et al. 

28 (2010) modeled the yield-loss due to O3 over the continuous tri-state area of Illinois, 

29 Iowa and Wisconsin. They determined that O3 concentrations significantly reduced 

30 soybean yield, which further reinforces previous results from FACE-type experiments 

31 and OTC experiments. 

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

33 exposure and reduced yield and quality of agricultural crops.
Draft – Do Not Cite or Quote 9-66 September 2011

Table 9-4 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 - h/day 

85 ppb

(Exposure duration)

Relative feed value n.s. 

*high variability among 

treatment groups (N/A)

Muntifering et al. 

(2006)

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)

Big Blue Stem 

(Andropogon gerardii)

OTC

Alabama, U.S.

4 months 12-h avg: 

CF = 14 ppb; 

Ambient = 29 ppb; 

Elevated = 71 ppb

(N/A)

Relative feed value n.s. (n.s.) Lewis et al. (2006)

Brassica napus

Growth chambers

Belgium

4 days CF & 176 ppb 

for 4 h/day

(N/A)

Glucosinolates −41 (N/A) Gielen et al. (2006)

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. 

(2009b)

Eastern Gamagrass 

(Tripsacum 

dactyloides)

OTC

Alabama, U.S.

4 months 12-h avg: 

CF = 14 ppb; 

Ambient = 29 ppb; 

Elevated = 71 ppb

(N/A)

Relative feed value −17 (−12) Lewis et al. (2006)

Lettuce 

(Lactuca sativa)

OTC

Carcaixent 

Experimental Station, 

Spain

30 days 12-h mean: 

CF = 10.2 ppb; 

NF = 30.1 ppb; 

NF+O3 = 62.7 ppb

(4 cultivars)

Lipid peroxidation; 

Root length

+77 (+38)

−22 (−14)

Calatayud et. al. 

(2002)

Peanut 

(Arachis hypogaea)

OTC

Raleigh, NC; U.S.

3 yr 12-h avg: 

CF = 22 ppb; 

Ambient = 46 ppb; 

Elevated = 75 ppb

(CO2: 375 ppm; 

548 ppm; 730 ppm)

Harvest biomass −40 (−10) Booker et al. 

(2007)

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)

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)

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)

Soybean

OTC

Italy

3 yr AOT40: 

CF = 0 ppm-h; 

Ambient = 3.4 ppm-h; 

Elevated = 9.0 ppm-h

(Well-watered & 

water-stressed)

Daily 

evapotranspiration

−28 (−14) Jaude et al. (2008)
Draft – Do Not Cite or Quote 9-67 September 2011

Species

Facility

Location

Exposure 

Duration

Ozone Exposurea

(Additional treatment) Variable(s) measured

percent change from 

CFb

(percent change from 

ambient)

Reference

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)

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)

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)

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., 

2004a)

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)

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. (2008a)

Strawberry 

(Fragaria x ananassa

Duch. Cv. Korona 

& Elsanta)

Growth chambers

Bonn, Germany

2 months 8-h avg: 

CF = 0 ppb; 

Elevated = 78 ppb

(N/A)

Total leaf area −16 (N/A) Keutgen et al. 

(2005)

Sweet Potato

Growth Chambers

Bonn, Germany

4 wk 8-h avg: 

CF = 0 ppb; 

Ambient < 40 ppb; 

Elevated = 255 ppb

(N/A)

Tuber weight −14 (−11.5) Keutgen et al. 

(2008)

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)

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)

a

Ozone exposure in ppb unless otherwise noted.

b

CF = Carbon-filtered air.

NF = Non-filtered air.
Draft – Do Not Cite or Quote 9-68 September 2011

Table 9-5 Modeled effects of ozone on crop yield loss at regional and global 

scales

Scale Index Ozone Impacts Reference

Global M7a; M12b; 

AOT40

Reduced by 7.3% to 12.3% for wheat, 5.4% to 15.6% for soybean, 2.8% to 3.7% for rice, 

and 2.4% to 4.1% for maize in year 2000.

Van Dingenen et al. 

(2009)

Global M12b; AOT40 O3-induced global yield reductions ranged from 8.5-14% for soybean, 3.9-15% for wheat, 

and 2.2-5.5% for maize in year 2000. Global crop production losses totaled 79-121 million 

metric tons, worth $11-18 billion annually (USD2000).

Avnery et al. (2011a)

U.S. M7; M12; AOT40 Reduced by 4.1% to 4.4% for wheat, 7.1% to 17.7% for soybean, 2.6% to 3.2% for rice, 

and 2.2% to 3.6% for maize in year 2000.

Van Dingenen et al. 

(2009)

U.S. SUM06 Caused a loss of 53.8 million to 438 million bushels in soybean production, which account 

for 1.7–14.2% of total U.S. soybean production in 2005

Tong et al. (2007)

East 

Asia

M7; M12 Reduced the yield of wheat, rice and corn by 1–9% and soybean by 23–27% in China, 

Japan and South Korea in 1990

Wang and Mauzerall 

(2004)

a

M7 is defined as 7-h mean O3 concentration (ppb).

b

M12 is defined as 12-h mean O3 concentration (ppb).

9.4.5 Water Cycling

1 Ozone can affect water use in plants and ecosystems through several mechanisms 

2 including damage to stomatal functioning and loss of leaf area. Section 9.3.2 reviewed 

3 possible mechanisms for effects of O3 exposure on stomatal functioning including build4 up of CO2 in substomatal cavity, impacts on signal transduction pathways, and direct O3

5 impact on guard cells. Regardless of the mechanism, O3 exposure has been shown to alter 

6 stomatal performance, which may affect plant and stand transpiration and therefore could 

7 affect hydrological cycling (Figure 9-7). 
Draft – Do Not Cite or Quote 9-69 September 2011

Figure 9-7 The potential effects of ozone exposure on watering cycling.

1 In the literature, there is not a clear consensus on the nature of leaf-level stomatal 

2 conductance response to O3 exposure. At the leaf level, O3 exposure is known to result in 

3 stomatal patchiness (Paoletti and Grulke, 2005; Omasa et al., 1987; Ellenson and 

4 Amundson, 1982), i.e., the heterogeneous aperture of stomata on the leaf surface, and, as 

5 a result, the collective response of groups of stomata on leaves and canopies determines 

6 larger-scale responses to O3. When measured at steady-state high light conditions, leaf7 level stomatal conductance is often found to be reduced when exposed to O3. For 

8 example, a meta-analysis of 55 studies found that O3 reduced stomatal conductance by 

9 11% (Wittig et al., 2007). However, these steady-state measurements were generally 

10 taken at saturating light conditions and steady-state vapor pressure deficit (VPD).

11 Saturating light and steady-state VPD conditions are not common in the field since many 

12 parts of the plant canopy are shaded throughout the day. When studied under varying 

13 environmental conditions, many studies have reported incomplete stomatal closure with 

14 elevated O3 exposure during the day (Mills et al., 2009; Grulke et al., 2007b; Matyssek et 

15 al., 1995; Wieser and Havranek, 1995) or at night (Grulke et al., 2004). This may be due 

16 to sluggish stomatal response. Sluggish stomatal response, defined as a delay in stomatal 

17 response to changing environmental factors relative to controls (Paoletti and Grulke, 

18 2010) has also been documented by several researchers (Grulke et al., 2007c; Matyssek et 

O3 exposure 

Altered canopy 

water loss 

Soil moisture

stem sap flow

Decrease stomatal 

conductance or 

sluggish stomatal 

response

Altered 

run off

Stream flows
Draft – Do Not Cite or Quote 9-70 September 2011

1 al., 1995; Pearson and Mansfield, 1993; Wallin and Skärby, 1992; Lee et al., 1990; 

2 Skarby et al., 1987; Keller and Häsler, 1984; Reich and Lassoie, 1984). Sluggish stomatal 

3 response associated with O3 exposure suggests an uncoupling of the normally tight 

4 relationship between carbon assimilation and stomatal conductance as measured under 

5 steady-state conditions (Gregg et al., 2006; Paoletti and Grulke, 2005). Several tree and 

6 ecosystem models, such as TREGRO, PnET and DLEM, rely on this tight relationship to 

7 simulate water and carbon dynamics. The O3-induced impairment of stomatal control 

8 may be more pronounced for plants growing under water stress (Wilkinson and Davies, 

9 2010; Grulke et al., 2007a; Paoletti and Grulke, 2005; Bonn et al., 2004; Kellomaki and 

10 Wang, 1997; Tjoelker et al., 1995; Reich and Lassoie, 1984). Since leaf-level stomatal 

11 regulation is usually assessed in a steady state rather than as a dynamic response to 

12 changing environmental conditions, steady state measurements cannot detect sluggish 

13 stomatal response. Because of sluggish stomatal responses, water loss from plants may be 

14 greater under dynamic environmental conditions over days and months.

15 In addition to the impacts on stomatal performance, O3-induced physiological changes, 

16 such as reduced leaf area index and accelerated leaf senescence could alter water use 

17 efficiency. It is well established from chamber and field studies that O3 exposure is 

18 correlated with lower foliar retention (Karnosky et al., 2003; Topa et al., 2001; Pell et al., 

19 1999; Grulke and Lee, 1997; Karnosky et al., 1996; Miller et al., 1972; Miller et al., 

20 1963). However, Lee et al. (2009a) did not find changes in needle area of ponderosa pine 

21 and reported that greater canopy conductance followed by water stress under elevated O3

22 may have been caused by stomatal dysfunction. At the Aspen FACE experiment, stand23 level water use, as indicated by sap flux per unit ground area, was not significantly 

24 affected by elevated O3 despite a 22% decrease in leaf area index and 20% decrease in 

25 basal area (Uddling et al., 2008). The lack of negative effect of elevated O3 on stand 

26 water use may be due to the substantially increased whole plant hydraulic conductance 

27 per unit leaf area under elevated O3, as indicated by the sap flux per unit total leaf area 

28 (kl) (Uddling et al., 2009). The increased kl may be caused by the sluggish of stomatal 

29 response. In pure aspen stands, the stomatal closure response to increasing vapor pressure 

30 deficit was less sensitive and mid-day leaf water potential was lower under elevated O3, 

31 suggesting O3 impaired stomatal control over transpiration (Uddling et al., 2009). Other 

32 potential factors contributing to the unchanged stand-level water use included the higher 

33 proportion of sun leaves, and similar or even increased fine root biomass under elevated 

34 O3 (Uddling et al., 2008). Elevated O3 could also affect evapotranspiration by altering 

35 tree crown interception of precipitation. Ozone has been shown to change branch 

36 architectural parameters, and the effects were species-dependent at the Aspen FACE 

37 experiment (Rhea et al., 2010). The authors found that there was a significant correlation 

38 between canopy architecture parameters and stem flow for birch but not aspen.
Draft – Do Not Cite or Quote 9-71 September 2011

1 It is difficult to scale up physiology measurements from leaves to ecosystems. Thus, the 

2 current understanding of how stomatal response at leaf scale is integrated at the scale of 

3 whole forest canopies, and therefore how it influences tree and forest stand water use is 

4 limited. Field studies by McLaughlin et al. (2007a; 2007b) provided valuable insight into 

5 the possible consequences of stomatal sluggishness for ecosystem water cycling. 

6 McLaughlin et al. (2007a; 2007b) indicated that O3 increased water use in a mixed 

7 deciduous forest in eastern Tennessee. McLaughlin et al. (2007a; 2007b) found that O3, 

8 with daily maximum levels ranging from 69.2 to 82.9 ppb, reduced stem growth by 30-

9 50% in the high-O3 year 2002. The decrease in growth rate was caused in part by 

10 amplification of diurnal cycles of water loss and recovery. Peak hourly O3 exposure 

11 increased the rate of water loss through transpiration as indicated by the increased stem 

12 sap flow. The authors suggested that a potential mechanism for the increased sap flow 

13 could be altered stomatal regulation from O3 exposure, but this was inferred through sap 

14 flow measurements and was not directly measured. The increased canopy water loss 

15 resulted in higher water uptake by the trees as reflected in the reduced soil moisture in the 

16 rooting zone. The change in tree water use led to further impacts on the hydrological 

17 cycle at the landscape level. Increased water use under high O3 exposure was reported to 

18 reduce late-season modeled streamflow in three forested watersheds in eastern Tennessee 

19 (McLaughlin et al., 2007b).

20 Felzer et al. (2009) used TEM-Hydro to assess the interactions of O3, climate, elevated 

21 CO2 and N limitation on the hydrological cycle in the eastern U.S. They found that 

22 elevated CO2 decreased evapotranspiration by 2-4% and increased runoff by 3-7%, as 

23 compared to the effects of climate alone. When O3 damage and N limitation were 

24 included, evapotranspiration was reduced by an additional 4-7% and runoff was increased 

25 by an additional 6-11% (Felzer et al., 2009). Based upon simulation with INTRAST and 

26 LINKAGES, Hanson et al. (2005) found that increasing O3 concentration by 20 ppb 

27 above the current ambient level yields a modest 3% reduction in water use. Those 

28 ecological models were generally built on the assumption that O3 induces stomatal 

29 closure and have not incorporated possible stomatal sluggishness due to O3 exposure. 

30 Because of this assumption, results of those models normally found that O3 reduced 

31 water use. 

9.4.5.1 Summary

32 Although the evidence was from a limited number of field and modeling studies, findings 

33 showed an association between O3 exposure and alteration of water use and cycling in 

34 vegetation and at the ecosystem level. There is not a clear consensus on the nature of 

35 leaf-level stomatal conductance response to O3 exposure. When measured under steady-
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1 state high light conditions, leaf-level stomatal conductance is often found to be reduced 

2 when plants are exposed to O3. However, measurements of stomatal conductance under 

3 dynamic light and VPD conditions indicate sluggish responses under elevated O3

4 exposure, which could potentially lead to increased water loss from vegetation. Field 

5 studies conducted by McLaughlin et al. (2007a; 2007b) suggested that peak hourly O3

6 exposure increased the rate of water loss from several tree species, and led to a reduction 

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

8 Sluggish stomatal responses during O3 exposure was suggested as a possible mechanism 

9 for increased water loss during peak O3 exposure. Currently, the O3-induced reduction in 

10 stomatal aperture is the biological assumption for most process-based models. Because of 

11 this assumption, results of those models normally found that O3 reduced water loss. For 

12 example, Felzer (2009) found that O3 damage and N limitation together reduced 

13 evapotranspiration and increased runoff.

14 Although the direction of the response differed among studies, the evidence is 

15 sufficient to conclude that there is likely to be a causal relationship between O3

16 exposure and the alteration of ecosystem water cycling.

9.4.6 Below-Ground Processes

17 Above-ground and below-ground processes are tightly interconnected. Because roots and 

18 soil organisms are not exposed directly to O3, below-ground processes are affected by O3

19 through alterations in the quality and quantity of C supply from photosynthates and 

20 litterfall (Andersen, 2003). Ozone can decrease leaf C uptake by reducing photosynthesis 

21 (Section 9.3). Ozone can also increase metabolic costs by stimulating the production of 

22 chemical compounds for defense and repair processes, and by increasing the synthesis of 

23 antioxidants to neutralize free radicals (see Section 9.3), both of which increase the 

24 consumption of carbon for above-ground processes. Therefore, O3 could significantly 

25 reduce the amount of C available for allocation to below-ground by decreasing C uptake 

26 while increasing C consumption of above-ground processes (Andersen, 2003). 

27 Since the 2006 O3 AQCD, there is additional evidence for O3 effects on below-ground 

28 processes. Ozone has been found to alter root growth, soil food web structure, 

29 decomposer activities, C turnover, water cycling and nutrient flow (Figure 9-8). Ozone 

30 effects on root development and root biomass production and soil food web structure are 

31 reviewed in sections 9.4.3.1 and 9.4.9.2, respectively. The focus in this section is on the 

32 response of litter input, decomposer activities, soil respiration, soil C formation and 

33 nutrient cycling. 
Draft – Do Not Cite or Quote 9-73 September 2011

Source: Modified from Andersen (2003)

Arrows denote C flux pathways that are affected by ozone. Dashed lines indicate where the impact of ozone is suspected but 

unknown. 

Figure 9-8 Conceptual diagram showing where ozone alters C, water and 

nutrient flow in a tree-soil system, including transfer between biotic 

and abiotic components below ground that influence soil physical 

and chemical properties.

9.4.6.1 Litter Carbon Chemistry, Litter Nutrient and Their 

Ecosystem Budgets

1 Consistent with previous findings, recent studies show that, although the responses are 

2 often species-dependent, O3 tends to alter litter chemistry (U.S. EPA, 2006b).Alterations 

CO2, H2O

Litter production 

and chemistry

Allocation of C

CO2 release

Roots/symbionts

CO2, H2O

Altered stomatal function

Nutrients

Soil foodweb

•Bacteria

•Fungi

•Micro & marco invertebrates

Soil physical & 

chemical properties

Organic matter

Root, leaf litter 

exudation

Soil moisture

Run off

Altered species competition

retention
Draft – Do Not Cite or Quote 9-74 September 2011

1 in chemical parameters, such as changes in C chemistry and nutrient concentrations, were 

2 observed in both leaf and root litter (9-7). 

3 At the Aspen FACE site, several studies investigated litter chemistry changes (Parsons et 

4 al., 2008; Johnson and Pregitzer, 2007; Chapman et al., 2005; Liu et al., 2005). In both 

5 aspen and birch leaf litter, elevated O3 increased the concentrations of soluble sugars, 

6 soluble phenolics and condensed tannins (Parsons et al., 2008; Liu et al., 2005). 

7 Compared to other treatments, aspen litter under elevated O3 had the highest fiber 

8 concentration, with the lowest concentration associated with the birch litter under the 

9 same conditions (Parsons et al., 2008). Chapman et al. (2005) measured chemical 

10 changes in fine root litter and found that elevated O3 decreased lignin concentration. O3-

11 induced chemistry changes were also reported from other experimental sites. Results 

12 from an OTC study in Finland suggested that elevated O3 increased the concentration of 

13 acid-soluble lignin, but had no significant impact on other chemicals such as total sugars, 

14 hemicelluloses, cellulose or total lignin in the litter of silver birch (Kasurinen et al., 

15 2006). Results from the free air canopy O3 exposure experiment at Kranzberg Forest 

16 showed that O3 increased starch concentrations but had no impact on cellulose and lignin 

17 in beech and spruce leaf litter (Aneja et al., 2007). The effect of O3 on three antioxidants 

18 (ascorbate, glutathione and α-tocopherol) in fine roots of beech was also assessed at 

19 Kranzberg Forest. The results indicated that O3 had no significant effect on α-tocopherol 

20 and ascorbate concentrations, but decreased glutathione concentrations in fine roots 

21 (Haberer et al., 2008). In addition to changing C chemistry, O3 also altered nutrient 

22 concentrations in green leaves and litter (Table 9-6). 

23 The combined effects of O3 on biomass productivity and chemistry changes may alter 

24 C chemicals and nutrient contents at the canopy or ecosystem level. For example, 

25 although O3 had different impacts on their concentrations, annual fluxes of C chemicals 

26 (soluble sugar, soluble phenolics, condensed tannins, lipid and hemicelluloses), macro 

27 nutrients (N, P, K and S) and micro nutrients (Mg, B, Cu and Zn) to soil were all reduced 

28 due to lower litter biomass productivity at Aspen FACE (Liu et al., 2007a; Liu et al., 

2005). At the Kranzberg Forest, N content of spruce canopy in a mixed culture and Ca2+ 29

30 content of beech canopy in a monoculture increased due to elevated O3 increased leaf 

31 concentrations of those nutrients although leaf production was not significantly altered by 

32 O3 (Rodenkirchen et al., 2009). 
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Table 9-6 The effect of elevated ozone on leaf/litter nutrient concentrations

Study Site Species Ozone 

Concentration Response Reference

Suonenjoki Research 

Station, Finland

Silver birch Ambient: 10-60 ppb

Elevated: 2×ambient

Decreased the concentration of P, Mn, Zn 

and B in leaf litter

Kasurinen et al. 

(2006)

Aspen FACE Aspen and birch Ambient: 50-60 ppb

Elevated: 1.5×ambient

Decreased the concentrations of P, S, Ca 

and Zn, but had no impact on the 

concentrations of N, K, Mg, Mn, B and Cu 

in leaf litter.

Liu et al. (2007a)

Aspen FACE Birch Ambient: 50-60 ppb

Elevated: 1.5×ambient

Increase N concentration in birch litter Parsons et al. (2008)

Kranzberg Forest, Germany Beech and spruce Ambient: 9-41 ppb

Elevated: 2×ambient

Increased N concentration in beach leaf, 

but not in spruce needle

Kozovits et al. (2005)

Kranzberg Forest, Germany Beech and spruce Ambient: 9-41 ppb

Elevated: 2×ambient

1) Had no significant effects on spruce 

needle chemistry; 2) increased Ca 

concentration in beech leaves in 

monoculture, but had no impacts on other 

nutrients

Rodenkirchen et al. 

(2009)

Salerno, Italy Holm oak Non-filtered OTC: 29 ppb

Filtered OTC: 17ppb

Ozone had no significant impacts on litter 

C, N, lignin and cellulose concentrations

Baldantoni et al. 

(2011)

Kuopio University Research 

Garden, Finland

Red Clover Ambient: 25.7 ppb

Elevated: 1.5×ambient

increased the total phenolic content of 

leaves and had minor effects on the 

concentrations of individual phenolic 

compounds

Saviranta et 

al.(2010)

9.4.6.2 Decomposer Metabolism and Litter Decomposition

1 The above- and below-ground physiological changes caused by O3 exposure cascade 

2 through the ecosystem and affect soil food webs. In the 2006 O3 AQCD, there were very 

3 few studies on the effect of O3 on the structure and function of soil food webs, except 

4 two studies conducted by Larson et al. (2002) and Phillips et al. (2002). Since the last O3

5 AQCD, new studies have provided more information on how O3 affects the metabolism 

6 of soil microbes and soil fauna. 

7 Chung et al.(2006) found that the activity of the cellulose-degrading enzyme 1,4-β8 glucosidase was reduced by 25% under elevated O3 at Aspen FACE. The decrease in 

9 cellulose-degrading enzymatic activity was associated with the lower cellulose 

10 availability under elevated O3 (Chung et al., 2006). However, a later study at the same 

11 site, which was conducted in the 10th year of the experiment, found that O3 had no 

12 impact on cellulolytic activity in soil (Edwards and Zak, 2011). In a lysimeter study of 

13 beech trees (Fagus sylvatica) in Germany, soil enzyme activity was found to be 

14 suppressed by O3 exposure (Esperschutz et al., 2009; Pritsch et al., 2009). Except for 

15 xylosidase, enzyme activities involved in plant cell wall degradation (cellobiohydrolase, 

16 beta-glucosidase and glucuronidase) were decreased in rhizosphere soil samples under 

17 elevated O3 (2 × ambient level) (Pritsch et al., 2009). Similarly, Chen et al. (2009) found 
Draft – Do Not Cite or Quote 9-76 September 2011

1 O3 exposure, with a 3-month AOT40 of 21.4-44.1 ppm-h, decreased the microbial 

2 metabolic capability in the rhizosphere and bulk soil of wheat, although the observed 

3 reduction in bulk soil was not significant.

4 Ozone-induced change in soil organisms’ activities could affect litter decomposition 

5 rates. Results of recent studies indicated that O3 slightly reduced or have no impacts on 

6 litter decomposition (Liu et al., 2009b; Parsons et al., 2008; Kasurinen et al., 2006)

7 (Baldantoni et al., 2011). The responses varied among species, sites and exposure length. 

8 Parsons et al. (2008) collected litter from aspen and birch seedlings at Aspen FACE site, 

9 and conducted a 23-month field litter incubation starting in 1999. They found that 

10 elevated O3 had different impacts on the decomposition of aspen and birch litter. 

11 Elevated O3 was found to reduce aspen litter decomposition. However, O3 accelerated 

12 birch litter decomposition under ambient CO2, but reduced it under elevated CO2

13 (Parsons et al., 2008). Liu et al. (2009b) conducted another litter decomposition study at 

14 Aspen FACE from 2003 to 2006, when stand leaf area index (LAI) reached its maximum. 

15 During the 935-day field incubation, elevated O3 was shown to reduce litter mass loss in 

16 the first year, but not in the second year. They suggested that higher initial tannin and 

17 phenolic concentrations under elevated O3 reduced microbial activity in the first year 

18 (Liu et al., 2009b). In an OTC experiment, Kasurinen et al. (2006) collected silver birch 

19 leaf litter from three consecutive growing seasons and conducted three separate litter-bag 

20 incubation experiments. Litter decomposition was not affected by O3 exposure in the first 

21 two incubations, but a slower decomposition rate was found in the third incubation. Their 

22 principle component analysis indicated that the litter chemistry changes caused by O3

23 (decreased Mn, P, B and increased C:N) might be partially responsible for the decreased 

24 mass loss of their third incubation. In another OTC experiment, Baldatoni et al. (2011)

25 found that O3 significantly reduced leaf litter decomposition of Quercus ilex L, although 

26 litter C, N, lignin and cellulose concentrations were not altered by O3 exposure.

9.4.6.3 Soil respiration and carbon formation 

27 Ozone could reduce the availability of photosynthates for export to roots, and increase 

28 root mortality and turnover rates. Ozone has also been shown to reduce above-ground 

29 litter productivity and alter litter chemistry, which would affect the quality and quantity 

30 of the C supply to soil organisms (Section 9.4.6.1). The complex interactions among 

31 those changes make it difficult to predict the response of soil C cycling under elevated 

32 O3. The 2006 O3 AQCD concluded that O3 had no consistent impact on soil respiration 

33 (U.S. EPA, 2006b). Ozone could increase or decrease soil respiration, depending on the 

34 approach and timing of the measurements. Ozone may also alter soil C formation. 

35 However, very few experiments directly measured changes in soil organic matter content 
Draft – Do Not Cite or Quote 9-77 September 2011

1 under O3 fumigation (U.S. EPA, 2006b). Recent studies on soil respiration and soil 

2 C content also found mixed responses. Most importantly, recent results from long-term 

3 fumigation experiments, such as the Aspen FACE experiment, suggest that ecosystem 

4 response to O3 exposure can change over time. Observations made during the late 

5 exposure years can be inconsistent with those during the early years, highlighting the 

6 need for caution when assessing O3 effects based on short-term studies (Table 9-7).

Table 9-7 The temporal variation of ecosystem responses to ozone exposure 

at Aspen FACE site

Endpoint Period of 

Measurement Response Reference

Litter decomposition 1999-2001 O3 reduced aspen litter decomposition. However, O3 accelerated 

birch litter decomposition under ambient CO2, but reduced it under 

elevated CO2

Parsons et al. (2008)

2003-2006 O3 reduced litter mass loss in the first year, but not in the second 

year.

Liu et al. (2009b)

Fine root production 1999 O3 had no significant impact on fine root biomass King et al. (2001)

2002, 2005 O3 increased fine root biomass Pregitzer et al. (2008)

Soil respiration 1998-1999 Soil respiration under +CO2+O3 treatment was lower than that 

under +CO2 treatment

King et al. (2001)

2003-2007 Soil respiration under +CO2+O3 treatment was 5-25% higher than 

under elevated CO2 treatment.

Pregitzer et al. (2006) (2008)

Soil C formation 1998-2001 O3 reduced the formation rates of total soil C by 51% and acidinsoluble soil C by 48%Loya et al. (2003)

2004-2008 No significant effect of O3 on the new C formed under elevated 

CO2

Talhelm et al. (2009)

Soil Respiration

7 Ozone has shown inconsistent impacts on soil respiration. A sun-lit controlled8 environment chamber study found that O3 had no significant effects on soil respiration, 

9 fine root biomass or any of the soil organisms in a reconstructed ponderosa pine/soil-litter 

10 system (Tingey et al., 2006). In an adult European beech/Norway spruce forest at 

11 Kranzberg Forest, the free air O3 fumigation (AOT40 of 10.2-117 ppm-h) increased soil 

12 respiration under both beech and spruce during a humid year (Nikolova et al., 2010) . The 

13 increased soil respiration under beech has been accompanied by the increase in fine root 

14 biomass and ectomycorrhizal fungi diversity and turnover (Grebenc and Kraigher, 2007). 

15 The stimulating effect on soil respiration disappeared under spruce in a dry year, which 

16 was associated with a decrease in fine root production in spruce under drought. This 

17 finding suggested that drought was a more dominant stress than O3 for spruce (Nikolova 

18 et al., 2010). Andersen et al. (2010) labeled the canopies of European beech and Norway 
Draft – Do Not Cite or Quote 9-78 September 2011

spruce with CO2 depleted in 13 1 C at the same site. They did not observe any significant 

2 changes in soil respiration for either species. 

3 The nearly 10 year long studies at Aspen FACE indicated that the response of soil 

4 respiration to O3 interacted with CO2 exposure and varied temporally (Table 9-7) 

5 (Pregitzer et al., 2008; Pregitzer et al., 2006; King et al., 2001). Ozone treatment alone 

6 generally had the lowest mean soil respiration rates, although those differences between 

7 control and elevated O3 were usually not significant. However, soil respiration rates were 

8 different with O3 alone and when acting in combination with elevated CO2. In the first 

9 five years (1998-2002), soil respiration under +CO2+O3 treatment was similar to that 

10 under control and lower than that under +CO2 treatment (Pregitzer et al., 2006; King et 

11 al., 2001). Since 2003, +CO2+O3 treatment started to show the greatest impact on soil 

12 respiration. Compared to elevated CO2, soil respiration rate under +CO2+O3 treatment 

13 was 15-25% higher from 2003-2004, and 5-10% higher from 2005-2007 (Pregitzer et al., 

14 2008; Pregitzer et al., 2006). Soil respiration was highly correlated with the biomass of 

15 roots with diameters of <2 mm and <1 mm, across plant community and atmospheric 

16 treatments. The authors suggested that the increase in soil respiration rate may be due to 

17 +CO2+O3 increased fine root (<1.0 mm) biomass production (Pregitzer et al., 2008).

18 Changes in leaf chemistry and productivity due to O3 exposure have been shown to affect 

19 herbivore growth and abundance (See Section 9.4.9.1). Canopy insects could affect soil 

20 carbon and nutrient cycling through frass deposition, or altering chemistry and quantity 

21 of litter input to the forest floor. A study at the Aspen FACE found that although elevated 

22 O3 affected the chemistry of frass and greenfall, these changes had small impact on 

23 microbial respiration and no effect on nitrogen leaching (Hillstrom et al., 2010a).

24 However, respiratory carbon loss and nitrate immobilization were nearly double in 

25 microcosms receiving herbivore inputs than those receiving no herbivore inputs 

26 (Hillstrom et al., 2010a).

Soil Carbon Formation

27 Ozone-induced reductions in plant growth can result in reduced C input to soil and 

28 therefore soil C content (Andersen, 2003). The simulations of most ecosystem models 

29 support this prediction (Ren et al., 2007a; Zhang et al., 2007a; Felzer et al., 2004). 

30 However, very few studies have directly measured soil C dynamics under elevated O3. 

31 After the first four years of fumigation (from 1998 to 2001) at the Aspen FACE site, 

32 Loya et al. (2003) found that forest stands exposed to both elevated O3 and CO2

33 accumulated 51% less total soil C, and 48% less acid-insoluble soil C compared to stands 

34 exposed only to elevated CO2. Soil organic carbon (SOC) was continuously monitored at 

35 the Aspen FACE site, and the later data showed that the initial reduction in new 
Draft – Do Not Cite or Quote 9-79 September 2011

1 C formation (soil C derived from plant litter since the start of the experiment) by O3

2 under elevated CO2 is only a temporary effect (Table 9-7) (Talhelm et al., 2009). The 

3 amount of new soil C in the elevated CO2 and the combined elevated CO2 and O3

4 treatments has converged since 2002. There was no significant effect of O3 on the new C 

5 formed under elevated CO2 over the last four years of the study (2004-2008). Talhelm

6 et al. (2009) suggested the observed reduction in the early years of the experiment might 

7 be driven by a suppression of C allocated to fine root biomass. During the early exposure 

8 years, O3 had no significant impact on fine root production (King et al., 2001). However, 

9 the effect of O3 on fine root biomass was observed later in the experiment. Ozone 

10 increased fine root production and the highest fine root biomass was observed under the 

11 combined elevated CO2 and O3 treatment in the late exposure years (Table 9-7) 

12 (Pregitzer et al., 2006). This increase in fine root production was due to changes in 

13 community composition, such as better survival of O3-tolerant aspen genotype, birch and 

14 maple, rather than changes in C allocation at the individual tree level (Pregitzer et al., 

15 2008; Zak et al., 2007). 

9.4.6.4 Nutrient cycling

16 Ozone can affect nutrient cycling by changing nutrient release from litter, nutrient uptake 

17 by plants, and soil microbial activity. Nitrogen is the limiting nutrient for most temperate 

18 ecosystems, and several studies examined N dynamics under elevated O3. Nutrient 

19 mineralization from decomposing organic matter is important for sustaining ecosystem 

20 production. Holmes et al. (2006) found that elevated O3 decreased gross N mineralization 

21 at the Aspen FACE site, indicating that O3 may reduce N availability. Other N cycling 

processes, such as NH4

+ 22 immobilization, gross nitrification, microbial biomass N and soil 

23 organic N, were not affected by elevated O3 (Holmes et al., 2006). Similarly, Kanerva 

24 et al. (2006) found total N, NO3-, microbial biomass N, potential nitrification and 

25 denitrification in their meadow mesocosms were not affected by elevated O3 (40-50 ppb). 

26 Ozone was found to decreased soil mineral N content at SoyFACE, which was likely 

27 caused by a reduction in plant material input and increased denitrification (Pujol Pereira 

28 et al., 2011). Ozone also showed small impact on other micro and macro nutrients. Liu 

29 et al. (2007a) assessed N, P, K, S, Ca, Mg, Mn, B, Zn and Cu release dynamics at Aspen 

30 FACE, and they found that O3 had no effects on most nutrients, except to decrease N and 

31 Ca release from litter. These studies reviewed above suggested that soil N cycling 

32 processes were not affected or slightly reduced by O3 exposure. However, in a lysimeter 

33 study with young beech trees Stoelken et al. (2010) found that elevated O3 stimulated N 

34 release from litter which was largely attributed to an enhanced mobilization of inert 

35 nitrogen fraction.
Draft – Do Not Cite or Quote 9-80 September 2011

1 Using the SImple NItrogen Cycle model (SINIC), Hong et al. (2006) evaluated the 

2 impacts of O3 exposure on soil N dynamics and streamflow nitrate flux. The detrimental 

3 effect of O3 on plant growth was found to reduce plant uptake of N and therefore increase 

4 nitrate leaching. Their model simulation indicated that ambient O3 exposure increased the 

mean annual stream flow nitrate export by 12% (0.042 g N/m2 5 /year) at the Hubbard 

6 Brook Experimental Watershed from 1964-1994 (Hong et al., 2006).

9.4.6.5 Dissolved Organic Carbon and Biogenic Trace Gases 

Emission

7 The O3-induced changes in plant growth, C and N fluxes to soil and microbial 

8 metabolism can alter other biogeochemical cycling processes, such as soil dissolved 

9 organic carbon (DOC) turnover and trace gases emission. 

10 Jones et al. (2009) collected fen cores from two peatlands in North Wales, UK and 

11 exposed them to one of four levels of O3 (AOT40 of 0, 3.69, 5.87 and 13.80 ppm-h for 

12 41 days). They found the concentration of porewater DOC in fen cores was significantly 

13 decreased by increased O3 exposure. A reduction of the low molecular weight fraction of 

14 DOC was concurrent with the observed decrease in DOC concentration. Their results 

15 suggested that O3 damage to overlying vegetation may decrease utilizable C flux to soil. 

16 Microbes, therefore, have to use labile C in the soil to maintain their metabolism, which, 

17 the authors hypothesized, leads to a decreased DOC concentration with a shift of the 

18 DOC composition to more aromatic, higher molecular weight organic compounds. 

19 Several studies since the 2006 O3 AQCD have examined the impacts of O3 on nitrous 

20 oxide (N2O) and methane (CH4) emission. Kanerva et al. (2007) measured the fluxes of 

21 N2O and CH4 in meadow mesocosms, which were exposed to elevated CO2 and O3 in 

22 OTCs in south-western Finland. They found that the daily N2O fluxes were decreased in 

23 the NF+O3 (non-filtered air + elevated O3, 40-50 ppb) after three seasons of exposure. 

24 Elevated O3 alone or combined with CO2 did not have any significant effect on the daily 

25 fluxes of CH4 (Kanerva et al., 2007). In another study conducted in central Finland, the 

26 4 year open air O3 fumigation (AOT40 of 20.8-35.5 ppm-h for growing season) slightly 

27 increased potential CH4 oxidation by 15% in the peatland microcosms, but did not affect 

28 the rate of potential CH4 production or net CH4 emissions, which is the net result of the 

29 potential CH4 production and oxidation (Morsky et al., 2008). However, several studies 

30 found that O3 could significantly reduce CH4 emission. Toet et al. (2011) exposed 

31 peatland mesocosms to O3 in OTCs for two years, and found that CH4 emissions were 

32 significantly reduced by about 25% during midsummer periods of both years. In an OTC 

33 study of rice paddy, Zheng et al. (2011) found that the daily mean CH4 emissions were 
Draft – Do Not Cite or Quote 9-81 September 2011

1 significantly lower under elevated O3 treatments than those in charcoal-filtered air and 

2 nonfiltered air treatments. They found that the seasonal mean CH4 emissions were 

3 negatively related with AOT40, but positively related to the relative rice yield, 

4 aboveground biomass and underground biomass.

9.4.6.6 Summary 

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

6 often site specific, O3 altered the quality and quantity of litter input to soil, microbial 

7 community composition, and C and nutrient cycling. Biogeochemical cycling of below8 ground processes is driven by C input from plants. Studies at the leaf and plant level have 

9 provided biologically plausible mechanisms, such as reduced photosynthetic rates, 

10 increased metabolic cost, and reduced root C allocation for the association of O3

11 exposure and the alteration of below-ground processes. 

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

13 reduced litter production and altered C chemistry, such as soluble sugars, soluble 

14 phenolics, condensed tannins, lignin, and macro/micro nutrient concentration in litter 

15 (Parsons et al., 2008; Kasurinen et al., 2006; Liu et al., 2005). The changes in substrate 

16 quality and quantity could alter microbial metabolism under elevated O3, and therefore 

17 soil C and nutrient cycling. Several studies indicated that O3 suppressed soil enzyme 

18 activities (Pritsch et al., 2009; Chung et al., 2006). However, the impact of O3 on litter 

19 decomposition was inconsistent and varied among species, sites and exposure length. 

20 Similarly, O3 had inconsistent impacts on dynamics of micro and macro nutrients. 

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

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

23 changed over time. For example, in the early part of the experiment (1998-2003), O3 had 

24 no impact on soil respiration but reduced the formation rates of total soil C under 

25 elevated CO2. However, after 10-11 yr of exposure, O3 was found to increase soil 

26 respiration but have no significant impact on soil C formation under elevated CO2. 

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

28 exposure and the alteration of below-ground biogeochemical cycles.

9.4.7 Community composition

29 The effects of O3 on species competition (AX9.3.3.4) and community composition 

30 (AX9.6.4) were summarized in the 2006 O3 AQCD. Plant species differ in their 
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1 sensitivity to O3. Fast growing plants with high stomatal conductance and high specific

2 leaf area (SLA) were more likely to be sensitive to O3 exposure. Further, different 

3 genotypes of a given species also vary in their sensitivity. This differential sensitivity 

4 could change the competitive interactions that lead to loss in O3 sensitive species or 

5 genotypes. In addition, O3 exposure has been found to alter reproductive processes in 

6 plants (See Section 9.4.3.3). Changes in reproductive success could lead to changes in 

7 species composition. However, since ecosystem-level responses result from the 

8 interaction of organisms with one another and with their physical environment, it takes 

9 longer for a change to develop to a level of prominence at which it can be identified and 

10 measured. A shift in community composition in forest and grassland ecosystems noted in 

11 the 2006 O3 AQCD has continued to be observed from experimental and gradient studies. 

12 Additionally, research since the last review has shown that O3 can alter community 

13 composition and diversity of soil microbial communities.

9.4.7.1 Forest

14 In the San Bernardino Mountains in southern California, O3 pollution caused a 

15 significant decline in ponderosa pine (Pinus ponderosa ) and Jeffrey pine (Pinus jeffreyi) 

16 (U.S. EPA, 2006b). Pine trees in the young mature age class group exhibited higher 

17 mortality rates compared with mature trees at a site with severe O3 visible foliar injury. 

18 The vulnerability of young mature pines was most likely caused by the fact that trees in 

19 this age class were emerging into the canopy, where higher O3 concentrations were 

20 encountered (McBride and Laven, 1999). Because of the loss of O3-sensitive pines, 

21 mixed forests of ponderosa pine, Jeffery Pine and white fir (Abies concolor) shifted to 

22 predominantly white fir (Miller, 1973). Ozone may have indirectly caused the decline in 

23 understory diversity in coniferous forests in the San Bernardino Mountains through an 

24 increase in pine litterfall. This increase in litterfall from O3 exposure results in an 

25 understory layer that may prohibit the establishment of native herbs, but not exotic annual 

26 Galium aparine (Allen et al., 2007). 

27 Ozone damage to conifer forests has also been observed in several other regions. In the 

28 Valley of Mexico, a widespread mortality of sacred fir (Abies religiosa) was observed in 

29 the heavily polluted area of the Desierto de los Leones National Park in the early 1980s 

30 (de Lourdes de Bauer and Hernandez-Tejeda, 2007; Fenn et al., 2002). Ozone damage 

31 was widely believed to be an important causal factor in the dramatic decline of sacred fir. 

32 In alpine regions of southern France and the Carpathians Mountains, O3 was also 

33 considered as the major cause of the observed decline in cembran pine (Pinus 

34 cembra)(Wieser et al., 2006). However, many environmental factors such as light, 

35 temperature, nutrient and soil moisture, and climate extremes such as unusual dry and 
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1 wet periods could interact with O3 and alter the response of forest to O3 exposure. For 

2 those pollution gradient studies, several confounding factors, such as drought, insect 

3 outbreak and forest management, may also contribute to or even be the dominant factors 

4 causing the mortality of trees (de Lourdes de Bauer and Hernandez-Tejeda, 2007; Wieser 

5 et al., 2006).

6 New evidence from long-term free O3 fumigation experiments provided additional 

7 support for the potential impacts of O3 on species competition and community 

8 composition changes in forest ecosystems. At the Aspen FACE site, community 

9 composition at both the genetic and species levels was altered after seven years of 

10 fumigation with O3 (Kubiske et al., 2007). In the pure aspen community, O3 fumigation 

11 reduced growth and increased mortality of sensitive clone 259, while the O3 tolerant 

12 clone 8L emerged as the dominant clone. Growth of clone 8L was even greater under 

13 elevated O3 compared to controls, probably due to O3 alleviated competitive pressure on 

14 clone 8L by reducing growth of other clones. In the mixed aspen-birch and aspen-maple 

15 communities, O3 reduced the competitive capacity of aspen compared to birch and maple 

16 (Kubiske et al., 2007). In a phytotron study, O3 fumigation reduced growth of beech but 

17 not spruce in mixed culture, suggesting a higher susceptibility of beech to O3 under 

18 interspecific competition (Kozovits et al., 2005).

9.4.7.2 Grassland and Agricultural Land

19 The response of managed pasture, often cultivated as a mixture of grasses and clover, to 

20 O3 pollution has been studied for many years. The tendency for O3-exposure to shift the 

21 biomass of grass-legume mixtures in favor of grass species, reported in the previous O3

22 AQCD has been generally confirmed by recent studies. In a mesocosm study, Trifolium 

23 repens and Lolium perenne mixtures were exposed to an episodic rural O3 regime within 

24 solardomes for 12 weeks. T. repens showed significant changes in biomass but not L. 

25 perenne, and the proportion of T. repens decreased in O3-exposed mixtures compared to 

26 the control (Hayes et al., 2009). The changes in community composition of grass-legume27 forb mixtures were also observed at the Le Mouret FACE experiment, Switzerland.

28 During the 5-year O3 fumigation (AOT40 of 13.3-59.5 ppm-h), the dominance of 

29 legumes in fumigated plots declined more quickly than those in the control plots (Volk et 

30 al., 2006). However, Stampfli and Fuhrer (2010) re-analyzed the species and soil data and 

31 suggested that Volk et al. (2006) overestimated the O3 effect. Stampfli and Fuhrer (2010)

32 found that the difference in the species dynamics between control and O3 treatment was 

33 more caused by heterogeneous initial conditions than O3 exposure. Several studies also 

34 suggested the mature/species-rich ecosystems were more resilient to O3 exposure. At 

35 another FACE experiment, located at Alp Flix, Switzerland, O3 fumigation (AOT40 of 
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1 15.2-64.9 ppm-h) showed no significant impact on community composition of this 

2 species-rich pasture (Bassin et al., 2007b). Although most studies demonstrated an 

3 increase in grass:forb ration with O3 exposure (Hayes et al., 2009; U.S. EPA, 2006b), a 

4 study on a simulated upland grassland community O3 reduce grass:forb ratio (Felicity et 

5 al., 2010), which may be due to grass species in this community, such as Anthoxanthum 

6 odoratum, was more sensitive to O3 than other most studied grass species such as L. 

7 perenne (Hayes et al., 2009). Pfleeger et al. (2010) collected seed bank soil from an 

8 agricultural field and examined how the plant community responded over several 

9 generations to elevated O3 exposures. Sixty plant species from 22 families emerged in the 

10 chambers over their four year study. Overall, they found that O3 appeared to have small 

11 impacts on seed germination and only a minor effect on species richness of pioneer plant 

12 communities.

13 Several review papers have discussed the physiological and ecological characteristics of 

14 O3-sensitive herbaceous plants. Hayes et al. (2007) assessed species traits associated with 

15 O3 sensitivity by the changes in biomass caused by O3 exposure. Plants of the therophyte 

16 (e.g., annual) life form were particularly sensitive to O3. Species with higher mature leaf 

17 N concentration tended to be more sensitive than those with lower leaf N concentration. 

18 Plants growing under high oxidative stress environments, such as high light or high 

19 saline, were more sensitive to O3. Using the same dataset from Hayes et al. (2007), Mills 

20 et al. (2007a) identified the O3 sensitive communities. They found that the largest number 

21 of these O3 sensitive communities were associated with grassland ecosystems. Among 

22 grassland ecosystems, alpine grassland, sub-alpine grassland, woodland fringe, and dry 

23 grassland were identified as the most sensitive communities.

9.4.7.3 Microbes

24 Several methods have been used to study microbial composition changes associated with 

25 elevated O3. Phospholipid fatty acid (PLFA) analysis is widely used to determine 

26 whether O3 elicits an overall effect on microbial community composition. However, 

27 since PLFA markers cover a broad range of different fungi, resolution of this method 

28 may be not fine enough to detect small changes in the composition of fungal 

29 communities. Methods, such as microscopic analyses and polymerase chain reaction–

30 denaturing gradient gel electrophoresis (PCR–DGGE), have better resolution to 

31 specifically analyze the fungal community composition. The resolution differences 

32 among those methods needs to be considered when assessing the O3 impact on microbial 

33 community composition.
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1 Kanerva et al. (2008) found that elevated O3 (40-50 ppb) decreased total, bacterial, 

2 actinobacterial and fungal PLFA biomass values as well as fungal:bacterial PLFA 

3 biomass ratio in their meadow mesocosms in south-western Finland. The relative 

4 proportions of individual PLFAs between the control and elevated O3 treatments were 

5 significantly different, suggesting that O3 modified the structure of the microbial 

6 community. Morsky et al. (2008) exposed boreal peatland microcosms to elevated O3, 

7 with growing season AOT40 of 20.8-35.3 ppm-h, in an open-air O3 exposure field in 

8 Central Finland. They also found that microbial composition was altered after three 

9 growing seasons with O3 fumigation, as measured by PLFA. Ozone tended to increase 

10 the presence of Gram-positive bacteria and the biomass of fungi in the peatland 

11 microcosms. Ozone also resulted in higher microbial biomass, which co-occurred with 

12 the increases in concentrations of organic acids and leaf density of sedges (Morsky et al., 

13 2008). In a lysimeter experiment in Germany, O3 was found to alter the PLFA profiles in 

14 the upper 0-20 cm rhizosphere soil of European beech. Elevated O3 reduced bacterial 

15 abundance but had no detectable effect on fungal abundance (Pritsch et al., 2009). Using 

16 microscopic analyses, Kasurinen et al. (2005) found that elevated O3, with 5 or 6 months 

17 of AOT40 of 20.6-30.9 ppm-h, decreased the proportions of black and liver-brown 

18 mycorrhizas and increased that of light brown/orange mycorrhizas. In an herbaceous 

19 plant study, SSCP (single-strand conformation polymorphism) profiles indicated that O3

20 stress (about 75 ppb) had a very small effect on the structural diversity of the bacterial 

21 community in rhizospheres (Dohrmann and Tebbe, 2005). At the Aspen FACE site, O3

22 had no significant effect on fungal relative abundance, as indicated by PLFA profile. 

23 However, elevated O3 altered fungal community composition, according to the 

24 identification of 39 fungal taxonomic units from soil using polymerase chain reaction–

25 denaturing gradient gel electrophoresis (PCR-DGGE) (Chung et al., 2006). In another 

26 study at Aspen FACE, phylogenetic analysis suggested that O3 exposure altered 

27 agaricomycete community. The ectomycorrhizal communities developing under elevated 

28 O3 had higher proportions of Cortinarius and Inocybe species, and lower proportions of 

29 Laccaria and Tomentella (Edwards and Zak, 2011). Ozone was found to change 

30 microbial community composition in an agricultural system. Chen et al. (2010b) found 

31 elevated O3 (100-150 ppb) had significant effects on soil microbial composition 

32 expressed as PLFA percentage in a rice paddy in China. 

9.4.7.4 Summary 

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

34 composition was assessed. Ozone was found to cause a significant decline in ponderosa 

35 and Jeffrey pine in the San Bernardino Mountains in southern California. Ozone exposure 
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1 also tended to shift the grass-legume mixtures in favor of grass species (U.S. EPA, 

2 2006b). Since the 2006 O3 AQCD, more evidence has shown that O3 exposure changed 

3 the competitive interactions and could lead to loss of O3 sensitive species or genotypes. 

4 Studies at plant level found that the severity of O3 damage on growth, reproduction and 

5 foliar injury varied among species, which provided the biological plausibility for the 

6 alteration of community composition. Additionally, research since the last review has 

7 shown that O3 can alter community composition and diversity of soil microbial 

8 communities.

9 The decline of conifer forests under O3 exposure was continually observed in several 

10 regions. Ozone damage was believed to be an important causal factor in the dramatic 

11 decline of sacred fir in the valley of Mexico (de Lourdes de Bauer and Hernandez12 Tejeda, 2007), as well as cembran pine in southern France and Carpathian Mountains 

13 (Wieser et al., 2006). Results from the Aspen FACE site indicated that O3 could alter 

14 community composition of broadleaf forests as well. At the Aspen FACE site, O3

15 reduced growth and increased mortality of a sensitive aspen clone, while the O3 tolerant 

16 clone emerged as the dominant clone in the pure aspen community. In the mixed aspen17 birch and aspen-maple communities, O3 reduced the competitive capacity of aspen 

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

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

20 grass species, was reported in the 2006 O3 AQCD and has been generally confirmed by 

21 recent studies. However, in a high elevation mature/species-rich grass-legume pasture, O3

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

23 2007b).

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

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

26 observed in both natural and agricultural ecosystems, although no general patterns could 

27 be identified (Kanerva et al., 2008; Morsky et al., 2008; Kasurinen et al., 2005). 

28 The evidence is sufficient to conclude that there is likely a causal relationship 

29 between O3 exposure and the alteration of community composition.

9.4.8 Factors that Modify Functional and Growth Response

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

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

32 CO2, influence or alter plant response to O3. These modifying factors were 

33 comprehensively reviewed in AX9.3 of the 2006 O3 AQCD and thus, this section serves 
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1 mainly as a brief summary of the previous findings. A limited number of new studies 

2 published since the 2006 O3 AQCD add to our understanding of the role of these 

3 interactions in modifying O3-induced plant responses. Many of these modifying factors 

4 and interactions are integrated into discussions elsewhere in this chapter and the reader is 

5 directed to those sections.

9.4.8.1 Genetics

6 It is well known that species vary greatly in their responsiveness to O3. Even within a 

7 given species, individual genotypes or populations can also vary significantly with 

8 respect to O3 sensitivity (U.S. EPA, 2006b). Therefore, caution should be taken when 

9 considering a species’ degree of sensitivity to O3. Plant response to O3 is determined by 

10 genes that are directly related to oxidant stress and to an unknown number of genes that 

11 are not specifically related to oxidants, but instead control leaf and cell wall thickness, 

12 stomatal conductance, and the internal architecture of the air spaces. It is rarely the case 

13 that single genes are responsible for O3 tolerance. Studies using molecular biological 

14 tools and transgenic plants have positively verified the role of various genes and gene 

15 products in O3 tolerance and are continuing to increase the understanding of O3 toxicity 

16 and differences in O3 sensitivity. See Section 9.3.3.2 of this document for a discussion of 

17 recent studies related to gene expression changes in response to O3.

9.4.8.2 Environmental Biological Factors

18 As stated in the 2006 O3 AQCD, the biological factors within the plant’s environment 

19 that may influence its response to O3 encompass insects and other animal pests, diseases, 

20 weeds, and other competing plant species. Ozone may influence the severity of a disease 

21 or infestation by a pest or weed, either by direct effects on the causal species, or 

22 indirectly by affecting the host, or both. In addition, the interaction between O3, a plant, 

23 and a pest, pathogen, or weed may influence the response of the target host species to O3

24 (U.S. EPA, 2006b). Several recent studies on the effects of O3 on insects via their 

25 interactions with plants are discussed in Section 9.4.9.1. In addition, O3 has also been 

26 shown to alter soil fauna communities (Section 9.4.9.2).

27 In contrast to detrimental biological interactions, there are mutually beneficial 

28 relationships or symbioses involving higher plants and bacteria or fungi. These include 

29 (1) the nitrogen-fixing species Rhizobium and Frankia that nodulate the roots of legumes 

30 and alder and (2) the mycorrhizae that infect the roots of many crop and tree species, all 
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1 of which may be affected by exposure of the host plants to O3. Some discussion of 

2 mycorrhizae can be found in Section 9.4.6.

3 In addition to the interactions involving animal pests, O3 also has indirect effects on 

4 higher herbivorous animals, e.g., livestock, due to O3-induced changes in feed quality. 

5 Recent studies on the effects of O3 on nutritive quality of plants are discussed in Sections 

6 9.4.4.2. 

7 Intra- and interspecific competition are also important factors in determining vegetation 

8 response to O3. Plant competition involves the ability of individual plants to acquire the 

9 environmental resources needed for growth and development: light, water, nutrients, and 

10 space. Intraspecific competition involves individuals of the same species, typically in 

11 monoculture crop situations, while interspecific competition refers to the interference 

12 exerted by individuals of different species on each other when they are in a mixed 

13 culture. This topic was previously reviewed in AX9.3.3.4 of the 2006 O3 AQCD. Recent 

14 studies on competition and its implications for community composition are discussed in 

15 Section 9.4.7.

9.4.8.3 Physical Factors

16 Physical or abiotic factors play a large role in modifying plant response to O3, and have 

17 been extensively discussed in previous O3 AQCDs. This section summarizes those 

18 findings as well as recent studies published since the last review.

19 Although some studies have indicated that O3 impact significantly increases with 

20 increased ambient temperature (Ball et al., 2000; Mills et al., 2000), other studies have 

21 indicated that temperature has little effect (Balls et al., 1996; Fredericksen et al., 1996). A 

22 recent study by Riikonen et al. (2009) at the Ruohoniemi open air exposure field in 

23 Kuopio, Finland found that the effects of temperature and O3 on total leaf area and 

24 photosynthesis of Betula pendula were counteractive. Elevated O3 reduced the saplings’

25 ability to utilize the warmer growth environment by increasing the stomatal limitation for 

26 photosynthesis and by reducing the redox state of ascorbate in the apoplast in the 

27 combination treatment as compared to temperature alone (Riikonen et al., 2009). 

28 Temperature affects the rates of all physiological processes based on enzyme catalysis 

29 and diffusion; each process and overall growth (the integral of all processes) has a 

30 distinct optimal temperature range. It is important to note that a plant’s response to 

31 changes in temperature will depend on whether it is growing near its optimum 

32 temperature for growth or near its maximum temperature (Rowland-Bamford, 2000). 

33 However, temperature is very likely an important variable affecting plant O3 response in 
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1 the presence of the elevated CO2 levels contributing to global climate change. In contrast, 

2 some evidence suggests that O3 exposure sensitizes plants to low temperature stress 

3 (Colls and Unsworth, 1992) and, also, that O3 decreases below-ground carbohydrate 

4 reserves, which may lead to responses in perennial species ranging from rapid demise to 

5 impaired growth in subsequent seasons (i.e., carry-over effects) (Andersen et al., 1997).

6 Light, a component of the plant’s physical environment, is an essential “resource” of 

7 energy content that drives photosynthesis and C assimilation. It has been suggested that 

8 increased light intensity may increase the O3 sensitivity of light-tolerant species while 

9 decreasing that of shade-tolerant species, but this appears to be an oversimplification with 

10 many exceptions. Several studies suggest that the interaction between O3 sensitivity and 

11 light environment is complicated by the developmental stage as well as the light 

12 environment of individual leaves in the canopy (Kitao et al., 2009; Topa et al., 2001; 

13 Chappelka and Samuelson, 1998). 

14 Although the relative humidity of the ambient air has generally been found to increase the 

15 effects of O3 by increasing stomatal conductance (thereby increasing O3 flux into the 

16 leaves), abundant evidence also indicates that the ready availability of soil moisture 

17 results in greater O3 sensitivity (Mills, 2002). The partial “protection” against the effects 

18 of O3 afforded by drought has been observed in field experiments (Low et al., 2006) and 

19 modeled in computer simulations (Broadmeadow and Jackson, 2000). Conversely, 

20 drought may exacerbate the effects of O3 on plants (Pollastrini et al., 2010; Grulke et al., 

21 2003b). There is also some evidence that O3 can predispose plants to drought stress 

22 (Maier-Maercker, 1998). Hence, the nature of the response is largely species-specific and 

23 will depend to some extent upon the sequence in which the stressors occur.

9.4.8.4 Interactions with other Pollutants

Ozone-Nitrogen Interactions

24 Elevated O3 exposure and N deposition often co-occur. However, the interactions of O3

25 exposure and N deposition on vegetation are complex and less well understood compared 

26 to their independent effects. Consistent with the conclusion of the 2006 O3 AQCD, the 

27 limited number of studies published since the last review indicated that the interactive 

28 effects of N and O3 varied among species and ecosystems (Table 9-8). To better 

29 understand these interactions in ecosystems across the U.S., more information is needed 

30 considering combined O3 exposure and N deposition related effects.

31 Nitrogen deposition could stimulate relative growth rate (RGR), and lead to increased 

32 stomatal conductance. Therefore, plants might become more susceptible to O3 exposure. 
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1 Alternatively, N deposition may increase the availability of photosynthates for use in 

2 detoxification and plants could become more tolerant to O3 (Bassin et al., 2007a). Only a 

3 few recent studies have investigated the interactive effects of O3 and N in the U.S. Grulke 

4 et al. (2005) measured stomatal conductance of California black oak (Quercus kelloggii) 

5 at a long-term N-enrichment site located in the San Bernardino Mountains, which is 

6 accompanied by high O3 exposure (80 ppb, 24-h avg. over a six month growing season). 

7 The authors found that N amendment led to poor stomatal control in full sun in 

8 midsummer of the average precipitation years, but enhanced stomatal control in shade 

9 leaves of California black oak. In an OTC study, Handley and Grulke (2008) found that 

10 O3 lowered photosynthetic ability and water-use efficiency, and increased leaf chlorosis 

11 and necrosis of California black oak. Nitrogen fertilization tended to reduce plant 

12 sensitivity to O3 exposure; however, the interaction was not statistically significant.

13 Studies conducted outside the U.S. are also summarized in Table 9-8. Generally, the 

14 responses were species specific. The O3-induced reduction in photosynthetic rate and 

15 biomass loss were greater in the relatively high N treatment for watermelon (Citrillus 

16 lanants) (Calatayud et al., 2006) and Japanese beech (Fagus crenata) seedlings 

17 (Yamaguchi et al., 2007). However, there was no significant interactive effect of O3 and 

18 N on biomass production for Quercus serrata seedlings (Watanabe et al., 2007), young 

19 Norway spruce (Picea abies) trees (Thomas et al., 2005), and young European beech 

20 (Fagus sylvatica) trees (Thomas et al., 2006).
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Table 9-8 Response of plants to the interactive effects of elevated ozone 

exposure and N enrichment

Site Species Ozone exposure N addition Responses References

San 

Bernardino 

Mountains, 

U.S.

California black 

oak (Quercus 

kelloggii)

80 ppb 0, and 50 kg N/ 

ha/yr

N-amended trees had lower late 

summer C gain and greater foliar 

chlorosis in the drought year, and poor 

stomatal control and lower leaf water 

use efficiency and in midsummer of the 

average precipitation year. 

Grulke et al. (2005)

San 

Bernardino 

Mountains, 

U.S.

California black 

oak (Quercus 

kelloggii)

0, 75, and 150 ppb 0, and 50 kg N/ 

ha/yr

N fertilization tended to reduce plant 

sensitivity to O3 exposure; however 

the interaction was not statistically 

significant.

Handley and Grulke 

(2008)

Switzerland Spruce trees 

(Picea abies)

Filtered (19.4-28.1 ppb); 

ambient (37.6-47.4 ppb)

0, 20, 40 and 80 

kg N/ ha/yr

Higher N levels alleviated the negative 

impact of O3 on root starch concentrations

Thomas et al. (2005)

Switzerland Beech trees 

(Fagus sylvatica)

Filtered (19.4-28.1 ppb); 

ambient (37.6-47.4 ppb)

0, 20, 40 and 

80 kg N/ ha/yr

N addition amplified the negative 

effects of O3 on leaf area and shoot 

elongation. 

Thomas et al. (2006)

Switzerland Alpine pasture Ambient ( AOT40 of 11.1-

12.6 ppm-h); 1.2 ambient 

(AOT40 of 15.2-29.5 ppm- h) and 1.6 ambient (28.4-

64.9 ppm-h)

0, 5, 10’ 25, 50 

kg N/ ha/yr

The positive effects of N addition on 

canopy greenness were counteracted 

by accelerated leaf senescence in the 

highest O3 treatment.

Bassin et al. (2007b)

Switzerland Alpine pasture Ambient ( AOT40 of 11.1- 12.6 ppm-h); 1.2 ambient 

(AOT40 of 15.2-29.5 ppmh) and 1.6 ambient (28.4- 64.9 ppm-h)

0, 5, 10, 25, 50 

kg N/ha/yr

Only a small number of species 

showed significant O3 and N 

interactive effects on leaf chlorophyll 

concentration, leaf weight and change 

in 18O, and the patterns were not 

consistent.

Bassin et al. (2009)

Switzerland Alpine pasture Ambient ( AOT40 of 11.1-

12.6 ppm-h); 1.2 ambient 

(AOT40 of 15.2-29.5 ppmh) and 1.6 ambient (28.4-

64.9 ppm-h)

0, 5, 10’ 25, 50 

kg N/ ha/yr

The positive effects of N addition on 

canopy greenness were counteracted 

by accelerated leaf senescence in the 

highest O3 treatment.

Bassin et al. (2007b)

Switzerland Alpine pasture Ambient ( AOT40 of 11.1-

12.6 ppm-h); 1.2 ambient 

(AOT40 of 15.2-29.5 ppm- h) and 1.6 ambient (28.4-

64.9 ppm-h)

0, 5, 10’ 25, 50 

kg N/ ha/yr

Highest N addition resulted in carbon 

loss, but there was no interaction 

between O3 and N treatments.

Volk et al. (2011)

Spain Watermelon 

(Citrillus lanants)

O3 free (AOT40 of 0 ppmh), ambient (AOT40 of 5.1- 6.3 ppm-h) and elevated 

O3 (AOT40 of 32.5-35.6 

ppm-h)

140, 280, and 

436 kg N/ ha/yr

High N concentration enhanced the 

detrimental effects of O3 on Chlorophyll a fluorescence 

parameters, lipid peroxidation, and the 

total yield.

Calatayud et al. (2006)

Spain Trifolium striatum Filtered (24-h avg. of 8-22 

ppb); ambient (29-34 ppb), 

elevated O3 (35-56 ppb) 

10, 30, and 60 

kg N/ ha/yr

O3 reduced total aerial biomass. N 

fertilization counterbalanced O3-

induced effects only when plants were 

exposed to moderate O3 levels 

(ambient) but not under elevated O3

concentrations.

Sanz et al. (2007)

Japan Japanese beech 

seedlings (Fagus 

crenata) 

Filtered (24-h avg. of 10.3- 13.2 ppb); ambient (42.0-

43.3 ppb), 1.5 ambient 

(62.6-63.9 ppb) and 2.0 

ambient (82.7-84.7 ppb)

0, 20 and 50 kg 

N/ ha/yr

The O3-induced reduction in net 

photosynthesis and whole-plant dry 

mass were greater in the relatively 

high N treatment than that in the low N 

treatment. 

Yamaguchi et al.

(2007)

Japan Quercus serrata

seedlings

Filtered (24-h avg. of 10.3-

13.2 ppb); ambient (42.0- 43.3 ppb), 1.5 ambient 

(62.6-63.9 ppb) and 2.0 

ambient (82.7-84.7 ppb)

0, 20 and 50 kg 

N/ ha/yr

No significant interactive effects of O3

and N load on the growth and net 

photosynthetic rate were detected.

Watanabe et al. (2007)

Ozone-Carbon Dioxide Interactions

1 Several decades of research has shown that exposure to elevated CO2 increases 

2 photosynthetic rates (Bernacchi et al., 2006; Bernacchi et al., 2005; Tissue et al., 1999; 
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1 Tissue et al., 1997; Will and Ceulemans, 1997), decreases stomatal conductance 

2 (Ainsworth and Rogers, 2007; Paoletti et al., 2007; Bernacchi et al., 2006; Leakey et al., 

3 2006; Medlyn et al., 2001) and generally increases the growth of plants(McCarthy et al., 

4 2009; Norby et al., 2005). This is in contrast to the decrease in photosynthesis and growth 

5 in many plants that are exposed to elevated O3. The interactive effects on vegetation have 

6 been the subject of research in the past two decades due to the implications on 

7 productivity and water use of ecosystems. This area of research was discussed in detail in 

8 AX9.3.8.1 of the 2006 O3 AQCD and the conclusions made then are still relevant (U.S. 

9 EPA, 2006b).

10 The bulk of the available evidence shows that, under the various experimental conditions 

11 used (which almost exclusively employed abrupt or “step” increases in CO2

12 concentration, as discussed below), increased CO2 levels (ambient + 200 to 400 ppm) 

13 may protect plants from the adverse effects of O3 on growth. This protection may be 

14 afforded in part by CO2 acting together with O3 in inducing stomatal closure, thereby 

15 reducing O3 uptake, and in part by CO2 reducing the negative effects of O3 on Rubisco 

16 and its activity in CO2-fixation. Although both CO2-induced and O3-induced decreases in 

17 stomatal conductance have been observed primarily in short-term studies, recent data 

18 show a long-term and sustained reduction in stomatal conductance under elevated CO2

19 for a number of species (Ainsworth and Long, 2005; Ellsworth et al., 2004; Gunderson et 

20 al., 2002). Instances of increased stomatal conductance have also been observed in 

21 response to O3 exposure, suggesting partial stomatal dysfunction after extended periods 

22 of exposure (Paoletti and Grulke, 2010; Grulke et al., 2007a; Maier-Maercker, 1998).

23 Important caveats must be raised with regard to the findings presented in published 

24 research. The first caveat concerns the distinctly different natures of the exposures to O3

25 and CO2 experienced by plants in the field. Changes in the ambient concentrations of 

26 these gases have very different dynamics. In the context of climate change, CO2 levels 

27 increase relatively slowly (globally 2 ppm/year) and may change little over several 

28 seasons of growth. On the other hand, O3 presents a fluctuating stressor with 

29 considerable hour-to-hour, day-to-day and regional variability (Polle and Pell, 1999). 

30 Almost all of the evidence presented comes from experimentation involving plants 

31 subjected to an abrupt step increase to a higher, steady CO2 concentration. In contrast, the 

32 O3 exposure concentrations usually varied from day to day. Luo and Reynolds (1999), 

33 Hui et al. (2002), and Luo (2001) noted the difficulties in predicting the likely effects of a 

34 gradual CO2 increase from experiments involving a step increase or those using a range 

35 of CO2 concentrations. It is also important to note that the levels of elevated CO2 in 

36 many of the studies will not be experienced in the field for 30 or 40 years, but elevated 

37 levels of O3 can occur presently in several areas of the U.S. Therefore, the CO2 × O3

38 interaction studies may be less relevant for current ambient conditions.
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1 Another caveat concerns the interactions of O3 and CO2 with other climatic variables, 

2 such as temperature and precipitation. In light of the key role played by temperature in 

3 regulating physiological processes and modifying plant response to increased CO2 levels 

4 (Morison and Lawlor, 1999; Long, 1991) and the knowledge that relatively modest 

5 increases in temperature may lead to dramatic consequences in terms of plant 

6 development (Lawlor, 1998), it is important to consider that studying CO2 and O3

7 interactions alone may not create a complete understanding of effects on plants under 

8 future climate change.

9.4.9 Insects and Other Wildlife

9.4.9.1 Insects 

9 Insects may respond indirectly to changes to plants (i.e., increased reactive oxygen 

10 species, altered phytochemistry, altered nutrient content) that occur under elevated O3

11 conditions, or O3 can have a direct effect on insect performance (Menendez et al., 2009). 

12 Effects of O3 on insects occur at the species level (i.e., growth, survival, reproduction, 

13 development, feeding behavior) and at the population and community-level (i.e., 

14 population growth rate, community composition). In general, effects of O3 on insects are 

15 highly context- and species-specific (Lindroth, 2010; Bidart-Bouzat and Imeh-Nathaniel, 

16 2008). Furthermore, plant responses to O3 exposure and herbivore attack have been 

17 demonstrated to share signaling pathways, complicating characterization of these 

18 stressors (Lindroth, 2010; Menendez et al., 2010, 2009). Although both species-level and 

19 population and community-level responses to elevated O3 are observed in field and 

20 laboratory studies discussed below, there is no consensus on how insects respond to 

21 feeding on O3–exposed plants. 

Species-Level Responses

22 In considering insect growth, survival and reproduction in elevated O3 conditions, several 

23 studies have indicated an effect while others have found no correlation. The performance 

24 of five herbivore species (three moths and two weevils) was assessed in an OTC 

25 experiment at 2 × ambient concentration (Peltonen et al., 2010). Growth of larvae of the 

26 Autumnal moth, Epirrita autumna, was significantly decreased in the O3 treatment while 

27 no effects were observed in the other species. In an aphid oviposition preference study 

28 using birch buds grown in a three year OTC experiment, O3 had neither a stimulatory or 

29 deterring effect on egg-laying (Peltonen et al., 2006). Furthermore, changes in birch bud 

30 phenolic compounds associated with the doubled ambient concentrations of O3 did not 
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1 correlate with changes in aphid oviposition (Peltonen et al., 2006). Reproduction in 

2 Popillia japonica, that were fed soybeans and grown under elevated O3 appeared to be 

3 unaffected (O'Neill et al., 2008). In a meta-analysis of effects of elevated O3 on 22 

4 species of trees and 10 species of insects, the rates of survival, reproduction and food 

5 consumption were typically unaffected while development times were reduced and pupal 

6 masses were increased (Valkama et al., 2007).

7 At the Aspen FACE site insect performance under elevated (50-60 ppb) O3 conditions 

8 (approximately 1.5 × background ambient levels of 30-40 ppb O3) have been considered 

9 for several species. Cumulative fecundity of aphids (Cepegillettea betulaefoliae), that 

10 were reared on O3-exposed paper birch (Betula papyrifera) trees, was lower than aphids 

11 from control plots (Awmack et al., 2004). No effects on growth, development, adult 

12 weight, embryo number and birth weight of newborn nymphs were observed. In a study 

13 conducted using three aspen genotypes, performance of the aspen beetle (Chrysomela 

14 crochi) decreased across all parameters measured (development time, adult mass and 

15 survivorship) under elevated O3 (Vigue and Lindroth, 2010). There was an increase in the 

16 development time of male and female aspen beetle larvae although the percentages varied 

17 across genotypes. Decreased beetle adult mass and survivorship was observed across all 

18 genotypes under elevated O3 conditions. Another study from the Aspen FACE site, did 

19 not find any significant effects of elevated O3 on performance (longevity, fecundity, 

20 abundance) of the invasive weevil (Polydrusus sericeus) (Hillstrom et al., 2010b). 

21 Since the 2006 O3 AQCD, several studies have considered the effect of elevated O3 on 

22 feeding behavior of insects. In a feeding preference study, the common leaf weevil 

23 (Phyllobius pyri) consumed significantly more leaf discs from one aspen clone when 

24 compared to a second clone under ambient air conditions (Freiwald et al., 2008). In a 

25 moderately elevated O3 environment (1.5 × ambient), this preference for a certain aspen 

26 clone was less evident, however, leaves from O3-exposed trees were significantly 

27 preferred to leaves grown under ambient conditions. Soybeans grown under enriched O3

28 had significantly less loss of leaf tissue to herbivory in August compared to earlier in the 

29 growing season (July) when herbivory was not affected (Hamilton et al., 2005). Other 

30 plant-herbivore interactions have shown no effects of elevated O3 on feeding. Feeding 

31 behavior of Japanese beetles (P. japonica) appeared to be unchanged when beetles were 

32 fed soybean leaves grown under elevated O3 conditions (O'Neill et al., 2008). At the 

33 Aspen FACE site, feeding by the invasive weevil (Polydrusus sericeus), as measured by 

34 leaf area consumption, was not significantly different between foliage that was grown 

35 under elevated O3 versus ambient conditions (Hillstrom et al., 2010b).
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Population-Level and Community-Level Responses 

1 Recent data on insects provide evidence of population-level and community-level 

2 responses to O3. Elevated levels of O3 can affect plant phytochemistry and nutrient 

3 content which in turn can alter population density and structure of the associated 

4 herbivorous insect communities and impact ecosystem processes (Cornelissen, 2011; 

5 Lindroth, 2010). In 72-hour exposures to elevated O3, mean relative growth rate of the 

6 aphid Diuraphis noxia increased with ozone concentration suggesting that more rapid 

population growth may occur when atmospheric O3 7 is elevated (Summers et al., 1994, 

8 735955). In a long-term study of elevated O3 on herbivore performance at the Aspen 

9 FACE site, individual performance and population-level effects of the aphid 

10 C. betulaefoliae were assessed. Elevated O3 levels had a strong positive effect on the 

11 population growth rates of the aphids; although effects were not detected by measuring 

12 growth, development, adult weight, embryo number or birth weight of newborn nymphs 

13 (Awmack et al., 2004). Conversely, a lower rate of population growth was observed in 

14 aphids previously exposed to O3 in an OTC (Menendez et al., 2010). No direct effects of 

15 O3 were observed; however, nymphs born from adults exposed to and feeding on O3

16 exposed plants were less capable of infesting new plants when compared to nymphs in 

17 the control plots (Menendez et al., 2010). Elevated O3 reduced total arthropod abundance 

18 by 17% at Aspen FACE, largely as a result of the negative effects on parasitoids, 

19 although phloem-feeding insects may benefit (Hillstrom and Lindroth, 2008). Herbivore 

20 communities affected by O3 and N were sampled along an air pollution gradient in the 

21 Los Angeles basin (Jones and Paine, 2006). Abundance, diversity, and richness of 

22 herbivores were not affected. However, a shift in community structure, from phloem23 feeding to chewing dominated communities, was observed along the gradient. No 

24 consistent effect of elevated O3 on herbivory or insect population size was detected at 

25 SoyFACE (O'Neill et al., 2010; Dermody et al., 2008).

26 Evidence of modification of insect populations and communities in response to elevated 

27 O3 includes genotypic and phenotypic changes. In a study conducted at the Aspen FACE 

28 site, elevated O3 altered the genotype frequencies of the pea aphid (Acyrthosiphon pisum) 

29 grown on red clover (Trifolium pratense) over multiple generations (Mondor et al., 

30 2005). Aphid color was used to distinguish between the two genotypes. Ozone increased 

31 the genotypic frequencies of pink-morph:green-morph aphids from 2:1 to 9:1, and 

32 depressed wing-induction responses more strongly in the pink than the green genotype 

33 (Mondor et al., 2005). Growth and development of individual green and pink aphids 

34 reared as a single genotype or mixed genotypes were unaffected by elevated O3 (Mondor 

35 et al., 2010). However, growth of pea aphid populations is not readily predictable using 

36 individual growth rates.
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9.4.9.2 Wildlife 

Herpetofauna

1 Since the 2006 O3 AQCD, direct effects of O3 exposure including physiological changes 

2 and alterations of ecologically important behaviors such as feeding and thermoregulation 

3 have been observed in wildlife. These studies have been conducted in limited laboratory 

exposures, and the levels of O3 4 treatment (e.g. 0.2-0.8 ppm) were often unrealistically 

5 higher than the ambient levels. Amphibians may be especially vulnerable to airborne 

6 oxidants due to the significant gas exchange that occurs across the skin (Andrews et al., 

7 2008; Dohm et al., 2008). Exposure to 0.2 ppm to 0.8 ppm O3 for 4 hours resulted in a 

8 decrease of oxygen consumption and depressed lung ventilation in the California tree 

9 frog Pseudacris cadaverina (Mautz and Dohm, 2004). Following a single 4-h exposure to 

10 O3, reduced pulmonary macrophage phagocytosis was observed at 1 and 24 hours post 

11 exposure in the marine toad (Bufo marinus) indicating an effect on immune system 

12 function (Dohm et al., 2005). There was no difference in macrophage function at 

13 48 hours post exposure in exposed and control individuals. 

14 Behavioral effects of O3 observed in amphibians include responses to minimize the 

15 surface area of the body exposed to the air and a decrease in feeding rates (Dohm et al., 

16 2008; Mautz and Dohm, 2004). The adoption of a low-profile “water conservation 

17 posture” during O3 exposure was observed in experiments with the California tree frog 

18 (Mautz and Dohm, 2004). Marine toads, Bufo marinus, exposed to 0.06 μL/L O3 for 

19 4 hours ate significantly fewer mealworms at 1 hour and 48 hours post exposure than 

20 control toads (Dohm et al., 2008). In the same study, escape/exploratory behavior as 

21 measured by total distance moved was not adversely affected in the O3-exposed 

22 individuals as compared to the controls (Dohm et al., 2008). 

23 Water balance and thermal preference in herpetofauna are altered with elevated O3. 

24 Marine toads exposed to 0.8 ppm O3 for 4 hours exhibited behavioral hypothermia when 

25 temperature selection in the toads was assessed at 1, 24 and 48 hours post exposure 

26 (Dohm et al., 2001). Ozone-exposed individuals lost almost 5g more body mass on 

27 average than controls due to evaporative water loss. At 24 hours after exposure, the 

28 individuals that had lost significant body mass selected lower body temperatures(Dohm 

29 et al., 2001). Behavioral hypothermia was also observed in reptiles following 4-h 

30 exposures to 0.6 ppm O3. Exposure of the Western Fence Lizard (Sceloporus 

occidentalis) at 25o 31 C induced behavioral hypothermia that recovered to control 

32 temperatures by 24 hours (Mautz and Dohm, 2004). The behavioral hypothermic 

response persisted in lizards exposed to O3 at 35o 33 C at 24 hours post exposure resulting in 

a mean body temperature of 3.3o 34 C over controls. 
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Soil Fauna Communities

1 Ozone has also been shown to alter soil fauna communities (Meehan et al., 2010; 

2 Kasurinen et al., 2007; Loranger et al., 2004). Abundance of Acari (mites and ticks) 

3 decreased by 47% under elevated O3 at Aspen FACE site, probably due to the higher 

4 secondary metabolites and lower N concentrations in litter and foliage under elevated O3

5 (Loranger et al., 2004). In another study from the Aspen FACE site, leaf litter collected 

6 from aspen grown under elevated O3 conditions were higher in fiber and lignin 

7 concentrations than trees grown under ambient conditions. These chemical characteristics 

8 of the leaves were associated with increased springtail population growth following 

9 10 weeks in a laboratory microcosm (Meehan et al., 2010). Consumption rates of 

10 earthworms fed on leaf litter for 6 weeks from trees grown under elevated O3 conditions 

11 and ambient air did not vary significantly between treatments (Meehan et al., 2010). In 

12 another study on juvenile earthworms Lumbricus terrestris, individual growth was 

13 reduced when worms were fed high-O3 birch litter from trees exposed for three years to 

14 elevated O3 in an OTC system (Kasurinen et al., 2007). In the same study no significant 

15 growth or mortality effects were observed in isopods. 

9.4.9.3 Indirect Effects on Wildlife

16 In addition to the direct effects of O3 exposure on physiological and behavioral endpoints 

17 observed in the laboratory, there are indirect effects to wildlife. These effects include 

18 changes in biomass and nutritive quality of O3-exposed plants (reviewed in Section 9.4.4) 

19 that are consumed by wildlife. Reduced digestibility of O3-exposed plants may alter 

20 dietary intake and foraging strategies in herbivores. In a study using native highbush 

21 blackberry (Rubus argutus) relative feed value of the plants decreased in bushes exposed 

22 to double ambient concentrations of O3 (Ditchkoff et al., 2009). Indirect effects of 

23 elevated O3 on wildlife include changes in chemical signaling important in ecological 

24 interactions reviewed below.

Chemical Signaling in Ecological Interactions

25 Ozone has been shown to degrade or alter biogenic VOC signals important to ecological 

26 interactions including; (1) attraction of pollinators and seed dispersers; (2) defense 

27 against herbivory; and (3) predator-prey interactions (Pinto et al., 2010; McFrederick et 

28 al., 2009; Yuan et al., 2009; Pinto et al., 2007a; Pinto et al., 2007b). Each signal released 

29 by emitters has an atmospheric lifetime and a unique chemical signature comprised of 

30 different ratios of individual hydrocarbons that is susceptible to atmospheric oxidants 

31 such as O3 (Yuan et al., 2009; Wright et al., 2005). Under elevated O3 conditions, these 
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1 olfactory cues may travel shorter distances before losing their specificity (McFrederick et 

2 al., 2009; McFrederick et al., 2008). Additional non-phytogenic VOC-mediated 

3 interrelationships with the potential to be modified by O3 include territorial marking, 

4 pheromones for attraction of mates and various social interactions including scent trails, 

5 nestmate recognition and signals involved in aggregation behaviors (McFrederick et al., 

6 2009). For example, the alcohols, ketones and aldehydes comprising sex pheromones in 

7 moths could be especially vulnerable to degradation by O3, since some males travel >100 

8 m to find mates (Carde and Haynes, 2004). In general, effects of O3 on scent-mediated 

9 ecological interactions are highly context- and species-specific (Lindroth, 2010; Bidart10 Bouzat and Imeh-Nathaniel, 2008).

Pollination and Seed Dispersal

11 Phytogenic VOC’s attract pollinators and seed dispersers to flowers and fruits (Dudareva 

12 et al., 2006; Theis and Raguso, 2005). These floral scent trails in plant-insect interactions 

13 may be destroyed or transformed by O3 (McFrederick et al., 2008). Using a Lagrangian 

14 model, the rate of destruction of phytogenic VOC’s was estimated in air parcels at 

15 increasing distance from a source in response to increased regional levels of O3, hydroxyl 

16 and nitrate radicals (McFrederick et al., 2008). Based on the model, the ability of 

17 pollinators to locate highly reactive VOCs from emitting flowers may have decreased 

18 from kilometers during pre-industrial times to <200 m at current ambient conditions 

19 (McFrederick et al., 2008). Scents that travel shorter distances (0-10 m) are less 

20 susceptible to air pollutants, while highly reactive scents that travel longer distances (10 

21 to 100’s of meters), are at a higher risk for degradation (McFrederick et al., 2009). For 

22 example, male euglossine bees can detect bait stations from a distance of at least one 

23 kilometer (Dobson, 1994). 

Defense Against Herbivory

24 Ozone can alter the chemical signature of VOCs emitted by plants and these VOCs are 

25 subsequently detected by herbivores (Blande et al., 2010; Iriti and Faoro, 2009; Pinto et 

26 al., 2007a; Vuorinen et al., 2004; Jackson et al., 1999; Cannon, 1990). These 

27 modifications can make the plant either more attractive or repellant to phytophagous 

28 insects (Pinto et al., 2010). For example, under elevated O3, the host plant preference by 

29 forest tent caterpillars increased for birch compared to aspen (Agrell et al., 2005). Ozone30 induced emissions from red spruce needles were found to repel spruce budworm larvae 

31 (Cannon, 1990). Transcriptional profiles of field grown soybean (Glycine max) grown in 

32 elevated O3 conditions were altered due to herbivory by Japanese beetles. The herbivory 
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1 resulted in a higher number of transcripts in the leaves of O3-exposed plants and up2 regulation of antioxidant metabolism associated with plant defense (Casteel et al., 2008). 

3 Ozone may modify signals involved in plant-to-plant interactions and plant defense 

4 against pathogens (Blande et al., 2010; Pinto et al., 2010; McFrederick et al., 2009; Yuan 

5 et al., 2009). In a recent study with lima beans, 80 ppb O3 degraded several herbivore6 induced VOCs, reducing the distance over which plant-to-plant signaling occurred 

7 (Blande et al., 2010).

Predator-Prey Interactions

8 Elevated O3 conditions are associated with disruption of pheromone-mediated 

9 interactions at higher trophic levels (e.g., predators and parasitoids of herbivores). In a 

10 study from the Aspen FACE site, predator escape behaviors of the aphid (Chatophorus 

11 stevensis) were enhanced on O3-fumigated aspen trees although the mechanism of this 

12 response remains unknown (Mondor et al., 2004). The predatory mite Phytoseiulus 

13 persimilis can distinguish between the VOC signature of ozonated lima bean plants and 

14 ozonated lima bean plants simultaneously damaged by T. urticae (Vuorinen et al., 2004)

15 however, other tritrophic interactions have shown no effect (Pinto et al., 2007b). 

16 There are few studies that consider host location behaviors of parasites under elevated 

17 O3. In closed chambers fumigated with O3, the searching efficiency and proportion of the 

18 host larval fruit flies parasitized by Asobara tabida, declined when compared to filtered 

19 air controls (Gate et al., 1995). The host location behavior and rate of parasitism of the 

20 wasp (Coesia plutellae) on Plutella xylostella-infested potted cabbage plants was tested 

21 under ambient and doubled O3 conditions in an open-air fumigation system (Pinto et al., 

22 2008). The number of wasps found in the field and the percentages of parasitized larvae 

23 were not significantly different from controls under elevated O3.

24 Elevated O3 has the potential to perturb specialized food-web communication in 

25 transgenic crops. In insect-resistant oilseed rape Brassica napus grown under 100 ppb O3

26 in a growth chamber, reduced feeding damage by Putella xylostella led to deceased 

27 attraction of the endoparasitoid (Costesia vestalis), however this tritrophic interaction 

28 was influenced by the degree of herbivore feeding (Himanen et al., 2009a; Himanen et 

29 al., 2009b). Under chronic O3-exposure, the insect resistance trait BT cry1Ac in 

30 transgenic B. napus was higher than the control (Himanen et al., 2009c). There was a 

31 negative relative growth rate of the Bt target herbivore, P. xylostella, in all O3 treatments.
Draft – Do Not Cite or Quote 9-100 September 2011

9.4.9.4 Summary 

1 New information on O3 effects on insects and other wildlife is limited to a few species 

2 and there is no consensus on how these organisms respond to elevated O3. Studies 

3 published since the last review show impacts of elevated O3 on both species-level 

4 responses (reproduction, growth, feeding behavior) and community and ecosystem-level 

5 responses (population growth, abundance, shift in community structure) in some insects 

6 and soil fauna. Changes in ecologically important behaviors such as feeding and 

7 thermoregulation have recently been observed with O3 exposure in amphibians and 

8 reptiles, however, these responses occur at concentrations of O3 much higher that 

9 ambient levels.

10 New information available since the last review considers the effects of O3 on chemical 

11 signaling in insect and wildlife interactions. Specifically, studies on O3 effects on 

12 pollination and seed dispersal, defenses against herbivory and predator-prey interactions 

all consider the ability of O3 13 to alter the chemical signature of VOCs emitted during these 

14 pheromone-mediated events. The effects of O3 on chemical signaling between plants, 

15 herbivores and pollinators as well as interactions between multiple trophic levels is an 

16 emerging area of study that may result in further elucidation of O3 effects at the species, 

17 community and ecosystem-level. 

9.5 Effects-Based Air Quality Exposure Indices and Dose 

Modeling

9.5.1 Introduction

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

19 damage (e.g., reduced growth). They are summary measures of monitored ambient O3

20 concentrations over time, intended to provide a consistent metric for reviewing and 

21 comparing exposure-response effects obtained from various studies. Such indices may 

22 also provide a basis for developing a biologically-relevant air quality standard for 

23 protecting vegetation and ecosystems. Effects on plant growth and/or yield have been a 

24 major focus of the characterization of O3 impacts on plants for purposes of the air quality 

25 standard setting process (U.S. EPA, 2007b, 1996e, 1986). The relationship of O3 and 

26 plant responses can be characterized quantitatively as “dose-response” or “exposure27 response.” The distinction is in how the pollutant concentration is expressed: “dose” is 

28 the pollutant concentration absorbed by the leaf over some time period, and is very 

29 difficult to measure directly, whereas “exposure” is the ambient air concentration 
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1 measured near the plant over some time period, and summarized for that period using an 

2 index. Exposure indices have been most useful in considering the form of secondary O3

3 NAAQS, in large part because they only require ambient air quality data rather than more 

4 complex indirect calculations of dose to the plant. The attributes of exposure indices that 

5 are most relevant to plant damage are the weighting of O3 concentrations and the daily 

6 and seasonal time-periods. Several different types of exposure indices are discussed in 

7 Section 9.5.2.

8 Form a theoretical perspective, a measure of plant O3 uptake or dose from ambient air 

9 (either rate of uptake or cumulative seasonal uptake) might be a better predictor of O3

10 damage to plants than an exposure index and may be useful in improving risk assessment. 

11 An uptake estimate would have to integrate all those environmental factors that influence 

12 stomatal conductance, including but not limited to temperature, humidity, and soil water 

13 status (Section 9.5.4). Therefore, uptake values are generally obtained with simulation 

14 models that require knowledge of species- and site-specific values for the variables 

15 mentioned. However, a limitation of modeling dose is that environmental variables are 

16 poorly characterized. In addition, it has also been recognized that O3 detoxification 

17 processes and the temporal dynamics of detoxification must be taken into account in dose 

18 modeling (Heath et al., 2009) (Section 9.5.4). Because of this, research has focused 

19 historically on predictors of O3 damage to plants based only on exposure as a summary 

20 measure of monitored ambient pollutant concentration over some integral of time, rather 

21 than dose (U.S. EPA, 1996c; Costa et al., 1992; Lee et al., 1988b; U.S. EPA, 1986; 

22 Lefohn and Benedict, 1982; O'Gara, 1922).

9.5.2 Description of Exposure Indices Available in the Literature

23 Mathematical approaches for summarizing ambient air quality information in biologically 

24 meaningful forms for O3 vegetation effects assessment purposes have been explored for 

25 more than 80 years (U.S. EPA, 1996b; O'Gara, 1922). In the context of national standards 

26 that protect for “known or anticipated” effects on many plant species in a variety of 

27 habitats, exposure indices provide a numerical summary of very large numbers of 

28 ambient observations of concentration over extended periods. Like any summary statistic, 

29 exposure indices retain information on some, but not all, characteristics of the original 

30 observations. Several indices have been developed to attempt to incorporate some of the 

31 biological, environmental, and exposure factors that influence the magnitude of the 

32 biological response and contribute to observed variability (Hogsett et al., 1988). In the 

33 1996 O3 AQCD, the exposure indices were arranged into five categories; (1) One event, 

34 (2) Mean, (3) Cumulative, (4) Concentration weighted, and (5) Multicomponent, and 

35 were discussed in detail (Lee et al., 1989). Figure 9-9 illustrates how several of the 
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1 indices weight concentration and accumulate exposure. For example, the SUM06 index 

2 (panel a) is a threshold-based approach wherein concentrations below 0.06 ppm are given 

3 a weight of zero and concentrations above 0.06 ppm are given a weight of 1.0 that is 

4 summed usually over 3 to 6 months . The Sigmoid approach (panel b), which is similar to 

5 the W126 index, is a non-threshold approach wherein all concentrations are given a 

6 weight that increases from zero to 1.0 with increasing concentration and summed.

Source: Used with permission from Air and Waste Management Association (Tingey et al., 1991)

(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 concentrationdependent 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.

Figure 9-9 Diagrammatic representation of several exposure indices

illustrating how they weight concentration and accumulate 

exposure. 

7 Various factors with known or suspected bearing on the exposure-response relationship, 

8 including concentration, time of day, respite time, frequency of peak occurrence, plant 
Draft – Do Not Cite or Quote 9-103 September 2011

1 phenology, predisposition, etc., have been weighted with various functions in a large set 

2 of indices. The resulting indices were evaluated by ranking them according to the 

3 goodness-of-fit of a regression model of growth or yield response (Lee et al., 1989). The 

4 statistical evaluations for each of these indices were completed using growth or yield 

5 response data from many earlier exposure studies (e.g., NCLAN). This retrospective 

6 approach was necessary because there were no studies specifically designed to test the 

7 goodness of fit of the various indices. The goodness of fit of a set of linear and nonlinear 

8 models for exposure-response was ranked as various proposed indices were used in turn 

9 to quantify exposure. This approach provided evidence for the best indices. The results of 

10 retrospective analyses are described below.

11 Most of the early retrospective studies reporting regression approaches used data from the 

12 NCLAN program or data from Corvallis, Oregon or California (Costa et al., 1992; Lee et 

13 al., 1988b; Lefohn et al., 1988; Musselman et al., 1988; Lee et al., 1987; U.S. EPA, 

14 1986). These studies were previously reviewed by the EPA (U.S. EPA, 1996c; Costa et 

15 al., 1992) and were in general agreement that the best fit to the data resulted from using 

16 cumulative concentration-weighted exposure indices (e.g. W126, SUM06). Lee et al. 

17 (1987) suggested that exposure indices that included all the 24-h data performed better 

18 than those that used only 7 hours of data; this was consistent with the conclusions of 

19 Heagle et al. (1987) that plants receiving exposures for an additional 5-h/day showed 

20 10% greater yield loss than those exposed for 7-h/day. In an analysis using the National 

21 Crop Loss Assessment Network (NCLAN) data, Lee et al. (1988) found several indices 

22 which only cumulated and weighted higher concentrations (e.g., W126, SUM06, SUM08, 

23 and AOT40) performed very well. Amongst this group no index had consistently better 

24 fits than the other indices across all studies and species (Heagle et al., 1994b; Lefohn et 

25 al., 1988; Musselman et al., 1988). Lee et al. (1988) found that adding phenology 

26 weighting to the index somewhat improved the performance of the indices. The “best”

27 exposure index was a phenologically weighted cumulative index, with sigmoid weighting 

28 on concentration and a gamma weighting function as a surrogate for plant growth stage. 

29 This index provided the best statistical fit when used in the models under consideration, 

30 but it required data on species and site conditions, making specification of weighting 

31 functions difficult for general use. 

32 Other factors, including predisposition time (Hogsett et al., 1988; McCool et al., 1988)

33 and crop development stage (Tingey et al., 2002; Heagle et al., 1991) contributed to 

34 variation in the biological response and suggested the need for weighting O3

35 concentrations to account for predisposition time and phenology. However, the roles of 

36 predisposition and phenology in plant response vary considerably with species and 

37 environmental conditions; therefore, specification of a weighting function for general use 

38 in characterizing plant exposure has not been possible.
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1 European scientists took a similar approach in developing indices describing growth and 

2 yield loss in crops and tree seedlings, using OTCs with modified ambient exposures, but 

3 many fewer species and study locations were employed in the European studies. There is 

4 evidence from some European studies that a lower (Pleijel et al., 1997) or higher (Finnan 

5 et al., 1997; Finnan et al., 1996) cutoff value in indices with a threshold may provide a 

6 better statistical fit to the experimental data. Finnan et al. (1997) used seven exposure 

7 studies of spring wheat to confirm that cumulative exposure indices emphasizing higher 

8 O3 concentrations were best related to plant response and that cumulative exposure 

9 indices using weighting functions, including cutoff concentrations, allometric and 

10 sigmoidal, provided a better fit and that the ranking of these indices differed depending 

11 on the exposure-response model used. Weighting those concentrations associated with 

12 sunshine hours in an attempt to incorporate an element of plant uptake did not improve 

13 the index performance (Finnan et al., 1997). A more recent study using data from several 

14 European studies of Norway spruce, analyzed the relationship between relative biomass 

15 accumulation and several cumulative, weighted indices, including the AOT40 (area over 

16 a threshold of 40ppb) and the SUM06 (Skarby et al., 2004). All the indices performed 

17 relatively well in regressing biomass and exposure index, with the AOT20 and AOT30 

doing slightly better than others (r2 18 = 0.46-0.47). In another comparative study of four 

19 independent data sets of potato yield and different cumulative uptake indices with 

different cutoff values, a similarly narrow range of r2 was observed (r2 20 = 0.3-0.4) (Pleijel 

21 et al., 2004b).

22 In Europe, the cutoff concentration-weighted index AOT40 was selected in developing 

23 exposure-response relationships based on OTC studies of a limited number of crops and 

24 trees (Grunhage and Jager, 2003). The United Nations Economic Commission for Europe 

25 (UNECE, 1988) adopted the critical levels approach for assessment of O3 risk to 

26 vegetation across Europe. As used by the UNECE, the critical levels are not like the air 

27 quality regulatory standards used in the U.S., but rather function as planning targets for 

28 reductions in pollutant emissions to protect ecological resources. Critical levels for O3 are 

29 intended to prevent long-term deleterious effects on the most sensitive plant species 

30 under the most sensitive environmental conditions, but not intended to quantify O3

31 effects. A critical level was defined as “the concentration of pollutant in the atmosphere 

32 above which direct adverse effects on receptors, such as plants, ecosystems, or materials 

33 may occur according to present knowledge” (UNECE, 1988). The nature of the “adverse 

34 effects” was not specified in the original definition, which provided for different levels 

35 for different types of harmful effect (e.g., visible injury or loss of crop yield). There are 

36 also different critical levels for crops, forests, and semi-natural vegetation. The caveat, 

37 “according to present knowledge” is important because critical levels are not rigid; they 

38 are revised periodically as new scientific information becomes available. For example, 

39 the original critical level for O3 specified concentrations for three averaging times, but 
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1 further research and debate led to the current critical level being stated as the cumulative 

2 exposure (concentration × hours) over a cutoff concentration of 40 ppb (AOT40) (Fuhrer 

3 et al., 1997). 

4 More recently in Europe, a decision was made to work towards a flux-based approach 

5 (see section 9.5.4) for the critical levels (“Level II”), with the goal of modeling O3 flux6 effect relationships for three vegetation types: crops, forests, and semi-natural vegetation 

7 (Grunhage and Jager, 2003). Progress has been made in modeling flux (U.S. EPA, 2006b)

8 and the Mapping Manual is being revised (Ashmore et al., 2004a, b; Grennfelt, 2004; 

9 Karlsson et al., 2003). The revisions may include a flux-based approach for three crops: 

10 wheat, potatoes, and cotton. However, because of a lack of flux-response data, a 

11 cumulative, cutoff concentration-based (AOTx) exposure index will remain in use for the 

12 near future for most crops and for forests and semi-natural herbaceous vegetation 

13 (Ashmore et al., 2004b)

14 In both the U.S. and Europe, the adequacy of these numerical summaries of exposure in 

15 relating biomass and yield changes have, for the most part, all been evaluated using data 

16 from studies not necessarily designed to compare one index to another (Skarby et al., 

17 2004; Lee et al., 1989; Lefohn et al., 1988). Very few studies in the U.S. have addressed 

18 this issue since the 2006 O3 AQCD. McLaughlin et al. (2007a) reported that the 

19 cumulative exposure index of AOT60 related well to reductions in growth rates at forest 

20 sites in the southern Appalachian Mountains. However, the authors did not report an 

21 analysis to compare multiple indices. Overall, given the available data from previous O3

22 AQCDs and the few recent studies, the cumulative, concentration-weighted indices 

23 perform better than the peak or mean indices. It is still not possible, however, to 

24 distinguish the differences in performance among the cumulative, concentration-weighted 

25 indices. 

26 The main conclusions from the 1996 and 2006 O3 AQCDs regarding an index based on 

27 ambient exposure are still valid. No information has come forth since the 2006 O3 AQCD 

28 to alter those conclusions significantly. These key conclusions can be restated as follows:

29  O3 effects in plants are cumulative;

30  higher O3 concentrations appear to be more important than lower 

31 concentrations in eliciting a response;

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

33 and

34  exposure indices that accumulate the O3 hourly concentrations and 

35 preferentially weight the higher concentrations have better statistical fits to 

36 growth/yield response than do the mean and peak indices.
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1 Following the 2006 criteria review process (U.S. EPA, 2006b), the EPA proposed an 

2 alternative form of the secondary NAAQS for O3 using a cumulative, concentration3 weighted exposure index to protect vegetation from damage (72 FR 37818). The EPA 

4 considered two specific concentration-weighted indices: the cutoff concentration 

5 weighted SUM06 and the sigmoid-weighted W126 exposure index (U.S. EPA, 2007b). 

6 These two indices performed equally well in predicting the exposure-response 

7 relationships observed in the crop and tree seedlings studies (Lee et al., 1989). At a 

8 workshop convened to consider the science supporting these indices (Heck and Cowling, 

9 1997) there was a consensus that these cumulative concentration-weighted indices being 

10 considered were equally capable of predicting plant response. Below are the definitions 

11 of the two cumulative index forms considered in the previous staff paper review (U.S. 

12 EPA, 2007b):

13  SUM06: Sum of all hourly O3 concentrations greater than or equal to 

14 0.06 ppm observed during a specified daily and seasonal time window (Figure 

15 9-9a).

16  W126: Sigmoidally weighted sum of all hourly O3 concentrations observed 

17 during a specified daily and seasonal time window (Similar to Figure 9-9b). 

18 The sigmoidal weighting of hourly O3 concentration is given in the equation 

19 below, where C is the hourly O3 concentration in ppm:

Equation 9-1

20 The SUM06 and W126 indices have a variety of relevant time windows that may be 

21 applied and are discussed in Section 9.5.3. 

22 It should be noted that there are some important differences between the SUM06 and 

23 W126. When considering the response of vegetation to ozone exposures represented by 

24 the threshold (e.g., SUM06) and non-threshold (e.g., W126) indices, the W126 metric 

25 does not have a cut-off in the weighting scheme as does SUM06 and thus it includes 

26 consideration of potentially damaging exposures below 60 ppb. The W126 metric also 

27 adds increasing weight to hourly concentrations from about 40 ppb to about 100 ppb. 

28 This is unlike cut-off metrics such as the SUM06 where all concentrations above 60 ppb 

29 are treated equally. This is an important feature of the W126 since as hourly 

30 concentrations become higher, they become increasingly likely to overwhelm plant 

31 defenses and are known to be more detrimental to vegetation (See Section 9.5.3.1).

c C e

w 126 1 4403

1 − + =
Draft – Do Not Cite or Quote 9-107 September 2011

9.5.3 Important Components of Exposure Indices 

1 In the previous O3 AQCDs it was established that higher hourly concentrations have 

2 greater effects on vegetation than lower concentrations (U.S. EPA, 2006b, 1996c). 

3 Further, it was determined that the diurnal and seasonal duration of exposure is important 

4 for plant response. Weighting of hourly concentrations and the diurnal and seasonal time 

5 window of exposure are the most important variables in a cumulative exposure index and 

6 will be discussed below. However, these variables must be taken in the context of plant 

7 phenology, diurnal conductance rates, plant canopy structure, and detoxification 

8 mechanisms of vegetation as well as the climate and meteorology, all of which are 

9 determinants of plant response. These more specific factors will be discussed in the 

10 uptake and dose modeling section 9.5.4. 

9.5.3.1 Role of Concentration

11 The significant role of peak O3 concentrations was established based on several 

12 experimental studies (U.S. EPA, 1996c). Several studies (Oksanen and Holopainen, 

13 2001; Yun and Laurence, 1999; Nussbaum et al., 1995) have added support for the 

14 important role that peak concentrations, as well as the pattern of occurrence, plays in 

15 plant response to O3. Oksanen and Holopainen (2001) found that the peak concentrations 

16 and the shape of the O3 exposure (i.e., duration of the event) were important determinants 

17 of foliar injury in European white birch saplings, but growth reductions were found to be 

18 more related to total cumulative exposure. Based on air quality data from 10 U.S. cities, 

19 three 4-week exposure treatments having the same SUM06 value were constructed by 

20 Yun and Laurence (1999). The authors used different exposure regimes to explore effects 

21 of treatments with variable versus uniform peak occurrence during the exposure period. 

22 The authors reported that the variable peak exposures were important in causing injury, 

23 and that the different exposure treatments, although having the same SUM06, resulted in 

24 very different patterns of foliar injury. Nussbaum et al. (1995) also found peak 

25 concentrations and the pattern of occurrence to be critical in determining the measured 

26 response. The authors recommended that to describe the effect on total forage yield, peak 

27 concentrations >0.11 ppm must be emphasized by using an AOT with higher threshold 

28 concentrations.

29 A greater role for peak concentrations in effects on plant growth might be inferred based 

30 on air quality analyses for the southern California area (Tingey et al., 2004; Lee et al., 

31 2003a). In the late 1960s and 1970s, extremely high O3 concentrations had impacted the 

32 San Bernardino National Forest. However, over the past 20+ years, significant reductions 

33 in O3 exposure have occurred (Bytnerowicz et al., 2008; Lee et al., 2003a; Lefohn and 
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1 Shadwick, 2000; Davidson, 1993). An illustration of this improvement in air quality is 

2 shown by the 37-year history of O3 air quality at the Crestline site in the San Bernardino 

3 Mountains (Figure 9-10) (Lee et al., 2003a). Ozone exposure increased from 1963 to 

4 1979 concurrent with increased population and vehicular miles, followed by a decline to 

5 the present mirroring decreases in precursor emissions. The pattern in exposure was 

6 evident in various exposure indices including the cumulative concentration weighted 

7 (SUM06), as well as maximum peak event (1-h peak), and the number of days having 

8 hourly averaged O3 concentrations greater than or equal to 95 ppb. The number of days 

9 having hourly averaged O3 concentrations greater than or equal to 95 ppb declined 

10 significantly from 163 days in 1978 to 103 days in 1997. The changes in ambient O3 air 

11 quality for the Crestline site were reflected in the changes in frequency and magnitude of 

12 the peak hourly concentration and the duration of exposure (Figure 9-10). Considering 

13 the role of exposure patterns in determining response, the seasonal and diurnal patterns in 

14 hourly O3 concentration did not vary appreciably from year to year over the 37-year 

15 period (Lee et al., 2003a).

16 The potential importance of exposure to peak concentrations comes both from results of 

17 measures of tree conditions on established plots and from results of model simulations. 

18 Across a broad area of the San Bernardino National Forest, the Forest Pest Management 

19 (FPM) method of injury assessment indicated an improvement in crown condition from 

20 1974 to 1988; and the area of improvement in injury assessment is coincident with an 

21 improvement in O3 air quality (Miller and Rechel, 1999). A more recent analysis of 

22 forest changes in the San Bernardino National Forest, using an expanded network of 

23 monitoring sites, has verified significant changes in growth, mortality rates, basal area, 

24 and species composition throughout the area since 1974 (Arbaugh et al., 2003). A model 

25 simulation of ponderosa pine growth over the 40-year period in the San Bernardino 

26 National Forest showed a significant impact of O3 exposure on tree growth and indicates 

27 improved growth with reduced O3 concentrations. This area has also experienced 

28 elevated N deposition and based on a number of environmental indicators, it appears that 

29 this area is experiencing N saturation (Fenn et al., 1996). To account for this potential 

30 interaction, the model simulations were conducted under conditions of unlimited soil N. 

31 The actual interactions are not known. The improvement in growth over the years was 

32 attributed to improved air quality, but no distinction was made regarding the relative role 

33 of mid-range and higher hourly concentrations, only that improved growth tracked 

34 decreasing SUM06, maximum peak concentration, and number of days of hourly O3

35 >95 ppb (Tingey et al., 2004). A summary of air quality data from 1980 to 2000 for the 

36 San Bernardino National Forest area of the number of “mid-range” hourly concentrations 

37 indicated no dramatic changes over this 20-year period, ranging from about 1,500 to 

38 2,000 hours per year (Figure 9-11). There was a slow increase in the number of mid39 range concentrations from 1980 to 1986, which corresponds to the period after 
Draft – Do Not Cite or Quote 9-109 September 2011

1 implementation of the air quality standard. Another sharper increase was observed in the 

2 late 1990s. This pattern of occurrence of mid-range hourly concentrations suggests a 

3 lesser role for these concentration ranges compared to the higher values in either of the 

4 ground-level tree injury observations of the model simulation of growth over the 40-year 

5 period.

6

7

Source: Used with permission from Elsevier Science Ltd. (Lee et al., 2003a).

Annual ROG and NOX emissions data for San Bernardino County were obtained from Alexis et al. (2001a) and the California Air 

Resource Board’s emission inventory available at http://www.arb.ca.gov/aqd/aqdpage.htm (Cal/EPA, 2010). 

Figure 9-10 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 (NOX) for San Bernardino 

County. 
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Figure 9-11 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.

9.5.3.2 Diurnal and Seasonal Exposure

Diurnal Exposure

1 The diurnal patterns of maximal leaf/needle conductance and occurrence of higher 

2 ambient concentrations can help determine which hours during the day over a season 

3 should be included in an exposure index. Stomatal conductance is species and phenology 

4 dependent and is linked to both diurnal and seasonal meteorological activity as well as to 

5 soil/site conditions (e.g., VPD, soil moisture). Daily patterns of leaf/needle conductance 

6 are often highest in midmorning, whereas higher ambient O3 concentrations generally 

7 occurred in early to late afternoon when stomata were often partially closed and 

8 conductances were lower. Total O3 flux depends on atmospheric and boundary layer 

9 resistances, both of which exhibit variability throughout the day. Experimental studies 

10 with tree species demonstrated the decoupling of ambient O3 exposure, peak occurrence, 
Draft – Do Not Cite or Quote 9-111 September 2011

1 and gas exchange, particularly in areas of drought (Panek, 2004). Several studies have 

2 suggested that ponderosa pine trees in the southern and northern Sierra Nevada 

3 Mountains may not be as susceptible to high O3 concentrations as to lower 

4 concentrations, due to reduced needle conductance and O3 uptake during the period when 

5 the highest concentrations occur (Panek et al., 2002; Panek and Goldstein, 2001; Bauer et 

6 al., 2000; Arbaugh et al., 1998). Panek et al. (2002) compared direct O3 flux 

7 measurements into a canopy of ponderosa pine and demonstrated a lack of correlation of 

8 daily patterns of conductance and O3 occurrence, especially in the late season drought 

9 period; the authors concluded that a consideration of climate or season was essential, 

10 especially considering the role of soil moisture and conductance/uptake. In contrast, 

11 Grulke et al. (2002) reported high conductance when O3 concentrations were high in the 

12 same species, but under different growing site conditions. The longer-term biological 

13 responses reported by Miller and Rechel (1999) for ponderosa pine in the same region, 

14 and the general reduction in recent years in ambient O3 concentrations, suggest that 

15 stomatal conductance alone may not be a sufficient indicator of potential vegetation 

16 injury or damage. Another consideration for the effect of O3 uptake is the diurnal pattern 

17 of detoxification capacity of the plant. The detoxification capacity may not follow the 

18 same pattern as stomatal conductance (Heath et al., 2009).

19 The use of a 12-h (8:00 a.m. to 8:00 p.m.) daylight period for a W126 cumulating 

20 exposure was based primarily on evidence that the conditions for uptake of O3 into the 

21 plant occur mainly during the daytime hours. In general, plants have the highest stomatal 

22 conductance during the daytime and in many areas atmospheric turbulent mixing is 

23 greatest during the day as well (Uddling et al., 2010; U.S. EPA, 2006b). However, 

24 notable exceptions to maximum daytime conductance are cacti and other plants with 

25 crassulacean acid metabolism (CAM photosynthesis) which only open their stomata at 

26 night. This section will focus on plants with C3 and C4 photosynthesis, which generally 

27 have maximum stomatal conductance during the daytime.

28 Recent reviews of the literature reported that a large number of species had varying 

29 degrees of nocturnal stomatal conductance (Caird et al., 2007; Dawson et al., 2007; 

30 Musselman and Minnick, 2000). The reason for night-time water loss through stomata is 

31 not well understood and is an area of active research (e.g. (Christman et al., 2009; 

32 Howard et al., 2009) Night-time stomata opening may be enhanced by O3 damage that 

33 could result in loss of stomatal control, and less complete closure of stomata, than under 

34 low O3 conditions (Caird et al., 2007; Grulke et al., 2007b). In general, the rate of 

35 stomatal conductance at night is much lower than during the day (Caird et al., 2007). 

36 Atmospheric turbulence at night is also often low, which results in stable boundary layers 

37 and unfavorable conditions for O3 uptake into vegetation (Finkelstein et al., 2000). 

38 Nevertheless, nocturnal turbulence does intermittently occur and may result in 
Draft – Do Not Cite or Quote 9-112 September 2011

1 nonnegligible O3 flux into the plants. In addition, plants might be more susceptible to O3

2 exposure at night than during the daytime, because of potentially lower plant defenses 

3 (Heath et al., 2009; Loreto and Fares, 2007; Musselman et al., 2006; Musselman and 

4 Minnick, 2000). For significant nocturnal stomatal flux and O3 effects to occur, specific 

5 conditions must exist. A susceptible plant with nocturnal stomatal conductance and low 

6 defenses must be growing in an area with relatively high night-time O3 concentrations 

7 and appreciable nocturnal atmospheric turbulence. It is unclear how many areas there are 

8 in the U.S. where these conditions occur. It may be possible that these conditions exist in 

9 mountainous areas of southern California, front-range of Colorado (Turnipseed et al., 

10 2009) and the Great Smoky Mountains of North Carolina and Tennessee. Tobiessen et al. 

11 (1982) found that shade intolerant tree species showed opening of stomata in the dark and 

12 did not find this in shade tolerant species. This may indicate shade intolerant trees may be 

13 more likely to be susceptible to O3 exposure at night. More information is needed in 

14 locations with high night-time O3 to assess the local O3 patterns, micrometeorology and 

15 responses of potentially vulnerable plant species. 

16 Several field studies have attempted to quantify night-time O3 uptake with a variety of 

17 methods. However, many of these studies have not linked the night-time flux to measured 

18 effects on plants. Grulke et al. (2004) showed that the stomatal conductance at night for 

19 ponderosa pine in the San Bernardino National Forest (CA) ranged from one tenth to one 

20 fourth that of maximum daytime stomatal conductance. In June, at a high-elevation site, it 

21 was calculated that 11% of the total daily O3 uptake of pole-sized trees occurred at night. 

22 In late summer, however, O3 uptake at night was negligible. However, this study did not 

23 consider the turbulent conditions at night. Finklestein et al. (2000) investigated O3

24 deposition velocity to forest canopies at three different sites. The authors found the total 

25 flux (stomatal and non-stomatal) to the canopy to be very low during night-time hours as 

26 compared to day-time hours. However, the authors did note that higher nocturnal 

27 deposition velocities at conifer sites may be due to some degree of stomatal opening at 

28 night (Finkelstein et al., 2000). Work by Mereu et al. (2009) in Italy on Mediterranean 

29 species indicated that nocturnal uptake was from 10 to 18% of total daily uptake during a 

30 weak drought and up to 24% as the drought became more pronounced. The proportion of 

31 night-time uptake was greater during the drought due to decreases in daytime stomatal 

32 conductance (Mereu et al., 2009). In a study conducted in California, Fares et al. (Fares et 

33 al., 2011) reported that calculated mean percentages of nocturnal uptake were 5%, 12.5%, 

34 6.9% of total O3 uptake for lemon, mandarin, and orange, respectively. In another recent 

35 study at the Aspen FACE site in Wisconsin, calculated leaf-level stomatal O3 flux was 

36 near zero from the night-time hours of 8:00 p.m. to 5:00 a.m. (Uddling et al., 2010). This 

37 was likely due to low horizontal wind speed (>1 m/s) and low O3 concentrations 

38 (<25 ppb) during those same night-time hours (Figure 9-12).
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Source: Used with permission from Elsevier Ltd (Uddling et al., 2010).

Figure 9-12 Mean diurnal. (a) conductance through boundary layer and 

stomata (gbs), (b) Ozone concentration, and leaf-level 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 nonstomatal ozone flux, respectively.

1 A few studies have tested the biological effects of night-time O3 exposure on vegetation 

2 in controlled chambers. Biomass of ponderosa pine seedlings was significantly reduced 

3 when seedlings were exposed to either daytime or nighttime episodic profiles (Lee and 

4 Hogsett, 1999). However, the biomass reductions were much greater with daytime peak 

5 concentrations than with nighttime peak concentrations. Similarly, birch cuttings grown 

6 in field chambers that were exposed to O3 at night only, daytime only, and 24 hours 

7 showed similar reductions in biomass in night only and day only treatments. Birch 

8 seedling showed greater reductions in growth in 24-h exposures than those exposed to O3

9 at night or day only (Matyssek et al., 1995). Field mustard (Brassica rapa) plants 
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1 exposed to O3 during the day or night showed little significant difference in the amounts 

2 of injury or reduced growth response to O3 treatment, although the stomatal conductance 

3 was 70-80% lower at night (Winner et al., 1989). These studies show that effects can be 

4 seen with night-time exposures to O3 but when atmospheric conditions are stable at night, 

5 it is uncertain how these exposures may affect plants and trees with complex canopies in 

6 the field.

Seasonal Exposure

7 Vegetation across the U.S. has widely varying periods of physiological activity during the 

8 year due to variability in climate and phenology. In order for a particular plant to be 

9 vulnerable to O3 pollution, it must have foliage and be physiologically active. Annual 

10 crops are typically grown for periods of two to three months. In contrast, perennial 

11 species may be photosynthetically active longer (up to 12 months each year for some 

12 species) depending on the species and where it is grown. In general, the period of 

13 maximum physiological activity and thus, potential O3 uptake for vegetation coincides 

14 with some or all of the intra-annual period defined as the O3 season, which varies on a 

15 state-by-state basis (Figure 3-19). This is because the high temperature and high light 

16 conditions that typically promote the formation of tropospheric O3 also promote 

17 physiological activity in vegetation. There are very limited exceptions to this pattern 

18 where O3 can form in the winter in areas in the western U.S. with intense natural gas 

19 exploration (Pinto, 2009), but this is typically when plants are dormant and there is little 

20 chance of O3 uptake. The selection of any single window of time for a national standard 

21 to consider hourly O3 concentrations represents a compromise, given the significant 

22 variability in growth patterns and lengths of growing season among the wide range of 

23 vegetation species that may experience adverse effects associated with O3 exposure.

24 Various intra-annual averaging and accumulation time periods have been considered for 

25 the protection of vegetation. The 2010 proposal for secondary O3 standard (75 FR 2938, 

26 p. 3003) proposed to use the maximum consecutive 3-month period within the O3 season. 

27 The U.S. Forest Service and federal land managers have used a 24-h W126 accumulated 

28 for 6 months from April through September (2000). However, some monitors in the U.S. 

29 are operational for as little as four months and would not have enough data for a 6-month 

30 seasonal window. The exposure period in the vast majority of O3 exposure studies 

31 conducted in the U.S. has been much shorter than 6 months. Most of the crop studies 

32 done through NCLAN had exposures less than three months with an average of 77 days. 

33 Open-top chamber studies of tree seedlings, compiled by the EPA, had an average 

34 exposure of just over three months or 99 days. In more recent FACE experiments, 

35 SoyFACE exposed soybeans for an average of approximately 120 days per year and the 

36 Aspen FACE experiment exposed trees to an average of approximately 145 days per year 
Draft – Do Not Cite or Quote 9-115 September 2011

1 of elevated O3, which included the entire growing season at those particular sites. Despite 

2 the possibility that plants may be exposed to ambient O3 longer than 3 months in some 

3 locations, there is a lack of exposure experiments conducted for longer than 3 months.

A

B

Figure 9-13 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). Highest 6 month W126

0

10

20

30

40

50

60

70

80

90

100

110

Highest 3 month W126

0 10 20 30 40 50 60 70

Highest 6 month W126

0

10

20

30

40

50

60

70

80

90

100

110

Highest 3 month W126

0 10 20 30 40 50 60 70
Draft – Do Not Cite or Quote 9-116 September 2011

1 In an analysis of the 3- and 6-month maximum W126 values calculated for over 1,200 

2 AQS (Air Quality System) and CASTNET (Clean Air Status and Trend Network) EPA 

3 monitoring sites for the years 2008-2009, it was found that these 2 accumulation periods 

4 resulted in highly correlated metrics (Figure 9-13). The two accumulation periods were 

5 centered on the yearly maximum for each monitoring site, and it is possible that this 

6 correlation would be weaker if the two periods were not temporally aligned. In the U.S., 

7 W126 cumulated over 3 months, and W126 cumulated over 6 months are proxies of one 

8 another, as long as the period in which daily W126 is accumulated corresponds to the 

9 seasonal maximum. Therefore, it is expected that either statistic will predict vegetation 

10 response equally well. In other words, the strength of the correlation between maximum 

11 3-month W126 and maximum 6-month W126 is such that there is no material difference 

12 in their predictive value for vegetation response.

9.5.4 Ozone Uptake/Dose Modeling for Vegetation

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

14 based on estimating the O3 concentration from the atmosphere that enters the leaf (i.e., 

15 flux or deposition). Interest has been increasing in recent years, particularly in Europe, in 

16 using mathematically tractable flux models for O3 assessments at the regional, national, 

17 and European scale (Matyssek et al., 2008; Paoletti and Manning, 2007; M and M, 2004; 

18 Emberson et al., 2000b; Emberson et al., 2000a). Some researchers have claimed that 

19 using flux models can be used to better predict vegetation responses to O3 than exposure20 based approaches (Matyssek et al., 2008). However, other research has suggested that 

21 flux models do not predict vegetation responses to O3 better than exposure-based models, 

22 such as AOT40 (Gonzalez-Fernandez et al., 2010). While some efforts have been made in 

23 the U.S. to calculate O3 flux into leaves and canopies (Fares et al., 2010a; Turnipseed et 

24 al., 2009; Uddling et al., 2009; Bergweiler et al., 2008; Hogg et al., 2007; Grulke et al., 

25 2004; Grantz et al., 1997; Grantz et al., 1995), little information has been published 

26 relating these fluxes to effects on vegetation. The lack of flux data in the U.S. and the 

27 lack of understanding of detoxification processes have made this technique less viable for 

28 vulnerability and risk assessments in the U.S. 

29 Flux calculations are data intensive and must be carefully implemented. Reducing 

30 uncertainties in flux estimates for areas with diverse surface or terrain conditions to 

31 within ±50% requires “very careful application of dry deposition models, some model 

32 development, and support by experimental observations” (Wesely and Hicks, 2000). As 

33 an example, the annual average deposition velocity of O3 among three nearby sites in 

34 similar vegetation was found to vary by ±10%, presumably due to terrain (Brook et al., 

35 1997). Moreover, the authors stated that the actual variation was even greater, because 
Draft – Do Not Cite or Quote 9-117 September 2011

1 stomatal uptake was unrealistically assumed to be the same among all sites, and flux is 

2 strongly influenced by stomatal conductance (Brook et al., 1997; Massman and Grantz, 

3 1995; Fuentes et al., 1992; Reich, 1987; Leuning et al., 1979). This uptake-based 

4 approach to quantify the vegetation impact of O3 requires inclusion of those factors that 

5 control the diurnal and seasonal O3 flux to vegetation (e.g., climate patterns, species 

6 and/or vegetation-type factors and site-specific factors). The models have to distinguish 

7 between stomatal and non-stomatal components of O3 deposition to adequately estimate 

8 actual concentration reaching the target tissue of a plant to elicit a response (Uddling et 

9 al., 2009). Determining this O3 uptake via canopy and stomatal conductance relies on 

10 models to predict flux and ultimately the “effective” flux (Grunhage et al., 2004; 

11 Massman, 2004; Massman et al., 2000). “Effective flux” has been defined as the balance 

12 between O3 flux and detoxification processes (Heath et al., 2009; Musselman and 

13 Massman, 1999; Grunhage and Haenel, 1997; Dammgen et al., 1993). The time14 integrated “effective flux” is termed “effective dose.” The uptake mechanisms and the 

15 resistances in this process, including stomatal conductance and biochemical defense 

16 mechanisms, are discussed below. The flux-based index is the goal for the “Level II”

17 critical level for assessment of O3 risk to vegetation and ecosystems across Europe 

18 (Ashmore et al., 2004a).

19 An important consideration in both O3 exposure and uptake is how the O3 concentration 

20 at the top of low vegetation such as, crops and tree seedlings may be lower than the 

21 height at which the measurement is taken. Ambient monitor inlets in the U.S. are 

22 typically at heights of 3 to 5 meters. During daytime hours, the vertical O3 gradient can 

23 be relatively small because turbulent mixing maintains the downward flux of O3. For 

24 example, Horvath et al. (1995) calculated a 7% decrease in O3 going from a height of 4 

25 meters down to 0.5 meters above the surface during unstable (or turbulent) conditions in 

26 a study over low vegetation in Hungary [see section AX3.3.2. of the 2006 O3 AQCD 

27 (U.S. EPA, 2006b)]. There have been several studies indicating decreased O3

28 concentrations under tree canopies (Kolb et al., 1997; Samuelson and Kelly, 1997; Joss 

29 and Graber, 1996; Fredericksen et al., 1995; Lorenzini and Nali, 1995; Enders, 1992; 

30 Fontan et al., 1992; Neufeld et al., 1992). In contrast, for forests, measured data may 

31 underestimate O3 concentration at the top of the canopy. The difference between 

32 measurement height and canopy height is a function of several factors, the intensity of 

33 turbulent mixing in the surface layer and other meteorological factors, canopy height and 

34 total deposition to the canopy. Some researchers have used deposition models to estimate 

35 O3 concentration at canopy-top height based on concentrations at measurement height 

36 (Emberson et al., 2000a). However, deposition models usually require meteorological 

37 data inputs that are not always available or well characterized across large spatial scales.
Draft – Do Not Cite or Quote 9-118 September 2011

1 Soil moisture is a critical factor in controlling O3 uptake through its effect on plant water 

2 status and stomatal conductance. In an attempt to relate uptake, soil moisture, and 

3 ambient air quality to identify areas of potential risk, available O3 monitoring data for 

4 1983 to 1990 were used along with literature-based seedling exposure-response data from 

5 regions within the southern Appalachian Mountains that might have experienced O3

6 exposures sufficient to inhibit growth (Lefohn et al., 1997). In a small number of areas 

7 within the region, O3 exposures and soil moisture availability were sufficient to possibly 

8 cause growth reductions in some O3 sensitive species (e.g., black cherry). The 

9 conclusions were limited, however, because of the uncertainty in interpolating O3

10 exposures in many of the areas and because the hydrologic index used might not reflect 

11 actual water stress.

12 The non-stomatal component of plant defenses are the most difficult to quantify, but 

13 some studies are available (Heath et al., 2009; Barnes et al., 2002; Plochl et al., 2000; 

14 Chen et al., 1998; Massman and Grantz, 1995). Massman et al. (2000) developed a 

15 conceptual model of a dose-based index to determine how plant injury response to O3

16 relates to the traditional exposure-based parameters. The index used time-varying17 weighted fluxes to account for the fact that flux was not necessarily correlated with plant 

18 injury or damage. The model applied only to plant foliar injury and suggested that 

19 application of flux-based models for determining plant damage (yield or biomass) would 

20 require a better understanding and quantification of the relationship between injury and 

21 damage.

9.5.5 Summary

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

23 damage (i.e., reduced growth). They are summary measures of monitored ambient O3

24 concentrations over time intended to provide a consistent metric for reviewing and 

25 comparing exposure-response effects obtained from various studies. No new information 

26 is available since 2006 that alters the basic conclusions put forth in the 2006 and 1996 O3

27 AQCDs. These AQCDs focused on the research used to develop various exposure indices 

28 to help quantify effects on growth and yield in crops, perennials, and trees (primarily 

29 seedlings). The performance of indices was compared through regression analyses of 

30 earlier studies designed to support the estimation of predictive O3 exposure-response 

31 models for growth and/or yield of crops and tree (seedling) species.

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

33 based on determining the O3 concentration from the atmosphere that enters the leaf (i.e., 

34 flux or deposition). Interest has been increasing in recent years, particularly in Europe, in 
Draft – Do Not Cite or Quote 9-119 September 2011

1 using mathematically tractable flux models for O3 assessments at the regional, national, 

2 and European scale (Matyssek et al., 2008; Paoletti and Manning, 2007; M and M, 2004; 

3 Emberson et al., 2000b; Emberson et al., 2000a). While some efforts have been made in 

4 the U.S. to calculate O3 flux into leaves and canopies (Turnipseed et al., 2009; Uddling et 

5 al., 2009; Bergweiler et al., 2008; Hogg et al., 2007; Grulke et al., 2004; Grantz et al., 

6 1997; Grantz et al., 1995), little information has been published relating these fluxes to 

7 effects on vegetation. There is also concern that not all O3 stomatal uptake results in a

8 yield reduction, which depends to some degree on the amount of internal detoxification 

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

10 detoxification potential may, in fact, show little relationship between O3 stomatal uptake 

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

12 lack of understanding of detoxification processes have made this technique less viable for 

13 vulnerability and risk assessments in the U.S. 

14 The main conclusions from the 1996 and 2006 O3 AQCDs regarding indices based on 

15 ambient exposure are still valid. These key conclusions can be restated as follows:

16  O3 effects in plants are cumulative;

17  higher O3 concentrations appear to be more important than lower 

18 concentrations in eliciting a response;

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

20 and

21  exposure indices that cumulate hourly O3 concentrations and preferentially 

22 weight the higher concentrations have better statistical fits to growth/yield 

23 response data than do the mean and peak indices.

24 Various weighting functions have been used, including threshold-weighted (e.g., 

25 SUM06) and continuous sigmoid-weighted (e.g., W126) functions. Based on statistical 

26 goodness-of-fit tests, these cumulative, concentration-weighted indices could not be 

27 differentiated from one another using data from previous exposure studies. Additional 

28 statistical forms for O3 exposure indices have been discussed in Lee et al. (1988b). The 

29 majority of studies published since the 2006 O3 AQCD do not change earlier 

30 conclusions, including the importance of peak concentrations, and the duration and 

31 occurrence of O3 exposures in altering plant growth and yield.

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

33 cumulate and differentially weight the higher hourly average concentrations and also 

34 include the mid-level values continue to offer the most defensible approach for use in 

35 developing response functions and comparing studies, as well as for defining future 

36 indices for vegetation protection.
Draft – Do Not Cite or Quote 9-120 September 2011

9.6 Ozone Exposure-Plant Response Relationships

9.6.1 Introduction

1 The adequate characterization of the effects of O3 on plants for the purpose of setting air 

2 quality standards is contingent not only on the choice of the index used (i.e. SUM06, 

3 W126) to summarize O3 concentrations (Section 9.5), but also on quantifying the 

4 response of the plant variables of interest at specific values of the selected index. The 

5 many factors that determine the response of plants to O3 exposure have been discussed in 

6 previous sections. They include species, genotype and other genetic characteristics 

7 (Section 9.3), biochemical and physiological status (Section 9.3), previous and current

8 exposure to other stressors (Section 9.4.8), and characteristics of the exposure itself 

9 (Section 9.5). Establishing a secondary air quality standard requires the capability to 

10 generalize those observations, in order to obtain predictions that are reliable enough 

11 under a broad variety of conditions, taking into account these factors. This section 

12 reviews results that have related specific quantitative observations of O3 exposure with 

13 quantitative observations of plant responses, and the predictions of responses that have 

14 been derived from those observations through empirical models. 

15 For four decades, exposure to O3 at ambient concentrations found in many areas of the 

16 U.S. has been known to cause detrimental effects in plants (U.S. EPA, 2006b, 1996b, 

17 1984, 1978a). Results published after the 2006 O3 AQCD continue to support this 

18 finding, and the following sections deal with the quantitative characterizations of the 

19 relationship, and what new insights may have appeared since 2006. Detrimental effects 

20 on plants include visible injury, decreases in the rate of photosynthesis, reduced growth, 

21 and reduced yield of marketable plant parts. Most published exposure-response data have 

22 reported O3 effects on the yield of crops and the growth of tree seedlings, and those two 

23 variables have been the focus of the characterization of ecological impacts of O3 for the 

24 purpose of setting secondary air quality standards. In order to support quantitative 

25 modeling of exposure-response relationships, data should preferably include more than 

26 three levels of exposure, and some control of potential confounding or interacting factors 

27 should be present in order to model the relationship with sufficient accuracy. Letting 

28 potential confounders, such as other stressors, vary freely when generating O3 exposure29 response data might improve the ‘realism’ of the data, but it also greatly increases the 

30 amount of data necessary to extract a clear quantitative description of the relationship. 

31 Conversely however, experimental settings should not be so exhaustively restrictive as to 

32 make generalization outside of them problematic. During the last four decades, many of 

33 the studies of the effects of O3 on growth and yield of plants have not included enough 

34 levels of O3 to parameterize more than the simplest linear model. The majority of these 
Draft – Do Not Cite or Quote 9-121 September 2011

1 studies have only contrasted two levels, ambient and elevated, or sometimes three by 

2 adding carbon filtration in OTC studies, with little or no consideration of quantitatively 

3 relating specific values of exposure to specific values of growth or yield. This is not to 

4 say that studies that did not include more than two or three levels of O3 exposure, or 

5 studies that were conducted in uncontrolled environments, do not provide exposure6 response information that is highly relevant to reviewing air quality standards. In fact, 

7 they can be essential in verifying the agreement between predictions obtained through the 

8 empirical models derived from experiments such as NCLAN, and observations. The 

9 consensus of model predictions and observations from a variety of studies conducted in 

10 other locations, at other times, and using different exposure methods, greatly increases 

11 confidence in the reliability of both. Furthermore, if they are considered in the aggregate, 

12 studies with few levels of exposure or high unaccounted variability can provide 

13 additional independent estimates of decrements in plant growth and yield, at least within 

14 a few broad categories of exposure. 

15 Extensive exposure-response information on a wide variety of plant species has been 

16 produced by two long-term projects that were designed with the explicit aim of obtaining 

17 quantitative characterizations of the response of such an assortment of crop plants and 

18 tree seedlings to O3 under North American conditions: the NCLAN project for crops, and 

19 the EPA National Health and Environmental Effects Research Laboratory, Western 

20 Ecology Division tree seedling project (NHEERL/WED). The NCLAN project was 

21 initiated by the EPA in 1980 primarily to improve estimates of yield loss under field 

22 conditions and to estimate the magnitude of crop losses caused by O3 throughout the U.S. 

23 (Heck et al., 1991; Heck et al., 1982). The cultural conditions used in the NCLAN studies 

24 approximated typical agronomic practices, and the primary objectives were: (1) to define 

25 relationships between yields of major agricultural crops and O3 exposure as required to 

26 provide data necessary for economic assessments and development of O3 NAAQS; (2) to 

27 assess the national economic consequences resulting from O3 exposure of major 

28 agricultural crops; and (3) to advance understanding of cause-and-effect relationships that 

29 determine crop responses to pollutant exposures.

30 NCLAN experiments yielded 54 exposure-response curves for 12 crop species, some of 

31 which were represented by multiple cultivars at several of 6 locations throughout the U.S. 

32 The NHEERL/WED project was initiated by EPA in 1988 with the same objectives for 

33 tree species, and yielded 49 exposure-responses curves for multiple genotypes of 11 tree 

34 species grown for up to three years in Oregon, Michigan, and the Great Smoky 

35 Mountains National Park. Both projects used OTCs to expose plants to three to five 

36 levels of O3. Eight of the 54 crop datasets were from plants grown under a combination 

37 of O3 exposure and experimental drought conditions. Figure 9-14 through 9-17 

38 summarize some of the NCLAN and NHEERL/WED results.
Draft – Do Not Cite or Quote 9-122 September 2011

1 It should be noted that data from FACE experiments might also be used for modeling 

2 exposure-response. They only use two levels of O3 (ambient concentration at the site and 

3 a multiple of it), but given that the value of both levels of exposure changes every year, 

4 and that they are typically run for many consecutive years, aggregating data over time 

5 produces twice as many levels of O3 as there are years. As described in Section 9.2.4, 

6 FACE experiments seek to impose fewer constraints on the growth environment then 

7 OTCs. As a consequence, FACE studies have to contend with larger variability, 

8 especially year-to-year variability, but the difference in experimental conditions between 

9 the two methodologies makes comparisons between their results especially useful.

10 Growth and yield of at least one crop (soybean) has been investigated in yearly 

11 experiments since 2001 at a FACE facility in Illinois (University of Illinois, 2010; 

12 Morgan et al., 2006), however almost all analyses of SoyFACE published so far have 

13 been based on subsets of one or two years, and have only contrasted ambient versus 

14 elevated O3 as categorical variables. They have not modeled the response of growth and 

15 yield to O3 exposure continuously over the range of exposure values that have occurred 

16 over time. The only exception is a study by Betzelberger et al. (2010), who used a linear 

17 regression model on data pooled over 2 years. Likewise, trees of three species (trembling 

18 aspen, paper birch, and sugar maple) were grown between 1998 and 2009 in a FACE 

19 experiment located in Rhinelander, Wisconsin (Pregitzer et al., 2008; Dickson et al., 

20 2000). The Aspen FACE experiment has provided extensive data on responses of trees 

21 beyond the seedling stage under long-term exposure, and also on ecosystem-level 

22 responses (Section 9.4), but the only attempt to use those data in a continuous model of 

23 the response of tree growth to O3 exposure (Percy et al., 2007) suffered severe 

24 methodological problems, some of which are discussed in Section 9.6.3. Finally, one 

25 experiment was able to exploit a naturally occurring gradient of O3 concentrations to fit a 

26 linear regression model to the growth of cottonwood (Gregg et al., 2006, 2003). Factors 

27 such as genotype, soil type and soil moisture were under experimental control, and the 

28 authors were able to partition out the effects of potential confounders such as 

29 temperature, atmospheric N deposition, and ambient CO2.

30 A serious difficulty in assessing results of exposure-response research is the multiplicity 

31 of O3 metrics that have been used in reporting. As described in Section 9.5, metrics that 

32 entail either weighting or thresholding of hourly values cannot be converted into one 

33 another, or into unweighted metrics such as hourly average. When computing O3

34 exposure using weighted or thresholded metrics, the computation of each metric has to 

35 start with the original hourly data. Comparisons of exposure-response models can only be 

36 made between studies that used the same metric, and the value of exposure at which a 

37 given plant response is expected using one metric of exposure cannot be exactly 

38 converted to another metric. Determining the exposure value at which an effect would be 
Draft – Do Not Cite or Quote 9-123 September 2011

1 observed in a different metric can only be accomplished by first computing the 

2 experimental exposures in this metric from the hourly data, then estimating (fitting) 

3 model coefficients again. This problem is irremediable, although useful comparisons 

4 might be made using categorical exposures such as ‘current ambient exposure’ or ‘2050 

5 projected exposure’, which can serve as a common reference for quantitative values 

6 expressed in various metrics. Studies that contained growth or yield exposure-response 

7 data at few levels of exposure, and/or using metrics other than W126 are summarized in 

8 Tables 9-18 and 9-19.

9.6.2 Estimates of Crop Yield Loss and Tree Seedling Biomass Loss in the 

1996 and 2006 Ozone AQCDs

9 The 1996 and 2006 O3 AQCDs relied extensively on analyses of NCLAN and 

10 NHEERL/WED by Lee et al. (1994; 1989, 1988b, 1987), Hogsett et al. (1997), Lee and 

11 Hogsett (1999), Heck et al. (1984), Rawlings and Cure (1985), Lesser et al. (1990), and 

12 Gumpertz and Rawlings (1992). Those analyses concluded that a three-parameter 

13 Weibull model –

β

η α







 







−

=

W126

Y e

Equation 9-2 

14 is the most appropriate model for the response of absolute yield and growth to O3

15 exposure, because of the interpretability of its parameters, its flexibility (given the small 

16 number of parameters), and its tractability for estimation. In addition, removing the 

17 intercept α results in a model of relative yield (yield relative to [yield at exposure=0]) 

18 without any further reparameterization. Formulating the model in terms of relative yield 

19 or relative yield loss (yield loss=[1 − relative yield]) is essential in comparing exposure20 response across species, genotypes, or experiments for which absolute values of the 

21 response may vary greatly. In the 1996 and 2006 O3 AQCDs, the two-parameter model 

22 of relative yield was used in deriving common models for multiple species, multiple 

23 genotypes within species, and multiple locations.

24 Given the disparate species, genotypes, and locations that were included in the NCLAN 

25 and NHEERL/WED projects, and in the absence of plausible distributional assumptions 

26 with respect to those variables, a three step process using robust methods was used to 

27 obtain parameter estimates that could be generalized. The models that were derived for 

28 each species or group of species were referred to as median composite functions. In the 

29 first step, the three parameters of the Weibull model were computed for absolute yield or 
Draft – Do Not Cite or Quote 9-124 September 2011

1 biomass data from each NCLAN and NHEERL/WED experiment (54 crop datasets and 

2 49 tree seedling datasets), using nonlinear regression. When data were only available for 

3 three levels of exposure because of experimental problems, the shape parameter β was 

4 constrained to 1, reducing the model to an exponential decay model. In the second step, α 

5 was dropped, and predicted values of relative yield or biomass were then computed for 

6 12-hr W126 exposures between 0 and 60 ppm-h. At each of these W126 exposure values, 

7 the 25th, 50th, and 75th percentiles of the response were identified among the predicted 

8 curves of relative response. For example, for the 34 NCLAN studies of 12 crop species 

9 grown under non-droughted conditions for a complete cropping cycle (Figure 9-14), the 3 

10 quartiles of the response were identified at every integer value of W126 between 0 and 

11 60. The third step fitted a two-parameter Weibull model to those percentiles, yielding the 

12 median composite function for the relative yield or biomass response to O3 exposure for 

13 each grouping of interest (e.g., all crops, all trees, all datasets for one species), as well as 

14 composite functions for the other quartiles. In the 1996 and 2006 O3 AQCDs this 

15 modeling of crop yield loss and tree seedling biomass loss was conducted using the 

16 SUM06 metric for exposure. This section updates those results by using the 12-hr W126 

17 as proposed in 2007 (72 FR 37818) and 2010 (75 FR 2938, p. 3003). Figures 9-14 

18 through 9-17 present quantiles of predicted relative yield or biomass loss at seven values 

19 of the 12-h W126 for some representative groupings of NCLAN and NHEERL/WED 

20 results. Tables 9-10 through 9-12 give the 90-day 12-h W126 O3 exposure values at 

21 which 10 and 20% yield or biomass losses are predicted in 50 and 75% of crop or tree 

22 species using the composite functions.
Draft – Do Not Cite or Quote 9-125 September 2011

Source of Weibull parameters: Lee and Hogsett (1996).

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.

Figure 9-14 Quantiles of predicted relative yield loss for 34 NCLAN crop 

experiments.

10th Pctile

25th Pctile

50th Pctile

75th Pctile

90th Pctile

0

10

20

30

40

50

60

70

80

90

100

0 10 20 30 40 50 60

Percent Yield Loss (%)

12 hr W126 (ppm-hr)

34 crop datasets
Draft – Do Not Cite or Quote 9-126 September 2011

Source of Weibull parameters: Lee and Hogsett (1996).

Figure 9-15 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 well-watered 

conditions for the full duration of 1 cropping cycle. The number of 

studies available for each species is indicated on each plot. 

25th Pctile

50th Pctile

75th Pctile

0

10

20

30

40

50

60

70

80

90

100

0 10 20 30 40 50 60

7 Wheat datasets

25th Pctile

50th Pctile

75th Pctile

0

10

20

30

40

50

60

70

80

90

100

0 10 20 30 40 50 60

Percent Yield Loss (%)

11 Soybean datasets

50th Pctile

0

10

20

30

40

50

60

70

80

90

100

0 10 20 30 40 50 60

12 hr W126 (ppm-hr)

2 Corn datasets

25th Pctile

50th Pctile

75th Pctile

0

10

20

30

40

50

60

70

80

90

100

0 10 20 30 40 50 60

Percent Yield Loss (%)

12 hr W126 (ppm-hr

5 Cotton datasets
Draft – Do Not Cite or Quote 9-127 September 2011

Source of Weibull parameters: Lee and Hogsett (1996).

Figure 9-16 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. 

10th Pctile

25th Pctile

50th Pctile

75th Pctile

90th Pctile

0

10

20

30

40

50

60

70

80

90

100

0 10 20 30 40 50 60

Percent Biomass Loss (%)

90 day 12 hr W126 (ppm-hr)

49 tree seedlings datasets
Draft – Do Not Cite or Quote 9-128 September 2011

Source of Weibull parameters: Lee and Hogsett (1996).

Figure 9-17 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. 

25th Pctile

50th Pctile

75th Pctile

0

10

20

30

40

50

60

70

80

90

100

0 10 20 30 40 50 60

Percent Biomass Loss (%)

14 Aspen datasets

25th Pctile

50th Pctile

75th Pctile

0

10

20

30

40

50

60

70

80

90

100

0 10 20 30 40 50 60

11 Ponderosa pine datasets

50th Pctile

0

10

20

30

40

50

60

70

80

90

100

0 10 20 30 40 50 60

Percent Biomass Loss (%)

90 day 12 hr W126 (ppm-hr)

7 Douglas fir datasets

50th Pctile

0

10

20

30

40

50

60

70

80

90

100

0 10 20 30 40 50 60

90 day 12 hr W126 (ppm-hr)

3 Tulip poplar datasets
Draft – Do Not Cite or Quote 9-129 September 2011

Table 9-9 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)

Table 9-10 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)
Draft – Do Not Cite or Quote 9-130 September 2011

Table 9-11 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 (ppm-h)

90 day 12 h W126 for 20% yield 

loss (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)

9.6.3 Validation of 1996 and 2006 Ozone AQCD Models and Methodology 

Using the 90 day 12-h W126 and Current FACE Data 

1 Since the completion of the NCLAN and NHEERL/WED projects, almost no studies 

2 have been published that could provide a basis for estimates of exposure-response that 

3 can be compared to those of the 1996 and 2006 O3 AQCDs. Most experiments, 

4 regardless of exposure methodology, include only two levels of exposure. In addition, 

5 very few studies have included measurements of exposure using the W126 metric, or the 

6 hourly O3 concentration data that would allow computing exposure using the W126. Two 

7 FACE projects, however, were conducted over multiple years, and by adding to the 

8 number of exposure levels over time, may support independent model estimation and 

9 prediction using the same model and the same robust process as summarized in Section 

10 9.6.2. Hourly O3 data were available from both FACE projects. 

11 The SoyFACE project is situated near Champaign, IL, and comprises 32 octagonal rings 

12 (20m-diameter), 4 of which in a given year are exposed to ambient conditions, and 4 of 

13 which are exposed to elevated O3 as a fixed proportion of the instantaneous ambient 

14 concentration (Betzelberger et al., 2010; University of Illinois, 2010; Morgan et al., 2006; 

15 Morgan et al., 2004). Since 2002, yield data have been collected for up to 8 genotypes of 

16 soybean grown in subplots within each ring. The Aspen FACE project is situated in 

17 Rhinelander, WI, and comprises 12 rings (30m-diameter), 3 of which are exposed to 

18 ambient conditions, and 3 of which are exposed to O3 as a fixed proportion of the 

19 instantaneous ambient concentration (Pregitzer et al., 2008; Karnosky et al., 2005; 

20 Dickson et al., 2000). In the summer of 1997, half the area of each ring was planted with 

21 small (five to seven leaf sized) clonally propagated plants of five genotypes of trembling 
Draft – Do Not Cite or Quote 9-131 September 2011

1 aspen, which were left to grow in those environments until 2009. Biomass data are 

2 currently available for the years 1997-2005 (King et al., 2005). Ozone exposure in these 

3 two FACE projects can be viewed as a categorical variable with two levels: ambient, and 

4 elevated. However, this overlooks the facts that yearly ambient and elevated exposure 

5 both vary with every year, and that the proportionality between them also changes. This 

6 change has two sources: first, the dispensing of O3 into the elevated exposure rings varies 

7 from the proportionality set point to some extent, and for SoyFACE, the set point 

8 changed between years. Second, the proportionality does not propagate predictably from 

9 the hourly data to the yearly value when using threshold or concentration-weighted 

10 cumulative metrics (such as AOT40, SUM06 or W126). Hourly average elevated 

11 exposures that are, for example, a constant 1.5 times greater than ambient do not result in 

12 AOT40, SUM06 or W126 values that are some constant multiple of the ambient values of 

13 those indices. The greater the fraction of elevated hourly values that are above the 

14 threshold or heavily weighted, compared to the fraction of hourly ambient values that are, 

15 the greater the difference between ambient and elevated yearly exposure, as measured 

16 using weighted cumulative indices. When elevated exposure is a multiple of ambient 

17 hourly intervals, the number of hours for which elevated exposure meets the threshold for 

18 inclusion can vary widely, even though the hourly mean for the year retains the 

19 proportionality. As a consequence, the number of exposure levels in multi-year 

20 experiments is twice the number of years. In the case of SoyFACE for the period between 

21 2002 and 2008, ambient exposure in the highest year was approximately equal to elevated 

22 exposure in the lowest year, with 14 levels of O3 exposure evenly distributed from lowest 

23 to highest. The particular conditions of the Aspen FACE experiment resulted in 12 

24 exposure levels between 1998 and 2003, but they were not as evenly distributed between 

25 minimum and maximum over the 6-year period.

26 There are necessary differences in the modeling of exposure-response in annual plants 

27 such as soybean, and in perennial plants such as aspen trees, when exposure takes place 

28 over multiple years. In annual plants, responses recorded at the end of the life cycle, i.e., 

29 yearly, are analyzed in relationship to that year’s exposure. Yield of soybeans is affected 

30 by exposure during the year the crop was growing, and a new crop is planted every year. 

31 Thus an exposure-response relationship can be modeled from yearly responses matched 

32 to yearly exposures, with those exposure-response data points having been generated in 

33 separate years. For perennial organisms, which are not harvested yearly and continue to 

34 grow from year to year, such pairing of exposure and response cannot be done without 

35 accounting for time. Not only does the size of the organism at the beginning of each year 

36 of exposure increase, but size is also dependent on the exposure from previous years. 

37 Therefore the relationship of response and exposure must be analyzed either one year at a 

38 time, or by standardizing the response as a yearly increment relative to size at the 

39 beginning of each year. Furthermore, the relevant measurement of exposure is 
Draft – Do Not Cite or Quote 9-132 September 2011

1 cumulative, or cumulative yearly average exposure, starting in the year exposure was 

2 initiated, up to the end of the year of interest. When analyzing the growth of trees over 

3 several years, it would be evidently incorrect to pair the exposure level in every discrete 

4 year with absolute size of the trees that year, and posit a direct relationship between them, 

5 without taking increasing age into consideration. In the Aspen FACE experiment, for 

6 example, one could not establish an exposure-response relationship by matching 

7 12 yearly exposures and 12 yearly tree sizes, while disregarding age as if size did not also 

8 depend on it. This is the basis of the 2007 study of Aspen FACE data by Percy et al. 

9 (2007), which compares the size of trees of various ages as if they were all the same age, 

10 and was therefore not informative.

9.6.3.1 Comparison of NCLAN-Based Prediction and SoyFACE 

Data 

11 For this ISA, EPA conducted a comparison between yield of soybean as predicted by the 

12 composite function three-step process (Section 9.6.2) using NCLAN data, and 

13 observations of yield in SoyFACE. The median composite function for relative yield was 

14 derived for the 11 NCLAN soybean Weibull functions for non-droughted studies, and 

15 comparisons between the predictions of the median composite and SoyFACE 

16 observations were conducted as follows. 

17 For the years 2007 and 2008, SoyFACE yield data were available for 7 and 6 genotypes, 

18 respectively. The EPA used those data to compare the relative change in yield observed 

19 in SoyFACE in a given year between ambient O3 and elevated O3, versus the relative 

20 change in yield predicted by the NCLAN-based median composite function between 

21 those same two values of O3 exposure. The two parameter median composite function for 

22 relative yield of soybean based on NCLAN data was used to predict yield response at the 

23 two values of exposure observed in SoyFACE in each year, and the change between yield 

24 under ambient and elevated was compared to the change observed in SoyFACE for the 

25 relevant year (Table 9-12). This approach results in a direct comparison of predicted 

26 versus observed change in yield. Because the value of relative response between any two 

27 values of O3 exposure is independent of the intercept α, this comparison does not require 

28 prediction of the absolute values of the responses.

29 Since comparisons of absolute values might be of interest, the predictive functions were 

30 also scaled to the observed data: SoyFACE data were used to compute an intercept α 

31 while the shape and scale parameters (β and η) were held at their value in the NCLAN 

32 predictive model. This method gives a comparison of prediction and observation that 

33 takes all the observed information into account to provide the best possible estimate of 
Draft – Do Not Cite or Quote 9-133 September 2011

1 the intercept, and thus the best possible scaling (Table 9-13 and Figure 9-18). For the 

2 comparison of NCLAN and SoyFACE, this validation was possible for 2007 and 2008, 

3 where data for 7 and 6 soybean genotypes, respectively, were available. The median 

4 composite function for relative yield was derived for the 11 NCLAN soybean Weibull 

5 functions for nondroughted studies, and the values of median yield under ambient 

6 exposure at SoyFACE in 2007 and 2008 were used to obtain an estimate of the intercept 

7 α for the NCLAN median function in each of the two years.

8 Table 9-12 presents the results of ambient/elevated relative yield comparisons between 

9 the NCLAN-derived predictions and SoyFACE observations. Table 9-13 and figure 9-18 

10 present the results of comparisons between NCLAN-derived predictions and SoyFACE 

11 observations of yield, with the predictive function scaled to provide absolute yield values. 

Table 9-12 Comparison between change in yield observed in the SoyFACE 

experiment between elevated and ambient ozone, and change 

predicted at the same values of ozone by the median composite 

function for NCLAN (two-parameter relative yield model)

Year 90-day 12-h W126 (ppm-h) observed at SoyFACE

Yield in

Elevated O3 Relative to Ambient O3 (%)

Ambient Elevated Predicted by NCLAN Observed at SoyFACE

2007 4.39 46.23 75 76

2008 3.23 28.79 85 88

Table 9-13 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 with intercept scaled to SoyFACE data)

Year 90-day 12-h W126 (ppm-h) 

observed at SoyFACE Yield predicted by NCLAN (g/m2

) Yield observed at SoyFACE 

(g/m2)

Ambient Elevated Ambient Elevated Ambient Elevated

2007 4.39 46.23 309.2 230.6 305.2 230.6

2008 3.23 28.79 350.3 298.2 344.8 304.4
Draft – Do Not Cite or Quote 9-134 September 2011

Source of data: Betzelberger et al. (2010); Morgan et al. (2006); Lee and Hogsett (1996).

Note: Black dots are the 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 9-18 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, and compared to the NCLAN-based function. The process 

3 described in Section 9.6.2 was applied to SoyFACE data for individual genotypes, 

4 aggregated over the years during which each was grown; one genotype from 2003 to 

5 2007, and six genotypes in 2007 and 2008. First, the three parameter Weibull model 

6 described in Section 9.6.2 was estimated using nonlinear regression on exposure-yield 

7 data for each genotype separately, over the years for which data were available, totaling 

8 seven curves. The 25th, 50th, and 75th percentiles of the predicted values for the two 

9 parameter relative yield curves were then identified at every integer of W126 between 0 

10 and 60, and a two-parameter Weibull model estimated by regression for the three 

11 quartiles. The comparison between these composite functions for the quartiles of relative 

12 yield loss in SoyFACE and the corresponding composite functions for NCLAN is 

13 presented in Figure 9-19.

0

50

100

150

200

250

300

350

400

0 10 20 30 40 50 60 70

Yield (g/m2)

90 day 12 hr W126 (ppm-hr)

2007, 7 genotypes

0

50

100

150

200

250

300

350

400

0 10 20 30 40 50 60 70

Yield (g/m2)

90 day 12 hr W126 (ppm-hr)

2008, 6 genotypes
Draft – Do Not Cite or Quote 9-135 September 2011

Source of data: Betzelberger et al. (2010);Morgan et al. (2006); Lee and Hogsett (1996).

Figure 9-19 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 9-13 and 9-14, and in Figure 9-18, the agreement between predictions 

2 based on NCLAN data and SoyFACE observations was notably close in single-year 

3 comparisons. Together with the very high agreement between median composite models 

4 for NCLAN and SoyFACE (Figure 9-19), it provides very strong mutual confirmation of 

5 those two projects’ results with respect to the response of yield of soybeans to O3

6 exposure. It is readily apparent from these results that the methodology described in 

7 Section 9.6.2 for obtaining predictions of yield or yield loss from NCLAN data is 

8 strongly validated by SoyFACE results. As described in Section 9.2, the exposure 

9 technologies used in the two projects were in sharp contrast, specifically with respect to 

10 the balance each achieved between control of potential interacting factors or confounders, 

11 and fidelity to natural conditions. The comparisons that EPA conducted therefore 

12 demonstrate that the methodology used in developing the composite functions is resistant 

13 to the influence of nuisance variables and that predictions are reliable. They may also 

14 suggest that the aspects in which the two exposure technologies differ have less influence 

15 on exposure-response than initially supposed. These results are also in agreement with 

16 comparative studies reviewed in 9.2.6. 

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

0 10 20 30 40 50 60 70

Percent Yield Loss (%)

90 day 12 hr W126 (ppm-hr)
Draft – Do Not Cite or Quote 9-136 September 2011

9.6.3.2 Comparison of NHEERL/WED-Based Prediction of Tree 

Biomass Response and Aspen FACE Data

1 EPA also conducted two comparisons between prediction of above-ground biomass loss 

2 based on NHEERL/WED results and observations from Aspen FACE. The median 

3 composite function was developed from NHEERL/WED data for 11 studies that used 

4 wild-type seedlings of aspen as well as four clonally propagated genotypes. All plants 

5 were grown in OTCs for one growing season before being destructively harvested. Aspen 

6 FACE data were from clonally propagated trees of five genotypes grown from 1998 to 

7 2003, with above-ground biomass calculated using allometric equations derived from 

8 data for trees harvested destructively in 2000 and 2002 (King et al., 2005). 

9 The two parameter median composite function for relative biomass was used to predict 

10 biomass response under the observed elevated exposure, relative to its value under 

11 observed ambient exposure, for each separate year of Aspen FACE. EPA first compared 

12 Aspen FACE observations of the change in biomass between ambient and elevated 

13 exposure with the corresponding prediction at the same values of exposure. Comparisons 

14 between observed and predicted absolute biomass values were then conducted for each 

15 year by scaling the predictive function to yearly Aspen FACE data as described for 

16 soybean data in Section 9.6.3.1. In all cases, yearly 90 day 12-hour W126 values for 

17 Aspen FACE were computed as the cumulative average from the year of planting up to 

18 the year of interest. A comparison of composite functions between NHEERL/WED and 

19 Aspen FACE, similar to the one performed for NCLAN and SoyFACE, was not possible: 

20 as discussed in the introduction to Section 9.6, the pairing of 12 exposure values from 

21 separate years and 12 values of biomass cannot be the basis for a model of exposure22 response, because the trees continued growing for the six-year period of exposure. 

23 Because the same trees were used for the entire duration, and continued to grow, data 

24 could not be aggregated over years. Table 9-14 presents the results of ambient/elevated 

25 relative biomass comparisons between the NHEERL/WED-derived predictions and 

26 Aspen FACE observations. Table 9-15 and Figure 9-20 present the results of the 

27 comparison between NHEERL/WED-derived predictions and Aspen FACE observations 

28 for absolute biomass, using Aspen FACE data to scale the NHEERL/WED-derived 

29 composite function.
Draft – Do Not Cite or Quote 9-137 September 2011

Table 9-14 Comparison between change in above-ground biomass elevated 

and ambient ozone in Aspen FACE experiment in 6 year, and 

change predicted at the same values of ozone by the median 

composite function for NHEERL/WED (two-parameter relative 

biomass model)

Table 9-15 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 with intercept scaled to Aspen FACE data)

Year

90 day 12-h W126 (ppm-h)

Cumulative Average observed at Aspen FACE

Biomass Predicted by 

NHEERL/WED (g/m2

) Biomass Observed at Aspen FACE (g/m2)

Ambient Elevated Ambient Elevated Ambient Elevated

1998 3.19 30.08 276.0 203.2 274.7 204.9

1999 2.61 33.85 958.7 668.3 955.3 673.3

2000 2.43 30.16 1382.4 1022.8 1400.3 998.6

2001 2.55 31.00 1607.0 1173.7 1620.7 1154.9

2002 2.51 30.27 2079.0 1532.1 2125.9 1468.41

2003 2.86 29.12 2640.1 1981.2 2695.2 1907.8

Year 90-day 12-h W126 (ppm-h)

Cumulative Average observed at Aspen FACE

Above-Ground Biomass in 

Elevated O3 relative To Ambient O3 (%)

Ambient Elevated Predicted by NHEERL/WED Observed at Aspen FACE

1998 3.19 30.08 74 75

1999 2.61 33.85 70 70

2000 2.43 30.16 74 71

2001 2.55 31.00 73 71

2002 2.51 30.27 74 69

2003 2.86 29.12 75 71
Draft – Do Not Cite or Quote 9-138 September 2011

Source of data: King et al. (2005), Lee and Hogsett (1996).

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 9-20 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. 

1 As in the comparisons between NCLAN and SoyFACE, the agreement between 

2 predictions based on NHEERL/WED data and Aspen FACE observations was very close. 

3 The results of the two projects strongly reinforce each other with respect to the response 

4 of aspen biomass to O3 exposure. The methodology used for obtaining the median 

5 composite function is shown to be capable of deriving a predictive model despite 

6 potential confounders, and despite the added measurement error that is expected from 

7 calculating biomass using allometric equations. In addition, the function based on 

8 one year of growth was shown to be applicable to subsequent years. 

9 The results of experiments that used different exposure methodologies, different 

10 genotypes, locations, and durations converged to the same values of response to O3

11 exposure for each of two very dissimilar plant species, and predictions based on the 

12 earlier experiments were validated by the data from current ones. However, in these 

13 comparisons, the process used in establishing predictive functions involved aggregating 

14 data over variables such as time, locations, and genotypes, and the use of a robust statistic 

1998

1999

2000

2001

2002

2003

0

500

1000

1500

2000

2500

3000

0 10 20 30 40 50 60 70

Biomass (g/m2)

90 day 12 hr W126 (yearly cumulative average, ppm-hr)
Draft – Do Not Cite or Quote 9-139 September 2011

1 (quartiles) for that aggregation. The validating data, from SoyFACE and Aspen FACE, 

2 were in turn aggregated over the same variables. The accuracy of predictions is not 

3 expected to be conserved for individual values of those variables over which aggregation 

4 occurred. For example, the predicted values for soybean, based on data for five 

5 genotypes, are not expected to be valid for each genotype separately. As shown in the 

6 validation, however, aggregation that occurred over different values of the same variable 

7 did not affect accuracy: composite functions based on one set of genotypes were 

8 predictive for another set, as long as medians were used for both sets. A study of 

9 cottonwood (Populus deltoides) conducted using a naturally occurring gradient of O3

10 exposure (Gregg et al., 2006, 2003) may provide an illustration of the response of an 

11 individual species whose response is far from the median response for an aggregation of 

12 species. 

9.6.3.3 Exposure-Response in a Gradient Study

13 Gregg et al. (2003) grew saplings of one clonally propagated genotype of cottonwood 

14 (Populus deltoides) in seven locations within New York City and in the surrounding 

15 region between July and September in 1992, 1993 and 1994, and harvested them 72 days 

16 after planting. Owing to regional gradients of atmospheric O3 concentration, the 

17 experiment yielded eight levels of exposure (Figure 9-21), and the authors were able to 

18 rule out environmental variables other than O3 to account for the large differences in 

19 biomass observed after one season of growth. The deficit in growth increased 

20 substantially faster with increasing O3 exposure than has been observed in aspen, another 

21 species of the same genus (Populus tremuloides, Section 9.6.3.2). Using a three 

22 parameter Weibull model (Figure 9-21), the biomass of cottonwood at a W126 exposure 

23 of 15 ppm-h, relative to biomass at 5 ppm-h, is estimated to be 0.18 (18% of growth at 

24 5 ppm-h). The relative biomass of trembling aspen within the same 5-15 ppm-h range of 

25 exposure is estimated to be 0.92, using the median composite model for aspen whose 

26 very close agreement with Aspen FACE data was shown in Section 9.6.3.2. Using a 

27 median composite function for all deciduous trees in the NHEERL/WED project (6 

28 species in 21 studies) also gives predictions that are very distant from the cottonwood 

29 response observed in this experiment. For all deciduous tree species in NHEERL/WED, 

30 biomass at a W126 exposure of 15 ppm-h, relative to biomass at 5 ppm-h, was estimated 

31 to be 0.87.
Draft – Do Not Cite or Quote 9-140 September 2011

Source: Modified with permission from Nature Publishing Group (Gregg et al., 2003).

Figure 9-21 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. 

1 These cottonwood data confirm that, as should be expected, some individual tree species 

2 are substantially more sensitive than the median of NHEERL/WED (Figure 9-16). As 

3 shown in Section 9.6.2, the median models available for trembling aspen and soybean 

4 have verifiable predictive ability for those particular species. This suggests that the 

5 corresponding NCLAN- and NHEERL/WED-based models for multiple crop and tree 

6 species can provide reliable estimates of losses for similar assortments of species. 

7 However, their predictive ability would likely be poor for individual species not tested. 

8 An alternative hypothesis for the difference between the response of cottonwood in this 

9 experiment and deciduous tree species in NHEERL/WED, or the difference between the 

10 response of cottonwood and aspen in NHEERL/WED and Aspen FACE, could be the 

11 presence of confounding factors in the environments where the experiment was 

12 conducted. However, variability in temperature, moisture, soil fertility, and atmospheric 

13 deposition of N were all ruled out by Gregg et al. (2003) as contributing to the observed 

14 response to O3. In addition, this hypothesis would imply that the unrecognized 

15 confounder(s) were either absent from both OTC and FACE studies, or had the same 

16 value in both. This is not impossible, but the hypothesis that cottonwood is very sensitive 

17 to O3 exposure is more parsimonious, and sufficient.

0

10

20

30

40

50

60

70

80

90

100

0 10 20 30 40 50 60 70

Biomass (g)

72 day 12 hr W126 (ppm-hr)
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9.6.3.4 Meta-analyses of growth and yield studies

1 Since the 2006 O3 AQCD, five studies have used meta-analytic methods to integrate 

2 results from experimental studies of crops or tree species relevant to the U.S. It is 

3 possible to obtain exposure-response data for growth and yield from those meta-analyses, 

4 but because all of them provided summary measurements of O3 exposure as hourly 

5 averages of various lengths of exposures, comparisons with exposure-response results 

6 where exposure is expressed as W126 are problematic. Table 9-16 summarizes the 

7 characteristics of the five meta-analyses. They all included studies conducted in the U.S. 

8 and other locations worldwide, and all of them expressed responses as comparative 

9 change between levels of exposure to O3, with carbon filtered air (CF) among those 

10 levels. Using hourly average concentration to summarize exposure, CF rarely equates 

11 absence of O3, although it almost always near zero when exposure is summarized as 

12 W126, SUM06, or AOT40. 

Table 9-16 Meta-analyses of growth or yield studies published since 2005

Study Number of articles 

included

Years of 

publication 

surveyed

Crop, species or 

genera Response 

Number 

of O3

levels

Duration of 

exposure 

Ainsworth (2008) 12 1980-2007 rice Yield 2 unreported

Feng et al. (2008b) 53 1980-2007 wheat Yield 5 > 10 days

Feng and 

Kobayashi (2009) All crops together : 81 1980-2007 Potato, barley, wheat, rice, bean, soybean Yield 3 > 10 days

Grantz et al. (2006) 16 1992-2004

34 herbaceous dicots

21 herbaceous monocots

5 tree species

Relative 

Growth Rate 2 2-24 weeks

Wittig et al. (2009)

All responses:263 

Articles that included 

biomass:unreported

1970-2006 4 gymnosperm tree genera

11 angiosperm tree genera

Total 

biomass 4 > 7 days

13 The only effect of O3 exposure on yield of rice reported in Ainsworth (2008) was a 

14 decrease of 14% with exposure increasing from CF to 62 ppb average concentration. 

15 Feng et al. (2008b) were able to separate exposure of wheat into four classes with average 

16 concentrations of 42, 69, 97, and 153 ppb, in data where O3 was the only treatment. Mean 

17 responses relative to CF were yield decreases of 17, 25, 49, and 61% respectively. Feng 

18 et al. (2008b) observed that wheat yield losses were smaller under conditions of drought, 

19 and that Spring wheat and Winter wheat appeared similarly affected. However, mean 

20 exposure in studies of Winter wheat was substantially higher than in studies of Spring 

21 wheat (86 versus 64 ppb), which suggests that the yield of Spring wheat was in fact more 

22 severely affected, since yield was approximately the same, even though Spring wheat was 

23 exposed to lower concentrations. Exposures of the six crops considered in Feng and 

24 Kobayashi (2009) were classified into two ranges, each compared to CF air. In the lower 
Draft – Do Not Cite or Quote 9-142 September 2011

1 range of exposure (41-49 ppb), potato studies had the highest average exposure (45 ppb) 

2 and wheat and rice the lowest (41 ppb). In the higher range (51-75 ppb), wheat studies 

3 had the highest average exposure (65 ppb), and potato, barley and rice the lowest (63 

4 ppb). In other words, across the studies included, all crops were exposed to very similar 

5 levels of O3. At approximately 42 ppb, the yield of potato, barley, wheat, rice, bean, and 

6 soybean declined by 5.3, 8.9, 9.7, 17.5, 19, and 7.7% respectively, relative to CF air. At 

7 approximately 64 ppb O3, declines were 11.9, 12.5, 21.1, 37.5, 41.4, and 21.6%. Grantz 

8 et al. (2006) reported Relative Growth Rate (RGR) rather than growth, and did not report 

9 O3 exposure values in a way that would allow calculation of mean exposure for each of 

10 the three categories of plants for which RGR changes are reported. All studies used only 

11 two levels of exposure, with CF air as the lower one, and most used elevated exposure in 

12 the range of 40 to 70 ppb. Decline in RGR was 8.2% for the 34 herbaceous dicots, 4.5% 

13 for the 21 herbaceous monocots, and 17.9% for the 5 tree species. Finally, Wittig et al. 

14 (2009) divided the studies analyzed into three classes of comparisons: CF versus ambient, 

15 CF versus elevated, and ambient versus elevated, but reported comparisons between three 

16 average levels of exposure besides CF: 40 ppb, 64 ppb, and 97 ppb. Corresponding 

17 decreases in total biomass relative to CF were 7, 17, and 17%. 

18 These meta-analyses provide very strong confirmation of EPA’s conclusions from 

19 previous O3 AQCDs: compared to lower levels of ambient O3, current levels in many 

20 locations are having a substantial detrimental effect on the growth and yield of a wide 

21 variety of crops and natural vegetation. They also confirm strongly that decreases in 

22 growth and yield continue at exposure levels higher than current ambient levels. 

23 However, direct comparisons with the predictions of exposure-response models that use 

24 concentration-weighted cumulative metrics are difficult. 

9.6.3.5 Additional exposure-response data 

25 The studies summarized in Tables 9-18 and 9-19 contain growth or yield exposure26 response data at too few levels of exposure for exposure-response models, and/or used 

27 metrics other than W126. These tables update Tables AX9-16 through AX9-19 of the 

28 2006 O3 AQCD.

9.6.4 Summary

29 None of the information on effects of O3 on vegetation published since the 2006 O3

30 AQCD has modified the assessment of quantitative exposure-response relationships that 

31 was presented in that document. This assessment updates the 2006 exposure-response 
Draft – Do Not Cite or Quote 9-143 September 2011

1 models by computing them using the W126 metric, cumulated over 90 days. Almost all 

2 of the experimental research on the effects of O3 on growth or yield of plants published 

3 since 2006 used only two levels of exposure. In addition, hourly O3 concentration data 

4 that would allow calculations of exposure using the W126 metric are generally 

5 unavailable. However, two long-term experiments, one with a crop species (soybean), 

6 one with a tree species (aspen), have produced data that can be used to validate the 

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

8 derive them. 

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

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

11 Network (NCLAN) and EPA National Health and Environmental Effects Research 

12 Laboratory, Western Ecology Division (NHEERL-WED) projects, using OTCs to expose 

13 crops and trees seedling to O3. In recent years, yield and growth results for two of the 

14 species that had provided extensive exposure-response information in those projects have 

15 become available from studies that used FACE technology, which is intended to provide 

16 conditions much closer to natural environments (Pregitzer et al., 2008; Morgan et al., 

17 2006; Morgan et al., 2004; Dickson et al., 2000). The robust methods that were used 

18 previously with exposure measured as SUM06 were applied to the NCLAN and 

19 NHEERL-WED data with exposure measured as W126, in order to derive single-species 

20 median models for soybean and aspen from studies involving different genotypes, years, 

21 and locations. The resulting models were used to predict the change in yield of soybean 

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

23 experiments. Results from these new experiments were exceptionally close to predictions 

24 from the models. The accuracy of model predictions for two widely different plant 

25 species provides support for the validity of the corresponding multiple-species models for 

26 crops and trees in the NCLAN and NHEERL-WED projects. However, variability among 

27 species in those projects indicates that the range of sensitivity is likely quite wide. This 

28 was confirmed by a recent experiment with cottonwood in a naturally occurring gradient 

29 of exposure (Gregg et al., 2006), which established the occurrence of species with 

30 responses substantially more severe under currently existing conditions than are predicted 

31 by the median model for multiple species.

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

33 yield of soybean, wheat, rice and other crops under broad categories of exposure, relative 

34 to charcoal-filtered air (Ainsworth, 2008; Feng et al., 2008b; Morgan et al., 2003). 

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

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

37 263 studies that reported effects on tree biomass. However, these analyses have proved 
Draft – Do Not Cite or Quote 9-144 September 2011

1 difficult to compare with exposure-response models, especially given that exposure was 

2 not expressed in the same W126 metric. 

Table 9-17 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)

Bean (Phaseolus 

vulgaris I. cv 

Borlotto)

OTC; groundplanted

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)

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)

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. 

(2009b)

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. 

(2007c)

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)

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)

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)

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)

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)

Peanut (Arachis 

hypogaea)

OTC

Raleigh, NC

3 yr

12-h avg: 

CF (22 ppb), 

Ambient (46 ppb), 

Elevated (75ppb)

(CO2: 375 ppm; 548 ppm; 

730 ppm)

Yield (seed weight, 

g/m) -33 (-8) Burkey et al. 

(2007)

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)
Draft – Do Not Cite or Quote 9-145 September 2011

Species

Facility

Location

Exposure 

Duration

O3 Exposure (Additional Treatment) Response 

Measured

percent change 

from CF

(percent change 

from ambient)

Reference

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)

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)

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)

Seminatural grass

FACE

Le Mouret, 

Switzerland

5 yr

Seasonal AOT40: Ambient 

(0.1-7.2 ppm-h); 

Elevated. (1.8-24.1 ppm-h)

(N/A)

Relative annual yield N/A (2×faster 

decrease in yield/yr)

Volk et al. 

(2006)

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) Bou Jaude

et al. (2008)

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)

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)

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. 

(2004a)

Soybean (Glycine 

max cv. Tracaja)

Chambers; pots

Brazil

20 days 12-h avg: CF & 30 ppb

(N/A) Biomass -18 (N/A) Bulbovas et al. (2007)

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)

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. 

(2008a)

Strawberry (Fragaria x ananassa Duch. 

Cv Korona & 

Elsanta)

Growth chambers

Bonn, Germany

2 months

8-h avg: CF (0 ppb) & 

Elevated (78 ppb)

(N/A)

Fruit yield 

(weight/plant) -16 (N/A) Keutgen et al. 

(2005)

Sugarbeet (Beta 

vulgaris cv. Patriot)

OTC

Belgium

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)

Sugarcane 

(Saccharum spp)

CSTR

San Joaquin Valley, 

CA

2007; 

11-13 wk.

12-h avg: CF (4 ppb); 

Ambient (58); 

Elevated (147)

(N/A)

Total biomass 

(g/plant) -40 (-30) Grantz and Vu 

(2009)
Draft – Do Not Cite or Quote 9-146 September 2011

Species

Facility

Location

Exposure 

Duration

O3 Exposure (Additional Treatment) Response 

Measured

percent change 

from CF

(percent change 

from ambient)

Reference

Sweet Potato

Growth chambers

Bonn, Germany

4 wk

8-h avg: CF (0 ppb), 

Ambient (<40 ppb) Elevated 

(255 ppb)

(N/A)

Tuber weight -14 (-11.5) Keutgen et al. 

(2008)

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)

Trifolium 

Subterraneum

OTC; 2.5-L pots

Madrid, Spain

29 days

12-h avg: CF (<7.9±6.3); 

Ambient (34.4±10.8); 

Elevated (56.4±22.3)

(N: 5, 15 & 30 kg/ha)

Above-ground 

biomass -45 (-35) Sanz et al. 

(2005)

Watermelon 

(Citrullus lanatus)

OTC

Valencia, Spain

2000, 2001. 

90 days

AOT40: CF (0 ppm-h)

Ambient (5.7 ppm-h), Elevated 

(34.1 ppm-h)

(N:0, 14.0 & 29.6 g/pot)

total fruit yield (kg) n.s. (54) Calatayud 

et al. (2006)

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)

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.

Table 9-18 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

Parlier, CA

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. (2010b)

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 

above-ground biomass 

(35 species) after 4 

years

CF (459 g/m2

), CF+ 

(457 g/m2

), CF2+ 

(398 g/m2

)

Pfleeger et al. (2010)

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)

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)

Red Oak (Quercus rubrum)

Forest sites

Look Rock & Twin Creeks 

Forests, TN

2001-2003, 

April-October

AOT60: 

2001 (11.5), 

2002 (24.0), 

2003 (11.7)

(Observational 

study with 

multiple 

environmental 

variables)

Annual circumference 

increment (change 

relative to 2001 in year 

2002;2003)

-42.8%; +1% McLaughlin et al. 

(2007a)
Draft – Do Not Cite or Quote 9-147 September 2011

Species

Facility

Location

Exposure 

Duration

O3 Exposure (Additional 

Treatment)

Response 

Measured Response Reference

Pine species

Forest sites

Look Rock Forest, TN

2001-2003, 

April-October

AOT60: 

2001 (11.5), 

2002 (24.0), 

2003 (11.7)

(Observational 

study with 

multiple 

environmental 

variables)

Annual circumference 

increment (change 

relative to 2001 in year 

2002;2003)

-62.5%; -2.9% McLaughlin et al. 

(2007a)

Hickory species

Forest sites

Look Rock Forest, TN

2001-2003, 

April-October

AOT60: 

2001 (11.5), 

2002 (24.0), 

2003 (11.7)

(Observational 

study with 

multiple 

environmental 

variables)

Annual circumference 

increment (change 

relative to 2001 in year 

2002;2003)

-14%; +30% McLaughlin et al. 

(2007a)

Chestnut Oak (Quercus 

prinus)

Forest sites

Look Rock Forest, TN

2001-2003, 

April-October

AOT60: 

2001 (11.5), 

2002 (24.0), 

2003 (11.7)

(Observational 

study with 

multiple 

environmental 

variables)

Annual circumference 

increment (change 

relative to 2001 in year 

2002;2003)

+44%; +55% McLaughlin et al. 

(2007a)

Black Cherry (Prunus rigida)

Forest sites

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 2003 in year 

2002)

-75% McLaughlin et al. 

(2007a)

Shortleaf pine (Pinus 

echinata)

Forest sites

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 2003 in year 

2002)

-16.8% McLaughlin et al. 

(2007a)

Hemlock (Tsuga 

canadensis)

Forest sites

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 2003 in year 

2002)

-21.9% McLaughlin et al. 

(2007a)

Red Maple (Acer rubrum)

Forest sites

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 2003 in year 

2002)

-59.6% McLaughlin et al. 

(2007a)

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. 

(2007a)
Draft – Do Not Cite or Quote 9-148 September 2011

Species

Facility

Location

Exposure 

Duration

O3 Exposure (Additional 

Treatment)

Response 

Measured Response Reference

Sugar Maple (Acer 

saccharum)

Forest sites

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 2003 in year 

2002)

-63.8% McLaughlin et al. 

(2007a)

Trembling aspen (Populus 

tremuloides), 5 genotypes

Aspen FACE

Rhinelander, WI

1998-2004, 

May-October

Cumulative avg 

90-day 12-h 

W126. 

Ambient 3.1 

ppm-h Elevated: 

27.2 ppm-h

(Competition with 

birch, maple)

main stem volume 

after 7 years

Ambient: 6.22 dm3; 

Elevated: 4.73 dm3 Kubiske et al. (2006)

Hybrid Poplar (Populus 

trichocarpa x Populus 

deltoides)

OTC

Seattle, WA

2003, 

3 months

Daily mean 

(μg/g): 

CF(<9),

Elevated (85-128)

Total biomass CF to elevated: 

-12.9%

Woo and Hinckley 

(2005)

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.

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