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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. This chapter of the ISA includes scientific research from 

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

12 of biological organization (e.g., organ, organism, population, community, ecosystem). 

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

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

15 update and 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 provide information on endpoints or processes 

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

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

27 Europe and Asia. This document includes discussion of studies of vegetation and 

28 ecosystems outside of North America only if those studies contribute to the general 

29 understanding of O3 effects across species and ecosystems. For example, studies outside 

30 North America are discussed that consider physiological and biochemical processes that 

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

2 ecosystem processes are discussed in the chapter. 

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

4 and ecosystems at various levels of biological organization and ends with policy-relevant 

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

6 illustrative diagram of the major pathway through which O3 enters plants and the major 

7 endpoints O3 may affect. First, Section 9.2 presents a brief overview of various 

8 methodologies that have been, and continue to be, central to quantifying O3 effects on 

9 vegetation (see AX9.1 of the 2006 O3 AQCD for more detailed discussion) (U.S. EPA, 

10 2006b). Section 9.3 through Section 9.4 begin with a discussion of effects at the cellular 

11 and subcellular level followed by consideration of the O3 effects on plant and ecosystem 

12 processes (Figure 9-1). In Section 9.3, research is reviewed from the molecular to the 

13 biochemical and physiological levels in impacted plants, offering insight into the mode of 

14 action of O3. Section 9.4 provides a review of the effects of O3 exposure on major 

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

16 and leaf gas exchange in woody and herbaceous plants in the U.S., as well as a brief 

17 discussion of O3 effects on agricultural crop yield and quality. Section 9.4 also integrates 

18 the effects of O3 on individual plants in a discussion of available research for assessing 

19 the effect of O3 on ecosystems, along with available studies that could inform 

20 assessments of various ecosystem services (See Section 9.4.1.2). The development of 

21 indices of O3 exposure and dose modeling is discussed in Section 9.5. Finally, exposure22 response relationships for a number of tree species, native vegetation, and crop species 

23 and cultivars are reviewed, tabulated, and compared in Section 9.6 to form the basis for 

24 an assessment of the potential risk to vegetation from current ambient levels of O3.
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Figure 9-1 An illustrative diagram of the major pathway through which ozone 

enters plants and the major endpoints that ozone 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. The majority of methodologies currently used have been 

3 discussed in detail in the 1996 O3 AQCD and 2006 O3 AQCD. This section will serve as 

4 a short overview of the methodologies and the reader is referred to the previous O3

5 AQCDs for more in-depth 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 the 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 meters in 

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

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

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

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

8 1976), a frustum at the top to reduce ambient air incursions, and a plastic rain-cap to 

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

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

11 of the 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., 2008a). 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. In studies that included exposure to ambient concentrations 

7 of O3 in both OTCs, and open-air, chamberless control plots, responses in the OTCs were 

8 the same as in open-air plots. Examples include studies of tolerant and sensitive white 

9 clover clones (Trifolium repens) to ambient O3 in greenhouse, open top, and ambient 

10 plots (Heagle et al., 1996), Black Cherry (Prunus serotina) (Neufeld et al., 1995), and 

11 three species of conifers (Neufeld et al., 2000). Experimental comparisons between 

12 exposure methodologies are reviewed in Section 9.2.6. 

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

14 recent studies (Lee et al., 2009a). Concern over the need to establish realistic plant-litter15 soil relationships as a prerequisite to studies of the effects of O3 and CO2 enrichment on 

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

17 partially environmentally controlled, sun-lit chambers (“terracosms”) incorporating 

18 lysimeters (1 meter deep) containing forest soil in which the appropriate horizon structure 

19 was retained.

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

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

22 2007). These closed chambers are approximately 2.44 meters ×1.52 meters with a growth 

volume of approximately 3.7 m3 23 in each chamber. These chambers admit 90% of full 

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

9.2.4 Plume and FACE-Type Systems

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

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

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

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

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

30 The most common plume system used in the U.S. is a modification of the free-air carbon 

31 dioxide/ozone enrichment (FACE) system (Hendrey et al., 1999; Hendrey and Kimball, 

32 1994). Although originally designed to provide chamberless field facilities for studying 

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

34 dispensing of O3 (Karnosky et al., 1999). This method has been employed in Illinois 
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1 (SoyFACE) to study soybeans (Morgan et al., 2004; Rogers et al., 2004) and in 

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

3 papyrifera) and maple (Acer saccharum) (Karnosky et al., 1999). Volk et al. (2003)

4 described a similar system for exposing grasslands that uses 7-m diameter plots. Another 

5 similar FACE system has been used in Finland (Saviranta et al., 2010; Oksanen, 2003).

6 The FACE systems in the U.S. discharge the pollutant gas (O3 and/or CO2) through 

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

8 pipes. Computer-controlled feedback from the monitoring of gas concentration regulates 

9 the feed rate of enriched air to the dispersion pipes. Feedback of wind speed and 

10 directional information ensures that the discharges only occur upwind of the treatment 

11 plots, and that discharge is restricted or closed down during periods of low wind speed or 

12 calm conditions. The diameter of the arrays and their height (25-30 m) in some FACE 

13 systems requires large throughputs of enriched air per plot, particularly in forest tree 

14 systems. The cost of the throughputs tends to limit the number of enrichment treatments, 

15 although Hendrey et al. (1999) argued that the cost on an enriched volume basis is 

16 comparable to that of chamber systems.

17 A different FACE-type facility has been developed for the Kranzberg Ozone Fumigation 

18 Experiment (KROFEX) in Germany beginning in 2000 (Nunn et al., 2002; Werner and 

19 Fabian, 2002). The experiment aims to study the effects of O3 on mature stands of beech 

20 (Fagus sylvatica) and spruce (Picea abies) trees in a system that functions independently 

21 of wind direction. The enrichment of a large volume of the ambient air immediately 

22 above the canopy takes place via orifices in vertical tubes suspended from a horizontal 

23 grid supported above the canopy.

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

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

26 the appropriate numbers, sizes, and orientations of the discharge orifices to avoid 

27 “hot-spots” resulting from the direct impingement of jets of pollutant-enriched air on 

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

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

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

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

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

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

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

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

36 in areas where O3 concentrations are phytotoxic.
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9.2.5 Ambient Gradients

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

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

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

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

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

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

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

8 also limited to small plants.

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

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

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

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

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

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

15 increased elevation.

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

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

18 and Jeffrey pine (Jones and Paine, 2006; Arbaugh et al., 2003; Grulke, 1999; U.S. EPA, 

19 1977). However, it is difficult to separate the effects of N and O3 in some instances in 

20 these studies (Arbaugh et al., 2003). An O3 gradient in Wisconsin has been used to study 

21 foliar injury in a series of trembling aspen clones (Populus tremuloides) differing in O3

22 sensitivity (Maňkovská et al., 2005; Karnosky et al., 1999). Also in the Midwest, an 

23 east-west O3 gradient around southern Lake Michigan was used to look at growth and 

24 visible foliar injury in (P. serotina) and common milkweed (Asclepias syriaca) (Bennett 

25 et al., 2006). 

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

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

28 (Populus deltoides) saplings grown in an urban to rural gradient of O3 by using seven 

29 locations in the New York City area. The secondary nature of the reactions of O3

30 formation and NOX titration reactions within the city center resulted in significantly 

31 higher cumulative O3 exposures in more rural sites. Potential modifying factors such as 

32 soil composition, moisture, or temperature were either controlled or accounted for in 

33 analysis. As shown in Section 9.6.3.3, the response of this species to O3 exposure was 

34 much stronger than most species. The natural gradient exposures were reproduced in 

35 parallel using OTCs, and yielded similar results. Also, the U.S. Forest Service Forest 

36 Inventory and Analysis (FIA) program uses large-scale O3 exposure patterns across the 
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1 continental U.S. to study occurrences of foliar injury due to O3 exposure (Smith et al., 

2 2003) (Section 9.4.2). Finally, McLaughlin et al. (2007a); 2007b) used spatial and 

3 temporal O3 gradients to study forest growth and water use in the southern Appalachians. 

4 These studies found varying O3 exposures between years and between sites.

9.2.6 Comparative Studies

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

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

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

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

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

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

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

12 also control potentially interacting variables and confounders, or else account for them. 

13 However, while increasing control of environmental variables makes it easier to discern 

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

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

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

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

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

19 noisiness of the response data, and ability to make general inferences. 

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

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

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

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

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

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

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

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

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

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

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

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

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

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

34 response of yield to O3 for plants grown in pots or in the ground, suggesting that even 
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1 such a significant change in environment does not alter the proportional response to O3, 

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

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

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

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

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

7 previously observed variation among trembling aspen clones using OTCs (Maňkovská et 

8 al., 2005; Karnosky et al., 1999). In the SoyFACE experiment in Illinois, soybean 

9 (Pioneer 93B15 cultivar) yield loss data from a two-year study was published (Morgan et 

10 al., 2006). This cultivar is a recent selection and, like most modern cultivars, has been 

11 selected under an already high current O3 exposure. It was found to have average 

12 sensitivity to O3 compared to 22 other cultivars tested at SoyFACE. In this experiment, 

13 ambient hourly O3 concentrations were increased by approximately 20% and measured 

14 yields were decreased by 15% in 2002 as a result of the increased O3 exposure (Morgan 

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

16 calculated the expected yield loss from a linear relationship constructed from chamber 

17 data using seven-hour seasonal averages (Ashmore, 2002). They calculated an 8% 

18 expected yield loss from the 2002 O3 exposure using that linear relationship. As reported 

19 in Section 9.2.5, Gregg et al. (2006, 2003) found similar O3 effects on cottonwood 

20 sapling biomass growth along an ambient O3 gradient in the New York City area and a 

21 parallel OTC study.

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

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

24 annual crops, and biomass growth of trees. They are presented in Section 9.6.3 of this 

25 document.

9.3 Mechanisms Governing Vegetation Response to Ozone

9.3.1 Introduction

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

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

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

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

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

31 apoplast, it is thought that the cell must be able to detect the presence of O3 or its 
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1 breakdown products in order to initiate the rapid changes in signaling pathways and gene 

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

3 how O3 and/or its breakdown products are detected in the apoplast and how that leads to 

4 signaling of oxidative stress in plants, much progress has been made in examining several 

5 different mechanisms that may contribute to detecting the presence of O3 and its 

6 breakdown products, and also initiating a signal transduction cascade, which will be 

7 described in Section 9.3.3.1. The next section focuses on changes in gene and protein 

8 expression measured in plants exposed to O3, with particular emphasis on results from 

9 transcriptome (all RNA molecules produced in a cell) and proteome (all proteins 

10 produced in a cell) analyses (Section 9.3.3.2). Subsequently, the role of phytohormones 

11 such as salicylic acid (SA), ethylene (ET), jasmonic acid (JA), and abscisic acid (ABA) 

12 and their interactions in both signal transduction processes and in determining plant 

13 response to O3 is discussed in Section 9.3.3.3. After O3 uptake, some plants can respond 

14 to the oxidative stress with detoxification to minimize damage. These mechanisms of 

15 detoxification, with particular emphasis on antioxidant enzymes and metabolites, are 

16 reviewed in Section 9.3.4. The next section focuses on changes in primary and secondary 

17 metabolism in plants exposed to O3, looking at photosynthesis, respiration and several 

18 secondary metabolites, some of which may also act as antioxidants and protect the plant 

19 from oxidative stress (Section 9.3.5). For many of these topics, information from the 

20 2006 O3 AQCD has been summarized, as this information is still valid and supported by 

21 more recent findings. For other topics, such as genomics and proteomics, which have 

22 arisen due to the availability of new technologies, the information is based solely on new 

23 publications with no reference to the 2006 O3 AQCD.

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

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

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

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

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

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

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

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

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

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

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

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

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

37 processes. 
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1 Many of the studies described in this section focus on changes in the expression of genes 

2 in O3-treated plants. Changes in gene expression (i.e., either upregulation or 

3 downregulation of gene expression) do not always translate into changes in protein 

4 quantity and/or activity, as there are many levels of post-transcriptional and post5 translational modifications which impact protein quantity and activity. Many studies do 

6 not evaluate whether the observed changes in gene expression lead to changes at the 

7 protein level and, therefore, it is not always clear if the changes in gene expression 

8 represent a meaningful biological response to O3 exposure. However, with the advent of 

9 proteomics, some very recent studies have evaluated changes in protein expression for 

10 large numbers of proteins in O3 treated plants, and the findings from these studies support 

11 the previous results regarding changes in gene expression studies as a result of O3

12 exposure. The next step in the process is to determine the implications of the measured 

13 changes occurring at the cellular level to whole plants and ecosystems, which is an 

14 important topic of study which has not been widely addressed.

15 The most noteworthy new body of research since the 2006 O3 AQCD is on the 

16 understanding of molecular mechanisms underlying how plants are affected by O3; many 

17 of the recent studies reviewed here focus on changes in gene expression in plants exposed 

18 to elevated O3. The findings summarized in the 2006 O3 AQCD included decreases in 

19 transcript levels of photosynthesis associated genes, and increases in transcript levels of 

20 genes encoding for pathogenesis-related proteins, enzymes needed for ethylene synthesis, 

21 antioxidant enzymes and defense genes such as phenylalanine ammonia lyase in plants 

22 exposed to O3. These findings have been supported by the new studies, and the advent of 

23 new technologies has allowed for a more comprehensive understanding of the 

24 mechanisms governing how plants are affected by O3. 

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

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

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

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

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

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

31 more realistic exposure conditions. 

9.3.2 Ozone Uptake into the Leaf

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

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

34 to be valid and is only summarized here. 
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1 Stomata provide the principal pathway for O3 to enter and affect plants (Massman and 

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

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

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

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

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

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

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

hydroxyl radicals (HO·) and peroxy radicals (HO2

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

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

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

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

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

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

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

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

17 (U.S. EPA, 2006b). The 2006 O3 AQCD provided a comprehensive summary of these 

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

19 apoplast that initial detoxification reactions by antioxidant metabolites and enzymes take 

20 place, and these initial reactions are critical to reduce concentrations of the oxidative 

21 breakdown products of O3; these reactions are described in more detail in Section 9.3.4 of 

22 this document. 

9.3.2.1 Changes in Stomatal Function

23 Ozone-induced changes in stomatal conductance have been reviewed in detail in previous 

24 O3 AQCDs. The findings summarized in these documents demonstrate that stomatal 

25 conductance is often reduced in plants exposed to O3, resulting either from a direct 

26 impact of O3 on the stomatal complex which causes closure, or as a response to 

27 increasing CO2 concentrations in the substomatal cavity as carbon fixation is reduced. 

28 Although the nature of these effects depends upon many different factors, including the 

29 plant species, concentration and duration of the O3 exposure, and prevailing 

30 meteorological conditions, stomatal conductance is often negatively affected by plant 

31 exposure to O3 (Wittig et al., 2007). Decreases in conductance have been shown to result 

32 from direct as well as indirect effects on stomata (Wittig et al., 2007). Results from the 

33 use of Arabidopsis mutants and new technologies, which allow for analysis of guard cell 

34 function in whole plants rather than in isolated guard cells or epidermal peels, suggest 

35 that O3 may also have a direct impact on stomatal guard cells, leading to alterations in 
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1 stomatal conductance. The use of a new simultaneous O3 exposure/gas exchange device 

2 has demonstrated that exposure of Arabidopsis ecotypes Col-0 and Ler to 150 ppb O3

3 resulted in a 60-70% decline in stomatal conductance within 9-12 minutes of beginning 

4 the exposure. Twenty to thirty minutes later, stomatal conductance had returned to its 

5 initial value, even with continuing exposure to O3, indicating a rapid direct effect of O3

6 on stomatal function (Kollist et al., 2007). This transient decrease in stomatal 

7 conductance was not observed in the abscisic acid insensitive (ABI2) Arabidopsis 

8 mutant. As the ABI2 protein is thought to regulate the signal transduction process 

9 involved in stomatal response downstream of ROS production, the authors suggest that 

10 the transient decrease in stomatal conductance in the Col-0 and Ler ecotypes results from 

11 the biological action of ROS in transducing signals, rather than direct physical damage to 

12 guard cells by ROS (Kollist et al., 2007). This rapid transient decrease in stomatal 

13 conductance was also not observed when exposing the Arabidopsis mutant slac1 (slow 

14 anion channel-associated 1) to 200 ppb O3 (Vahisalu et al., 2008). The SLAC1 protein 

15 was shown to be essential for guard cell slow anion channel functioning and for stomatal 

16 closure in response to O3. Based on additional studies using a variety of Arabidopsis 

17 mutants impaired in various aspects of stomatal function, Vahisalu et al. (2008) suggest 

18 that the presence of ROS in the guard cell apoplast (formed either by O3 breakdown or 

19 through ROS production from NADPH oxidase activity) leads to the activation of a 

20 signaling pathway in the guard cells, which includes SLAC1, and results in stomatal 

21 closure.

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

23 transduction pathway involved in regulating stomatal responses to pollutant stress. A 

24 number of studies in this review provide some evidence that exposure to O3 increases the 

cytosolic free calcium concentration ([Ca2+ 25 ]cyt) in guard cells, which may result in an 

inhibition of the plasma membrane inward-rectifying K+ 26 channels in guard cells, which 

allow for the K+ 27 uptake needed for stomatal opening (McAinsh et al., 2002; Torsethaugen 

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

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

presence of H2O2 activated Ca2+ 30 -permeable channels, which mediate increases in 

[Ca2+ 31 ]cyt in guard cell plasma membranes of Arabidopsis. They also determined that 

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

stomatal closure via activation of the membrane Ca2+ 33 channels. Therefore, it is possible 

that H2O2, a byproduct of O3 breakdown in the apoplast, could disrupt the Ca2+ 34 -ABA 

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

36 2002). The studies described here provide some evidence to suggest that O3 and its 

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

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

39 described as a delay in stomatal response to changing environmental conditions in 
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1 sensitive species exposed to higher concentrations and/or longer-term O3 exposures 

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

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

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

5 McAinsh et al., 2002).

Note: 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 9-2 The microarchitecture of a dicot leaf.
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Note: (a) Ozone reacts at the double bonds to form carbonyl groups. (b) Under certain circumstances, peroxides are generated.

Figure 9-3 Possible reactions of ozone within water.

Note: (a) The typical Crigee mechanism is shown in which several reaction paths from the initial product are shown. 

(b) Typical reaction of ascorbic acid with ozone.

Source: Adapted from Mudd (1996).

Figure 9-4 The Crigee mechanism of ozone attack of a double bond.
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9.3.3 Cellular to Systemic Responses

9.3.3.1 Ozone Detection 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 detection of O3 by plants have been proposed; 

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

15 events that occur in plants exposed to O3 have been described in the guard cells of 

16 stomata. Reactive oxygen species were observed in the chloroplasts of guard cells in the 

17 O3 tolerant Col-0 Arabidopsis thaliana ecotype plants within 5 minutes of plant exposure 

18 to 350 ppb O3 (Joo et al., 2005). Reactive oxygen species from the breakdown of O3 in 

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

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

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

22 oxidases to produce ROS and, as a result, propagate the oxidative burst to neighboring 

23 cells (Joo et al., 2005). Therefore, G-proteins are recognized as important molecules 

24 involved in plant responses to O3 and may play a role in detecting the presence of ROS 

25 from the breakdown of O3 in the apoplast (Kangasjarvi et al., 2005; Booker et al., 2004a). 

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 detection 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 
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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 

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 MAPK activity in 

13 different Arabidopsis and tobacco lines resulted in increased plant sensitivity to O3 (Miles 

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

15 Arabidopsis ecotypes was correlated with greater upregulation of MAPK signaling 

16 pathways upon O3 exposure than in more sensitive Arabidopsis ecotypes (Li et al., 

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

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

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

20 these communication channels.

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

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

23 These mechanisms may vary by species or developmental stage of the plant, or may 

24 co-exist and be activated by different exposure conditions. Calcium and protein kinases 

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

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

27 the plant 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 
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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 Sadowski, 

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

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

5 2005; Mahalingam et al., 2005; Tamaoki et al., 2003), pepper (Capsicum annuum) (Lee 

6 and Yun, 2006), clover (Medicago truncatula) (Puckette et al., 2008), Phillyrea latifolia

7 (Paolacci et al., 2007), poplar (Street et al., 2011), and European beech (Fagus sylvatica) 

8 (Olbrich et al., 2010; Olbrich et al., 2009; Olbrich et al., 2005). In some cases, 

9 researchers compared transcriptomes of two or more cultivars, ecotypes or mutants that 

10 differed in their sensitivity to O3 (Puckette et al., 2008; Rizzo et al., 2007; Lee and Yun, 

11 2006; Li et al., 2006b; Tamaoki et al., 2003). Species, O3 exposure conditions 

12 (concentration, duration of exposure) and sampling times varied considerably in these 

13 studies. However, functional classification of the genes that were either upregulated or 

14 downregulated 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 generallyupregulated, while those related to photosynthesis 

17 and general metabolism were typically downregulated 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 sensitive and tolerant plants in response to O3 exposure. In pepper, 67% of the 

22 180 genes studied that were affected by O3 were differentially regulated in the sensitive 

23 and tolerant cultivars. At both 0 hours and 48 hours after a 3-day exposure at 150 ppb, O3

24 responsive genes were either upregulated or downregulated more markedly in the 

25 sensitive than in the tolerant cultivar (Lee and Yun, 2006). Transcriptome analysis also 

26 revealed differences in timing and magnitude of changes in gene expression between 

27 sensitive and tolerant clovers. Acute exposure (300 ppb O3 for 6 hours) led to the 

28 production of an oxidative burst in both clovers (Puckette et al., 2008). However, the 

29 sensitive-Jemalong cultivar exhibited a sustained ROS burst and a concomitant 

30 downregulation of defense response genes at 12 hours after the onset of exposure, while 

31 the tolerant JE 154 accession showed much more rapid and large-scale transcriptome 

32 changes than the Jemalong cultivar (Puckette et 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 
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1 proteases, and downregulation 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-hour 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 upregulated 

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

8 in the class of late upregulated 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 resulted 

11 in the upregulation 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), 

16 O3-induced changes in gene transcription were similar to those observed for herbaceous 

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

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

19 detoxifying ROS, wereupregulated. Some genes associated with primary metabolism, cell 

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

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

22 transcriptional changes described above was far greater in the saplings than in the adult 

23 trees exposed to the same O3 concentrations for the same time period (Olbrich et al., 

24 2010). 

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

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

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

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

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

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

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

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

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

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

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

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

37 occurred after the plants had been exposed to O3 for 14 days, and at this point in time, 
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1 several Calvin cycle enzymes were reduced in quantity, while the effects on the light 

2 reactions appeared later, at 21 days after beginning treatment. Some of the antioxidant 

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

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

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

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

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

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

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

10 (exposed to 200 ppb O3 for 24 hours) focusing on the integration of transcriptomics and 

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

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

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

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

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

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

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

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

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

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

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

22 in plants exposed to a stressor such as O3. 

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

24 several O3 concentrations, including filtered air, ambient O3 (mean concentration 47 ppb), 

25 ambient + 10 ppb and ambient + 20 ppb for 5 h/day for 50 days. Declines in the rate of 

26 photosynthesis and stomatal conductance were related to decreases in proteins involved 

27 in carbon fixation and electron transport and increased proteolysis of photosynthetic 

28 proteins such as the large subunit of ribulose-1,6-bisphosphate carboxylase/oxygenase 

29 (Rubisco). Enzymes that take part in energy metabolism, such as ATP synthesis, were 

30 also downregulated, while defense/stress related proteins were upregulated in O3-treated 

31 plants. In comparing the two wheat cultivars, Sarkar et al. (2010) found that while the 

32 qualitative changes in protein expression between the two cultivars were similar, the 

33 magnitude of these changes differed between the sensitive and tolerant wheat cultivars. 

34 Greater foliar injury and a smaller decline in stomatal conductance was observed in the 

35 sensitive cultivar as compared to the more tolerant cultivar, along with greater losses in 

36 photosynthetic enzymes and higher quantities of antioxidant enzymes. Results from a 

37 three-year exposure of European beech saplings to elevated O3 (AOT40 value was 

38 52.6 μL/L•h for 2006 when trees were sampled) supported the results from the short-term 
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1 exposure studies described above (Kerner et al., 2011). The O3 treatment of the saplings 

2 resulted in reductions in enzymes associated with the Calvin cycle, which could lead to 

3 reduced carbon fixation. Enzymes associated with carbon metabolism/catabolism were 

4 increased, and quantities of starch and sucrose were reduced in response to the O3

5 treatment in these trees, indicating a potential impact of O3 on overall carbon metabolism 

6 in long-term exposure conditions (Kerner et al., 2011). 

7 Transcriptome and proteome studies have provided valuable information about O3 effects 

8 on plants. These studies allow for simultaneous analysis of changes in the expression 

9 patterns of many different genes and proteins, and also provide information on how these 

10 molecules might interact with one another as a result of plant exposure to oxidative 

11 stress. Gene and protein expression patterns generally differ between O3-sensitive and 

12 tolerant plants, which could result from differential uptake or detoxification of O3 or from 

13 differential regulation of the transcriptome and proteome.
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Note: (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 upregulated 

genes (yellow, peaks at 0.5-1 hours), and the 3 hours (blue), 4.5 hours (red) and 9-12 hours (green) late upregulated genes and for 

the downregulated genes coding for photosynthesis proteins (brown). (B) Diagrammatic representation of redox regulation of the 

oxidative stress response.

Source: Reprinted with permission of Springer (Mahalingam et al., 2005).

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

the genetic response to ozone stress.

1 All of these studies describe common trends for changes in gene and protein expression 

2 which occur in a variety of plant species exposed to O3. While genes associated with 

3 carbon assimilation and general metabolism are typically downregulated, genes 

4 associated with signaling, catabolism, and defense are upregulated. The magnitude of 

5 these changes in gene and protein expression appears to be related to plant species, age 

6 and their sensitivity or tolerance to O3. averaged (normalized) transcript profiles for gene groups
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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. The 2006 O3 AQCD 

3 documents the O3-induced production of ET and its role in promoting the formation of 

4 leaf lesions. Transcriptome analysis and the use of a variety of mutants have allowed for 

5 further elucidation of the complex interactions between SA, ET, JA and the role of 

6 abscisic acid (ABA) in mediating plant response to O3 (Ludwikow and Sadowski, 2008). 

7 In addition to their roles in signaling pathways, phytohormones also appear to regulate, 

8 and be regulated by, the MAPK signaling cascades described previously. Most evidence 

9 suggests that while ET and SA are needed to develop O3-induced leaf lesions, JA acts 

10 antagonistically to SA and ET to limit the 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 

31 JA-associated genes, which would serve to inhibit the spread of lesions (Mahalingam et 

32 al., 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 for 

2 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 upregulated 

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 were downregulated, while SA was upregulated 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. (2010b) 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. It was determined that the presence of H2O2, which is formed from O3

20 degradation, increases the sensitivity of guard cells to ABA and, therefore, more readily 

21 results in stomatal closure. More recently, it was determined that synthesis of ABA was 

22 induced in O3-treated Arabidopsis plants (250-350 ppb O3 for 6 hours), with a more 

23 pronounced induction in the O3 sensitive rcd3 mutant as compared to the wildtype Col-0 

24 (Overmyer et al., 2008). The rcd3 mutant also exhibited a lack of O3-induced stomatal 

25 closure, and the RCD3 protein has been shown to be required for slow anion channels 

26 (Overmyer et al., 2008). Ludwikow et al. (2009) used Arabidopsis ABI1td mutants, in 

27 which a key negative regulator of ABA action (abscisic acid insensitive1 protein 

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

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

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

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

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

33 (Ludwikow et al., 2009).

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

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

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

37 (150 ppb for 5 hours) stimulated NO and NO-dependent ET production, while NO 

38 production itself did not depend on the presence of ET (Ederli et al., 2006). Analysis of 
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1 O3-treated Arabidopsis indicated the possibility of a dual role for NO in the initiation of 

2 cell death and later lesion containment (Ahlfors et al., 2009). 

3 While much work remains to be done to better elucidate how plants detect O3, what 

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

5 that the mechanism for O3 detection and signal transduction is very complex. Many of the 

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

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

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

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

10 plant responses to that stress. 

Note: 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.

Source: Reprinted with permission of Blackwell Publishing Ltd. (Kangasjarvi et al., 2005).

Figure 9-6 The oxidative cell death cycle.
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9.3.4 Detoxification

9.3.4.1 Overview of Ozone-induced Defense Mechanisms

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

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

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

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

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

6 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

- 7 , HO· and 

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

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

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

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

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

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

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

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

16 antioxidants (Foyer and Noctor, 2005b). 

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

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

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

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

21 2005b). As such, not only the quantity and types of antioxidant enzymes and metabolites 

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

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

24 2008). Molecules such as glutathione (GSH), thioredoxins and NADPH play very 

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

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

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

9.3.4.2 Role of Antioxidants in Plant Defense Responses

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

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

30 chloroplast, the cytosol and the apoplast (Noctor and Foyer, 1998). Ascorbate is 
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1 synthesized in the cell and transported to the apoplast. Apoplastic ascorbate can be 

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

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

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

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

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

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

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

9 foliar injury from O3 exposure in a field study on tall milkweed plants (Asclepsias 

10 exaltata L.) (Burkey et al., 2006; Souza et al., 2006). Cheng et al. (2007) exposed two 

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

12 differences in sensitivity between the two cultivars could not be explained by differential 

13 O3 uptake or by the fraction of reduced ascorbate present in the apoplast. However, total 

14 antioxidant capacity of the apoplast was 2-fold higher in the tolerant Essex cultivar as 

15 compared to the sensitive Forrest cultivar, indicating that there may be other compounds 

16 in the leaf apoplast that scavenge ROS. D'Haese et al. (2005) exposed the NC-S 

17 (sensitive) and NC-R (resistant) clones of white clover (Trifolium repens) to 60 ppb O3

18 for 7-h/day for 5 days in environmental chambers. Surprisingly, the NC-S clone had a 

19 higher constitutive concentration of apoplastic ascorbate with a higher redox status than 

20 the NC-R clone. However, the redox status of symplastic GSH was higher in NC-R, even 

21 though the concentration of GSH was not higher than in NC-S. In addition, total 

22 symplastic antioxidative capacity was not a determining factor in differential sensitivity 

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

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

25 a growing season and then exposing them to elevated O3 (100 ppb for 8-h/day for 

26 10 days) in OTC at the end of the field season. The NC-R clone had greater quantities of 

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

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

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

30 10 days in OTC (Burkey et al., 2003). While most of the apoplastic ascorbate was in the 

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

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

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

34 44 ppb O3) and elevated (7-h average 56 ppb O3) O3 for 60 days in open-air field 

35 conditions showed higher concentrations of reduced ascorbate in the apoplast in the 

36 tolerant Y16 variety than the more sensitive Y2 variety, however no varietal differences 

37 were seen in the decrease in reduced ascorbate quantity in response to O3 exposure (Feng 

38 et al., 2010). There is much evidence that supports an important role for ascorbate, 

39 particularly apoplastic ascorbate, in protecting plants from oxidative stressors such as O3; 
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1 however, it is also clear that there is much variation in the importance of ascorbate for 

2 different plant species and differing exposure conditions. Additionally, the work of 

3 several authors suggests that there may be other compounds in the apoplast which have 

4 the capacity to act as antioxidants. 

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

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

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

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

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

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

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

12 exhibited an increase in guard cell ascorbic acid, leading to a decrease in stomatal 

13 responsiveness to O3 and an increase in stomatal conductance and O3 uptake. Despite 

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

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

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

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

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

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

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

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

22 (200 ppb O3), with higher rates of photosynthesis and higher levels of reduced ascorbic 

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

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

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

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

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

28 peroxidase (POD) and superoxide dismutase (SOD), were comprehensively reviewed in 

29 the 2006 O3 AQCD. Based on the review of the literature, no conclusive and consistent 

30 effects of O3 on the quantity of GSH and CAT could be identified. Both apoplastic and 

31 cytosolic POD activity increased in response to O3 exposure, while various isoforms of 

32 SOD showed inconsistent changes in quantity in response to O3. Additional studies have 

33 been conducted to further elucidate the roles of these antioxidant enzymes and 

34 metabolites in protecting plants from oxidative stress. Superoxide dismutase and POD 

35 activities were measured in both the tolerant Bel B and sensitive Bel W3 tobacco 

36 cultivars exposed to ambient O3 concentrations for 2 weeks 3 times throughout a growing 

37 season (Borowiak et al., 2009). In this study, SOD and POD activity, including that of 

38 several different isoforms, increased in both the sensitive and tolerant tobacco cultivars 
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1 with exposure to O3, however the isoenzyme composition for POD differed between the 

2 sensitive and tolerant tobacco cultivars (Borowiak et al., 2009) Tulip poplar 

3 (Liriodendron tulipifera) trees exposed to increasing O3 concentrations (from 100 to 

4 300 ppb O3 during a 2-week period) showed increases in activities of SOD, ascorbate 

5 peroxidase (APX), glutathione reductase (GR), MDAR, DHAR, CAT and POD in the 

6 2-week period, although individual enzyme activities increased at different times during 

7 the 2-week period (Ryang et al., 2009).

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

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

10 declines in the activities and quantities of these enzymes after 75 days of exposure (Yan 

11 et al., 2010). Similarly, activities of SOD, APX, DHAR, MDAR, and GR increased in 

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

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

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

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

16 after germination (Singh et al., 2010a).

17 Antioxidant enzymes and metabolites have been shown to play an important role in 

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

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

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

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

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

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

24 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

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

26 been documented for many different plant species (Booker et al., 2009; Wittig et al., 

27 2007; U.S. EPA, 2006b). The 2006 O3 AQCD described the mechanism by which plant 

28 exposure to O3 reduces the quantity of Rubisco, and the more recent scientific literature 

29 confirms these findings. While several measures of the light reactions of photosynthesis 

30 are sensitive to exposure to O3 (see below), photosynthetic carbon assimilation is 

31 generally considered to be more affected by pollutant exposure, resulting in an overall 

32 decline in photosynthesis (Guidi and Degl'lnnocenti, 2008; Heath, 2008; Fiscus et al., 
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1 2005). Loss of carbon assimilation capacity has been shown to result primarily from 

2 declines in the quantity of Rubisco (Singh et al., 2009; Calatayud et al., 2007a). 

3 Experimental evidence suggests that both decreases in Rubisco synthesis and enhanced 

4 degradation of the protein contribute to the measured reduction in its quantity (U.S. EPA, 

5 2006b). Reduced carbon assimilation has been linked to reductions in biomass and yield 

6 (Wang et al., 2009b; He et al., 2007; Novak et al., 2007; Gregg et al., 2006; Keutgen et 

7 al., 2005). Recent studies evaluating O3 induced changes in the transcriptome and 

8 proteome of several different species confirm these findings. Levels of mRNA for the 

9 small subunit of Rubisco (rbcS) declined in European beech saplings exposed to 300 ppb 

10 O3 for 8-h/day for up to 26 days (Olbrich et al., 2005). Similar declines in rbcS mRNA 

11 were also measured in the beech saplings in a free air exposure system over a course of 

12 two growing seasons (Olbrich et al., 2009). Proteomics studies have also confirmed the 

13 effects of O3 on proteins involved in carbon assimilation. Reductions in quantities of the 

14 small and large subunit (rbcL) of Rubisco and Rubisco activase were measured in 

15 soybean plants exposed to 120 ppb O3 for 3 days in growth chambers (Ahsan et al., 

16 2010). Exposure of young poplar trees to 120 ppb O3 for 35 days in exposure chambers 

17 resulted in reductions of Rubisco, Rubisco activase, and up to 24 isoforms of Calvin 

18 cycle enzymes, most of which play a role in regenerating the CO2 acceptor molecule, 

19 ribulose-1.5-bisphosphate (Bohler et al., 2007). Reductions in protein quantity of both the 

20 small and large subunit of Rubisco were seen in wheat plants exposed to ambient 

21 (average concentration 47.3 ppb O3) and elevated O3 (ambient + 10 or 20 ppb O3) in 

22 open-top chambers for 5-h/day for 50 days (Sarkar et al., 2010). Lettuce plants exposed 

23 to 100 ppb O3 in growth chambers for 8-h/day for 3 weeks also showed reductions in 

24 transcript and protein levels of the small and large subunits of Rubisco and Rubisco 

25 activase (Goumenaki et al., 2010). The reductions in carbon assimilation have been 

26 associated with declines in both the mRNA of the small and large subunits of Rubisco, 

27 and with reductions in Rubisco activase mRNA and protein. Additionally, the reduction 

28 in Rubisco quantity has also been associated with the O3-induced oxidative modification 

29 of the enzyme, which is evidenced by the increases in carbonyl groups on the protein 

30 after plant exposure to O3. 

31 In addition to impacts on carbon assimilation, the deleterious effects of O3 on the 

32 photosynthetic light reactions have received more attention in recent years. Chlorophyll 

33 fluorescence provides a useful measure of changes to the photosynthetic process from 

34 exposure to oxidative stress. Decreases in the Fv/Fm ratio (a measure of the maximum 

35 efficiency of Photosystem II) in dark adapted leaves indicate a decline in the efficiency of 

36 the PSII photosystems and a concomitant increase in non-photochemical quenching 

37 (Guidi and Degl'lnnocenti, 2008; Scebba et al., 2006). Changes in these parameters have 

38 been correlated to differential sensitivity of plants to the pollutant. In a study to evaluate 

39 the response of 4 maple species to O3 (exposed to an 8-h avg of 51 ppb for ambient and 
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1 79 ppb for elevated treatment in OTC), the 2 species which were most sensitive based on 

2 visible injury and declines in CO2 assimilation also showed the greatest decreases in 

3 Fv/Fm in symptomatic leaves. In asymptomatic leaves, CO2 assimilation decreased 

4 significantly but there was no significant decline in Fv/Fm (Calatayud et al., 2007a). 

5 Degl'Innocenti et al. (2007) measured significant decreases in Fv/Fm in young and 

6 symptomatic leaves of a resistant tomato genotype (line 93.1033/1) in response to O3

7 exposure (150 ppb O3 for 3 hours in a growth chamber), but only minor decreases in 

8 asymptomatic leaves with no associated changes in net photosynthetic rate. In the O3

9 sensitive tomato cultivar Cuor Di Bue, the Fv/Fm ratio did not change, while the 

10 photosynthetic rate declined significantly in asymptomatic leaves (Degl'Innocenti et al., 

11 2007). In two soybean cultivars, Fv/Fm also declined significantly with plant exposure to 

12 O3 (Singh et al., 2009). It appears that in asymptomatic leaves, photoinhibition, as 

13 indicated by a decrease in Fv/Fm, is not the main reason for a decline in photosynthesis.

14 An evaluation of photosynthetic parameters of two white clover (Trifolium repens cv. 

15 Regal) clones that differ in their O3 sensitivity revealed that O3 (40-110 ppb O3 for 7-

16 h/day for 5 days) increased the coefficient of non-photochemical quenching (qNP) in both 

17 the resistant (NC-R) and sensitive (NC-S) clones, however qNP was significantly lower 

18 for the sensitive clone (Crous et al., 2006). Sensitive Acer clones had a lower coefficient 

19 of non-photochemical quenching, while exposure to O3 increased qNP in both sensitive 

20 and tolerant clones (Calatayud et al., 2007a). While exposure to O3 also increased qNP in 

21 tomato, there were no differences in the coefficient of photochemical quenching between 

22 cultivars thought to be differentially sensitive to O3 (Degl'Innocenti et al., 2007). Higher 

23 qNP as a result of exposure to O3 indicates a reduction in the proportion of absorbed light 

24 energy being used to drive photochemistry. A lower coefficient of non-photochemical 

25 quenching in O3 sensitive plants could indicate increased vulnerability to ROS generated 

26 during exposure to oxidative stress (Crous et al., 2006). 

27 Most of the research on O3 effects on photosynthesis has focused on C3 (Calvin cycle) 

28 plants because C4 (Hatch-Slack) plants have lower stomatal conductance and are, 

29 therefore, thought to be less sensitive to O3 stress. However, some studies have been 

30 conducted to evaluate the effects of O3 on C4 photosynthesis. In older maize leaves, 

31 Leitao et al. (2007c); Leitao et al. (2007a) found that the activity, quantity and transcript 

32 levels of both Rubisco and phosphoenolpyruvate carboxylase (PEPc) decreased as a 

33 function of rising O3 concentration. In younger maize leaves, the quantity, activity, and 

34 transcript levels of the carboxylases were either increased or unaffected in plants exposed 

35 to 40 ppb O3 for 7- h/day for 28-33 days, but decreased at 80 ppb (Leitao et al., 2007b; 

36 Leitao et al., 2007c). In another study, Grantz and Vu (2009) reported that O3 exposures 

37 (4, 58, and 114 ppb, 12-hour mean) decreased sugarcane biomass production by more 

38 than one third and allocation to roots by more than two thirds. 
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9.3.5.2 Respiration and Dark Respiration

1 While much research emphasis regarding O3 effects on plants has focused on the negative 

2 impacts on carbon assimilation, other studies have measured impacts on catabolic 

3 pathways such as shoot respiration and photorespiration. Generally, shoot respiration has 

4 been found to increase in plants exposed to O3. Bean plants exposed to ambient (average 

5 12-h mean 43 ppb) and twice ambient (average 12-h mean 80 ppb) O3 showed increases 

6 in respiration. When mathematically partitioned, the maintenance coefficient of 

7 respiration was significantly increased in O3 treated plants, while the growth coefficient 

8 of respiration was not affected (Amthor, 1988). Loblolly pines were exposed to ambient 

9 (12-h daily mean was 45 ppb) and twice ambient (12 hours daily mean was 86 ppb) O3

10 for 12-h/day for approximately seven months per year for 3 and 4 years. While 

11 photosynthetic activity declined with the age of the needles and increasing O3

12 concentration, enzymes associated with respiration showed higher levels of activity with 

13 increasing O3 concentration (Dizengremel et al., 1994). In their review on the role of 

14 metabolic changes in plant redox status after O3 exposure, Dizengremel et al. (2009)

15 summarized multiple studies in which several different tree species were exposed to O3

16 concentrations ranging from ambient to 200 ppb O3 for at least several weeks. In all 

17 cases, the activity of enzymes, including phosphofructokinase, pyruvate kinase and 

18 fumarase, which are part of several catabolic pathways, were increased in O3 treated 

19 plants.

20 Photorespiration is a light-stimulated process which consumes O2 and releases CO2. 

21 While it has been regarded as a wasteful process, more recent evidence suggests that it 

22 may play a role in photoprotection during photosynthesis (Bagard et al., 2008). The few 

23 studies that have been conducted on O3 effects on photorespiration suggest that rates of 

24 photorespiration decline concomitantly with rates of photosynthesis. Soybean plants were 

25 exposed to ambient (daily averages 43-58 ppb) and 1.5 ambient O3 (daily averages 63-

26 83 ppb) O3 in OTCs for 12-h/day for 4 months. Rates of photosynthesis and 

27 photorespiration and photorespiratory enzyme activity declined only at the end of the 

28 growing season and did not appear to be very sensitive to O3 exposure (Booker et al., 

29 1997). Young hybrid poplars exposed to 120 ppb O3 for 13-h/day for 35 days in 

30 phytotron chambers showed that effects on photorespiration and photosynthesis were 

31 dependent upon the developmental stage of the leaf. While young leaves were not 

32 impacted, reductions in photosynthesis and photorespiration were measured in fully 

33 expanded leaves (Bagard et al., 2008).
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9.3.5.3 Secondary Metabolism

1 Transcriptome analysis of Arabidopsis plants has revealed modulation of several genes 

2 involved in plant secondary metabolism (Ludwikow and Sadowski, 2008). Phenylalanine 

3 ammonia lyase (PAL) has been the focus of many studies involving plant exposure to O3

4 due to its importance in linking the phenylpropanoid pathway of plant secondary 

5 metabolism to primary metabolism in the form of the shikimate pathway. Genes encoding 

6 several enzymes of the phenylpropanoid pathway and lignin biosynthesis were 

7 upregulated in transcriptome analysis of Arabidopsis plants (Col-0) exposed to 350 ppb 

8 O3 for 6 hours, while 2 genes involved in flavonoid biosynthesis were downregulated 

9 (Ludwikow et al., 2004). Exposure of Arabidopsis (Col-0) to lower O3 concentrations 

10 (150 ppb for 8-h/day for 2 days) resulted in the induction of 11 transcripts involved in 

11 flavonoid synthesis. In their exposure of 2-year-old Mediterranean shrub Phillyrea 

12 latifolia to 110 ppb O3 for 90 days, Paolacci et al. (2007) identified four clones that were 

13 upregulated and corresponded to genes involved in the synthesis of secondary 

14 metabolites, such as isoprenoids, polyamines and phenylpropanoids. Upregulation of 

15 genes involved in isoprene synthesis was also observed in Medicago trunculata exposed 

16 to 300 ppb O3 for 6 hours, while genes encoding enzymes of the flavonoid synthesis 

17 pathway were either upregulated or downregulated (Puckette et al., 2008). Exposure of 

18 red clover to 1.5 × ambient O3 (average concentrations of 32.4 ppb) for up to 9 weeks in 

19 an open field exposure system resulted in increases in leaf total phenolic content. 

20 However, the types of phenolics that were increased in response to O3 exposure differed 

21 depending upon the developmental stage of the plant. While almost all of the 31 different 

22 phenolic compounds measured increased in quantity initially during the exposure, after 

23 3 weeks the quantity of isoflavones decreased while other phenolics increased (Saviranta 

24 et al., 2010). Exposure of beech saplings to ambient and 2 × ambient O3 concentrations 

25 over 2 growing seasons resulted in the induction of several enzymes which contribute to 

26 lignin formation, while enzymes involved in flavonoid biosynthesis were downregulated

27 (Olbrich et al., 2009). Exposure of tobacco Bel W3 to 160 ppb O3 for 5 hours showed 

28 upregulation of almost all genes encoding for enzymes which are part of the 

29 prechorismate pathway (Janzik et al., 2005). Isoprenoids can serve as antioxidant 

30 compounds in plants exposed to oxidative stress (Paolacci et al., 2007). 

31 The prechorismate pathway is the pathway leading to the formation of chorismate, a 

32 precursor to the formation of the aromatic amino acids tryptophan, tyrosine and 

33 phenylalanine. These amino acids are precursors for the formation of many secondary 

34 aromatic compounds, and, therefore, the prechorismate pathway represents a branch35 point in the regulation of metabolites into either primary or secondary metabolism (Janzik 

36 et al., 2005). Exposure of the O3 sensitive Bel W3 tobacco cultivar at 160 ppb for 5 hours 

37 showed an increase in transcript levels of most of the genes encoding enzymes of the 
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1 prechorismate pathway. However, shikimate kinase (SK) did not show any change in 

2 transcript levels and only one of three isoforms of DAHPS (3-deoxy-D-arabino3 heptulosonat-7-phosphate synthase), the first enzyme in this pathway, was induced by O3

4 exposure (Janzik et al., 2005). Differential induction of DAHPS isoforms was also 

5 observed in European beech after 40 days of exposure to 150-190 ppb O3. At this time 

6 point in the beech experiment, transcript levels of shikimate pathway enzymes, including 

7 SK, were generally strongly induced after an only weak initial induction after the first 

8 40 days of exposure. Both soluble and cell-wall bound phenolic metabolites showed only 

9 minimal increases in response to O3 for the duration of the exposure period (Alonso et al., 

10 2007). Total leaf phenolics decreased with leaf age in Populus nigra exposed to 80 ppb 

11 O3 for 12-h/day for 14 days. Ozone increased the concentration of total leaf phenolics in 

12 newly expanded leaves, with the greatest increases occurring in compounds such as 

13 quercitin glycoside, which has a high antioxidant capacity (Fares et al., 2010b). While 

14 several phenylpropanoid pathway enzymes were induced in two poplar clones exposed to 

15 60 ppb O3 for 5-h/day for 15 days, the degree of induction differed between the two 

16 clones. In the tolerant I-214 clone, PAL activity increased 9-fold in O3-treated plants as 

17 compared to controls, while there was no significant difference in PAL activity in the 

18 sensitive Eridano clone (DiBaccio et al., 2008). 

19 Polyamines such as putrescine, spermidine and spermine play a variety of roles in plants 

20 and have been implicated in plant defense responses to both abiotic and biotic stresses. 

21 They exist in both a free form and conjugated to hydroxycinnamic acids. Investigations 

22 on the role of polyamines have found that levels of putrescine increase in response to 

23 oxidative stress. This increase stems largely from the increase in the activity of arginine 

24 decarboxylase (ADC), a key enzyme in the synthesis of putrescine (Groppa and 

25 Benavides, 2008). Langebartels et al. (1991) described differences in putrescine 

26 accumulation in O3-treated tobacco plants exposed to several O3 concentrations, ranging 

27 from 0-400 ppb for 5-7 hours. A large and rapid increase in putrescine occurred in the 

28 tolerant Bel B cultivar and only a small increase in the sensitive Bel W3 cultivar, which 

29 occurred only after the formation of necrotic leaf lesions. Van Buuren et al. (2002)

30 further examined the role of polyamines in these two tobacco cultivars during an acute 

31 (130 ppb O3 for 7-h in a growth chamber) exposure. They found that while free 

32 putrescine accumulated in undamaged tissue of both cultivars, conjugated putrescine 

33 predominantly accumulated in tissues undergoing cell death after plant exposure to O3

34 (Van Buuren et al., 2002). The authors suggest that while free putrescine may not play a 

35 role in conferring tolerance in the Bel B cultivar, conjugated putrescine may play a role in 

36 O3-induced programmed cell death in Bel W3 plants.

37 Isoprene is emitted by some plant species and represents the predominant biogenic source 

38 of hydrocarbon emissions in the atmosphere (Guenther et al., 2006). In the atmosphere, 
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1 the oxidation of isoprene by hydroxyl radicals can enhance O3 formation in the presence 

2 of NOX, thereby impacting the O3 concentration that plants are exposed to. While 

3 isoprene emission varies widely between species, it has been proposed to stabilize 

4 membranes and provide those plant species that produce it with a mechanism of 

5 thermotolerance (Sharkey et al., 2008). It has also been suggested that isoprene may act 

6 as an antioxidant compound to scavenge O3 (Loreto and Velikova, 2001). Recent studies 

7 using a variety of plant species have shown conflicting results in trying to understand the 

8 effects of O3 on isoprene emission. Exposure to acute doses of O3 (300 ppb for 3-h) in 

9 detached leaves of Phragmites australis resulted in stimulation of isoprene emissions 

10 (Velikova et al., 2005). Similar increases in isoprene emissions were measured in 

11 Populus nigra after exposure to 100 ppb O3 for 5 days continuously (Fares et al., 2008). 

12 Isoprene emission in attached leaves of Populus alba, which were exposed to 150 ppb O3

13 for 11-h/day for 30 days inside cuvettes, was inhibited, while isoprene emission and 

14 transcript levels of isoprene synthase mRNA were increased in the leaves exposed to 

15 ambient O3 (40 ppb), which were located above the leaves enclosed in the exposure 

16 cuvettes (Fares et al., 2006). Exposure of 2 genotypes of hybrid poplar to 120 ppb O3 for 

17 6-h/day for 8 days resulted in a significant reduction in isoprene emission in the O3-

18 sensitive but not the tolerant genotype (Ryan et al., 2009). Similarly, O3 treatment 

19 (80 ppb 12-h/day for 14 days) of Populus nigra showed that isoprene emission was 

20 reduced in the treated plants relative to the control plants (Fares et al., 2010b). Based on 

21 results of this and other studies, Fares et al. (2010b) concluded that the isoprenoid 

22 pathway may be induced in plants exposed to acute O3 doses, while at lower doses 

23 isoprene emission may be inhibited. Vickers et al. (2009) developed transgenic tobacco 

24 plants with the isoprene synthase gene from Populus alba and exposed them to 120 ppb 

25 O3 for 6-h/day for 2 days. They determined that the wildtype plants showed significantly 

26 more O3 damage, including the development of leaf lesions and a decline in 

27 photosynthetic rates, than the transgenic, isoprene-emitting plants. Transgenic plants also 

28 accumulated less H2O2 and had lower levels of lipid peroxidation following exposure to 

29 O3 than the wildtype plants (Vickers et al., 2009). These results indicate that isoprene 

30 may have a protective role for plants exposed to oxidative stress.

9.3.6 Summary

31 The results of recent studies on the effects of O3 stress on plants support and strengthen 

32 those reported in the 2006 O3 AQCD. The most significant new body of evidence since 

33 the 2006 O3 AQCD comes from research on molecular mechanisms of the biochemical 

34 and physiological changes observed in many plant species in response to O3 exposure. 

35 Recent studies have employed new techniques, such as those used in evaluating 
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1 transcriptomes and proteomes to perform very comprehensive analyses of changes in 

2 gene transcription and protein expression in plants exposed to O3. These newer molecular 

3 studies not only provide very important information regarding the many mechanisms of 

4 plant responses to O3, they also allow for the analysis of interactions between various 

5 biochemical pathways which are induced in response to O3. However, many of these 

6 studies have been conducted in artificial conditions with model plants, which are 

7 typically exposed to very high, short doses of O3. Therefore, additional work remains to 

8 elucidate whether these plant responses are transferable to other plant species exposed to 

9 more realistic ambient conditions. 

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

11 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 , )12 ־

radicals (HO·) and peroxy radicals (HO2

· 13 ). Plants could be detecting the presence of O3

14 and/or its breakdown products in a variety different ways, depending upon the plant 

15 species and the exposure parameters. Experimental evidence suggests that mitogen16 activated protein kinases and calcium are important components of the signal 

17 transduction pathways, which communicate signals to the nucleus and lead to changes in 

18 gene expression in response to O3. It is probable that there are multiple detection 

19 mechanisms and signal transduction pathways, and their activation may depend upon the 

20 plant species, its developmental stage and/or O3 exposure conditions. Initiation of signal 

21 transduction pathways in O3 treated plants has also been observed in stomatal guard cells. 

22 Reductions in stomatal conductance have been described for many plant species exposed 

23 to O3, and new experimental evidence suggests that this reduction may be due not only to 

24 a decrease in carboxylation efficiency, but also to a direct impact of O3 on stomatal guard 

25 cell function, leading to a changes in stomatal conductance. 

26 Alterations in gene transcription that have been observed in O3-treated plants are now 

27 evaluated more comprehensively using DNA microarray studies, which measure changes 

28 in the entire transcriptome rather than measuring the transcript levels of individual genes. 

29 These studies have demonstrated very consistent trends, even though O3 exposure 

30 conditions (concentration, duration of exposure), plant species and sampling times vary 

31 significantly. Genes involved in plant defense, signaling, and those associated with the 

32 synthesis of plant hormones and secondary metabolism are generally upregulated in 

33 plants exposed to O3, while those related to photosynthesis and general metabolism are 

34 typically downregulated. Proteome studies support these results by demonstrating 

35 concomitant increases or decreases in the proteins encoded by these genes. Transcriptome 

36 analysis has also illuminated the complex interactions that exist between several different 

37 phytohormones and how they modulate plant sensitivity and response to O3. 

38 Experimental evidence suggests that while ethylene and salicylic acid are needed to 
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1 develop O3-induced leaf lesions, jasmonic acid acts antagonistically to ethylene and 

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

3 regulating stomatal aperture, may also act antagonistically to the jasmonic acid signaling 

4 pathway. Changes in the quantity and activity of these phytohormones and the 

5 interactions between them reveal some of the complexity of plant responses to an 

6 oxidative stressor such as O3. 

7 Another critical area of interest is to better understand and quantify the capacity of the 

8 plant to detoxify oxygen radicals using antioxidant metabolites, such as ascorbate and 

9 glutathione, and the enzymes that regenerate them. Ascorbate remains an important focus 

10 of research, and, due to its location in the apoplast in addition to other cellular 

11 compartments, it is regarded as a first line of defense against oxygen radicals formed in 

12 the apoplast. Most studies demonstrate that antioxidant metabolites and enzymes increase 

13 in quantity and activity in plants exposed to O3, indicating that they play an important 

14 role in protecting plants from oxidative stress. However, attempts to quantify the 

15 detoxification capacity of plants have remained unsuccessful, as high quantities of 

16 antioxidant metabolites and enzymes do not always translate into greater protection of the 

17 plant. Considerable variation exists between plant species, different developmental 

18 stages, and the environmental and O3 exposure conditions which plants are exposed to. 

19 As indicated earlier, the described alterations in transcript levels of genes correlate with 

20 observed changes quantity and activity of the enzymes and metabolites involved in 

21 primary and secondary metabolism. In addition to the generalized upregulation of the 

22 antioxidant defense system, photosynthesis typically declines in O3 treated plants. 

23 Declines in C fixation due to reductions in quantity and activity of Rubisco were 

24 extensively described in the 2006 O3 AQCD. More recent studies support these results 

25 and indicate that declines in Rubisco activity may also result from reductions in Rubisco 

26 activase enzyme quantity. Other studies, which have focused on the light reactions of 

27 photosynthesis, demonstrate that plant exposure to O3 results in declines in electron 

28 transport efficiency and a decreased capacity to quench oxidizing radicals. Therefore, the 

29 overall declines in photosynthesis observed in O3-treated plants likely result from 

30 combined impacts on stomatal conductance, carbon fixation and the light reactions. 

31 While photosynthesis generally declines in plants exposed to O3, catabolic pathways such 

32 as respiration have been shown to increase. It has been hypothesized that increased 

33 respiration may result from greater energy needs for defense and repair. Secondary 

34 metabolism is generally upregulated in a variety of species exposed to O3 as a part of a 

35 generalized plant defense mechanism. Some secondary metabolites, such as flavonoids 

36 and polyamines, are of particular interest as they are known to have antioxidant 

37 properties. The combination of decreases in C assimilation and increases in catabolism 
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1 and the production of secondary metabolites would negatively impact plants by 

2 decreasing the energy available for growth and reproduction.

9.4 Nature of Effects on Vegetation and Ecosystems

9.4.1 Introduction

3 Ambient O3 concentrations have long been known to cause visible symptoms, decreases 

4 in photosynthetic rates, decreases in growth and yield of plants as well as many other 

5 effects on ecosystems (U.S. EPA, 2006b, 1996c, 1986, 1978a). Numerous studies have 

6 related O3 exposure to plant responses, with most effort focused on the yield of crops and 

7 the growth of tree seedlings. Many experiments exposed individual plants grown in pots 

8 or soil under controlled conditions to known concentrations of O3 for a segment of 

9 daylight hours for some portion of the plant’s life span. Information in this section also 

10 goes beyond individual plant-scale responses to consider effects at the broader ecosystem 

11 scale, including effects related to ecosystem services. 

12 This section will focus mainly on studies published since the release of the 2006 O3

13 AQCD. However, because much O3 research was conducted prior to the 2006 O3 AQCD, 

14 the present discussion of vegetation and ecosystem response to O3 exposure is largely 

15 based on the conclusions of the 1978, 1986, 1996, and 2006 O3 AQCDs.

9.4.1.1 Ecosystem Scale, Function, and Structure

16 Information presented in this section was collected at multiple spatial scales or levels of 

17 biological organization, ranging from the physiology of a given species to population, 

18 community, and ecosystem investigations. An ecological population is a group of 

19 individuals of the same species and a community is an assemblage of populations of 

20 different species interacting with one another that inhabit an area. For this assessment, 

21 “ecosystem” is defined as the interactive system formed from all living organisms and 

22 their abiotic (physical and chemical) environment within a given area (IPCC, 2007a). The 

23 boundaries of what could be called an ecosystem are somewhat arbitrary, depending on 

24 the focus of interest or study. Thus, the extent of an ecosystem may range from very 

25 small spatial scales or levels of biological organization to, ultimately, the entire Earth 

26 (IPCC, 2007a). All ecosystems, regardless of size or complexity, have interactions and 

27 physical exchanges between biota and abiotic factors, this includes both structural 
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1 (e.g., soil type and food web trophic levels) and functional (e.g., energy flow, 

2 decomposition, nitrification) attributes. 

3 Ecosystems can be described, in part, by their structure, i.e., the number and type of 

4 species present. Structure may refer to a variety of measurements including the species 

5 richness, abundance, community composition and biodiversity as well as landscape 

6 attributes. Competition among and within species and their tolerance to environmental 

7 stressors are key elements of survivorship. When environmental conditions are shifted, 

8 for example, by the presence of anthropogenic air pollution, these competitive 

9 relationships may change and tolerance to stress may be exceeded. Ecosystems may also 

10 be defined on a functional basis. “Function” refers to the suite of processes and 

11 interactions among the ecosystem components and their environment that involve 

12 nutrient and energy flow as well as other attributes including water dynamics and the flux 

13 of trace gases. Plants, via such processes as photosynthesis, respiration, C allocation, 

14 nutrient uptake and evaporation, affect energy flow, C, nutrient cycling and water 

15 cycling. The energy accumulated and stored by vegetation (via photosynthetic C capture) 

16 is available to other organisms. Energy moves from one organism to another through 

17 food webs, until it is ultimately released as heat. Nutrients and water can be recycled. Air 

18 pollution alters the function of ecosystems when elemental cycles or the energy flow are 

19 altered. This alteration can also be manifested in changes in the biotic composition of 

20 ecosystems. 

21 There are at least three levels of ecosystem response to pollutants: (1) the individual 

22 organism and its environment; (2) the population and its environment; and (3) the 

23 biological community composed of many species and their environment (Billings, 1978). 

24 Individual organisms within a population vary in their ability to withstand the stress of 

25 environmental change. The response of individual organisms within a population is based 

26 on their genetic constitution, stage of growth at time of exposure to stress, and the 

27 microhabitat in which they are growing (Levine and Pinto, 1998). The stress range within 

28 which organisms can exist and function determines the ability of the population to 

29 survive.

9.4.1.2 Ecosystem Services

30 Ecosystem structure and function may be translated into ecosystem services. Ecosystem 

31 services are the benefits people obtain from ecosystems (UNEP, 2003). Ecosystems 

32 provide many goods and services that are of vital importance for the functioning of the 

33 biosphere and provide the basis for the delivery of tangible benefits to human society. 
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1 Hassan et al. (2005) define these benefits to include supporting, provisioning, regulating, 

2 and cultural services: 

3  Supporting services are necessary for the production of all other ecosystem 

4 services. Some examples include biomass production, production of 

5 atmospheric O2, soil formation and retention, nutrient cycling, water cycling, 

6 and provisioning of habitat. Biodiversity is a supporting service that is 

7 increasingly recognized to sustain many of the goods and services that humans 

8 enjoy from ecosystems. These provide a basis for three higher-level categories 

9 of services. 

10  Provisioning services, such as products (Gitay et al., 2001), i.e., food 

11 (including game, roots, seeds, nuts and other fruit, spices, fodder), water, fiber 

12 (including wood, textiles), and medicinal and cosmetic products (such as 

13 aromatic plants, pigments). 

14  Regulating services that are of paramount importance for human society such 

15 as (1) C sequestration, (2) climate and water regulation, (3) protection from 

16 natural hazards such as floods, avalanches, or rock-fall, (4) water and air 

17 purification, and (5) disease and pest regulation. 

18  Cultural services that satisfy human spiritual and aesthetic appreciation of 

19 ecosystems and their components including recreational and other nonmaterial 

20 benefits.

21 In the sections that follow, available information on individual, population and 

22 community response to O3 will be discussed. Effects of O3 on productivity and 

23 C sequestration, water cycling, below-ground processes, competition and biodiversity, 

24 and insects and wildlife are considered below and in the context of ecosystem services 

25 where appropriate. 

9.4.2 Visible Foliar Injury and Biomonitoring

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

27 documented over several decades on many tree, shrub, herbaceous, and crop species 

28 (U.S. EPA, 2006b, 1996b, 1984, 1978a). Visible foliar injury symptoms are considered 

29 diagnostic as they have been verified experimentally in exposure-response studies, using 

30 exposure methodologies such as CSTRs, OTCs, and free-air fumigation (see Section 9.2

31 for more detail on exposure methodologies). Several pictorial atlases and guides have 

32 been published, providing details on diagnosis and identification of O3-induced visible 

33 foliar injury on many plant species throughout North America (Flagler, 1998; NAPAP, 

34 1987) and Europe (Innes et al., 2001; Sánchez et al., 2001). Typical visible injury 
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1 symptoms on broad-leaved plants include: stippling, flecking, surface bleaching, bifacial 

2 necrosis, pigmentation (e.g., bronzing), chlorosis, and/or premature senescence. Typical 

3 visible injury symptoms for conifers include: chlorotic banding, tip burn, flecking, 

4 chlorotic mottling, and/or premature senescence of needles. Although common patterns 

5 of injury develop within a species, these foliar lesions can vary considerably between and 

6 within taxonomic groups. Furthermore, the degree and extent of visible foliar injury 

7 development varies from year to year and site to site (Orendovici-Best et al., 2008; 

8 Chappelka et al., 2007; Smith et al., 2003), even among co-members of a population 

9 exposed to similar O3 levels, due to the influence of co-occurring environmental and 

10 genetic factors (Souza et al., 2006; Chappelka et al., 2003; Somers et al., 1998). 

11 Nevertheless, Chappelka et al. (2007) reported that the average incidence of O3-induced 

12 foliar injury was 73% on milkweed observed in the Great Smoky Mountains National 

13 Park in the years 1992-1996. 

14 Although the majority of O3-induced visible foliar injury occurrence has been observed 

15 on seedlings and small plants, many studies have reported visible injury of mature 

16 coniferous trees, primarily in the western U.S. (Arbaugh et al., 1998) and to mature 

17 deciduous trees in eastern North America (Schaub et al., 2005; Vollenweider et al., 2003; 

18 Chappelka et al., 1999a; Chappelka et al., 1999b; Somers et al., 1998; Hildebrand et al., 

19 1996). 

20 It is important to note that visible foliar injury occurs only when sensitive plants are 

21 exposed to elevated O3 concentrations in a predisposing environment. A major modifying 

22 factor for O3-induced visible foliar injury is the amount of soil moisture available to a 

23 plant during the year that the visible foliar injury is being assessed. This is because lack 

24 of soil moisture generally decreases stomatal conductance of plants and, therefore, limits 

25 the amount of O3 entering the leaf that can cause injury (Matyssek et al., 2006; Panek, 

26 2004; Grulke et al., 2003a; Panek and Goldstein, 2001; Temple et al., 1992; Temple et 

27 al., 1988). Consequently, many studies have shown that dry periods in local areas tend to 

28 decrease the incidence and severity of O3-induced visible foliar injury; therefore, the 

29 incidence of visible foliar injury is not always higher in years and areas with higher O3, 

30 especially with co-occurring drought (Smith et al., 2003). Other factors such as leaf age 

31 influence the severity of symptom expression with older leaves showing greater injury 

32 severity as a result of greater seasonal exposure (Zhang et al., 2010a).

33 Although visible injury is a valuable indicator of the presence of phytotoxic 

34 concentrations of O3 in ambient air, it is not always a reliable indicator of other negative 

35 effects on vegetation. The significance of O3 injury at the leaf and whole plant levels 

36 depends on how much of the total leaf area of the plant has been affected, as well as the 

37 plant’s age, size, developmental stage, and degree of functional redundancy among the 
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1 existing leaf area. Previous O3 AQCDs have noted the difficulty in relating visible foliar 

2 injury symptoms to other vegetation effects such as individual plant growth, stand 

3 growth, or ecosystem characteristics (U.S. EPA, 2006b, 1996b). As a result, it is not 

4 presently possible to determine, with consistency across species and environments, what 

5 degree of injury at the leaf level has significance to the vigor of the whole plant. 

6 However, in some cases, visible foliar symptoms have been correlated with decreased 

7 vegetative growth (Somers et al., 1998; Karnosky et al., 1996; Peterson et al., 1987; 

8 Benoit et al., 1982) and with impaired reproductive function (Chappelka, 2002; Black et 

9 al., 2000). Conversely, the lack of visible injury does not always indicate a lack of 

10 phytotoxic concentrations of O3 or a lack of non-visible O3 effects (Gregg et al., 2006, 

11 2003). 

9.4.2.1 Biomonitoring

12 The use of biological indicators to detect phytotoxic levels of O3 is a longstanding and 

13 effective methodology (Chappelka and Samuelson, 1998; Manning and Krupa, 1992). A 

14 plant bioindicator can be defined as a vascular or nonvascular plant exhibiting a typical 

15 and verifiable response when exposed to a plant stress such as an air pollutant (Manning, 

16 2003). To be considered a good indicator species, plants must (1) exhibit a distinct, 

17 verified response; (2) have few or no confounding disease or pest problems; and (3) 

18 exhibit genetic stability (U.S. EPA, 2006b). Such sensitive plants can be used to detect 

19 the presence of a specific air pollutant such as O3 in the ambient air at a specific location 

20 or region and, as a result of the magnitude of their response, provide unique information 

21 regarding specific ambient air quality. Bioindicators can be either introduced sentinels, 

22 such as the widely used tobacco (Nicotiana tabacum) variety Bel W3 (Calatayud et al., 

23 2007b; Laffray et al., 2007; Nali et al., 2007; Gombert et al., 2006; Kostka-Rick and 

24 Hahn, 2005; Heggestad, 1991) or detectors, which are sensitive native plant species 

25 (Chappelka et al., 2007; Souza et al., 2006). The approach is especially useful in areas 

26 where O3 monitors are not operated (Manning, 2003). For example, in remote wilderness 

27 areas where instrument monitoring is generally not available, the use of bioindicator 

28 surveys in conjunction with the use of passive samplers (Krupa et al., 2001) may be a 

29 useful methodology (Manning, 2003). However, it requires expertise in recognizing those 

30 signs and symptoms uniquely attributable to exposure to O3 as well as in their 

31 quantitative assessment.

32 Since the 2006 O3 AQCD, new sensitive plant species have been identified from field 

33 surveys and verified in controlled exposure studies (Kline et al., 2009; Kline et al., 2008). 

34 Several multiple-year field surveys have also been conducted at National Wildlife 
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1 Refuges in Maine, Michigan, New Jersey, and South Carolina (Davis, 2009, 2007a, b; 

2 Davis and Orendovici, 2006). 

3 The USDA Forest Service through the Forest Health Monitoring Program (FHM) (1990 -

4 2001) and currently the Forest Inventory and Analysis (FIA) Program has been collecting 

5 data regarding the incidence and severity of visible foliar injury on a variety of O3

6 sensitive plant species throughout the U.S. (Coulston et al., 2003; Smith et al., 2003). The 

7 plots where these data are taken are known as biosites. These biosites are located 

8 throughout the country and analysis of visible foliar injury within these sites follows a set 

9 of established protocols. For more details, see http://www.nrs.fs.fed.us/fia/topics/ozone/

10 (USDA, 2011). The network has provided evidence of O3 concentrations high enough to 

11 induce visible symptoms on sensitive vegetation. From repeated observations and 

12 measurements made over a number of years, specific patterns of areas experiencing 

13 visible O3 injury symptoms can be identified. (Coulston et al., 2003) used information 

14 gathered over a 6-year period (1994-1999) from the network to identify several species 

15 that were sensitive to O3 over entire regions, including sweetgum (Liquidambar 

16 styraciflua), loblolly pine (Pinus taeda), and black cherry (P. serotina). In a study of the 

17 west coast of the U.S, Campbell et al. (2007) reported O3 injury in 25-37% of biosites in 

18 California forested ecosystems from 2000-2005.

19 A study by Kohut (2007) assessed the estimated risk of O3-induced visible foliar injury 

20 on bioindicator plants (NPS, 2006) in 244 national parks in support of the National Park 

21 Service’s Vital Signs Monitoring Network (NPS, 2007). The risk assessment was based 

22 on a simple model relating response to the interaction of species, level of O3 exposure, 

23 and exposure environment. Kohut (2007) concluded that the estimated risk of visible 

24 foliar injury was high in 65 parks (27%), moderate in 46 parks (19%), and low in 131 

25 parks (54%). Some of the well-known parks with a high risk of O3-induced visible foliar 

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

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

28 Sleeping Bear Dunes, Great Smoky Mountains, Joshua Tree, Sequoia and Kings Canyon, 

29 and Yosemite.

30 Lichens have also long been used as biomonitors of air pollution effects on forest health 

31 (Nash, 2008). It has been suspected, based on field surveys in the San Bernardino 

32 Mountains surrounding the Los Angeles air basin, that declines in lichen diversity and 

33 abundance were correlated with measured O3 gradients (Gül et al., 2011). Several recent 

34 studies in North America (Geiser and Neitlich, 2007; Gombert et al., 2006; Jovan and 

35 McCune, 2006) and Europe (Nali et al., 2007; Gombert et al., 2006) have used lichens as 

36 biomonitors of atmospheric deposition (e.g., N and S) and O3 exposure. Nali et al. (2007)

37 found that epiphytic lichen biodiversity was not related to O3 geographical distribution. 
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1 In addition, a recent study by Riddell et al. (2010) found that lichen species, Ramalina 

2 menziesii, showed no decline in physiological response to low and moderate 

3 concentrations of O3 and may not be a good indicator for O3 pollution. Mosses have also 

4 been used as biomonitors of air pollution; however, there remains a knowledge gap in the 

5 understanding of the effects of ozone on mosses as there has been very little information 

6 available on this topic in recent years.

9.4.2.2 Summary

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

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

9 species (U.S. EPA, 2006b, 1996b, 1984, 1978a). Ozone-induced visible foliar injury 

10 symptoms on certain bioindicator plant species are considered diagnostic as they have 

11 been verified experimentally in exposure-response studies, using exposure methodologies 

12 such as continuous stirred tank reactors (CSTRs), OTCs, and free-air fumigation. 

13 Experimental evidence has clearly established a consistent association of visible injury 

14 with O3 exposure, with greater exposure often resulting in greater and more prevalent 

15 injury. Since the 2006 O3 AQCD, results of several multi-year field surveys of 

16 O3-induced visible foliar injury at National Wildlife Refuges in Maine, Michigan, New 

17 Jersey, and South Carolina have been published. New sensitive species showing visible 

18 foliar injury continue to be identified from field surveys and verified in controlled 

19 exposure studies.

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

21 longstanding and effective methodology. The USDA Forest Service through the Forest 

22 Health Monitoring (FHM) Program (1990-2001) and currently the Forest Inventory and 

23 Analysis (FIA) Program has been collecting data regarding the incidence and severity of 

24 visible foliar injury on a variety of O3 sensitive plant species throughout the U.S. The 

25 network has provided evidence that O3 concentrations were high enough to induce visible 

26 symptoms on sensitive vegetation. From repeated observations and measurements made 

27 over a number of years, specific patterns of areas experiencing visible O3 injury 

28 symptoms can be identified. As noted in the preceding section, a study of 244 national 

29 parks indicated that the estimated risk of visible foliar injury was high in 65 parks (27%), 

30 moderate in 46 parks (19%), and low in 131 parks (54%). 

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

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

33 vegetation across the U.S.
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9.4.3 Growth, Productivity and Carbon Storage in Natural Ecosystems

1 Ambient O3 concentrations have long been known to cause decreases in photosynthetic 

2 rates, decreases in growth, and decreases in yield (U.S. EPA, 2006b, 1996c, 1986, 

3 1978a). The O3-induced damages at the plant scale may translate to damages at the stand, 

4 then ecosystem scales, and cause changes in productivity and C storage. This section 

5 focuses on the responses of C cycling to seasonal or multi-year exposures to O3 at levels 

6 of organization ranging from individual plants to ecosystems. Quantitative responses 

7 include changes in plant growth, plant biomass allocation, ecosystem production and 

8 ecosystem C sequestration. Most information available on plant-scale responses was 

9 obtained from studies that used a single species especially tree seedlings and crops, while 

10 some used mixtures of herbaceous species. Ecosystem changes are difficult to evaluate in 

11 natural settings, due to the complexity of interactions, the number of potential 

12 confounders, and the large spatial and temporal scales. The discussion of ecosystem 

13 effects focuses on new studies at the large-scale FACE experiments and on ecological 

14 model simulations.

9.4.3.1 Plant Growth and Biomass Allocation

15 The previous O3 AQCDs concluded that there is strong evidence that exposure to O3

16 decreases photosynthesis and growth in numerous plant species (U.S. EPA, 2006b, 

17 1996b, 1984, 1978a). Studies published since the last review support those conclusions 

18 and are summarized below. 

19 In general, research conducted over several decades has indicated that exposure to O3 

20 alters stomatal conductance and reduces photosynthesis in a wide variety of plant species. 

21 In a review of more than 55 studies, Wittig et al. (2007) reported that current O3

22 concentrations in the northern hemisphere are decreasing stomatal conductance (13%) 

23 and photosynthesis (11%) across tree species. It was also found that younger trees (<4 

24 years) were affected less by O3 than older trees. Further, the authors also found that 

25 decreases in photosynthesis are consistent with the cumulative uptake of O3 into the leaf. 

26 In contrast, several studies reported that O3 exposure may result in loss of stomatal 

27 control, incomplete stomatal closure at night and a decoupling of photosynthesis and 

28 stomatal conductance, which may have implications for whole- plant water use 

29 (Section 9.4.5).

30 In a recently published meta-analysis, Wittig et al. (2009) quantitatively compiled peer 

31 reviewed studies from the past 40 years on the effect of current and future O3 exposures 

32 on the physiology and growth of forest species. They found that current ambient O3
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1 concentrations as reported in those studies significantly decreased annual total biomass 

2 growth (7%) across 263 studies. The authors calculated the ambient O3 concentrations 

3 across these studies to average 40 ppb. This average was calculated across the duration of 

4 each study and there were therefore many hourly exposures well above 40 ppb. The 

5 decreased growth effect was reported to be greater (11 to 17%) in elevated O3 exposures 

6 (97 ppb) (Wittig et al., 2009). This meta-analysis demonstrates the coherence of O3

7 effects across numerous studies and species that used a variety of experimental 

8 techniques, and these results support the conclusion of the previous AQCD that exposure 

9 to O3 decreases plant growth.

10 In two companion papers, McLaughlin et al. (2007a); (2007b) investigated the effects of 

11 ambient O3 on tree growth and hydrology at forest sites in the southern Appalachian 

12 Mountains. The authors reported that the cumulative effects of ambient levels of O3

13 decreased seasonal stem growth by 30-50% for most tree species in a high O3 year in 

14 comparison to a low O3 year (McLaughlin et al., 2007a). The authors also reported that 

15 high ambient O3 concentrations can disrupt whole-tree water use and in turn reduce late16 season streamflow (McLaughlin et al., 2007b); see Section 9.4.5 for more on water 

17 cycling. 

18 Since the 2006 O3 AQCD, several recent studies have reported results from the Aspen 

19 FACE “free air” O3 and CO2 exposure experiment in Wisconsin (Darbah et al., 2008; 

20 Riikonen et al., 2008; Darbah et al., 2007; Kubiske et al., 2007; Kubiske et al., 2006; 

21 King et al., 2005). At the Aspen FACE site, single-species and two-species stands of trees 

22 were grown in 12, 30-m diameter rings corresponding to three replications of a full 

23 factorial arrangement of two levels each of CO2 and O3 exposure. Over the first 

24 seven years of stand development, Kubiske et al. (2006) observed that elevated O3

25 decreased tree heights, diameters, and main stem volumes in the aspen community by 11, 

26 16, and 20%, respectively. In addition, Kubiske et al. (2007) reported that elevated O3

27 may change intra- and inter-species competition. For example, O3 treatments increased 

28 the rate of conversion from a mixed aspen-birch community to a birch dominated 

29 community. In a comparison presented in Section 9.6.3 of this document, EPA found that 

30 effects on biomass accumulation in aspen during the first seven years closely agreed with 

31 the exposure-response function based on data from earlier OTC experiments. 

32 Several studies at the Aspen FACE site also considered other growth-related effects of 

33 elevated O3. Darbah et al. (2008); Darbah et al. (2007) reported that O3 treatments 

34 decreased paper birch seed weight and seed germination and that this would likely lead to 

35 a negative impact of regeneration for that species. Riikonen et al. (2008) found that 

36 elevated O3 decreased the amount of starch in birch buds by 16%, and reduced aspen bud 

37 size, which may have been related to the observed delay in spring leaf development. The 
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1 results suggest that elevated O3 concentrations have the potential to alter C metabolism of 

2 overwintering buds, which may have carry-over effects in the subsequent growing season 

3 (Riikonen et al., 2008).

4 Effects on growth of understory vegetation were also investigated at Aspen FACE. 

5 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 6 N recovery were a 

7 result of overstory community responses to those treatments; however, the lack of 

8 apparent direct O3 treatment effects may have been due to high variability in the data. 

9 Total understory biomass increased with increasing light and was greatest under the open 

10 canopy of the aspen/maple community, as well as the more open canopy of the elevated 

11 O3 treatments (Bandeff et al., 2006). Similarly, data from a study by Awmack et al. 

12 (2007) suggest that elevated CO2 and O3 may have indirect growth effects on red 

13 (Trifolium pratense) and white (Trifolium repens) clover in the understory via overstory 

14 community effects; however, no direct effects of elevated O3 were observed. 

15 Overall, the studies at the Aspen FACE experiment are consistent with many of the OTC 

16 studies that were evaluated in previous O3 AQCDs demonstrating that O3 exposure 

17 decreases growth in numerous plant species. These results strengthen the understanding 

18 of O3 effects on forests and demonstrate the relevance of the knowledge gained from 

19 trees grown in open-top chamber studies.

20 For some annual species, particularly crops, the relevant measurement for an assessment 

21 of the risk of O3 exposure is yield or growth, e.g., production of grain or biomass. For 

22 plants grown in mixtures such as hayfields, and natural or semi-natural grasslands 

23 (including native nonagricultural species), affected factors other than production of 

24 biomass may be important. Such endpoints include biodiversity or species composition, 

25 and effects on those endpoints may be indirect, resulting, for example, from competitive 

26 interactions among plants in mixed-species communities. Most of the available data on 

27 non-crop herbaceous species are for grasslands, with many of the recent studies 

28 conducted in Europe. See Section 9.4.7 for a review of the recent literature on O3 effects 

29 on competition and biodiversity in grasslands.

Root growth

30 Although O3 does not penetrate soil, it could alter root development by decreasing 

31 C assimilation via photosynthesis leading to less C allocation to the roots (Andersen, 

32 2003). The response of root development to O3 exposure depends on available

33 photosynthate within the plant and could vary over time. Many biotic and abiotic factors, 

34 such as community dynamics and drought stress, have been found to alter root 
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1 development under elevated O3. Generally, there is clear evidence that O3 reduces C 

2 allocation to roots; however, results of a few recent individual studies have shown 

3 negative (Jones et al., 2010), non-significant (Andersen et al., 2010; Phillips et al., 2009)

4 and positive effects (Pregitzer et al., 2008; Grebenc and Kraigher, 2007) on root biomass 

5 and root: shoot ratio. 

6 An earlier study at the Aspen FACE experiment found that elevated O3 reduced coarse 

7 root and fine roots biomass in young stands of paper birch and trembling aspen (King et 

8 al., 2001). However, this reduction disappeared several years later. Ozone significantly 

9 increased fine-root (<1.0 mm) in the aspen community (Pregitzer et al., 2008). This 

10 increase in fine root production was due to changes in community composition, such as 

11 better survival of the O3-tolerant aspen genotype, birch, and maple, rather than changes in 

12 C allocation at the individual tree level (Pregitzer et al., 2008; Zak et al., 2007). In an 

13 adult European beech/Norway spruce forest in Germany, drought was found to nullify the 

14 O3-driven stimulation of fine root growth. Ozone stimulated fine-root production of 

15 beech during the humid year, but had no significant impact on fine root production in the 

16 dry year (Matyssek et al., 2010; Nikolova et al., 2010).

17 Using a non-destructive method, Vollsnes et al. (2010) studied the in vivo root 

18 development of subterranean clover (Trifolium subterraneum) before, during and after 

19 short-term O3 exposure. It was found that O3 reduced root tip formation, root elongation, 

20 the total root length, and the ratios between below- and above-ground growth within 

21 one week after exposure. Those effects persisted for up to three weeks; however, biomass 

22 and biomass ratios were not significantly altered at the harvest five weeks after exposure.

23 Several recent meta-analyses have generally indicated that O3 reduced C allocated to 

24 roots. In one meta-analysis, Grantz et al. (2006) estimated the effect of O3 on the 

25 root:shoot allometric coefficient (k), the ratio between the relative growth rate of the root 

26 and shoot. The results showed that O3 reduced the root:shoot allometric coefficient by 

27 5.6%, and the largest decline of the root:shoot allometric coefficient was observed in 

28 slow-growing plants. In another meta-analysis including 263 publications, Wittig et al. 

29 (2009) found that current O3 exposure had no significant impacts on root biomass and 

30 root:shoot ratio when compared to pre-industrial O3 exposure. However, if O3

31 concentrations rose to 81-101 ppb (projected O3 levels in 2100), both root biomass and 

32 root:shoot ratio were found to significantly decrease. Gymnosperms and angiosperms 

33 differed in their responses, with gymnosperms being less sensitive to elevated O3. In two 

34 other meta-analyses, Wang and Taub (2010) found elevated O3 reduced biomass 

35 allocation to roots by 8.3% at ambient CO2 and 6.0% at elevated CO2, and Morgan et al. 

36 (2003) found O3 reduced root dry weight of soybean. 
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9.4.3.2 Summary

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

2 ambient concentrations of O3 decrease photosynthesis and growth in numerous plant 

3 species across the U.S. Studies published since the last review continue to support that 

4 conclusion.

5 The meta-analyses by Wittig et al. (2009); Wittig et al. (2007) demonstrate the coherence 

6 of O3 effects on plant photosynthesis and growth across numerous studies and species 

7 using a variety of experimental techniques. Furthermore, recent meta-analyses have 

8 generally indicated that O3 reduced C allocation to roots (Wittig et al., 2009; Grantz et 

9 al., 2006). Since the 2006 O3 AQCD, several studies were published based on the Aspen

10 FACE experiment using “free air,” O3, and CO2 exposures in a planted forest in 

11 Wisconsin. Overall, the studies at the Aspen FACE experiment were consistent with 

12 many of the open-top chamber (OTC) studies that were the foundation of previous O3

13 NAAQS reviews. These results strengthen the understanding of O3 effects on forests and 

14 demonstrate the relevance of the knowledge gained from trees grown in open-top 

15 chamber studies.

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

17 O3 exposure and reduced growth of native woody and herbaceous vegetation.

9.4.3.3 Reproduction

18 Studies during recent decades have demonstrated O3 effects on various stages of plant 

19 reproduction. The impacts of O3 on reproductive development, as reviewed by Black et 

20 al. (2000), can occur by influencing (1) age at which flowering occurs, particularly in 

21 long-lived trees that often have long juvenile periods of early growth without flower and 

22 seed production; (2) flower bud initiation and development; (3) pollen germination and 

23 pollen tube growth; (4) seed, fruit, or cone yields; and (5) seed quality (Table 9-1) (U.S. 

24 EPA, 2006b). Several recent studies since the 2006 O3 AQCD further demonstrate the 

25 effects of O3 on reproductive processes in herbaceous and woody plant species. Although 

26 there have been documented effects of ozone on reproductive processes, a knowledge gap 

27 still exists pertaining to the exact mechanism of these responses. 

28 Ramo et al. (2007) exposed several meadow species to elevated O3 (40-50 ppb) and CO2

29 (+100 ppm), both individually and combined, over three growing seasons in ground30 planted mesocosms, using OTCs. Elevated O3 delayed flowering of Campanula 

31 rotundifolia and Vicia cracca. Ozone also reduced the overall number of produced 

32 flowers and decreased fresh weight of individual Fragaria vesca berries. 
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1 Black et al. (2007) exposed Brassica campestris to 70 ppb for two days during late 

2 vegetative growth or ten days during most of the vegetative phase. The two-day exposure 

3 had no effect on growth or reproductive characteristics, while the 10 day exposure 

4 reduced vegetative growth and reproductive site number on the terminal raceme, 

5 emphasizing the importance of exposure duration and timing. Mature seed number and 

6 weight per pod were unaffected due to reduced seed abortion, suggesting that, although 

7 O3 affected reproductive processes, indeterminate species such as B. campestris possess 

8 enough compensatory flexibility to avoid reduced seed production Black et al. (2007). 

9 In the determinate species, Plantago major, Black et al. (2010) found that O3 may have 

10 direct effects on reproductive development in populations of differing sensitivity. Only 

11 the first flowering spike was exposed to 120 ppb O3 for 7 hours per day on 9 successive 

12 days (corresponding to flower development) while the leaves and second spike were 

13 exposed to charcoal-filtered air. Exposure of the first spike to O3 affected seed number 

14 per capsule on both spikes even though spike two was not exposed. The combined seed 

15 weight of spikes one and two was increased by 19% in the two resistant populations, 

16 suggesting an overcompensation for injury; whereas, a decrease of 21% was observed in 

17 the most sensitive population (Black et al., 2010). The question remains as to whether 

18 these effects are true direct ozone-induced effects or compensatory responses.

19 Studies by Darbah et al. (2008); Darbah et al. (2007) of paper birch (Betula papyrifera) 

20 trees at the Aspen FACE site in Rhinelander, WI investigated the effects of elevated O3

21 and/or CO2 on reproductive fitness. Elevated O3 increased flowering, but decreased seed 

22 weight and germination success rate of seeds from the exposed trees. These results 

23 suggest that O3 can dramatically affect flowering, seed production, and seed quality of 

24 paper birch, ultimately affecting its reproductive fitness (Darbah et al., 2008; Darbah et 

25 al., 2007).
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Table 9-1 Ozone effects on plant reproductive processes

Species Condition Measures References

Apocynum androsaemifolium 

(spreading dogbane)

Flowering time Bergweiler and Manning (1999)

Buddleia davidii 

(butterfly bush)

Flowering time Findley et al. (1997)

Rubus cuneifolius 

(sand blackberry)

Pollen germination Chappelka (2002)

Plantago major 

(plantain)

Pollen tube elongation Stewart (1998)

Fragaria × ananassa 

(cultivated strawberry)

Fruit yield Drogoudi and Ashmore (2001); Drogoudi and Ashmore (2000)

Plantago major 

(plantain)

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: Derived from Table AX9-22 of the 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 from 

3 long-term FACE experiments provide more evidence of the association of O3 exposure 

4 and changes in productivity at the ecosystem level of organization. In addition to 

5 experimental studies, model studies also assessed the impact of O3 exposure on 

6 productivity and C sequestration from stand to global scales. 

7 In this section productivity of ecosystems is expressed in different ways depending on the 

8 model or the measurements of a study. The most common metric of productivity is Gross 

9 Primary Productivity. Gross Primary Productivity (GPP) is total carbon that enters the 

10 ecosystem through photosynthesis by plants. Plants return a larger portion of this carbon 

11 back to the atmosphere through respiration from roots and aboveground portions of plants 

12 (Rplant). Net primary production (NPP) is the difference between total carbon gain (GPP) 

and carbon loss through Rplant 13 . Net ecosystem productivity (NEP) is the difference 

14 between NPP and carbon loss through heterotrophic respiration (Rhet) (mostly 

15 decomposition of dead organic matter) (Lambers et al., 1998). Similarly net ecosystem 

16 exchange (NEE) is the net flux of carbon between the land and the atmosphere, typically 

17 measured using eddy covariance techniques. Positive values of NEE usually refer to 

18 carbon released to the atmosphere (i.e., a source), and negative values refer to carbon 

19 uptake (i.e., a sink). Other studies have calculated net carbon exchange (NCE). NCE is 

defined as NPP minus Rhet 20 , Ec (the carbon emission during the conversion of natural 

21 ecosystems to agriculture) and Ep (the sum of carbon emission from the decomposition of 
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1 agricultural products). For natural vegetation, Ec and Ep are equal to 0, so NCE is equal 

2 NEP (Felzer et al., 2005). In general, modeling studies take into account the effect of O3

on C fixation of a system and there is generally not an effect on Rplant, Rhet 3 , Ec or Ep. 

4 Therefore, decreases in GPP, NPP, NEP, NEE and NCE indicate a general decrease in 

5 productivity of an ecosystem.

6 Two types of models are most often used to study the ecological consequences of O3

7 exposure: (1) single plant growth models such as TREGRO and PnET–II (Hogsett et al.,

8 2008; Martin et al., 2001; Ollinger et al., 1997b), and (2) process-based ecosystem 

9 models such as PnET-CN, Dynamic Land Ecosystem Model (DLEM), Terrestrial 

10 Ecosystem Model (TEM), or Met Office Surface Exchange Scheme - Top-down 

11 Representation of Interactive Foliage and Flora Including Dynamics (MOSES-TRIFFID) 

12 (Felzer et al., 2009; Ren et al., 2007b; Sitch et al., 2007; Ollinger et al., 2002)

13 (Table 9-2). In these models, carbon uptake is simulated through photosynthesis 

14 (TREGRO, PnET –II, PnET- CN, DLEM and MOSES-TRIFFID) or gross primary 

15 production (TEM). Photosynthesis rate at leaf level is modeled by a function of stomatal 

16 conductance and other parameters in TREGRO, PnET –II, PnET- CN, DLEM and 

17 MOSES-TRIFFID. Photosynthesis at canopy level is calculated by summing either 

18 photosynthesis of different leaf types (TREGRO, DLEM, and MOSES-TRIFFID) or 

19 photosynthesis of different canopy layers (PnET –II, PnET- CN). The detrimental effect 

20 of O3 on plant growth is often simulated by multiplying photosynthesis rate by a 

21 coefficient that is dependent on stomatal conductance and cumulative O3 uptake

22 (Table 9-2). Different plant functional groups (PFTs, such as deciduous trees, coniferous 

23 trees or crops) show different responses to O3 exposure. PnET-iI, PnET-CN, TEM, 

24 DLEM and MOSES-TRIFFID estimate this difference by modifying net photosynthesis 

25 with coefficients that represent the O3 induced fractional reduction of photosynthesis for 

26 each functional group. The coefficients used in PnET-iI, PnET-CN, TEM, DLEM are 

27 derived from the functions of O3 exposure (AOT40) versus photosynthesis reduction 

28 from Reich (1987) and Tjoelker et al. (1995). The coefficients used in MOSES-TRIFFID 

29 are derived from the O3 dose-photosynthesis response function from Pleijel et al. (2004a)

30 and Karlsson et al. (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 31 /sec for 

woody PFT and 5 nmol/m2 32 /sec for grass PFT, and is different from AOT40, which has an 

33 O3 threshold level of 40 ppb for all PFTs. Experimental and model studies on ecosystem 

34 productivity and C sequestration at the forest stand scale as well as regional and global 

35 scales are reviewed in the following section.
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Table 9-2 Comparison of models used to simulate the ecological 

consequences of ozone 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 of 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 

time-step, 

single 

plant 

model 

PnET –

CN: 

Monthly 

time-step, 

ecosystem 

model

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, evenmass 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); 

Ollinger et 

al. (1997b); 

Pan et al. 

(2009)

TEM Monthly 

time-step, 

ecosystem 

model

Ecosystem: TEM is run at a 0.5*0.5 degree 

resolution. Each grid cell is classified by 

vegetation type and soil texture, and vegetation 

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 

time-step 

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. 

(2007b); 

(Ren et al., 

2007a); 

Zhang et al. 

(2007a)

MOSESTRIFFID30 minute 

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 Both experimental and modeling studies have provided new information on effects of O3

2 exposure at the stand or site level, i.e., at the local scale. The above- and below-ground 

3 biomass and net primary production (NPP) were measured at the Aspen FACE site after 7 

4 years of O3 exposure. Elevated O3 caused 23, 13 and 14% reductions in total biomass 

5 relative to the control in the aspen, aspen–birch and aspen–maple communities, 

6 respectively (King et al., 2005). At the Kranzberg Forest FACE experiment in Germany, 

O3 reduced annual volume growth by 9.5 m3 7 /ha in a mixed mature stand of Norway 

8 spruce and European beech (Pretzsch et al., 2010). At the grassland FACE experiment at 

9 Alp Flix, Switzerland, O3 reduced the seasonal mean rates of ecosystem respiration and 

10 GPP by 8%, but had no significant impacts on aboveground dry matter productivity or 

11 growing season net ecosystem production (NEP) (Volk et al., 2011). Ozone also altered 

12 C accumulation and turnover in soil, as discussed in Section 9.4.6.

13 Changes in forest stand productivity under elevated O3 were assessed by several model 

14 studies. TREGRO (Table 9-2) has been widely used to simulate the effects of O3 on the 

15 growth of several species in different regions in the U.S. Hogsett et al. (2008) used 

16 TREGRO to evaluate the effectiveness of various forms and levels of air quality 

17 standards for protecting tree growth in the San Bernardino Mountains of California. They 

18 found that O3 exposures at the Crestline site resulted in a mean 20.9% biomass reduction 

19 from 1980 to 1985 and 10.3% biomass reduction from 1995 to 2000, compared to the 

20 ‘‘background’’ O3 concentrations (O3 concentration in Crook County, Oregon). The level 

21 of vegetation protection projected was different depending on the air quality scenarios

22 under consideration. Specifically, when air quality was simulated to just meet the 

23 California 8-h average maximum of 70 ppb and the maximum 3 months 12-h SUM06 of 

24 25 ppm-h, annual growth reductions were limited to 1% or less, while air quality that just 

25 met a previous NAAQS (the second highest 1-h max [125 ppb]) resulted in 6-7% annual 

26 reduction in growth, resulting in the least protection relative to background O3 (Hogsett et 

27 al., 2008).

28 ZELIG is a forest succession gap model, and has been used to evaluate the dynamics of 

29 natural stand succession. Combining TREGRO with ZELIG, Weinstein et al. (2005)

30 simulated the effects of different O3 levels (0.5, 1.5, 1.75, and 2 times [×] ambient) on the 

31 growth and competitive interactions of white fir and ponderosa pine at three sites in 

32 California: Lassen National Park, Yosemite National Park, and Crestline. Their results 

33 suggested that O3 had little impact on white fir, but greatly reduced the growth of 

34 ponderosa pine. If current O3 concentrations continue over the next century, ambient O3

35 exposure (SUM06 of 110 ppm-h) at Crestline was predicted to decrease individual tree 

36 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 To evaluate the influence of interspecies competition on O3 effects, the linked TREGRO 

4 and ZELIG modeling system was used to predict the effects of O3 over 100 years on the 

5 basal area of species in a Liriodendron tulipifera-dominated forest in the Great Smoky 

6 Mountains National Park (Weinstein et al., 2001). Ambient O3 was predicted to decrease 

7 individual tree C budget by 28% and reduce the basal area of L. tulipifera by 10%, 

8 whereas a 1.5×-ambient exposure was predicted to cause a 42% decrease in the individual 

9 tree C budget and a 30% reduction in basal area. Individual tree C balance for Acer 

10 rubrum decreased 14% and 23% under ambient and 1.5×-ambient exposure, respectively. 

11 Prunus serotina was predicted to have less than a 2% decrease in tree C balance in all 

12 scenarios, but its basal area was greatly altered by the O3 effects on the other tree species. 

13 Basal area of A. rubrum and P. serotina was predicted to increase for some years, but 

14 then decrease by up to 30%, depending on the scenario. The effects of O3 on stand 

15 productivity and dynamics were also studied by other tree growth or stand models, such 

16 as ECOPHYS, INTRASTAND and LINKAGES. ECOPHYS is a functional-structural 

17 tree growth model. The model used the linear relationship between the maximum 

18 capacity of carboxylation and O3 dose to predict the relative effect of O3 on leaf 

19 photosynthesis (Martin et al., 2001). Simulations with ECOPHYS found that O3

20 decreased stem dry matter production, stem diameter and leaf dry matter production, 

21 induced earlier leaf abscission, and inhibited root growth (Martin et al., 2001). 

22 INTRASTAND is an hourly time step model for forest stand carbon and water budgets. 

23 LINKAGES is a monthly time step model simulating forest growth and community 

24 dynamics. Linking INTRASTAND with LINKAGES, Hanson et al. (2005) found that a 

25 simulated increase in O3 concentration in 2100 (a mean 20-ppb increase over the current 

26 O3 concentration) yields a 35% loss of net ecosystem C exchange (NEE) with respect to 

the current conditions (174 g C/m2 27 /year). 

Regional and global scales

28 Since the publication of the 2006 O3 AQCD, there is additional evidence suggesting that 

29 O3 exposure alters ecosystem productivity and biogeochemical cycling at the regional 

30 scale, i.e., at scales ranging from watershed to subcontinental divisions, and at continental 

31 and global scales. Most of those studies were conducted by using process-based 

32 ecosystem models (Table 9-2) and are briefly reviewed in the following sections.

33 Ollinger et al. (1997a) simulated the effect of O3 on hardwood forest productivity of 64 

34 hardwood sites in the northeastern U.S. with PnET-iI (Table 9-2). Their simulations 

35 indicated that O3 caused a 3-16% reduction in NPP from 1987 to 1992 (Table 9-3). The 
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1 interactive effects of O3, N deposition, elevated CO2 and land use history on C dynamics 

2 were estimated by PnET-CN (Table 9-2) (Ollinger et al., 2002). The results indicated that 

3 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 4 /year) under 

5 timber harvest site history. PnET-CN was also used to assess changes in C sequestration 

6 of U.S. Mid-Atlantic temperate forest. Pan et al. (2009) designed a factorial modeling 

7 experiment to separate the effects of changes in atmospheric composition, historical 

8 climatic variability and land-disturbances on the C cycle. They found that O3 acted as a 

9 negative factor, partially offsetting the growth stimulation caused by elevated CO2 and 

10 N deposition in U.S. Mid-Atlantic temperate forest. Ozone decreased NPP of most forest 

11 types by 7-8%. Among all the forest types, spruce-fir forest was most resistant to O3

12 damage, and NPP decreased by only 1% (Pan et al., 2009). 

13 Felzer et al. (2004) developed TEM 4.3 (Table 9-2) to simulate the effects of O3 on plant 

14 growth and estimated effects of O3 on NPP and C sequestration of deciduous trees, 

15 conifers and crops in the conterminous U.S. The results indicated that O3 reduced NPP 

16 and C sequestration in the U.S. (Table 9-3) with the largest decreases (over 13% in some 

17 locations) in NPP occurring in the Midwest agricultural lands during the mid-summer. 

18 TEM was also used to evaluate the magnitude of O3 damage at the global scale 

19 (Table 9-3) (Felzer et al., 2005). Simulations for the period 1860 to 1995 show that the 

20 largest reductions in NPP and net C exchange occurred in the mid western U.S., eastern 

21 Europe, and eastern China (Felzer et al., 2005). DLEM (Table 9-2) was developed to 

22 simulate the detrimental effect of O3 on ecosystems, and has been used to examine the O3

23 damage on NPP and C sequestration in Great Smoky Mountains National Park (Zhang et 

24 al., 2007a), grassland ecosystems and terrestrial ecosystems in China (Ren et al., 2007b; 

25 Ren et al., 2007a). Results of those simulations are listed in Table 9-3. 

26 Instead of using AOT40 as their O3 exposure metric as PnET, TEM and DLEM did, Sitch 

27 et al. (2007) incorporated a different O3 metric named CUOt (cumulative stomatal uptake 

28 of O3), derived from Pleijel et al. (2004a), into the MOSES-TRIFFID coupled model

29 (Table 9-2). In the CUOt metric, the fractional reduction of plant production is dependent 

30 on O3 uptake by stomata over a critical threshold for damage with this threshold level 

31 varying by plant functional type. Consistent with previous studies, their model simulation 

32 indicated that O3 reduced global gross primary production (GPP), C-exchange rate and 

33 C sequestration (Table 9-3). The largest reductions in GPP and land-C storage were 

34 projected over North America, Europe, China and India. In the model, reduced ecosystem 

35 C uptake due to O3 damage results in additional CO2 accumulation in the atmosphere and 

36 an indirect radiative forcing of climate change. Their simulations indicated that the 

indirect radiative forcing caused by O3 (0.62-1.09 W/m2 37 ) could have even greater impact 

on global warming than the direct radiative forcing of O3 (0.89 W/m2 38 ) (Sitch et al., 2007).
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1 Results from the various model studies presented in Table 9-3 are difficult to compare 

2 because of the various spatial and temporal scales used. However, all the studies showed 

3 that O3 exposure decreased ecosystem productivity and C sequestration. These results are 

4 consistent and coherent with experimental results obtained from studies at the leaf, plant 

5 and ecosystem scales (Sitch et al., 2007; Felzer et al., 2005). Many of the models use the 

6 same underlying function to simulate the effect of O3 exposure to C uptake. For example 

7 the functions of O3 exposure (AOT40) versus photosynthesis reduction for PnET-iI, 

8 PnET-CN, TEM, DLEM were all from Reich (1987) and Tjoelker et al. (1995). 

9 Therefore, it is not surprising that the results are similar. While these models can be 

10 improved and more evaluation with experimental data can be done, these models 

11 represent the state of the science for estimating the effect of O3 exposure on productivity 

12 and C sequestration.

9.4.3.5 Summary

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

14 effect of O3 exposure on ecosystem productivity and C sequestration. Recent studies from 

15 long-term FACE experiments, such as Aspen FACE, SoyFACE and the Kranzberg Forest 

16 (Germany), provide evidence of the association of O3 exposure and reduced productivity 

17 at the ecosystem level of organization. Studies at the leaf and plant scales show that O3

18 decreased photosynthesis and plant growth, which provides coherence and biological 

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

20 models, such as TREGRO, PnET, TEM and DLEM, are consistent with the FACE 

21 experimental evidence, which show that O3 reduced productivity of various ecosystems. 

22 Productivity is measured by various metrics such as GPP, NPP, NEP , NCE, NEE and 

23 individual tree biomass gain. All these metrics indicate a decrease in CO2 fixation by the 

24 systems that were studied.

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

26 response varies among plant communities. For example, O3 had little impact on white fir, 

27 but greatly reduced growth of ponderosa pine in southern California (Weinstein et al., 

28 2005). Ozone decreased net primary production (NPP) of most forest types in the Mid29 Atlantic region, but had small impacts on spruce-fir forest (Pan et al., 2009). 

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

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

32 simulations consistently found that O3 exposure caused negative impacts on these 

33 indicators, but the severity of these impacts was influenced by multiple interactions of 

34 biological and environmental factors. The suppression of ecosystem C sinks results in 
Draft – Do Not Cite or Quote 9-59 June 2012

1 more CO2 accumulation in the atmosphere. Globally, the indirect radiative forcing caused 

2 by O3 exposure through lowering the ecosystem C sink could have an even greater impact 

3 on global warming than the direct radiative forcing of O3 (Sitch et al., 2007). Ozone 

4 could also affect regional C budgets through interacting with multiple factors, such as 

5 N deposition, elevated CO2 and land use history. Model simulations suggested that O3

6 partially offset the growth stimulation caused by elevated CO2 and N deposition in both 

7 Northeast- and Mid-Atlantic-region forest ecosystems of the U.S. (Pan et al., 2009; 

8 Ollinger et al., 2002).

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

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

11 exposure and reduced carbon sequestration in terrestrial ecosystems.
Draft – Do Not Cite or Quote 9-60 June 2012

Table 9-3 Modeled effects of ozone on primary production, C exchange, 

and C sequestration.

Scale Model Index O3 Impacts Reference

GPP Global MOSESTRIFFIDCUOta 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 to 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 Tgb C from 

1960s to 1990s

Ren et al. 

(2007a)

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 MOSESTRIFFIDCUOt 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 MOSESTRIFFIDCUOt 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. 

(2007b)

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/sec.

b

Pg equals 1 × 1015 grams.

9.4.4 Crop Yield and Quality in Agricultural Systems

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

2 a large body of research has stemmed from that recognition. Previous O3 AQCDs have 

3 extensively reviewed this body of literature. Table 9-4 summarizes recent experimental 

4 studies of O3 effects on agricultural crops, exclusive of growth and yield. Growth and 

5 yield results are summarized in Table 9-17. 
Draft – Do Not Cite or Quote 9-61 June 2012

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 VanDingenen et al., 2009; Ainsworth, 2008; Feng et al., 2008; Hayes et al., 2007; Mills 

10 et al., 2007; 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., 2007a) 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 Table 9-17). For example, one study with cotton in a crop-weed 
Draft – Do Not Cite or Quote 9-62 June 2012

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)

34 reported a 20% seed yield loss due to a 23% increase in average daytime O3

35 concentration (56-69 ppb) within a single soybean cultivar across two growing seasons in 

36 Illinois, which lies within the range predicted by the meta-analysis. A further breakdown 

37 of the 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 9-63 June 2012

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 phenolic 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 merit 

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 9-64 June 2012

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 from 

8 41-49 ppb in the studies included, soybean (-7.7%), bean (-19.0%), barley (-8.9%), wheat 

9 (-9.7%), rice (-17.5%) and potato (-5.3%) all had yield losses compared to the baseline 

10 O3 concentrations (<26 ppb). At future O3 concentrations, averaging 63 ppb, soybean 

11 (-21.6%), bean (-41.4%), barley (-14%), wheat (-28%), rice (-17.5%) and potato (-11.9%) 

12 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., 2007a). 

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., 2007a).

24 Feng and Kobayashi (2009) compared their exposure-response results to those published 

25 by Mills et al. (2007a) and found the ranges of yield loss to be similar for soybean, rice 

26 and bean. However, Feng and Kobayashi (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. (2007a), 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 9-65 June 2012

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 

9 geographical scales are difficult to estimate. Fishman et al. (2010) have used satellite 

10 observations to estimate O3 concentrations in the contiguous tri-state area of Iowa, 

11 Illinois and Indiana and have combined that information with other measured 

12 environmental variables to model the historical impact of O3 concentrations on soybean 

13 yield across the 2002-2006 growing seasons. When soybean yield across Iowa, Indiana 

14 and Illinois was modeled as a function of seasonal temperature, soil moisture and O3

15 concentrations, O3 had the largest contribution to the variability in yield for the southern16 most latitudes included in the dataset. Fishman et al. (2010) determined that O3

17 concentrations significantly reduced soybean yield by 0.38 to 1.63% for every 

18 additional ppb of exposure across the 5 years. This value is consistent with previous 

19 chamber studies (Heagle, 1989) and results from SoyFACE (Morgan et al., 2006). 

20 Satellite estimates of tropospheric O3 concentrations exist globally (Fishman et al., 2008), 

21 therefore utilizing this historical modeling approach is feasible across a wider 

22 geographical area, longer time-span and perhaps for more crop 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; VanDingenen et al., 2009; Tong and Mauzerall, 2008; Wang and 

29 Mauzerall, 2004). Those studies found that O3 generally reduced crop yield and that 

30 different crops showed different sensitivity to O3 pollution (

31 Table 9-5). Ozone was calculated to induce a possible 45-82 million metric tons loss for 

32 wheat globally. Production losses for rice, maize and soybean were on the order of 

33 17-23 million metric tons globally (VanDingenen et al., 2009). The largest yield losses 

34 occur in high-production areas exposed to high O3 concentrations, such the Midwest and 
Draft – Do Not Cite or Quote 9-66 June 2012

1 the Mississippi Valley regions in the U.S., Europe, China and India (VanDingenen et al., 

2 2009; Tong et al., 2007).

9.4.4.2 Crop Quality

3 In general, it appears that increasing O3 concentrations above current ambient 

4 concentrations can cause species-dependent biomass losses, decreases in root biomass 

5 and nutritive quality, accelerated senescence and shifts in biodiversity. A study conducted 

6 with highbush blackberry has demonstrated decreased nutritive quality with increasing O3

7 concentration despite no change in biomass between charcoal-filtered control, ambient O3

8 and 2 × ambient O3 exposures (Ditchkoff et al., 2009). A study conducted with sedge 

9 using control (30 ppb), low (55 ppb), medium (80 ppb) and high (105 ppb) O3 treatments 

10 has demonstrated decreased root biomass and accelerated senescence in the medium and 

11 high O3 treatments (Jones et al., 2010). Alfalfa showed no biomass changes across 

12 two years of double ambient O3 concentrations (AOT40 of 13.9 ppm-h) using FACE 

13 fumigation (Maggio et al., 2009). However a modeling study has demonstrated that 84% 

14 of the variability in the relative feed value in high-yielding alfalfa was due to the 

15 variability in mean O3 concentration from 1998-2002 (Lin et al., 2007). Further, in a 

16 managed grassland FACE system, the reduction in total biomass harvest over five years 

17 decreased twice as fast in the elevated treatment (AOT40 of 13-59 ppm-h) compared to 

18 ambient (AOT40 of 1-20.7 ppm-h). Compared with the ambient control, loss in annual 

19 dry matter yield was 23% after 5 year. Further, functional groups were differentially 

20 affected, with legumes showing the strongest negative response (Volk et al., 2006). 

21 However, a later study by Stampfli and Fuhrer (2010) at the same site suggested that 

22 Volk et al. (2006) likely overestimated the effects of O3 on yield reduction because the 

23 overlapping effects of species dynamics caused by heterogeneous initial conditions and a 

24 change in management were not considered by these authors. An OTC study conducted 

25 with Trifolium subterraneum exposed to filtered (<15 ppb), ambient, and 40 ppb above 

26 ambient O3 demonstrated decreases in biomass in the highest O3 treatment as well as 10-

27 20% decreased nutritive quality which was mainly attributed to accelerated senescence 

28 (Sanz et al., 2005). A study conducted with Eastern gamagrass and big bluestem in OTCs 

29 suggested that big bluestem was not sensitive to O3, but gamagrass displayed decreased 

30 nutritive quality in the 2 × ambient O3 treatment, due to higher lignin content and 

31 decreased N (Lewis et al., 2006).
Draft – Do Not Cite or Quote 9-67 June 2012

9.4.4.3 Summary

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

2 a large body of research has subsequently stemmed from those initial findings. Previous 

3 O3 AQCDs have extensively reviewed this body of literature (U.S. EPA, 2006b). Current 

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

5 agricultural crops including, but not limited to, soybean, wheat, potato, watermelon, 

6 beans, turnip, onion, lettuce, and tomato. Continued increases in O3 concentration may 

7 further decrease yield in these sensitive crops. Despite the well-documented yield losses 

8 due to increasing O3 concentration, there is still a knowledge gap pertaining to the exact 

9 mechanisms of O3-induced yield loss. Research has linked increasing O3 concentration to 

10 decreased photosynthetic rates and accelerated senescence, which are related to yield.

11 New research is beginning to consider the mechanism of damage caused by prolonged, 

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

13 concentration (so-called acute exposure). Both types of O3 exposure cause damage to 

14 agricultural crops, but through very different mechanisms. Historically, most research on 

15 the mechanism of O3 damage used acute exposure studies. During the last decade, it has 

16 become clear that the cellular and biochemical processes involved in the response to 

17 acute O3 exposure are not involved in response to chronic O3 exposure, even though both 

18 cause yield loss in agriculturally important crops.

19 In addition, recent research has highlighted the effects of O3 on crop quality. Increasing 

20 O3 concentration decreases nutritive quality of grasses, decreases macro- and micro21 nutrient concentrations in fruits and vegetable crops, and decreases cotton fiber quality. 

22 These areas of research require further investigation to determine mechanisms and 

23 exposure-response relationships.

24 During the previous NAAQS reviews, there were very few studies that estimated O3

25 impacts on crop yields at large geographical scales. Recent modeling studies found that 

26 O3 generally reduced crop yield, but the impacts varied across regions and crop species. 

27 For example, the largest O3-induced crop yield losses occurred in high-production areas 

28 exposed to high O3 concentrations, such the Midwest and the Mississippi Valley regions 

29 of the U.S. (VanDingenen et al., 2009). Among crop species, the estimated yield loss for 

30 wheat and soybean were higher than for rice and maize (VanDingenen et al., 2009). 

31 Using satellite air-column observations with direct air-sampling O3 data, Fishman et al. 

32 (2010) modeled the yield-loss due to O3 over the continuous tri-state area of Illinois, 

33 Iowa and Wisconsin. They determined that O3 concentrations significantly reduced 

34 soybean yield, which further reinforces previous results from FACE-type experiments 

35 and OTC experiments. Evidence is sufficient to conclude that there is a causal 
Draft – Do Not Cite or Quote 9-68 June 2012

1 relationship between O3 exposure and reduced yield and quality of agricultural 

2 crops.
Draft – Do Not Cite or Quote 9-69 June 2012

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 (% change 

from ambient)

Reference

Alfalfa (Medicago 

sativa cv. Beaver)

Growth chambers

1, 2 or 

4 days

3 or 5 h/day 85 ppb

(Exposure duration)

Relative feed value n.s. 

*high variability 

among 

treatment 

groups (N/A)

Muntifering et 

al. (2006b)

Bean 

(Phaseolus 

vulgaris I. 

cv Borlotto)

OTC, groundplanted

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 weeks 

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)
Draft – Do Not Cite or Quote 9-70 June 2012

Species

Facility

Location

Exposure 

Duration

Ozone Exposurea

(Additional treatment)

Variable(s) measured Percent (%) 

change from 

CFb (% change 

from ambient)

Reference

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 weeks CF = 10 ppb; 

Ambient = 50 ppb; 

2× Ambient = 100 ppb

(CO2: 400 ppm & 700 ppm)

Pathogen infestation using 

percent 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) Bou Jaoudé 

et al. (2008a)

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, groundplanted

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.

3 months 12-h avg: 

CF = 18 ppb); 

Elevated = 72 ppb)

(CO2: 367 & 718)

Water-use efficiency n.s. (N/A) Booker et al. 

(2004b)

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; N/A (Significant 

negative 

correlation)

Piikki et al. 

(2008b)

1,000-seed weight regressed 

across all experiments

N/A (Significant 

negative 

correlation)
Draft – Do Not Cite or Quote 9-71 June 2012

Species

Facility

Location

Exposure 

Duration

Ozone Exposurea

(Additional treatment)

Variable(s) measured Percent (%) 

change from 

CFb (% change 

from ambient)

Reference

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 weeks 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) Dalstein and 

Vas (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. (2006a)

a

Ozone exposure in ppb unless otherwise noted.

b

CF = Carbon-filtered air.

NF = Non-filtered air.
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Table 9-5 Modeled effects of ozone on crop yield loss at regional and global 

scales

Scale Index O3 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 (in 

U.S. Dollars; 2000).

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. Figure 9-7 provides a 

3 simple illustration of potential effects of O3 exposure on water cycling. Section 9.3.2

4 reviewed possible mechanisms for effects of O3 exposure on stomatal functioning. This 

5 section on water cycling discusses how this alteration of stomatal functioning may affect 

6 water use in leaves, whole plants, a planted forest and watersheds. .
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Figure 9-7 The potential effects of ozone exposure on water 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 widths of stomata on the leaf surface, 

5 and, as a result, the collective response of groups of stomata on leaves and canopies 

6 determines larger-scale responses to O3. When measured at steady-state high light 

7 conditions, leaf-level stomatal conductance is often found to be reduced when exposed to 

8 O3. For example, a meta-analysis of 55 studies found that O3 reduced stomatal 

9 conductance by 11% (Wittig et al., 2007). However, these steady-state measurements 

10 were generally taken at saturating light conditions and steady-state vapor pressure deficit 

11 (VPD). Saturating light and steady-state VPD conditions are not common in the field 

12 since many parts of the plant canopy are shaded throughout the day. When studied under 

13 varying environmental conditions, many studies have reported incomplete stomatal 

14 closure with elevated O3 exposure during the day (Mills et al., 2009; Grulke et al., 2007b; 

15 Matyssek et al., 1995; Wieser and Havranek, 1995) or at night (Grulke et al., 2004). This 

O3 exposure 

Altered canopy 

water loss 

Soil moisture

stem sap flow

Decreased stomatal 

conductance, sluggish 

stomatal response, 

canopy leaf area loss

Altered 

run off

Stream flows
Draft – Do Not Cite or Quote 9-74 June 2012

1 may be due to sluggish stomatal response. Sluggish stomatal response, defined as a delay 

2 in stomatal response to changing environmental factors relative to controls (Paoletti and 

3 Grulke, 2010) has also been documented by several researchers (Grulke et al., 2007c; 

4 Matyssek et al., 1995; Pearson and Mansfield, 1993; Wallin and Skärby, 1992; Lee et al., 

5 1990; Skarby et al., 1987; Keller and Häsler, 1984; Reich and Lassoie, 1984). Sluggish 

6 stomatal response associated with O3 exposure suggests an uncoupling of the normally 

7 tight relationship between carbon assimilation and stomatal conductance as measured 

8 under steady-state conditions (Gregg et al., 2006; Paoletti and Grulke, 2005). Several tree 

9 and ecosystem models, such as TREGRO, PnET and DLEM, rely on this tight 

10 relationship to simulate water and carbon dynamics. The O3-induced impairment of 

11 stomatal control may be more pronounced for plants growing under water stress 

12 (Wilkinson and Davies, 2010; Grulke et al., 2007a; Paoletti and Grulke, 2005; Bonn et 

13 al., 2004; Kellomaki and Wang, 1997; Tjoelker et al., 1995; Reich and Lassoie, 1984). 

14 Since leaf-level stomatal regulation is usually assessed in a steady state rather than as a 

15 dynamic response to changing environmental conditions, steady state measurements 

16 cannot detect sluggish stomatal response. Because of sluggish stomatal responses, water 

17 loss from plants could be greater or reduced under dynamic environmental conditions 

18 over days and months. In situations where stomata fail to close under low light or water 

19 stressed conditions, water loss may be greater over time. In other situations, it is possible 

20 that slugglish stomata may fail to completely open in response to environmental stimuli 

21 and result in decreased water loss.

22 In addition to the impacts on stomatal performance, O3-induced physiological changes, 

23 such as reduced leaf area index and accelerated leaf senescence could alter water use 

24 efficiency. It is well established from chamber and field studies that O3 exposure is 

25 correlated with lower foliar retention (Karnosky et al., 2003; Topa et al., 2001; Pell et al., 

26 1999; Grulke and Lee, 1997; Karnosky et al., 1996; Miller et al., 1972; Miller et al., 

27 1963). However, Lee et al. (2009a) did not find changes in needle area of ponderosa pine 

28 and reported that greater canopy conductance followed by water stress under elevated O3

29 may have been caused by stomatal dysfunction. At the Aspen FACE experiment, stand30 level water use, as indicated by sap flux per unit ground area, was not significantly 

31 affected by elevated O3 despite a 22% decrease in leaf area index and 20% decrease in 

32 basal area (Uddling et al., 2008). The lack of negative effect of elevated O3 on stand 

33 water use may be due to the substantially increased leaf area-specific hydraulic 

34 conductance (Uddling et al., 2009). The increased leaf area-specific hydraulic 

35 conductance may be caused by the sluggish stomatal response. For example, in the pure 

36 aspen stands, the stomatal closure response to increasing vapor pressure deficit was less 

37 sensitive and mid-day leaf water potential was more negative under elevated O3

38 compared to controls. This suggests that O3 impaired stomatal control over transpiration 

39 (Uddling et al., 2009). Another potential factor contributing to the unchanged stand-level 
Draft – Do Not Cite or Quote 9-75 June 2012

1 water use included the higher proportion of sun leaves in trees under elevated O3

2 compared with control trees (Uddling et al., 2008). 

3 Elevated O3 could also affect evapotranspiration by altering tree crown interception of 

4 precipitation. Ozone was shown to change branch architectural parameters, and the 

5 effects were species-dependent at the Aspen FACE experiment (Rhea et al., 2010). The 

6 authors found that there was a significant correlation between canopy architecture 

7 parameters and stemflow (the flow of intercepted water down the stem of a tree) for birch 

8 but not aspen.

9 It is difficult to scale up physiology measurements from leaves to ecosystems. Thus, the 

10 current understanding of how stomatal response at the leaf scale is integrated at the scale 

11 of whole forest canopies, and therefore how it influences tree and forest stand water use 

12 is limited. Field studies by (McLaughlin et al., 2007a; 2007b) provided valuable insight 

13 into the possible consequences of stomatal sluggishness for ecosystem water cycling. 

14 McLaughlin et al. (2007a); (2007b) indicated that O3 increased water use in a mixed 

15 deciduous forest in eastern Tennessee. McLaughlin et al. (2007a); (2007b) found that O3, 

16 with daily maximum levels ranging from 69.2 to 82.9 ppb, reduced stem growth by 30-

17 50% in the high-O3 year 2002. The decrease in growth rate was caused in part by 

18 amplification of diurnal cycles of water loss and recovery. Peak hourly O3 exposure 

19 increased the rate of water loss through transpiration as indicated by the increased stem 

20 sap flow. The authors suggested that a potential mechanism for the increased sap flow 

21 could be altered stomatal regulation from O3 exposure, but this was inferred through sap 

22 flow measurements and was not directly measured. The increased canopy water loss 

23 resulted in higher water uptake by the trees as reflected in the reduced soil moisture in the 

24 rooting zone. The change in tree water use led to further impacts on the hydrological 

25 cycle at the landscape level. Increased water use under high O3 exposure was reported to 

26 reduce late-season modeled streamflow in three forested watersheds in eastern Tennessee 

27 (McLaughlin et al., 2007b).

28 Felzer et al. (2009) used TEM-Hydro to assess the interactions of O3, climate, elevated 

29 CO2 and N limitation on the hydrological cycle in the eastern U.S. They found that 

30 elevated CO2 decreased evapotranspiration by 2-4% and increased runoff by 3-7%, as 

31 compared to the effects of climate alone. When O3 damage and N limitation were 

32 included, evapotranspiration was reduced by an additional 4-7% and runoff was increased 

33 by an additional 6-11% (Felzer et al., 2009). Based upon simulation with INTRAST and 

34 LINKAGES, Hanson et al. (2005) found that increasing O3 concentration by 20 ppb 

35 above the current ambient level yields a modest 3% reduction in water use. Those 

36 ecological models were generally built on the assumption that O3 induces stomatal 

37 closure and have not incorporated possible stomatal sluggishness due to O3 exposure. 
Draft – Do Not Cite or Quote 9-76 June 2012

1 Because of this assumption, results of those models normally found that O3 reduced water 

2 use. 

9.4.5.1 Summary

3 Although the evidence was from a limited number of field and modeling studies, findings 

4 showed an association between O3 exposure and alteration of water use and cycling in 

5 vegetation, and at the watershed level. There is not a clear consensus on the nature of 

6 leaf-level stomatal conductance response to O3 exposure. When measured under steady7 state high light conditions, leaf-level stomatal conductance is often found to be reduced 

8 when plants are exposed to O3. However, measurements of stomatal conductance under 

9 dynamic light and VPD conditions indicate sluggish responses under elevated O3

10 exposure, which could potentially lead to increased water loss from vegetation in some 

11 situations. Field studies conducted by McLaughlin et al. (2007a); (2007b) suggested that 

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

13 led to a reduction in the late-season modeled stream flow in three forested watersheds in 

14 eastern Tennessee. Sluggish stomatal responses during O3 exposure was suggested as a 

15 possible mechanism for increased water loss during peak O3 exposure. Currently, the 

16 O3-induced reduction in stomatal aperture is the biological assumption for most process17 based models. Because of this assumption, results of those models normally found that 

18 O3 reduced water loss. For example, Felzer et al. (2009) found that O3 damage and 

19 N limitation together reduced evapotranspiration and increased runoff.

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

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

22 the alteration of ecosystem water cycling.

9.4.6 Below-Ground Processes

23 Above-ground and below-ground processes are tightly interconnected. Because roots and 

24 soil organisms are not exposed directly to O3, below-ground processes are affected by O3

25 through alterations in the quality and quantity of C supply from photosynthates and 

26 litterfall (Andersen, 2003). Ozone can decrease leaf C uptake by reducing photosynthesis 

27 (Section 9.3). Ozone can also increase metabolic costs by stimulating the production of 

28 chemical compounds for defense and repair processes, and by increasing the synthesis of 

29 antioxidants to neutralize free radicals (see Section 9.3), both of which increase the 

30 allocation of carbon for above-ground processes. Therefore, O3 could significantly reduce 
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1 the amount of C available for allocation to below-ground by decreasing C uptake while 

2 increasing C consumption of above-ground processes (Andersen, 2003). 

3 Since the 2006 O3 AQCD, there is additional evidence for O3 effects on below-ground 

4 processes. Ozone has been found to alter root growth, soil food web structure, 

5 decomposer activities, C turnover, water cycling and nutrient flow (Figure 9-8). Ozone 

6 effects on root development and root biomass production and soil food web structure are 

7 reviewed in Section 9.4.3.1 and Section 9.4.9.2, respectively. The focus in this section is 

8 on the response of litter input, decomposer activities, soil respiration, soil C formation 

9 and nutrient cycling. 

Note: Arrows denote C flux pathways that are affected by ozone. Dashed lines indicate where the impact of ozone is suspected but 

unknown. 

Source: Modified from Andersen (2003).

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.

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 9-78 June 2012

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 

3 in chemical parameters, such as changes in C chemistry and nutrient concentrations, were 

4 observed in both leaf and root litter (Table 9-6). 

5 At the Aspen FACE site, several studies investigated litter chemistry changes (Parsons et 

6 al., 2008; Johnson and Pregitzer, 2007; Chapman et al., 2005; Liu et al., 2005). In both 

7 aspen and birch leaf litter, elevated O3 increased the concentrations of soluble sugars, 

8 soluble phenolics and condensed tannins (Parsons et al., 2008; Liu et al., 2005). 

9 Compared to other treatments, aspen litter under elevated O3 had the highest fiber 

10 concentration, with the lowest concentration associated with the birch litter under the 

11 same conditions (Parsons et al., 2008). Chapman et al. (2005) measured chemical 

12 changes in fine root litter and found that elevated O3 decreased lignin concentration. 

13 O3-induced chemistry changes were also reported from other experimental sites. Results 

14 from an OTC study in Finland suggested that elevated O3 increased the concentration of 

15 acid-soluble lignin, but had no significant impact on other chemicals such as total sugars, 

16 hemicelluloses, cellulose or total lignin in the litter of silver birch (Kasurinen et al., 

17 2006). Results from the free air canopy O3 exposure experiment at Kranzberg Forest 

18 showed that O3 increased starch concentrations but had no impact on cellulose and lignin 

19 in beech and spruce leaf litter (Aneja et al., 2007). The effect of O3 on three antioxidants 

20 (ascorbate, glutathione and α-tocopherol) in fine roots of beech was also assessed at 

21 Kranzberg Forest. The results indicated that O3 had no significant effect on α-tocopherol 

22 and ascorbate concentrations, but decreased glutathione concentrations in fine roots 

23 (Haberer et al., 2008). In addition to changing C chemistry, O3 also altered nutrient 

24 concentrations in green leaves and litter (Table 9-6). 

25 The combined effects of O3 on biomass productivity and chemistry changes may alter 

26 C chemicals and nutrient contents at the canopy or stand level. For example, although O3

27 had different impacts on their concentrations, annual fluxes of C chemicals (soluble 

28 sugar, soluble phenolics, condensed tannins, lipid and hemicelluloses), macro nutrients 

29 (N, P, K and S) and micro nutrients (Mg, B, Cu and Zn) to soil were all reduced due to 

30 lower litter biomass productivity at Aspen FACE (Liu et al., 2007; Liu et al., 2005).. In a 

31 2-year growth chamber experiment in Germany, N content of a spruce canopy in a mixed 

32 culture and Ca content of a beech canopy in a monoculture was increased due to elevated 

33 O3, although leaf production was not significantly altered by O3 (Rodenkirchen et al., 

34 2009). 
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Table 9-6 The effect of elevated ozone on leaf/litter nutrient concentrations.

Study Site Species O3 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. (2007)a)

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

O3 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 two 

4 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 

8 1,4-β-glucosidase was reduced by 25% under elevated O3 at Aspen FACE. The decrease 

9 in 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 
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1 xylosidase, enzyme activities involved in plant cell wall degradation (cellobiohydrolase, 

2 beta-glucosidase and glucuronidase) were decreased in rhizosphere soil samples under 

3 elevated O3 (2 × ambient level) (Pritsch et al., 2009). Similarly, Chen et al. (2009) found 

4 O3 exposure, with a 3-month AOT40 of 21-44 ppm-h, decreased the microbial metabolic 

5 capability in the rhizosphere and bulk soil of wheat, although the observed reduction in 

6 bulk soil was not significant.

7 Ozone-induced change in soil organisms’ activities could affect litter decomposition 

8 rates. Results of recent studies indicated that O3 slightly reduced or had no impacts on 

9 litter decomposition (Liu et al., 2009b; Parsons et al., 2008; Kasurinen et al., 2006)

10 (Baldantoni et al., 2011). The responses varied among species, sites and exposure length. 

11 Parsons et al. (2008) collected litter from aspen and birch seedlings at Aspen FACE site, 

12 and conducted a 23-month field litter incubation starting in 1999. They found that 

13 elevated O3 had different impacts on the decomposition of aspen and birch litter. Elevated 

14 O3 was found to reduce aspen litter decomposition. However, O3 accelerated birch litter 

15 decomposition under ambient CO2, but reduced it under elevated CO2 (Parsons et al., 

16 2008). Liu et al. (2009b) conducted another litter decomposition study at Aspen FACE 

17 from 2003 to 2006, when stand leaf area index (LAI) reached its maximum. During the 

18 935-day field incubation, elevated O3 was shown to reduce litter mass loss in the first 

19 year, but not in the second year. They suggested that higher initial tannin and phenolic 

20 concentrations under elevated O3 reduced microbial activity in the first year (Liu et al., 

21 2009b). In an OTC experiment, Kasurinen et al. (2006) collected silver birch leaf litter 

22 from three consecutive growing seasons and conducted three separate litter-bag 

23 incubation experiments. Litter decomposition was not affected by O3 exposure in the first 

24 two incubations, but a slower decomposition rate was found in the third incubation. Their 

25 principle component analysis indicated that the litter chemistry changes caused by O3

26 (decreased Mn, P, B and increased C:N) might be partially responsible for the decreased 

27 mass loss of their third incubation. In another OTC experiment, Baldantoni et al. (2011)

28 found that O3 significantly reduced leaf litter decomposition of Quercus ilex L, although 

29 litter C, N, lignin and cellulose concentrations were not altered by O3 exposure.

9.4.6.3 Soil Respiration and Carbon Formation 

30 Ozone could reduce the availability of photosynthates for export to roots, and thus, 

31 indirectly increase root mortality and turnover rates. Ozone has also been shown to 

32 reduce above-ground litter productivity and alter litter chemistry, which would affect the 

33 quality and quantity of the C supply to soil organisms (Section 9.4.6.1). The complex 

34 interactions among those changes make it difficult to predict the response of soil 

35 C cycling under elevated O3. The 2006 O3 AQCD concluded that O3 had no consistent 
Draft – Do Not Cite or Quote 9-81 June 2012

1 impact on soil respiration (U.S. EPA, 2006b). Ozone could increase or decrease soil 

2 respiration, depending on the approach and timing of the measurements. Ozone may also 

3 alter soil C formation. However, very few experiments directly measured changes in soil 

4 organic matter content under O3 fumigation (U.S. EPA, 2006b). Recent studies on soil 

5 respiration and soil C content also found mixed responses. Most importantly, recent 

6 results from long-term fumigation experiments, such as the Aspen FACE experiment, 

7 suggest that ecosystem response to O3 exposure can change over time. Observations 

8 made during the late exposure years can be inconsistent with those during the early years, 

9 highlighting the need for caution when assessing O3 effects based on short-term studies 

10 (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. (2008); Pregitzer 

et al. (2006)

Soil C formation 1998-2001 O3 reduced the formation rates of total soil C 

by 51% and acid-insoluble 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

11 Ozone has shown inconsistent impacts on soil respiration. A sun-lit 

12 controlled-environment chamber study found that O3 had no significant effects on soil 

13 respiration, fine root biomass or any of the soil organisms in a reconstructed ponderosa 

14 pine/soil-litter system (Tingey et al., 2006). In an adult European beech/Norway spruce 

15 forest at Kranzberg Forest, the free air O3 fumigation (AOT40 of 10.2-117 ppm-h) 

16 increased soil respiration under both beech and spruce during a humid year (Nikolova et 

17 al., 2010). The increased soil respiration under beech has been accompanied by the 
Draft – Do Not Cite or Quote 9-82 June 2012

1 increase in fine root biomass and ectomycorrhizal fungi diversity and turnover (Grebenc 

2 and Kraigher, 2007). The stimulating effect on soil respiration disappeared under spruce 

3 in a dry year, which was associated with a decrease in fine root production in spruce 

4 under drought. This finding suggested that drought was a more dominant stress than O3

5 for spruce (Nikolova et al., 2010). Andersen et al. (2010) labeled the canopies of 

European beech and Norway spruce with CO2 depleted in 13 6 C at the same site. They did 

7 not observe any significant changes in soil respiration for either species. 

8 The nearly 10 year long studies at Aspen FACE indicated that the response of soil 

9 respiration to O3 interacted with CO2 exposure and varied temporally (Table 9-7) 

10 (Pregitzer et al., 2008; Pregitzer et al., 2006; King et al., 2001). Ozone treatment alone 

11 generally had the lowest mean soil respiration rates, although those differences between 

12 control and elevated O3 were usually not significant. However, soil respiration rates were 

13 different with O3 alone and when acting in combination with elevated CO2. In the first 

14 five years (1998-2002), soil respiration under +CO2+O3 treatment was similar to that 

15 under control and lower than that under +CO2 treatment (Pregitzer et al., 2006; King et 

16 al., 2001). Since 2003, +CO2+O3 treatment started to show the greatest impact on soil 

17 respiration. Compared to elevated CO2, soil respiration rate under +CO2+O3 treatment 

18 was 15-25% higher from 2003-2004, and 5-10% higher from 2005-2007 (Pregitzer et al., 

19 2008; Pregitzer et al., 2006). Soil respiration was highly correlated with the biomass of 

20 roots with diameters of <2 mm and <1 mm, across plant community and atmospheric 

21 treatments. The authors suggested that the increase in soil respiration rate may be due to 

22 +CO2+O3 increased fine root (<1.0 mm) biomass production (Pregitzer et al., 2008).

23 Changes in leaf chemistry and productivity due to O3 exposure have been shown to affect 

24 herbivore growth and abundance (See Section 9.4.9.1). Canopy insects could affect soil 

25 carbon and nutrient cycling through frass deposition, or altering chemistry and quantity 

26 of litter input to the forest floor. A study at the Aspen FACE found that although elevated 

27 O3 affected the chemistry of frass and greenfall, these changes had small impact on 

28 microbial respiration and no effect on nitrogen leaching (Hillstrom et al., 2010a). 

29 However, respiratory carbon loss and nitrate immobilization were nearly double in 

30 microcosms receiving herbivore inputs than those receiving no herbivore inputs 

31 (Hillstrom et al., 2010a).

Soil Carbon Formation

32 Ozone-induced reductions in plant growth can result in reduced C input to soil and 

33 therefore soil C content (Andersen, 2003). The simulations of most ecosystem models 

34 support this prediction (Ren et al., 2007b; Zhang et al., 2007a; Felzer et al., 2004). 

35 However, very few studies have directly measured soil C dynamics under elevated O3. 
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1 After the first four years of fumigation (from 1998 to 2001) at the Aspen FACE site, 

2 Loya et al. (2003) found that forest stands exposed to both elevated O3 and CO2

3 accumulated 51% less total soil C, and 48% less acid-insoluble soil C compared to stands 

4 exposed only to elevated CO2. Soil organic carbon (SOC) was continuously monitored at 

5 the Aspen FACE site, and the later data showed that the initial reduction in new

6 C formation (soil C derived from plant litter since the start of the experiment) by O3

7 under elevated CO2 is only a temporary effect (Table 9-7) (Talhelm et al., 2009). The 

8 amount of new soil C in the elevated CO2 and the combined elevated CO2 and O3

9 treatments has converged since 2002. There was no significant effect of O3 on the new C 

10 formed under elevated CO2 over the last four years of the study (2004-2008). Talhelm et 

11 al. (2009) suggested the observed reduction in the early years of the experiment might be 

12 driven by a suppression of C allocated to fine root biomass. During the early exposure 

13 years, O3 had no significant impact on fine root production (King et al., 2001). However, 

14 the effect of O3 on fine root biomass was observed later in the experiment. Ozone 

15 increased fine root production and the highest fine root biomass was observed under the 

16 combined elevated CO2 and O3 treatment in the late exposure years (Table 9-7) (Pregitzer 

17 et al., 2006). This increase in fine root production was due to changes in community 

18 composition, such as better survival of an O3-tolerant aspen genotype, birch and maple, 

19 rather than changes in C allocation at the individual tree level (Pregitzer et al., 2008; Zak 

20 et al., 2007). 

9.4.6.4 Nutrient Cycling

21 Ozone can affect nutrient cycling by changing nutrient release from litter, nutrient uptake 

22 by plants, and soil microbial activity. Nitrogen is the limiting nutrient for most temperate 

23 ecosystems, and several studies examined N dynamics under elevated O3. Nutrient 

24 mineralization from decomposing organic matter is important for sustaining ecosystem 

25 production. Holmes et al. (2006) found that elevated O3 decreased gross N mineralization 

26 at the Aspen FACE site, indicating that O3 may reduce N availability. Other N cycling 

processes, such as NH4

+ 27 immobilization, gross nitrification, microbial biomass N and soil 

28 organic N, were not affected by elevated O3 (Holmes et al., 2006). Similarly, Kanerva et 

29 al. (2006) found total N, NO3-, microbial biomass N, potential nitrification and 

30 denitrification in their meadow mesocosms were not affected by elevated O3 (40-50 ppb). 

31 Ozone was found to decrease soil mineral N content at SoyFACE, which was likely 

32 caused by a reduction in plant material input and increased denitrification (Pujol Pereira 

33 et al., 2011). Ozone also showed small impact on other micro and macro nutrients. Liu et 

34 al. (2007)a) assessed N, P, K, S, Ca, Mg, Mn, B, Zn and Cu release dynamics at Aspen 

35 FACE, and they found that O3 had no effects on most nutrients, except to decrease N and 
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1 Ca release from litter. These studies reviewed above suggest that soil N cycling processes 

2 are not affected or slightly reduced by O3 exposure. However, in a lysimeter study with 

3 young beech trees, Stoelken et al. (2010) found that elevated O3 stimulated N release 

4 from litter which was largely attributed to an enhanced mobilization of inert nitrogen 

5 fraction.

6 Using the SImple NItrogen Cycle model (SINIC), Hong et al. (2006) evaluated the 

7 impacts of O3 exposure on soil N dynamics and streamflow nitrate flux. The detrimental 

8 effect of O3 on plant growth was found to reduce plant uptake of N and therefore increase 

9 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 10 /year) at the Hubbard 

11 Brook Experimental Watershed from 1964-1994 (Hong et al., 2006).

9.4.6.5 Dissolved Organic Carbon and Biogenic Trace Gases 

Emission

12 The O3-induced changes in plant growth, C and N fluxes to soil and microbial 

13 metabolism can alter other biogeochemical cycling processes, such as soil dissolved 

14 organic carbon (DOC) turnover and trace gases emission. 

15 Jones et al. (2009) collected fen cores from two peatlands in North Wales, UK and 

16 exposed them to one of four levels of O3 (AOT40 of 0, 3.69, 5.87 and 13.80 ppm-h for 

17 41 days). They found the concentration of porewater DOC in fen cores was significantly 

18 decreased by increased O3 exposure. A reduction of the low molecular weight fraction of 

19 DOC was concurrent with the observed decrease in DOC concentration. Their results 

20 suggested that O3 damage to overlying vegetation may decrease utilizable C flux to soil. 

21 Microbes, therefore, have to use labile C in the soil to maintain their metabolism, which, 

22 the authors hypothesized, leads to a decreased DOC concentration with a shift of the 

23 DOC composition to more aromatic, higher molecular weight organic compounds. 

24 Several studies since the 2006 O3 AQCD have examined the impacts of O3 on nitrous 

25 oxide (N2O) and methane (CH4) emission. Kanerva et al. (2007) measured the fluxes of 

26 N2O and CH4 in meadow mesocosms, which were exposed to elevated CO2 and O3 in 

27 OTCs in south-western Finland. They found that the daily N2O fluxes were decreased in 

28 the NF+O3 (non-filtered air + elevated O3, 40-50 ppb) after three seasons of exposure. 

29 Elevated O3 alone or combined with CO2 did not have any significant effect on the daily 

30 fluxes of CH4 (Kanerva et al., 2007). In another study conducted in central Finland, the 

31 4 year open air O3 fumigation (AOT40 of 20.8-35.5 ppm-h for growing season) slightly 

32 increased potential CH4 oxidation by 15% in the peatland microcosms, but did not affect 

33 the rate of potential CH4 production or net CH4 emissions, which is the net result of the 
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1 potential CH4 production and oxidation (Morsky et al., 2008). However, several studies 

2 found that O3 could significantly reduce CH4 emission. Toet et al. (2011) exposed 

3 peatland mesocosms to O3 in OTCs for two years, and found that CH4 emissions were 

4 significantly reduced by about 25% during midsummer periods of both years. In an OTC 

5 study of rice paddy, Zheng et al. (2011) found that the daily mean CH4 emissions were 

6 significantly lower under elevated O3 treatments than those in charcoal-filtered air and 

7 nonfiltered air treatments. They found that the seasonal mean CH4 emissions were 

8 negatively related with AOT40, but positively related to the relative rice yield, 

9 aboveground biomass and underground biomass.

9.4.6.6 Summary

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

11 often site specific, O3 altered the quality and quantity of litter input to soil, microbial 

12 community composition, and C and nutrient cycling. Biogeochemical cycling of below13 ground processes is fueled by C input from plants. Studies at the leaf and plant level have 

14 provided biologically plausible mechanisms, such as reduced photosynthetic rates, 

15 increased metabolic cost, and reduced root C allocation for the association of O3 exposure 

16 and the alteration of below-ground processes. 

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

18 reduced litter production and altered C chemistry, such as soluble sugars, soluble 

19 phenolics, condensed tannins, lignin, and macro/micro nutrient concentration in litter 

20 (Parsons et al., 2008; Kasurinen et al., 2006; Liu et al., 2005). Under elevated O3, the 

21 changes in substrate quality and quantity could alter microbial metabolism and therefore 

22 soil C and nutrient cycling. Several studies indicated that O3 suppressed soil enzyme 

23 activities (Pritsch et al., 2009; Chung et al., 2006). However, the impact of O3 on litter 

24 decomposition was inconsistent and varied among species, sites and exposure length. 

25 Similarly, O3 had inconsistent impacts on dynamics of micro and macro nutrients. 

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

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

28 changed over time. For example, in the early part of the experiment (1998-2003), O3 had 

29 no impact on soil respiration but reduced the formation rates of total soil C under 

30 elevated CO2. However, after 10-11 years of exposure, O3 was found to increase soil 

31 respiration but have no significant impact on soil C formation under elevated CO2. 

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

33 exposure and the alteration of below-ground biogeochemical cycles.
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9.4.7 Community Composition

1 The effects of O3 on species competition (AX9.3.3.4) and community composition 

2 (AX9.6.4) were summarized in the 2006 O3 AQCD. Plant species differ in their 

3 sensitivity to O3. Further, different genotypes of a given species also vary in their 

4 sensitivity. This differential sensitivity could change the competitive interactions that 

5 lead to loss in O3 sensitive species or genotypes. In addition, O3 exposure has been found 

6 to alter reproductive processes in plants (See Section 9.4.3.3). Changes in reproductive 

7 success could lead to changes in species composition. However, since ecosystem-level 

8 responses result from the interaction of organisms with one another and with their 

9 physical environment, it takes longer for a change to develop to a level of prominence at 

10 which it can be identified and measured. A shift in community composition in forest and 

11 grassland ecosystems noted in the 2006 O3 AQCD has continued to be observed from 

12 experimental and gradient studies. Additionally, research since the last review has shown 

13 that O3 can alter community 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 significant 

15 decline in ponderosa pine (Pinus ponderosa ) and Jeffrey pine (Pinus jeffreyi) (U.S. EPA, 

16 2006b). Pine trees in the young mature age class group exhibited higher mortality rates 

17 compared with mature trees at a site with severe O3 visible foliar injury. The vulnerability 

18 of young mature pines was most likely caused by the fact that trees in this age class were 

19 emerging into the canopy, where higher O3 concentrations were encountered (McBride 

20 and Laven, 1999). Because of the loss of O3-sensitive pines, mixed forests of ponderosa 

21 pine, Jeffery Pine and white fir (Abies concolor) shifted to predominantly white fir 

22 (Miller, 1973). Ozone may have indirectly caused the decline in understory diversity in 

23 coniferous forests in the San Bernardino Mountains through an increase in pine litterfall. 

24 This increase in litterfall from O3 exposure results in an understory layer that may 

25 prohibit the establishment of native herbs, but not the exotic annual Galium aparine

26 (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 cembra) 
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1 (Wieser et al., 2006). However, many environmental factors such as light, temperature, 

2 nutrient and soil moisture, and climate extremes such as unusual dry and wet periods 

3 could interact with O3 and alter the response of forest to O3 exposure. For those pollution 

4 gradient studies, several confounding factors, such as drought, insect outbreak and forest 

5 management, may also contribute to or even be the dominant factors causing the 

6 mortality of trees (de Lourdes de Bauer and Hernandez-Tejeda, 2007; Wieser et al., 

7 2006).

8 Recent evidence from long-term free O3 fumigation experiments provided additional 

9 support for the potential impacts of O3 on species competition and community 

10 composition changes in forest ecosystems. At the Aspen FACE site, community 

11 composition at both the genetic and species levels was altered after seven years of 

12 fumigation with O3 (Kubiske et al., 2007). In the pure aspen community, O3 fumigation 

13 reduced growth and increased mortality of sensitive clone 259, while the O3 tolerant 

14 clone 8L emerged as the dominant clone. Growth of clone 8L was even greater under 

15 elevated O3 compared to controls, probably due to O3 alleviated competitive pressure on 

16 clone 8L by reducing growth of other clones. In the mixed aspen-birch and aspen-maple 

17 communities, O3 reduced the competitive capacity of aspen compared to birch and maple 

18 (Kubiske et al., 2007). In a phytotron study, O3 fumigation reduced growth of beech but 

19 not spruce in mixed culture, suggesting a higher susceptibility of beech to O3 under 

20 interspecific competition (Kozovits et al., 2005).

9.4.7.2 Grassland and Agricultural Land

21 The response of managed pasture, often cultivated as a mixture of grasses and clover, to 

22 O3 pollution has been studied for many years. The tendency for O3-exposure to shift the 

23 biomass of grass-legume mixtures in favor of grass species, reported in the previous O3

24 AQCD has been generally confirmed by recent studies. In a mesocosm study, Trifolium 

25 repens and Lolium perenne mixtures were exposed to an episodic rural O3 regime within 

26 solardomes for 12 weeks. T. repens showed significant changes in biomass but not L. 

27 perenne, and the proportion of T. repens decreased in O3-exposed mixtures compared to 

28 the control (Hayes et al., 2009). The changes in community composition of grass-legume29 forb mixtures were also observed at the Le Mouret FACE experiment, Switzerland. 

30 During the 5-year O3 fumigation (AOT40 of 13.3-59.5 ppm-h), the dominance of 

31 legumes in fumigated plots declined more quickly than those in the control plots (Volk et 

32 al., 2006). However, Stampfli and Fuhrer (2010) reanalyzed the species and soil data and 

33 suggested that Volk et al. (2006) overestimated the O3 effect. Stampfli and Fuhrer (2010)

34 found that the difference in the species dynamics between control and O3 treatment was 

35 more caused by heterogeneous initial conditions than O3 exposure. Several studies also 
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1 suggested that mature/species-rich ecosystems were more resilient to O3 exposure. At 

2 another FACE experiment, located at Alp Flix, Switzerland, O3 fumigation (AOT40 of 

3 15.2-64.9 ppm-h) showed no significant impact on community composition of this 

4 species-rich pasture (Bassin et al., 2007b). Although most studies demonstrated an 

5 increase in grass:forb ratio with O3 exposure (Hayes et al., 2009; U.S. EPA, 2006b), a 

6 study on a simulated upland grassland community showed that O3 reduced the grass:forb 

7 ratio (Hayes et al., 2010) which may be due to the grass species in this community. The 

8 grass species studied by Hayes et al. (2010), Anthoxanthum odoratum, was more 

9 sensitive to O3 than other grass species such as L. perenne (Hayes et al., 2009). Pfleeger 

10 et al. (2010) collected seed bank soil from an agricultural field and examined how the 

11 plant community responded over several generations to elevated O3 exposures. Sixty 

12 plant species from 22 families emerged in the chambers over their four year study. 

13 Overall, they found that O3 appeared to have small impacts on seed germination and only 

14 a minor effect on species richness of pioneer plant communities.

15 Several review papers have discussed the physiological and ecological characteristics of 

16 O3-sensitive herbaceous plants. Hayes et al. (2007) assessed species traits associated with 

17 O3 sensitivity by the changes in biomass caused by O3 exposure. Plants of the therophyte 

18 (e.g., annual) life form were particularly sensitive to O3. Species with higher mature leaf 

19 N concentration tended to be more sensitive than those with lower leaf N concentration. 

20 Plants growing under high oxidative stress environments, such as high light or high 

21 saline, were more sensitive to O3. Using the same dataset from Hayes et al. (2007), Mills 

22 et al. (2007b) identified the O3 sensitive communities. They found that the largest number 

23 of these O3 sensitive communities were associated with grassland ecosystems. Among 

24 grassland ecosystems, alpine grassland, sub-alpine grassland, woodland fringe, and dry 

25 grassland were identified as the most sensitive communities.

9.4.7.3 Microbes

26 Several methods have been used to study microbial composition changes associated with 

27 elevated O3. Phospholipid fatty acid (PLFA) analysis is widely used to determine whether 

28 O3 elicits an overall effect on microbial community composition. However, since PLFA 

29 markers cover a broad range of different fungi, resolution of this method may be not fine 

30 enough to detect small changes in the composition of fungal communities. Methods, such 

31 as microscopic analyses and polymerase chain reaction–denaturing gradient gel 

32 electrophoresis (PCR–DGGE), have better resolution to specifically analyze the fungal 

33 community composition. The resolution differences among those methods needs to be 

34 considered when assessing the O3 impact on microbial 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 the 

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 the 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 to be a causal relationship 

29 between O3 exposure and the alteration of community composition of some 

30 ecosystems.

9.4.8 Factors that Modify Functional and Growth Response

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

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

33 CO2, influence or alter plant response to O3. These modifying factors were 
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1 comprehensively reviewed in AX9.3 of the 2006 O3 AQCD and thus, this section serves 

2 mainly as a brief summary of the previous findings. A limited number of new studies 

3 published since the 2006 O3 AQCD add to the understanding of the role of these 

4 interactions in modifying O3-induced plant responses. Many of these modifying factors 

5 and interactions are integrated into discussions elsewhere in this chapter and the reader is 

6 directed to those sections.

9.4.8.1 Genetics

7 It is well known that species vary greatly in their responsiveness to O3. Even within a 

8 given species, individual genotypes or populations can also vary significantly with 

9 respect to O3 sensitivity (U.S. EPA, 2006b). Therefore, caution should be taken when 

10 considering a species’ degree of sensitivity to O3. Plant response to O3 is determined by 

11 genes that are directly related to oxidant stress and to an unknown number of genes that 

12 are not specifically related to oxidants, but instead control leaf and cell wall thickness, 

13 stomatal conductance, and the internal architecture of the air spaces. It is rarely the case 

14 that single genes are responsible for O3 tolerance. Studies using molecular biological 

15 tools and transgenic plants have positively verified the role of various genes and gene 

16 products in O3 tolerance and are continuing to increase the understanding of O3 toxicity 

17 and differences in O3 sensitivity. See Section 9.3.3.2 of this document for a discussion of 

18 recent studies related to gene expression changes in response to O3.

9.4.8.2 Environmental Biological Factors

19 As stated in the 2006 O3 AQCD, the biological factors within the plant’s environment 

20 that may influence its response to O3 encompass insects and other animal pests, diseases, 

21 weeds, and other competing plant species. Ozone may influence the severity of a disease 

22 or infestation by a pest or weed, either by direct effects on the causal species, or 

23 indirectly by affecting the host, or both. In addition, the interaction between O3, a plant, 

24 and a pest, pathogen, or weed may influence the response of the target host species to O3

25 (U.S. EPA, 2006b). Several recent studies on the effects of O3 on insects via their 

26 interactions with plants are discussed in Section 9.4.9.1. In addition, O3 has also been 

27 shown to alter soil fauna communities (Section 9.4.9.2).

28 In contrast to detrimental biological interactions, there are mutually beneficial 

29 relationships or symbioses involving higher plants and bacteria or fungi. These include 

30 (1) the nitrogen-fixing species Rhizobium and Frankia that nodulate the roots of legumes 

31 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 

6 Section 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. Riikonen et al. (2009) at the Ruohoniemi open air 

23 exposure field in Kuopio, Finland found that the effects of temperature and O3 on total 

24 leaf area and photosynthesis of Betula pendula were counteractive. Elevated O3 reduced 

25 the saplings’ ability to utilize the warmer growth environment by increasing the stomatal 

26 limitation for photosynthesis and by reducing the redox state of ascorbate in the apoplast 

27 in the 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. In 

13 another study, Grulke et al. (2008) reported that various lines of phenomenological and 

14 experimental evidence indicate that N deposition and O3 pollution contribute to the 

15 susceptibility of forests to wildfire in the San Bernadino Mountains by increasing stress 

16 due to drought, weakening trees, and predisposing them to bark beetle infestation 

17 (U.S. EPA, 2008; NOx/SOx ISA).

18 Studies conducted outside the U.S. are also summarized in Table 9-8. Generally, the 

19 responses were species specific. The O3-induced reduction in photosynthetic rate and 

20 biomass loss were greater in the relatively high N treatment for watermelon (Citrillus 

21 lanants) (Calatayud et al., 2006) and Japanese beech (Fagus crenata) seedlings 

22 (Yamaguchi et al., 2007). However, there was no significant interactive effect of O3 and 

23 N on biomass production for Quercus serrata seedlings (Watanabe et al., 2007), young 

24 Norway spruce (Picea abies) trees (Thomas et al., 2005), and young European beech 

25 (Fagus sylvatica) trees Thomas et al. (2006).

Table 9-8 Response of plants to the interactive effects of elevated ozone 

exposure and nitrogen 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)
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Site Species Ozone exposure N addition Responses References

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 ppm-h) 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 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 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 ppm-h), 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)
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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; 

3 Tissue et al., 1997; Will and Ceulemans, 1997), decreases stomatal conductance 

4 (Ainsworth and Rogers, 2007; Paoletti et al., 2007; Bernacchi et al., 2006; Leakey et al., 

5 2006; Medlyn et al., 2001) and generally increases the growth of plants (McCarthy et al., 

6 2009; Norby et al., 2005). This is in contrast to the decrease in photosynthesis and growth 

7 in many plants that are exposed to elevated O3. The interactive effects on vegetation have 

8 been the subject of research in the past two decades due to the implications on 

9 productivity and water use of ecosystems. This area of research was discussed in detail in 

10 AX9.3.8.1 of the 2006 O3 AQCD and the conclusions made then are still relevant (U.S. 

11 EPA, 2006b).

12 The bulk of the available evidence shows that, under the various experimental conditions 

13 used (which almost exclusively employed abrupt or “step” increases in CO2

14 concentration, as discussed below), increased CO2 levels (ambient + 200 to 400 ppm) 

15 may protect plants from the negative effects of O3 on growth. This protection may be 

16 afforded in part by CO2 acting together with O3 in inducing stomatal closure, thereby 

17 reducing O3 uptake, and in part by CO2 reducing the negative effects of O3 on Rubisco 

18 and its activity in CO2-fixation. Although both CO2-induced and O3-induced decreases in 

19 stomatal conductance have been observed primarily in short-term studies, recent data 

20 show a long-term and sustained reduction in stomatal conductance under elevated CO2

21 for a number of species (Ainsworth and Long, 2005; Ellsworth et al., 2004; Gunderson et 

22 al., 2002). Instances of increased stomatal conductance have also been observed in 

23 response to O3 exposure, suggesting partial stomatal dysfunction after extended periods 

24 of exposure (Paoletti and Grulke, 2010; Grulke et al., 2007a; Maier-Maercker, 1998).

25 Important caveats must be raised with regard to the findings presented in published 

26 research. The first caveat concerns the distinctly different natures of the exposures to O3

27 and CO2 experienced by plants in the field. Changes in the ambient concentrations of 

28 these gases have very different dynamics. In the context of climate change, CO2 levels 

29 increase relatively slowly (globally 2 ppm/year) and may change little over several 

30 seasons of growth. On the other hand, O3 presents a fluctuating stressor with considerable 

31 hour-to-hour, day-to-day and regional variability (Polle and Pell, 1999). Almost all of the 

32 evidence presented comes from experimentation involving plants subjected to an abrupt

33 step increase to a higher, steady CO2 concentration. In contrast, the O3 exposure 

34 concentrations usually varied from day to day. Luo and Reynolds (1999), Hui et al. 

35 (2002), and Luo (2001) noted the difficulties in predicting the likely effects of a gradual 

36 CO2 increase from experiments involving a step increase or those using a range of CO2
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1 concentrations. It is also important to note that the levels of elevated CO2 in many of the 

2 studies will not be experienced in the field for 30 or 40 years, but elevated levels of O3

3 can occur presently in several areas of the U.S. Therefore, the CO2 × O3 interaction 

4 studies may be less relevant for current ambient conditions.

5 Another caveat concerns the interactions of O3 and CO2 with other climatic variables, 

6 such as temperature and precipitation. In light of the key role played by temperature in 

7 regulating physiological processes and modifying plant response to increased CO2 levels 

8 (Morison and Lawlor, 1999; Long, 1991) and the knowledge that relatively modest 

9 increases in temperature may lead to dramatic consequences in terms of plant 

10 development (Lawlor, 1998), it is important to consider that studying CO2 and O3

11 interactions alone may not create a complete understanding of effects on plants under 

12 future climate change.

9.4.9 Insects and Other Wildlife

9.4.9.1 Insects

13 Insects may respond indirectly to changes in plants (i.e., increased reactive oxygen 

14 species, altered phytochemistry, altered nutrient content) that occur under elevated O3

15 conditions, or O3 can have a direct effect on insect performance (Menendez et al., 2009). 

16 Effects of O3 on insects occur at the species level (i.e., growth, survival, reproduction, 

17 development, feeding behavior) and at the population and community-level 

18 (i.e., population growth rate, community composition). In general, effects of O3 on 

19 insects are highly context- and species-specific (Lindroth, 2010; Bidart-Bouzat and Imeh20 Nathaniel, 2008). Furthermore, plant responses to O3 exposure and herbivore attack have 

21 been demonstrated to share signaling pathways, complicating characterization of these 

22 stressors (Lindroth, 2010; Menendez et al., 2010, 2009). Although both species-level and 

23 population and community-level responses to elevated O3 are observed in field and 

24 laboratory studies discussed below, there is no consensus on how insects respond to 

25 feeding on O3–exposed plants. 

Species-level responses

26 In considering insect growth, survival and reproduction in elevated O3 conditions, several 

27 studies have indicated an effect while others have found no correlation. The performance 

28 of five herbivore species (three moths and two weevils) was assessed in an OTC 

29 experiment at 2 × ambient concentration (Peltonen et al., 2010). Growth of larvae of the 
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1 Autumnal moth, Epirrita autumna, was significantly decreased in the O3 treatment while 

2 no effects were observed in the other species. In an aphid oviposition preference study 

3 using birch buds grown in a three year OTC experiment, O3 had neither a stimulatory or 

4 deterring effect on egg-laying (Peltonen et al., 2006). Furthermore, changes in birch bud 

5 phenolic compounds associated with the doubled ambient concentrations of O3 did not 

6 correlate with changes in aphid oviposition (Peltonen et al., 2006). Reproduction in 

7 Popillia japonica, that were fed soybeans and grown under elevated O3 appeared to be 

8 unaffected (O'Neill et al., 2008). In a meta-analysis of effects of elevated O3 on 22 

9 species of trees and 10 species of insects, the rates of survival, reproduction and food 

10 consumption were typically unaffected while development times were reduced and pupal 

11 masses were increased (Valkama et al., 2007).

12 At the Aspen FACE site insect performance under elevated (50-60 ppb) O3 conditions 

13 (approximately 1.5 × background ambient levels of 30-40 ppb O3) have been considered 

14 for several species. Cumulative fecundity of aphids (Cepegillettea betulaefoliae), that 

15 were reared on O3-exposed paper birch (Betula papyrifera) trees, was lower than aphids 

16 from control plots (Awmack et al., 2004). No effects on growth, development, adult 

17 weight, embryo number and birth weight of newborn nymphs were observed. In a study 

18 conducted using three aspen genotypes, performance of the aspen beetle (Chrysomela 

19 crochi) decreased across all parameters measured (development time, adult mass and 

20 survivorship) under elevated O3 (Vigue and Lindroth, 2010). There was an increase in the 

21 development time of male and female aspen beetle larvae although the percentages varied 

22 across genotypes. Decreased beetle adult mass and survivorship was observed across all 

23 genotypes under elevated O3 conditions. Another study from the Aspen FACE site did 

24 not find any significant effects of elevated O3 on performance (longevity, fecundity, 

25 abundance) of the invasive weevil (Polydrusus sericeus) (Hillstrom et al., 2010b). 

26 Since the 2006 O3 AQCD, several studies have considered the effect of elevated O3 on 

27 feeding behavior of insects. In a feeding preference study, the common leaf weevil 

28 (Phyllobius pyri) consumed significantly more leaf discs from one aspen clone when 

29 compared to a second clone under ambient air conditions (Freiwald et al., 2008). In a 

30 moderately elevated O3 environment (1.5 × ambient), this preference for a certain aspen 

31 clone was less evident, however, leaves from O3-exposed trees were significantly 

32 preferred to leaves grown under ambient conditions. Soybeans grown under enriched O3

33 had significantly less loss of leaf tissue to herbivory in August compared to earlier in the 

34 growing season (July) when herbivory was not affected (Hamilton et al., 2005). Other 

35 plant-herbivore interactions have shown no effects of elevated O3 on feeding. Feeding 

36 behavior of Japanese beetles (P. japonica) appeared to be unchanged when beetles were 

37 fed soybean leaves grown under elevated O3 conditions (O'Neill et al., 2008). At the 

38 Aspen FACE site, feeding by the invasive weevil (Polydrusus sericeus), as measured by 
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1 leaf area consumption, was not significantly different between foliage that was grown 

2 under elevated O3 versus ambient conditions (Hillstrom et al., 2010b). 

Population-level and community-level responses

3 Recent data on insects provide evidence of population-level and community-level 

4 responses to O3. Elevated levels of O3 can affect plant phytochemistry and nutrient 

5 content which in turn can alter population density and structure of the associated 

6 herbivorous insect communities and impact ecosystem processes (Cornelissen, 2011; 

7 Lindroth, 2010). In 72-hour exposures to elevated O3, mean relative growth rate of the 

8 aphid Diuraphis noxia increased with O3 concentration suggesting that more rapid 

9 population growth may occur when atmospheric O3 is elevated (Summers et al., 1994). In 

10 a long-term study of elevated O3 on herbivore performance at the Aspen FACE site, 

11 individual performance and population-level effects of the aphid C. betulaefoliae were 

12 assessed. Elevated O3 levels had a strong positive effect on the population growth rates of 

13 the aphids; although effects were not detected by measuring growth, development, adult 

14 weight, embryo number or birth weight of newborn nymphs (Awmack et al., 2004). 

15 Conversely, a lower rate of population growth was observed in aphids previously 

16 exposed to O3 in an OTC (Menendez et al., 2010). No direct effects of O3 were observed; 

17 however, nymphs born from adults exposed to and feeding on O3 exposed plants were 

18 less capable of infesting new plants when compared to nymphs in the control plots 

19 (Menendez et al., 2010). Elevated O3 reduced total arthropod abundance by 17% at 

20 Aspen FACE, largely as a result of the negative effects on parasitoids, although phloem21 feeding insects may benefit (Hillstrom and Lindroth, 2008). Herbivore communities 

22 affected by O3 and N were sampled along an air pollution gradient in the Los Angeles 

23 basin (Jones and Paine, 2006). Abundance, diversity, and richness of herbivores were not 

24 affected. However, a shift in community structure, from phloem-feeding to chewing 

25 dominated communities, was observed along the gradient. No consistent effect of 

26 elevated O3 on herbivory or insect population size was detected at SoyFACE (O'Neill et 

27 al., 2010; Dermody et al., 2008).

28 Evidence of modification of insect populations and communities in response to elevated 

29 O3 includes genotypic and phenotypic changes. In a study conducted at the Aspen FACE 

30 site, elevated O3 altered the genotype frequencies of the pea aphid (Acyrthosiphon pisum) 

31 grown on red clover (Trifolium pratense) over multiple generations (Mondor et al., 

32 2005). Aphid color was used to distinguish between the two genotypes. Ozone increased 

33 the genotypic frequencies of pink-morph:green-morph aphids from 2:1 to 9:1, and 

34 depressed wing-induction responses more strongly in the pink than the green genotype 

35 (Mondor et al., 2005). Growth and development of individual green and pink aphids 

36 reared as a single genotype or mixed genotypes were unaffected by elevated O3 (Mondor 
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1 et al., 2010). However, growth of pea aphid populations is not readily predictable using 

2 individual growth rates.

9.4.9.2 Wildlife 

Herpetofauna

3 Since the 2006 O3 AQCD, direct effects of O3 exposure including physiological changes 

4 and alterations of ecologically important behaviors such as feeding and thermoregulation 

5 have been observed in wildlife. These studies have been conducted in limited laboratory 

6 exposures, and the levels of O3 treatment (e.g., 0.2-0.8 ppm) were often unrealistically 

7 higher than the ambient levels. Amphibians may be especially vulnerable to airborne 

8 oxidants due to the significant gas exchange that occurs across the skin (Andrews et al., 

9 2008; Dohm et al., 2008). Exposure to 0.2 ppm to 0.8 ppm O3 for 4 hours resulted in a 

10 decrease of oxygen consumption and depressed lung ventilation in the California tree 

11 frog Pseudacris cadaverina (Mautz and Dohm, 2004). Following a single 4-h inhalation 

12 exposure to 0.8 ppm O3, reduced pulmonary macrophage phagocytosis was observed at 1 

13 and 24 hours postexposure in the marine toad (Bufo marinus) indicating an effect on 

14 immune system function (Dohm et al., 2005). There was no difference in macrophage 

15 function at 48 hours postexposure in exposed and control individuals. 

16 Behavioral effects of O3 observed in amphibians include responses to minimize the 

17 surface area of the body exposed to the air and a decrease in feeding rates (Dohm et al., 

18 2008; Mautz and Dohm, 2004). The adoption of a low-profile “water conservation 

19 posture” during O3 exposure was observed in experiments with the California tree frog 

20 (Mautz and Dohm, 2004). Marine toads, Bufo marinus, exposed to 0.06 μL/L O3 for 

21 4 hours ate significantly fewer mealworms at 1 hour and 48 hours postexposure than 

22 control toads (Dohm et al., 2008). In the same study, escape/exploratory behavior as 

23 measured by total distance moved was not negatively affected in the O3-exposed 

24 individuals as compared to the controls (Dohm et al., 2008). 

25 Water balance and thermal preference in herpetofauna are altered with elevated O3. 

26 Marine toads exposed to 0.8 ppm O3 for 4 hours exhibited behavioral hypothermia when 

27 temperature selection in the toads was assessed at 1, 24 and 48 hours postexposure

28 (Dohm et al., 2001). Ozone-exposed individuals lost almost 5g more body mass on 

29 average than controls due to evaporative water loss. At 24 hours after exposure, the 

30 individuals that had lost significant body mass selected lower body temperatures (Dohm 

31 et al., 2001). Behavioral hypothermia was also observed in reptiles following 4-h 

32 exposures to 0.6 ppm O3. Exposure of the Western Fence Lizard (Sceloporus 
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occidentalis) at 25o 1 C induced behavioral hypothermia that recovered to control 

2 temperatures by 24 hours (Mautz and Dohm, 2004). The behavioral hypothermic 

response persisted in lizards exposed to O3 at 35o 3 C at 24 hours postexposure resulting in a 

mean body temperature of 3.3o 4 C over controls. 

Soil fauna communities

5 Ozone has also been shown to alter soil fauna communities (Meehan et al., 2010; 

6 Kasurinen et al., 2007; Loranger et al., 2004). Abundance of Acari (mites and ticks) 

7 decreased by 47% under elevated O3 at Aspen FACE site, probably due to the higher 

8 secondary metabolites and lower N concentrations in litter and foliage under elevated O3

9 (Loranger et al., 2004). In another study from the Aspen FACE site, leaf litter collected 

10 from aspen grown under elevated O3 conditions was higher in fiber and lignin 

11 concentrations than litter from trees grown under ambient conditions. These chemical 

12 characteristics of the leaves were associated with increased springtail population growth 

13 following 10 weeks in a laboratory microcosm (Meehan et al., 2010). Consumption rates 

14 of earthworms fed on leaf litter for 6 weeks from trees grown under elevated O3

15 conditions and ambient air did not vary significantly between treatments (Meehan et al., 

16 2010). In another study on juvenile earthworms Lumbricus terrestris, individual growth 

17 was reduced when worms were fed high-O3 birch litter from trees exposed for three years 

18 to elevated O3 in an OTC system (Kasurinen et al., 2007). In the same study no 

19 significant growth or mortality effects were observed in isopods. 

9.4.9.3 Indirect Effects on Wildlife

20 In addition to the direct effects of O3 exposure on physiological and behavioral endpoints 

21 observed in the laboratory, there are indirect effects to wildlife. These effects include 

22 changes in biomass and nutritive quality of O3-exposed plants (reviewed in Section 9.4.4) 

23 that are consumed by wildlife. Reduced digestibility of O3-exposed plants may alter 

24 dietary intake and foraging strategies in herbivores. In a study using native highbush 

25 blackberry (Rubus argutus) relative feed value of the plants decreased in bushes exposed 

26 to double ambient concentrations of O3 (Ditchkoff et al., 2009). Indirect effects of 

27 elevated O3 on wildlife include changes in chemical signaling important in ecological 

28 interactions reviewed below.
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Chemical signaling in ecological interactions

1 Ozone has been shown to degrade or alter biogenic VOC signals important to ecological 

2 interactions including; (1) attraction of pollinators and seed dispersers; (2) defense 

3 against herbivory; and (3) predator-prey interactions (Pinto et al., 2010; McFrederick et 

4 al., 2009; Yuan et al., 2009; Pinto et al., 2007a; Pinto et al., 2007b). Each signal released 

5 by emitters has an atmospheric lifetime and a unique chemical signature comprised of 

6 different ratios of individual hydrocarbons that are susceptible to atmospheric oxidants 

7 such as O3 (Yuan et al., 2009; Wright et al., 2005). Under elevated O3 conditions, these 

8 olfactory cues may travel shorter distances before losing their specificity (McFrederick et 

9 al., 2009; McFrederick et al., 2008). Additional non-phytogenic VOC-mediated 

10 interrelationships with the potential to be modified by O3 include territorial marking, 

11 pheromones for attraction of mates and various social interactions including scent trails, 

12 nestmate recognition and signals involved in aggregation behaviors (McFrederick et al., 

13 2009). For example, the alcohols, ketones and aldehydes comprising sex pheromones in 

14 moths could be especially vulnerable to degradation by O3, since some males travel >100 

15 meters to find mates (Carde and Haynes, 2004). In general, effects of O3 on scent16 mediated ecological interactions are highly context- and species-specific (Lindroth, 2010; 

17 Bidart-Bouzat and Imeh-Nathaniel, 2008).

Pollination and seed dispersal

18 Phytogenic VOC’s attract pollinators and seed dispersers to flowers and fruits (Dudareva 

19 et al., 2006; Theis and Raguso, 2005). These floral scent trails in plant-insect interactions 

20 may be destroyed or transformed by O3 (McFrederick et al., 2008). Using a Lagrangian 

21 model, the rate of destruction of phytogenic VOC’s was estimated in air parcels at 

22 increasing distance from a source in response to increased regional levels of O3, hydroxyl 

23 and nitrate radicals (McFrederick et al., 2008). Based on the model, the ability of 

24 pollinators to locate highly reactive VOCs from emitting flowers may have decreased 

25 from kilometers during pre-industrial times to <200 meters at current ambient conditions 

26 (McFrederick et al., 2008). Scents that travel shorter distances (0-10 m) are less 

27 susceptible to air pollutants, while highly reactive scents that travel longer distances (10 

28 to 100’s of meters), are at a higher risk for degradation (McFrederick et al., 2009). For 

29 example, male euglossine bees can detect bait stations from a distance of at least one 

30 kilometer (Dobson, 1994). 
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Defense against herbivory

1 Ozone can alter the chemical signature of VOCs emitted by plants and these VOCs are 

2 subsequently detected by herbivores (Blande et al., 2010; Iriti and Faoro, 2009; Pinto et 

3 al., 2007a; Vuorinen et al., 2004; Jackson et al., 1999; Cannon, 1990). These 

4 modifications can make the plant either more attractive or repellant to phytophagous 

5 insects (Pinto et al., 2010). For example, under elevated O3, the host plant preference by 

6 forest tent caterpillars increased for birch compared to aspen (Agrell et al., 2005). 

7 Ozone-induced emissions from red spruce needles were found to repel spruce budworm 

8 larvae (Cannon, 1990). Transcriptional profiles of field grown soybean (Glycine max) 

9 grown in elevated O3 conditions were altered due to herbivory by Japanese beetles. The 

10 herbivory resulted in a higher number of transcripts in the leaves of O3-exposed plants 

11 and upregulation of antioxidant metabolism associated with plant defense (Casteel et al., 

12 2008). 

13 Ozone may modify signals involved in plant-to-plant interactions and plant defense 

14 against pathogens (Blande et al., 2010; Pinto et al., 2010; McFrederick et al., 2009; Yuan 

15 et al., 2009). In a recent study with lima beans, 80 ppb O3 degraded several 

16 herbivore-induced VOCs, reducing the distance over which plant-to-plant signaling 

17 occurred (Blande et al., 2010).

Predator-prey interactions

18 Elevated O3 conditions are associated with disruption of pheromone-mediated 

19 interactions at higher trophic levels (e.g., predators and parasitoids of herbivores). In a 

20 study from the Aspen FACE site, predator escape behaviors of the aphid (Chatophorus 

21 stevensis) were enhanced on O3-fumigated aspen trees although the mechanism of this 

22 response remains unknown (Mondor et al., 2004). The predatory mite Phytoseiulus 

23 persimilis can distinguish between the VOC signature of ozonated lima bean plants and 

24 ozonated lima bean plants simultaneously damaged by T. urticae (Vuorinen et al., 2004)

25 however, other tritrophic interactions have shown no effect (Pinto et al., 2007b). 

26 There are few studies that consider host location behaviors of parasites under elevated O3. 

27 In closed chambers fumigated with O3, the searching efficiency and proportion of the 

28 host larval fruit flies parasitized by Asobara tabida declined when compared to filtered 

29 air controls (Gate et al., 1995). The host location behavior and rate of parasitism of the 

30 wasp (Coesia plutellae) on Plutella xylostella-infested potted cabbage plants was tested 

31 under ambient and doubled O3 conditions in an open-air fumigation system (Pinto et al., 

32 2008). The number of wasps found in the field and the percentages of parasitized larvae 

33 were not significantly different from controls under elevated O3.
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1 Elevated O3 has the potential to perturb specialized food-web communication in 

2 transgenic crops. In insect-resistant oilseed rape Brassica napus grown under 100 ppb O3

3 in a growth chamber, reduced feeding damage by Putella xylostella led to deceased 

4 attraction of the endoparasitoid (Costesia vestalis), however this tritrophic interaction 

5 was influenced by the degree of herbivore feeding (Himanen et al., 2009a; Himanen et 

6 al., 2009b). Under chronic O3-exposure, the insect resistance trait BT cry1Ac in 

7 transgenic B. napus was higher than the control (Himanen et al., 2009c). There was a 

8 negative relative growth rate of the Bt target herbivore, P. xylostella, in all O3 treatments.

9.4.9.4 Summary 

9 Recent information on O3 effects on insects and other wildlife is limited to a few species 

10 and there is no consensus on how these organisms respond to elevated O3. Studies 

11 published since the last review show impacts of elevated O3 on both species-level 

12 responses (reproduction, growth, feeding behavior) and community and ecosystem-level 

13 responses (population growth, abundance, shift in community structure) in some insects 

14 and soil fauna. Changes in ecologically important behaviors such as feeding and 

15 thermoregulation have recently been observed with O3 exposure in amphibians and 

16 reptiles, however, these responses occur at concentrations of O3 much higher than 

17 ambient levels.

18 Recent information available since the last review considers the effects of O3 on chemical 

19 signaling in insect and wildlife interactions. Specifically, studies on O3 effects on 

20 pollination and seed dispersal, defenses against herbivory and predator-prey interactions 

21 all consider the ability of O3 to alter the chemical signature of VOCs emitted during these 

22 pheromone-mediated events. The effects of O3 on chemical signaling between plants, 

23 herbivores and pollinators as well as interactions between multiple trophic levels is an 

24 emerging area of study that may result in further elucidation of O3 effects at the species, 

25 community and ecosystem-level. 

9.5 Effects-Based Air Quality Exposure Indices and Dose 

Modeling

9.5.1 Introduction

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

27 damage (e.g., reduced growth). They are summary measures of monitored ambient O3
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1 concentrations over time, intended to provide a consistent metric for reviewing and 

2 comparing exposure-response effects obtained from various studies. Such indices may 

3 also provide a basis for developing a biologically-relevant air quality standard for 

4 protecting vegetation and ecosystems. Effects on plant growth and/or yield have been a 

5 major focus of the characterization of O3 impacts on plants for purposes of the air quality 

6 standard setting process (U.S. EPA, 2007b, 1996e, 1986). The relationship of O3 and 

7 plant responses can be characterized quantitatively as “dose-response” or “exposure8 response.” The distinction is in how the pollutant concentration is expressed: “dose” is 

9 the pollutant concentration absorbed by the leaf over some time period, and is very 

10 difficult to measure directly, whereas “exposure” is the ambient air concentration 

11 measured near the plant over some time period, and summarized for that period using an 

12 index. Exposure indices have been most useful in considering the form of the secondary 

13 O3 NAAQS, in large part because they only require ambient air quality data rather than 

14 more complex indirect calculations of dose to the plant. The attributes of exposure 

15 indices that are most relevant to plant damage are the weighting of O3 concentrations and 

16 the daily and seasonal time-periods. Several different types of exposure indices are 

17 discussed in Section 9.5.2.

18 From a theoretical perspective, a measure of plant O3 uptake or dose from ambient air 

19 (either rate of uptake or cumulative seasonal uptake) might be a better predictor of O3

20 damage to plants than an exposure index and may be useful in improving risk assessment. 

21 An uptake estimate would have to integrate all those environmental factors that influence 

22 stomatal conductance, including but not limited to temperature, humidity, and soil water

23 status (Section 9.5.4). Therefore, uptake values are generally obtained with simulation 

24 models that require knowledge of species- and site-specific values for the variables 

25 mentioned. However, a limitation of modeling dose is that environmental variables are 

26 poorly characterized. In addition, it has also been recognized that O3 detoxification 

27 processes and the temporal dynamics of detoxification must be taken into account in dose 

28 modeling (Heath et al., 2009) (Section 9.5.4). Because of this, research has focused 

29 historically on predictors of O3 damage to plants based only on exposure as a summary 

30 measure of monitored ambient pollutant concentration over some integral of time, rather 

31 than dose (U.S. EPA, 1996c; Costa et al., 1992; Lee et al., 1988b; U.S. EPA, 1986; 

32 Lefohn and Benedict, 1982; O'Gara, 1922).

9.5.2 Description of Exposure Indices Available in the Literature

33 Mathematical approaches for summarizing ambient air quality information in biologically 

34 meaningful forms for O3 vegetation effects assessment purposes have been explored for 

35 more than 80 years (U.S. EPA, 1996b; O'Gara, 1922). In the context of national standards 
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1 that protect for “known or anticipated” effects on many plant species in a variety of 

2 habitats, exposure indices provide a numerical summary of very large numbers of 

3 ambient observations of concentration over extended periods. Like any summary statistic, 

4 exposure indices retain information on some, but not all, characteristics of the original 

5 observations. Several indices have been developed to attempt to incorporate some of the 

6 biological, environmental, and exposure factors that influence the magnitude of the 

7 biological response and contribute to observed variability (Hogsett et al., 1988). In the 

8 1996 O3 AQCD, the exposure indices were arranged into five categories; (1) One event, 

9 (2) Mean, (3) Cumulative, (4) Concentration weighted, and (5) Multicomponent, and 

10 were discussed in detail (Lee et al., 1989). Figure 9-9 illustrates how several of the 

11 indices weight concentration and accumulate exposure. For example, the SUM06 index 

12 (panel a) is a threshold-based approach wherein concentrations below 0.06 ppm are given 

13 a weight of zero and concentrations at or above 0.06 ppm are given a weight of 1.0 that is 

14 summed, usually over 3 to 6 months. The Sigmoid approach (panel b), which is similar to 

15 the W126 index (Lefohn et al., 1988; Lefohn and Runeckles, 1987), is a non-threshold 

16 approach wherein all concentrations are given a weight that increases from zero to 1.0 

17 with increasing concentration and summed. 
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(a) SUM06: the upper graphic illustrates an episodic exposure profile; the shaded area under some of the peaks illustrates the 

concentrations greater than or equal to 0.06 ppm that are accumulated in the index. The insert shows the concentration weighting (0 

or 1) function. The lower portion of the graphic illustrates how concentration is accumulated over the exposure period. (b) SIGMOID: 

the upper graphic illustrates an episodic exposure profile; the variable shaded area under the peaks illustrates the 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.

Source: Reprinted with permission of Air and Waste Management Association (Tingey et al., 1991).

Figure 9-9 Diagrammatic representation of several exposure indices 

illustrating how they weight concentration and accumulate 

exposure.

1 This section will primarily discuss SUM06, W126 and AOTx exposure metrics. Below 

2 are the definitions of the three cumulative index forms:

3  SUM06: Sum of all hourly O3 concentrations greater than or equal to 

4 0.06 ppm observed during a specified daily and seasonal time window 

5 (Figure 9-9a).

6  AOTx: Sum of the differences between hourly O3 concentrations greater than 

7 a specified threshold during a specified daily and seasonal time window. For 

8 example, AOT40 is sum of the differences between hourly concentarions 

9 above 0.04 ppm. 

10  W126: Sigmoidally weighted sum of all hourly O3 concentrations observed 

11 during a specified daily and seasonal time window (Lefohn et al., 1988; 

12 Lefohn and Runeckles, 1987), similar to Figure 9-9b). The sigmoidal 
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1 weighting of hourly O3 concentration is given in the equation below, where C

2 is the hourly O3 concentration in ppm:

Equation 9-1

3 These indices have a variety of relevant time windows that may be applied and are 

4 discussed in Section 9.5.3. 

5 Various factors with known or suspected bearing on the exposure-response relationship, 

6 including concentration, time of day, respite time, frequency of peak occurrence, plant 

7 phenology, predisposition, etc., have been weighted with various functions in a large set 

8 of indices. The resulting indices were evaluated by ranking them according to the 

9 goodness-of-fit of a regression model of growth or yield response (Lee et al., 1989). The 

10 statistical evaluations for each of these indices were completed using growth or yield 

11 response data from many earlier exposure studies (e.g., NCLAN). This retrospective 

12 approach was necessary because there were no studies specifically designed to test the 

13 goodness-of-fit of the various indices. The goodness-of-fit of a set of linear and nonlinear 

14 models for exposure-response was ranked as various proposed indices were used in turn 

15 to quantify exposure. This approach provided evidence for the best indices. The results of 

16 retrospective analyses are described below.

17 Most of the early retrospective studies reporting regression approaches used data from the 

18 NCLAN program or data from Corvallis, Oregon or California (Costa et al., 1992; Lee et 

19 al., 1988b; Lefohn et al., 1988; Musselman et al., 1988; Lee et al., 1987; U.S. EPA, 

20 1986). These studies were previously reviewed by the EPA (U.S. EPA, 1996c; Costa et 

21 al., 1992) and were in general agreement that the best fit to the data resulted from using 

22 cumulative concentration-weighted exposure indices (e.g., W126, SUM06). Lee et al. 

23 (1987) suggested that exposure indices that included all the 24-h data performed better 

24 than those that used only 7 hours of data; this was consistent with the conclusions of 

25 Heagle et al. (1987) that plants receiving exposures for an additional 5-h/day showed 

26 10% greater yield loss than those exposed for 7-h/day. In an analysis using the National 

27 Crop Loss Assessment Network (NCLAN) data, Lefohn et al. (1988) found several 

28 indices which only cumulated and weighted higher concentrations (e.g., W126, SUM06, 

29 SUM08, and AOT40) performed very well. Amongst this group no index had 

30 consistently better fits than the other indices across all studies and species (Heagle et al., 

31 1994b; Lefohn et al., 1988; Musselman et al., 1988). Lefohn et al. (1988) found that 

32 adding phenology weighting to the index somewhat improved the performance of the 

c C e

w 126 1 4403

1 − + =
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1 indices. The “best” exposure index was a phenologically weighted cumulative index, 

2 with sigmoid weighting on concentration and a gamma weighting function as a surrogate 

3 for plant growth stage. This index provided the best statistical fit when used in the models 

4 under consideration, but it required data on species and site conditions, making 

5 specification of weighting functions difficult for general use. 

6 Other factors, including predisposition time (Hogsett et al., 1988; McCool et al., 1988)

7 and crop development stage (Tingey et al., 2002; Heagle et al., 1991) contributed to 

8 variation in the biological response and suggested the need for weighting O3

9 concentrations to account for predisposition time and phenology. However, the roles of 

10 predisposition and phenology in plant response vary considerably with species and 

11 environmental conditions; therefore, specification of a weighting function for general use 

12 in characterizing plant exposure has not been possible.

13 European scientists took a similar approach in developing indices describing growth and 

14 yield loss in crops and tree seedlings, using OTCs with modified ambient exposures, but 

15 many fewer species and study locations were employed in the European studies. There is 

16 evidence from some European studies that a lower (Pleijel et al., 1997) or higher (Finnan 

17 et al., 1997; Finnan et al., 1996) cutoff value in indices with a threshold may provide a 

18 better statistical fit to the experimental data. Finnan et al. (1997) used seven exposure 

19 studies of spring wheat to confirm that cumulative exposure indices emphasizing higher 

20 O3 concentrations were best related to plant response and that cumulative exposure 

21 indices using weighting functions, including cutoff concentrations, allometric and

22 sigmoidal, provided a better fit and that the ranking of these indices differed depending 

23 on the exposure-response model used. Weighting those concentrations associated with 

24 sunshine hours in an attempt to incorporate an element of plant uptake did not improve 

25 the index performance (Finnan et al., 1997). A more recent study using data from several 

26 European studies of Norway spruce, analyzed the relationship between relative biomass 

27 accumulation and several cumulative, weighted indices, including the AOT40 (area over 

28 a threshold of 40ppb) and the SUM06 (Skarby et al., 2004). All the indices performed 

29 relatively well in regressing biomass and exposure index, with the AOT20 and AOT30 

doing slightly better than others (r2 30 = 0.46-0.47). In another comparative study of four 

31 independent data sets of potato yield and different cumulative uptake indices with 

different cutoff values, a similarly narrow range of r2 was observed (r2 32 = 0.3-0.4) (Pleijel 

33 et al., 2004b).

34 In Europe, the cutoff concentration-weighted index AOT40 was selected in developing 

35 exposure-response relationships based on OTC studies of a limited number of crops and 

36 trees (Grunhage and Jager, 2003). The United Nations Economic Commission for Europe 

37 (UNECE, 1988) adopted the critical levels approach for assessment of O3 risk to 
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1 vegetation across Europe. As used by the UNECE, the critical levels are not like the air 

2 quality regulatory standards used in the U.S., but rather function as planning targets for 

3 reductions in pollutant emissions to protect ecological resources. Critical levels for O3 are 

4 intended to prevent long-term deleterious effects on the most sensitive plant species 

5 under the most sensitive environmental conditions, but not intended to quantify O3

6 effects. A critical level was defined as “the concentration of pollutant in the atmosphere 

7 above which direct adverse effects on receptors, such as plants, ecosystems, or materials 

8 may occur according to present knowledge” (UNECE, 1988). The nature of the “adverse 

9 effects” was not specified in the original definition, which provided for different levels 

10 for different types of harmful effect (e.g., visible injury or loss of crop yield). There are 

11 also different critical levels for crops, forests, and semi-natural vegetation. The caveat, 

12 “according to present knowledge” is important because critical levels are not rigid; they 

13 are revised periodically as new scientific information becomes available. For example, 

14 the original critical level for O3 specified concentrations for three averaging times, but 

15 further research and debate led to the current critical level being stated as the cumulative 

16 exposure (concentration × hours) over a cutoff concentration of 40 ppb (AOT40) (Fuhrer 

17 et al., 1997). 

18 More recently in Europe, a decision was made to work towards a flux-based approach 

19 (see Section 9.5.4) for the critical levels (“Level II”), with the goal of modeling O3

20 flux-effect relationships for three vegetation types: crops, forests, and semi-natural 

21 vegetation (Grunhage and Jager, 2003). Progress has been made in modeling flux (U.S. 

22 EPA, 2006b) and the Mapping Manual is being revised (Ashmore et al., 2004a, b; 

23 Grennfelt, 2004; Karlsson et al., 2003). The revisions may include a flux-based approach 

24 for three crops: wheat, potatoes, and cotton. However, because of a lack of flux-response 

25 data, a cumulative, cutoff concentration-based (AOTx) exposure index will remain in use 

26 for the near future for most crops and for forests and semi-natural herbaceous vegetation 

27 (Ashmore et al., 2004b).

28 In both the U.S. and Europe, the adequacy of these numerical summaries of exposure in 

29 relating biomass and yield changes have, for the most part, all been evaluated using data 

30 from studies not necessarily designed to compare one index to another (Skarby et al., 

31 2004; Lee et al., 1989; Lefohn et al., 1988). Very few studies in the U.S. have addressed 

32 this issue since the 2006 O3 AQCD. McLaughlin et al. (2007a) reported that the 

33 cumulative exposure index of AOT60 related well to reductions in growth rates at forest 

34 sites in the southern Appalachian Mountains. However, the authors did not report an 

35 analysis to compare multiple indices. Overall, given the available data from previous O3

36 AQCDs and the few recent studies, the cumulative, concentration-weighted indices 

37 perform better than the peak or mean indices. It is still not possible, however, to 
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1 distinguish the differences in performance among the cumulative, concentration-weighted 

2 indices. 

3 The main conclusions from the 1996 and 2006 O3 AQCDs regarding an index based on 

4 ambient exposure are still valid. No information has come forth since the 2006 O3 AQCD 

5 to alter those conclusions. These key conclusions can be restated as follows:

6  O3 effects in plants are cumulative;

7  higher O3 concentrations appear to be more important than lower 

8 concentrations in eliciting a response;

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

10 and

11  quantifying exposure with indices that accumulate the O3 hourly 

12 concentrations and preferentially weight the higher concentrations improves 

13 the explanatory power of exposure/response models for growth and yield, over 

14 using indices based on mean and peak exposure values.

15 Following the 2006 criteria review process (U.S. EPA, 2006b), the EPA proposed an 

16 alternative form of the secondary NAAQS for O3 using a cumulative, concentration17 weighted exposure index to protect vegetation from damage (72 FR 37818). The EPA 

18 considered two specific concentration-weighted indices: the cutoff concentration 

19 weighted SUM06 and the sigmoid-weighted W126 exposure index (U.S. EPA, 2007b). 

20 These two indices performed equally well in predicting the exposure-response 

21 relationships observed in the crop and tree seedlings studies (Lee et al., 1989). At a 

22 workshop convened to consider the science supporting these indices (Heck and Cowling, 

23 1997) there was a consensus that these cumulative concentration-weighted indices being 

24 considered were equally capable of predicting plant response. It should be noted that 

25 there are some important differences between the SUM06 and W126. When considering 

26 the response of vegetation to O3 exposures represented by the threshold (e.g., SUM06) 

27 and non-threshold (e.g., W126) indices, the W126 metric does not have a cut-off in the 

28 weighting scheme as does SUM06 and thus it includes consideration of potentially 

29 damaging exposures below 60 ppb. The W126 metric also adds increasing weight to 

30 hourly concentrations from about 40 ppb to about 100 ppb (Lefohn et al., 1988; Lefohn 

31 and Runeckles, 1987). This is unlike cut-off metrics such as the SUM06 where all 

32 concentrations above 60 ppb are treated equally. This is an important feature of the W126 

33 since as hourly concentrations become higher, they become increasingly likely to 

34 overwhelm plant defenses and are known to be more detrimental to vegetation (See 

35 Section 9.5.3.1). 
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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 should be looked at in the context of 

7 plant 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 forest 

22 changes in the San Bernardino National Forest, using an expanded network of monitoring 

23 sites, has verified significant changes in growth, mortality rates, basal area, and species 

24 composition throughout the area since 1974 (Arbaugh et al., 2003). A model simulation 

25 of ponderosa pine growth over the 40-year period in the San Bernardino National Forest 

26 showed a significant impact of O3 exposure on tree growth and indicates improved 

27 growth with reduced O3 concentrations. This area has also experienced elevated 

28 N deposition and based on a number of environmental indicators, it appears that this area 

29 is experiencing N saturation (Fenn et al., 1996). To account for this potential interaction, 

30 the model simulations were conducted under conditions of unlimited soil N. The actual 

31 interactions are not known. The improvement in growth over the years was attributed to 

32 improved air quality, but no distinction was made regarding the relative role of 

33 “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 

39 “mid-range” concentrations from 1980 to 1986, which corresponds to the period after 
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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.

Note: 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/html/ds.htm (Cal EPA, 2010). 

Source: Reprinted with permission of Elsevier Science Ltd. (Lee et al., 2003a).

Figure 9-10 Trends in May to September: 12-hour SUM06, Peak 1-hour 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 occur in early to late afternoon when stomata are often partially closed and conductances 

8 are lower. Total O3 flux depends on atmospheric and boundary layer resistances, both of 

9 which exhibit variability throughout the day. Experimental studies with tree species 

10 demonstrated the decoupling of ambient O3 exposure, peak occurrence, and gas 

11 exchange, particularly in areas of drought (Panek, 2004). Several studies have suggested 

12 that ponderosa pine trees in the southern and northern Sierra Nevada Mountains may not 

13 be as susceptible to high O3 concentrations as to lower concentrations, due to reduced
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1 needle conductance and O3 uptake during the period when the highest concentrations 

2 occur (Panek et al., 2002; Panek and Goldstein, 2001; Bauer et al., 2000; Arbaugh et al., 

3 1998). Panek et al. (2002) compared direct O3 flux measurements into a canopy of 

4 ponderosa pine and demonstrated a lack of correlation of daily patterns of conductance 

5 and O3 occurrence, especially in the late season drought period; the authors concluded 

6 that a consideration of climate or season was essential, especially considering the role of 

7 soil moisture and conductance/uptake. In contrast, Grulke et al. (2002) reported high 

8 conductance when O3 concentrations were high in the same species, but under different 

9 growing site conditions. The longer-term biological responses reported by Miller and 

10 Rechel (1999) for ponderosa pine in the same region, and the general reduction in recent 

11 years in ambient O3 concentrations, suggest that stomatal conductance alone may not be a 

12 sufficient indicator of potential vegetation injury or damage. Another consideration for 

13 the effect of O3 uptake is the diurnal pattern of detoxification capacity of the plant. The 

14 detoxification capacity may not follow the same pattern as stomatal conductance (Heath 

15 et al., 2009).

16 The use of a 12-h (8:00 a.m. to 8:00 p.m.) daylight period for a W126 cumulating 

17 exposure was based primarily on evidence that the conditions for uptake of O3 into the 

18 plant occur mainly during the daytime hours. In general, plants have the highest stomatal 

19 conductance during the daytime and in many areas atmospheric turbulent mixing is 

20 greatest during the day as well (Uddling et al., 2010; U.S. EPA, 2006b). However, 

21 notable exceptions to maximum daytime conductance are cacti and other plants with 

22 crassulacean acid metabolism (CAM photosynthesis) which only open their stomata at 

23 night. This section will focus on plants with C3 and C4 photosynthesis, which generally 

24 have maximum stomatal conductance during the daytime.

25 Recent reviews of the literature reported that a large number of species had varying 

26 degrees of nocturnal stomatal conductance (Caird et al., 2007; Dawson et al., 2007; 

27 Musselman and Minnick, 2000). The reason for night-time water loss through stomata is 

28 not well understood and is an area of active research (e.g., Christman et al., 2009; 

29 Howard et al., 2009). Night-time stomatal opening may be enhanced by O3 damage that 

30 could result in loss of stomatal control, and less complete closure of stomata, than under 

31 low O3 conditions (Caird et al., 2007; Grulke et al., 2007b). In general, the rate of 

32 stomatal conductance at night is much lower than during the day (Caird et al., 2007). 

33 Atmospheric turbulence at night is also often low, which results in stable boundary layers 

34 and unfavorable conditions for O3 uptake into vegetation (Finkelstein et al., 2000). 

35 Nevertheless, nocturnal turbulence does intermittently occur and may result in 

36 non-negligible O3 flux into the plants. In addition, plants might be more susceptible to O3

37 exposure at night than during the daytime, because of potentially lower plant defenses 

38 (Heath et al., 2009; Loreto and Fares, 2007; Musselman et al., 2006; Musselman and 
Draft – Do Not Cite or Quote 9-117 June 2012

1 Minnick, 2000). For significant nocturnal stomatal flux and O3 effects to occur, specific 

2 conditions must exist. A susceptible plant with nocturnal stomatal conductance and low 

3 defenses must be growing in an area with relatively high night-time O3 concentrations 

4 and appreciable nocturnal atmospheric turbulence. It is unclear how many areas there are 

5 in the U.S. where these conditions occur. It may be possible that these conditions exist in 

6 mountainous areas of southern California, front-range of Colorado (Turnipseed et al., 

7 2009) and the Great Smoky Mountains of North Carolina and Tennessee. Tobiessen 

8 (1982) found that shade intolerant tree species showed opening of stomata in the dark and 

9 did not find this in shade tolerant species. This may indicate shade intolerant trees may be 

10 more likely to be susceptible to O3 exposure at night. More information is needed in 

11 locations with high night-time O3 to assess the local O3 patterns, micrometeorology and 

12 responses of potentially vulnerable plant species. 

13 Several field studies have attempted to quantify night-time O3 uptake with a variety of 

14 methods. However, many of these studies have not linked the night-time flux to measured 

15 effects on plants. Grulke et al. (2004) showed that the stomatal conductance at night for 

16 ponderosa pine in the San Bernardino National Forest (CA) ranged from one tenth to one 

17 fourth that of maximum daytime stomatal conductance. In June, at a high-elevation site, it 

18 was calculated that 11% of the total daily O3 uptake of pole-sized trees occurred at night. 

19 In late summer, however, O3 uptake at night was negligible. However, this study did not 

20 consider the turbulent conditions at night. Finkelstein et al. (2000) investigated O3

21 deposition velocity to forest canopies at three different sites. The authors found the total 

22 flux (stomatal and non-stomatal) to the canopy to be very low during night-time hours as 

23 compared to day-time hours. However, the authors did note that higher nocturnal 

24 deposition velocities at conifer sites may be due to some degree of stomatal opening at 

25 night (Finkelstein et al., 2000). Work by Mereu et al. (2009) in Italy on Mediterranean 

26 species indicated that nocturnal uptake was from 10 to 18% of total daily uptake during a 

27 weak drought and up to 24% as the drought became more pronounced. The proportion of 

28 night-time uptake was greater during the drought due to decreases in daytime stomatal 

29 conductance (Mereu et al., 2009). In a study conducted in California, (Fares et al., 2011)

30 reported that calculated mean percentages of nocturnal uptake were 5%, 12.5%, 6.9% of 

31 total O3 uptake for lemon, mandarin, and orange, respectively. In another recent study at 

32 the Aspen FACE site in Wisconsin, calculated leaf-level stomatal O3 flux was near zero 

33 from the night-time hours of 8:00 p.m. to 5:00 a.m. (Uddling et al., 2010). This was likely 

34 due to low horizontal wind speed (>1 meter/sec) and low O3 concentrations (<25 ppb) 

35 during those same night-time hours (Figure 9-12).
Draft – Do Not Cite or Quote 9-118 June 2012

Note: Subscripts “max” and “min” refer to stomatal fluxes calculated neglecting and accounting for potential non-stomatal ozone flux, 

respectively.

Source: Reprinted with permission of Elsevier Ltd. (Uddling et al., 2010).

Figure 9-12 Diurnal (a) conductance through boundary layer and stomata (gbs), 

(b) ozone concentration, and leaf-level stomatal ozone flux (Fst0l) 

in control plots from mid-June through August in (c) 2004 and 

(d) 2005 in the Aspen FACE experiment.

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 

10 exposed to O3 during the day or night showed little significant difference in the amounts 
Draft – Do Not Cite or Quote 9-119 June 2012

1 of injury or reduced growth response to O3 treatment, although the stomatal conductance 

2 was 70-80% lower at night (Winner et al., 1989). These studies show that effects can be 

3 seen with night-time exposures to O3 but when atmospheric conditions are stable at night, 

4 it is uncertain how these exposures may affect plants and trees with complex canopies in 

5 the field.

Seasonal exposure

6 Vegetation across the U.S. has widely varying periods of physiological activity during the 

7 year due to variability in climate and phenology. In order for a particular plant to be 

8 vulnerable to O3 pollution, it must have foliage and be physiologically active. Annual 

9 crops are typically grown for periods of two to three months. In contrast, perennial 

10 species may be photosynthetically active longer (up to 12 months each year for some 

11 species) depending on the species and where it is grown. In general, the period of 

12 maximum physiological activity and thus, potential O3 uptake for vegetation coincides 

13 with some or all of the intra-annual period defined as the O3 season, which varies on a 

14 state-by-state basis (Figure 3-24). This is because the high temperature and high light 

15 conditions that typically promote the formation of tropospheric O3 also promote 

16 physiological activity in vegetation. There are very limited exceptions to this pattern 

17 where O3 can form in the winter in areas in the western U.S. with intense natural gas 

18 exploration (Pinto, 2009), but this is typically when plants are dormant and there is little 

19 chance of O3 uptake. Given the significant variability in growth patterns and lengths of 

20 growing season among the wide range of vegetation species that may experience adverse 

21 effects associated with O3 exposure, no single time window of exposure can work 

22 perfectly for all types of vegetation.

23 Various intra-annual averaging and accumulation time periods have been considered for 

24 the protection of vegetation. The 2007 proposal for the secondary O3 standard (75 FR 

25 37818) proposed to use the maximum consecutive 3-month period within the O3 season. 

26 The U.S. Forest Service and federal land managers have used a 24-h W126 accumulated 

27 for 6 months from April through September (U.S. Forest Service, 2000). However, some 

28 monitors in the U.S. are operational for as little as four months and would not have 

29 enough data for a 6-month seasonal window. The exposure period in the vast majority of 

30 O3 exposure studies conducted in the U.S. has been much shorter than 6 months. Most of 

31 the crop studies done through NCLAN had exposures less than three months with an 

32 average of 77 days. Open-top chamber studies of tree seedlings, compiled by the EPA, 

33 had an average exposure of just over three months or 99 days. In more recent FACE 

34 experiments, SoyFACE exposed soybeans for an average of approximately 120 days per 

35 year and the Aspen FACE experiment exposed trees to an average of approximately 

36 145 days per year of elevated O3, which included the entire growing season at those 
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1 particular sites. Despite the possibility that plants may be exposed to ambient O3 longer 

2 than 3 months in some locations, there is generally a lack of exposure experiments 

3 conducted for longer than 3 months.

A

B

Note: 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).

Figure 9-13 Maximum 3-month, 12-h W126 plotted against maximum 6-month, 

12-h W126. 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
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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 

15 (i.e., flux or deposition). Interest has been increasing in recent years, particularly in 

16 Europe, in using mathematically tractable flux models for O3 assessments at the regional, 

17 national, and European scale (Matyssek et al., 2008; Paoletti and Manning, 2007; ICP 

18 M&M, 2004; Emberson et al., 2000b; Emberson et al., 2000a). Some researchers have 

19 claimed that using flux models can be used to better predict vegetation responses to O3

20 than exposure-based approaches (Matyssek et al., 2008). However, other research has 

21 suggested that flux models do not predict vegetation responses to O3 better than 

22 exposure-based models, such as AOT40 (Gonzalez-Fernandez et al., 2010). While some 

23 efforts have been made in the U.S. to calculate O3 flux into leaves and canopies (Fares et 

24 al., 2010a; Turnipseed et al., 2009; Uddling et al., 2009; Bergweiler et al., 2008; Hogg et 

25 al., 2007; Grulke et al., 2004; Grantz et al., 1997; Grantz et al., 1995), little information 

26 has been published relating these fluxes to effects on vegetation. The lack of flux data in 

27 the U.S. and the lack of understanding of detoxification processes have made this 

28 technique less viable for 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 9-122 June 2012

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 

14 time-integrated “effective flux” is termed “effective dose.” The uptake mechanisms and 

15 the 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 geographical 

38 scales. 
Draft – Do Not Cite or Quote 9-123 June 2012

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 recent 

26 information is available since 2006 that alters the basic conclusions put forth in the 2006 

27 and 1996 O3 AQCDs. These AQCDs focused on the research used to develop various 

28 exposure indices to help quantify effects on growth and yield in crops, perennials, and 

29 trees (primarily seedlings). The performance of indices was compared through regression 

30 analyses of earlier studies designed to support the estimation of predictive O3 exposure31 response 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 

34 (i.e., flux or deposition). Interest has been increasing in recent years, particularly in 
Draft – Do Not Cite or Quote 9-124 June 2012

1 Europe, in using mathematically tractable flux models for O3 assessments at the regional, 

2 national, and European scale (Matyssek et al., 2008; Paoletti and Manning, 2007; ICP 

3 M&M, 2004; Emberson et al., 2000b; Emberson et al., 2000a). While some efforts have 

4 been made in the U.S. to calculate O3 flux into leaves and canopies (Turnipseed et al., 

5 2009; Uddling et al., 2009; Bergweiler et al., 2008; Hogg et al., 2007; Grulke et al., 2004; 

6 Grantz et al., 1997; Grantz et al., 1995), little information has been published relating 

7 these fluxes to effects on vegetation. There is also concern that not all O3 stomatal uptake 

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

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

10 of detoxification potential may, in fact, show little relationship between O3 stomatal 

11 uptake and plant response (Musselman and Massman, 1999). The lack of data in the U.S. 

12 and the lack of understanding of detoxification processes have made this technique less 

13 viable for 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  quantifying exposure with indices that accumulate the O3 hourly 

22 concentrations and preferentially weight the higher concentrations improves 

23 the explanatory power of exposure/response models for growth and yield, over 

24 using indices based on mean and peak exposure values.

25 Various weighting functions have been used, including threshold-weighted 

26 (e.g., SUM06) and continuous sigmoid-weighted (e.g., W126) functions. Based on 

27 statistical goodness-of-fit tests, these cumulative, concentration-weighted indices could 

28 not be differentiated from one another using data from previous exposure studies. 

29 Additional statistical forms for O3 exposure indices have been discussed in Lee et al. 

30 (1988b). The majority of studies published since the 2006 O3 AQCD do not change 

31 earlier conclusions, including the importance of peak concentrations, and the duration 

32 and occurrence of O3 exposures in altering plant growth and yield.

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

34 cumulate and differentially weight the higher hourly average concentrations and also 

35 include the “mid-level” values continue to offer the most defensible approach for use in 
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1 developing response functions and comparing studies, as well as for defining future 

2 indices for vegetation protection.

9.6 Ozone Exposure-Plant Response Relationships

9.6.1 Introduction

3 The adequate characterization of the effects of O3 on plants for the purpose of setting air 

4 quality standards is contingent not only on the choice of the index used (i.e., SUM06, 

5 W126) to summarize O3 concentrations (Section 9.5), but also on quantifying the 

6 response of the plant variables of interest at specific values of the selected index. The 

7 many factors that determine the response of plants to O3 exposure have been discussed in 

8 previous sections. They include species, genotype and other genetic characteristics 

9 (Section 9.3), biochemical and physiological status (Section 9.3), previous and current 

10 exposure to other stressors (Section 9.4.8), and characteristics of the exposure itself 

11 (Section 9.5). Establishing a secondary air quality standard entails the capability to 

12 generalize those observations, in order to obtain predictions that are reliable enough 

13 under a broad variety of conditions, taking into account these factors. This section 

14 reviews results that have related specific quantitative observations of O3 exposure with 

15 quantitative observations of plant responses, and the predictions of responses that have 

16 been derived from those observations through empirical models. 

17 For four decades, exposure to O3 at ambient concentrations found in many areas of the 

18 U.S. has been known to cause detrimental effects in plants (U.S. EPA, 2006b, 1996b, 

19 1984, 1978a). Results published after the 2006 O3 AQCD continue to support this 

20 finding, and the following sections deal with the quantitative characterizations of the 

21 relationship, and what new insights may have appeared since 2006. Detrimental effects 

22 on plants include visible injury, decreases in the rate of photosynthesis, reduced growth,

23 and reduced yield of marketable plant parts. Most published exposure-response data have 

24 reported O3 effects on the yield of crops and the growth of tree seedlings, and those two 

25 variables have been the focus of the characterization of ecological impacts of O3 for the 

26 purpose of setting secondary air quality standards. In order to support quantitative 

27 modeling of exposure-response relationships, data should preferably include more than 

28 three levels of exposure, and some control of potential confounding or interacting factors 

29 should be present in order to model the relationship with sufficient accuracy. Letting 

30 potential confounders, such as other stressors, vary freely when generating O3 exposure31 response data might improve the ‘realism’ of the data, but it also greatly increases the 

32 amount of data necessary to extract a clear quantitative description of the relationship. 
Draft – Do Not Cite or Quote 9-126 June 2012

1 Conversely however, experimental settings should not be so exhaustively restrictive as to 

2 make generalization outside of them problematic. During the last four decades, many of 

3 the studies of the effects of O3 on growth and yield of plants have not included enough 

4 levels of O3 to parameterize more than the simplest linear model. The majority of these 

5 studies have only contrasted two levels, ambient and elevated, or sometimes three by 

6 adding carbon filtration in OTC studies, with little or no consideration of quantitatively 

7 relating specific values of exposure to specific values of growth or yield. This is not to 

8 say that studies that did not include more than two or three levels of O3 exposure, or 

9 studies that were conducted in uncontrolled environments, do not provide exposure10 response information that is highly relevant to reviewing air quality standards. In fact, 

11 they can be essential in verifying the agreement between predictions obtained through the 

12 empirical models derived from experiments such as NCLAN, and observations. The 

13 consensus of model predictions and observations from a variety of studies conducted in 

14 other locations, at other times, and using different exposure methods, greatly increases 

15 confidence in the reliability of both. Furthermore, if they are considered in the aggregate, 

16 studies with few levels of exposure or high unaccounted variability can provide 

17 additional independent estimates of decrements in plant growth and yield, at least within 

18 a few broad categories of exposure. 

19 Extensive exposure-response information on a wide variety of plant species has been 

20 produced by two long-term projects that were designed with the explicit aim of obtaining 

21 quantitative characterizations of the response of such an assortment of crop plants and 

22 tree seedlings to O3 under North American conditions: the NCLAN project for crops, and 

23 the EPA National Health and Environmental Effects Research Laboratory, Western 

24 Ecology Division tree seedling project (NHEERL/WED). The NCLAN project was 

25 initiated by the EPA in 1980 primarily to improve estimates of yield loss under field 

26 conditions and to estimate the magnitude of crop losses caused by O3 throughout the U.S. 

27 (Heck et al., 1991; 1982). The cultural conditions used in the NCLAN studies 

28 approximated typical agronomic practices, and the primary objectives were: (1) to define 

29 relationships between yields of major agricultural crops and O3 exposure as required to 

30 provide data necessary for economic assessments and development of O3 NAAQS; (2) to 

31 assess the national economic consequences resulting from O3 exposure of major 

32 agricultural crops; and (3) to advance understanding of cause-and-effect relationships that 

33 determine crop responses to pollutant exposures.

34 NCLAN experiments yielded 54 exposure-response curves for 12 crop species, some of 

35 which were represented by multiple cultivars at several of 6 locations throughout the U.S. 

36 The NHEERL/WED project was initiated by EPA in 1988 with similar objectives for tree 

37 species, and yielded 49 exposure-responses curves for multiple genotypes of 11 tree 

38 species grown for up to three years in Oregon, Michigan, and the Great Smoky 
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1 Mountains National Park. Both projects used OTCs to expose plants to three to five 

2 levels of O3. Eight of the 54 crop datasets were from plants grown under a combination 

3 of O3 exposure and experimental drought conditions. Figure 9-14 through Figure 9-17

4 summarize some of the NCLAN and NHEERL/WED results.

5 It should be noted that data from FACE experiments might also be used for modeling 

6 exposure-response. They only use two levels of O3 (ambient concentration at the site and 

7 a multiple of it), but given that the value of both levels of exposure changes every year, 

8 and that they are typically run for many consecutive years, aggregating data over time 

9 produces twice as many levels of O3 as there are years. As described in Section 9.2.4, 

10 FACE experiments seek to impose fewer constraints on the growth environment than 

11 OTCs. As a consequence, FACE studies have to contend with larger variability, 

12 especially year-to-year variability, but the difference in experimental conditions between 

13 the two methodologies makes comparisons between their results especially useful.

14 Growth and yield of at least one crop (soybean) has been investigated in yearly 

15 experiments since 2001 at a FACE facility in Illinois (UIUC, 2010; Morgan et al., 2006). 

16 However, almost all analyses of SoyFACE published so far have been based on subsets 

17 of one or two years, and have only contrasted ambient versus elevated O3 as categorical 

18 variables. They have not modeled the response of growth and yield to O3 exposure 

19 continuously over the range of exposure values that have occurred over time. The only 

20 exception is a study by Betzelberger et al. (2010), who used a linear regression model on 

21 data pooled over 2 years. Likewise, trees of three species (trembling aspen, paper birch, 

22 and sugar maple) were grown between 1998 and 2009 in a FACE experiment located in 

23 Rhinelander, Wisconsin (Pregitzer et al., 2008; Dickson et al., 2000). The Aspen FACE 

24 experiment has provided extensive data on responses of trees beyond the seedling stage 

25 under long-term exposure, and also on ecosystem-level responses (Section 9.4), but the 

26 only attempt to use those data in a continuous model of the response of tree growth to O3

27 exposure (Percy et al., 2007) suffered from severe methodological problems, some of 

28 which are discussed in Section 9.6.3. Finally, one experiment was able to exploit a 

29 naturally occurring gradient of O3 concentrations to fit a linear regression model to the 

30 growth of cottonwood (Gregg et al., 2006, 2003). Factors such as genotype, soil type and 

31 soil moisture were under experimental control, and the authors were able to partition out 

32 the effects of potential confounders such as temperature, atmospheric N deposition, and 

33 ambient CO2.

34 A serious difficulty in assessing results of exposure-response research is the multiplicity 

35 of O3 metrics that have been used in reporting. As described in Section 9.5, metrics that 

36 entail either weighting or thresholding of hourly values cannot be algebraically converted 

37 into one another, or into unweighted metrics such as hourly average. When computing O3
Draft – Do Not Cite or Quote 9-128 June 2012

1 exposure using weighted or thresholded metrics, each metric has to be computed 

2 separately from the original hourly data. Comparisons of exposure-response models can 

3 only be made between studies that used the same metric, and the value of exposure at 

4 which a given plant response is expected using one metric of exposure cannot be exactly 

5 converted to another metric. Determining the exposure value at which an effect would be 

6 observed in a different metric can only be accomplished by first computing the 

7 experimental exposures in this metric from the hourly data, then estimating (fitting) 

8 model coefficients again. This problem is irremediable, although useful comparisons 

9 might be made using categorical exposures such as ‘current ambient exposure’ or ‘2050 

10 projected exposure’, which can serve as a common reference for quantitative values 

11 expressed in various metrics. Studies that contained growth or yield exposure-response 

12 data at few levels of exposure, and/or using metrics other than W126 are summarized in 

13 Table 9-18 and Table 9-19.

9.6.2 Estimates of Crop Yield Loss and Tree Seedling Biomass Loss in the 

1996 and 2006 Ozone AQCDs

14 The 1996 and 2006 O3 AQCDs relied extensively on analyses of NCLAN and 

15 NHEERL/WED by Lee et al. (1994); (1989, 1988b, 1987), Hogsett et al. (1997), Lee and 

16 Hogsett (1999), Heck et al. (1984), (Rawlings and Cure, 1985), (Lesser et al., 1990), and 

17 (Gumpertz and Rawlings, 1992). Those analyses concluded that a three-parameter 

18 Weibull model –

β

η α







 







−

=

W126

Y e

Equation 9-2

19 is the most appropriate model for the response of absolute yield and growth to O3

20 exposure, because of the interpretability of its parameters, its flexibility (given the small 

21 number of parameters), and its tractability for estimation. In addition, removing the 

22 intercept α results in a model of relative yield (yield relative to [yield at exposure=0]) 

23 without any further reparameterization. Formulating the model in terms of relative yield 

24 or relative yield loss (yield loss=[1 − relative yield]) is essential in comparing exposure25 response across species, genotypes, or experiments for which absolute values of the 

26 response may vary greatly. In the 1996 and 2006 O3 AQCDs, the two-parameter model of 

27 relative yield was used in deriving common models for multiple species, multiple 

28 genotypes within species, and multiple locations.
Draft – Do Not Cite or Quote 9-129 June 2012

1 Given the disparate species, genotypes, and locations that were included in the NCLAN 

2 and NHEERL/WED projects, and in the absence of plausible distributional assumptions 

3 with respect to those variables, a three step process using robust methods was used to 

4 obtain parameter estimates that could be generalized. The models that were derived for 

5 each species or group of species were referred to as median composite functions. In the 

6 first step, the three parameters of the Weibull model were computed for absolute yield or 

7 biomass data from each NCLAN and NHEERL/WED experiment (54 crop datasets and 

8 49 tree seedling datasets), using nonlinear regression. When data were only available for 

9 three levels of exposure because of experimental problems, the shape parameter β was 

10 constrained to 1, reducing the model to an exponential decay model. In the second step, α 

11 was dropped, and predicted values of relative yield or biomass were then computed for 

12 12-hour W126 exposures between 0 and 60 ppm-h. At each of these W126 exposure 

13 values, the 25th, 50th, and 75th percentiles of the response were identified among the 

14 predicted curves of relative response. For example, for the 34 NCLAN studies of 12 crop 

15 species grown under non-droughted conditions for a complete cropping cycle 

16 (Figure 9-14), the 3 quartiles of the response were identified at every integer value of 

17 W126 between 0 and 60. The third step fitted a two-parameter Weibull model to those 

18 percentiles, yielding the median composite function for the relative yield or biomass 

19 response to O3 exposure for each grouping of interest (e.g., all crops, all trees, all datasets 

20 for one species), as well as composite functions for the other quartiles. In the 1996 and 

21 2006 O3 AQCDs this modeling of crop yield loss and tree seedling biomass loss was 

22 conducted using the SUM06 metric for exposure. This section updates those results by 

23 using the 12-hour W126 as proposed in 2007 (72 FR 37818) and 2010 (75 FR 2938, page 

24 3003). Figure 9-14 through Figure 9-17 present quantiles of predicted relative yield or 

25 biomass loss at seven values of the 12-h W126 for some representative groupings of 

26 NCLAN and NHEERL/WED results. Table 9-9 through Table 9-11 give the 90-day 12-h 

27 W126 O3 exposure values at which 10 and 20% yield or biomass losses are predicted in 

28 50 and 75% of crop or tree species using the composite functions.
Draft – Do Not Cite or Quote 9-130 June 2012

Note: 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.

Source of Weibull parameters: Lee and Hogsett (1996).

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 9-131 June 2012

Notes: 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. 

Source of Weibull parameters: Lee and Hogsett (1996).

Figure 9-15 Quantiles of predicted relative yield loss for 4 crop species in

NCLAN experiments.

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 9-132 June 2012

Note: 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. 

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.

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 9-133 June 2012

Note: 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. 

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.

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 9-134 June 2012

Table 9-9 Ozone exposures at which 10 and 20% yield loss is predicted for 50 

and 75% of crop species.

Predited Yield Loss for 

Crop Speciesa

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

a

Based on composite functions for the 50th and 75th percentiles of 34 Weibull curves for relative yield loss data from 34 nondroughted NCLAN studies of 12 crop species; curves were standardized to 90-day W126.

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 (Droughted versus Watered conditions).

Predicted Yield Loss for Crop Speciesa 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

a

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.

Source of parameters for the 16 curves: Lee and Hogsett (1996).
Draft – Do Not Cite or Quote 9-135 June 2012

Table 9-11 Ozone exposures at which 10 and 20% biomass loss is predicted 

for 50 and 75%of tree species.

Predicted Biomass Loss for 

Tree Speciesa 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

a

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.

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, regardless 

4 of exposure methodology, include only two levels of exposure. In addition, very few 

5 studies have included measurements of exposure using the W126 metric, or the hourly O3

6 concentration data that would allow computing exposure using the W126. Two FACE 

7 projects, however, were conducted over multiple years, and by adding to the number of 

8 exposure levels over time, can support independent model estimation and prediction 

9 using the same model and the same robust process as summarized in Section 9.6.2. 

10 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; UIUC, 2010; Morgan et al., 2006; Morgan et al., 

15 2004). Since 2002, yield data have been collected for up to 8 genotypes of soybean 

16 grown in subplots within each ring. The Aspen FACE project is situated in Rhinelander, 

17 WI, and comprises 12 rings (30m-diameter), 3 of which are exposed to ambient 

18 conditions, and 3 of which are exposed to O3 as a fixed proportion of the instantaneous 

19 ambient concentration (Pregitzer et al., 2008; Karnosky et al., 2005; Dickson et al., 

20 2000). In the summer of 1997, half the area of each ring was planted with small (five to 

21 seven leaf sized) clonally propagated plants of five genotypes of trembling aspen, which 

22 were left to grow in those environments until 2009. Biomass data are currently available 
Draft – Do Not Cite or Quote 9-136 June 2012

1 for the years 1997-2005 (King et al., 2005). Ozone exposure in these two FACE projects 

2 can be viewed as a categorical variable with two levels: ambient, and elevated. However, 

3 this overlooks the facts that not only do both ambient and elevated exposure vary from 

4 year to year, but the proportionality between them also changes yearly. This change has 

5 two sources: first, the dispensing of O3 into the elevated exposure rings varies from the 

6 set point for the ambient/elevated proportionality to some extent, and for SoyFACE, the 

7 set point changed between years. Second, when using threshold or concentration8 weighted cumulative metrics (such as AOT40, SUM06 or W126), the proportionality 

9 does not propagate regularly from the hourly data to the yearly value. For example, 

10 hourly average elevated exposures that are a constant 1.5 times greater than ambient do 

11 not result in AOT40, SUM06 or W126 values that are some constant multiple of the 

12 ambient values of those indices. Depending on the fraction of hourly values that are 

13 above the threshold or heavily weighted, the same average yearly exposure will result in 

14 different exposure values when using thresholded or weighted metrics. In some years, 

15 elevated exposures in FACE experiments experience many more values above the 

16 threshold, or more heavily weighted than the ambient exposures; thus in those years, the 

17 distance between ambient and elevated exposure values increases relative to other years. 

18 As a consequence, the number of exposure levels in multi-year experiments is twice the 

19 number of years. In the case of SoyFACE for the period between 2002 and 2008, ambient 

20 exposure in the highest year was approximately equal to elevated exposure in the lowest 

21 year, with 14 levels of O3 exposure evenly distributed from lowest to highest. The 

22 particular conditions of the Aspen FACE experiment resulted in 12 exposure levels 

23 between 1998 and 2003, but they were not as evenly distributed between minimum and 

24 maximum over the 6-year period.

25 There are necessary differences in the modeling of exposure-response in annual plants 

26 such as soybean, and in perennial plants such as aspen trees, when exposure takes place 

27 over multiple years. In annual plants, responses recorded at the end of the life cycle, 

28 i.e., yearly, are analyzed in relationship to that year’s exposure. Yield of soybeans is 

29 affected by exposure during the year the crop was growing, and a new crop is planted 

30 every year. Thus an exposure-response relationship can be modeled from yearly 

31 responses matched to yearly exposures, with those exposure-response data points having 

32 been generated in separate years. For perennial organisms, which are not harvested yearly 

33 and continue to grow from year to year, such pairing of exposure and response cannot be 

34 done without accounting for time. Not only does the size of the organism at the beginning 

35 of each year of exposure increase, but size is also dependent on the exposure from 

36 previous years. Therefore the relationship of response and exposure must be analyzed 

37 either one year at a time, or by standardizing the response as a yearly increment relative 

38 to size at the beginning of each year. Furthermore, the relevant measurement of exposure 

39 is cumulative, or cumulative yearly average exposure, starting in the year exposure was 
Draft – Do Not Cite or Quote 9-137 June 2012

1 initiated, up to the end of the year of interest. When analyzing the growth of trees over 

2 several years, it would be evidently incorrect to pair the exposure level in every discrete 

3 year with absolute size of the trees that year, and posit a direct relationship between them, 

4 without taking increasing age into consideration. In the Aspen FACE experiment, for 

5 example, one could not establish an exposure-response relationship by matching 

6 12 yearly exposures and 12 yearly tree sizes, while disregarding age as if size did not also 

7 depend on it. This is the basis of the 2007 study of Aspen FACE data by Percy et al. 

8 (2007), which compares the size of trees of various ages as if they were all the same age, 

9 and was therefore not informative.

9.6.3.1 Comparison of NCLAN-Based Prediction and SoyFACE 

Data 

10 For this ISA, EPA conducted a comparison between yield of soybean as predicted by the 

11 composite function three-step process (Section 9.6.2) using NCLAN data, and 

12 observations of yield in SoyFACE. The median composite function for relative yield was 

13 derived for the 11 NCLAN soybean Weibull functions for non-droughted studies, and 

14 comparisons between the predictions of the median composite and SoyFACE 

15 observations were conducted as follows. 

16 For the years 2007 and 2008, SoyFACE yield data were available for 7 and 6 genotypes, 

17 respectively. The EPA used those data to compare the relative change in yield observed 

18 in SoyFACE in a given year between ambient O3 and elevated O3, versus the relative 

19 change in yield predicted by the NCLAN-based median composite function between 

20 those same two values of O3 exposure. The two parameter median composite function for 

21 relative yield of soybean based on NCLAN data was used to predict yield response at the 

22 two values of exposure observed in SoyFACE in each year, and the change between yield 

23 under ambient and elevated was compared to the change observed in SoyFACE for the 

24 relevant year (Table 9-12). This approach results in a direct comparison of predicted 

25 versus observed change in yield. Because the value of relative response between any two 

26 values of O3 exposure is independent of the intercept α, this comparison does not require 

27 prediction of the absolute values of the responses.

28 Since comparisons of absolute values might be of interest, the predictive functions were 

29 also scaled to the observed data: SoyFACE data were used to compute an intercept α 

30 while the shape and scale parameters (β and η) were held at their value in the NCLAN 

31 predictive model. This method gives a comparison of prediction and observation that 

32 takes all the observed information into account to provide the best possible estimate of 

33 the intercept, and thus the best possible scaling (
Draft – Do Not Cite or Quote 9-138 June 2012

1 Table 9-13 and Figure 9-18). For the comparison of NCLAN and SoyFACE, this 

2 validation was possible for 2007 and 2008, where data for 7 and 6 soybean genotypes, 

3 respectively, were available. The median composite function for relative yield was 

4 derived for the 11 NCLAN soybean Weibull functions for nondroughted studies, and the 

5 values of median yield under ambient exposure at SoyFACE in 2007 and 2008 were used 

6 to obtain an estimate of the intercept α for the NCLAN median function in each of the 

7 two years. Table 9-12 presents the results of ambient/elevated relative yield comparisons 

8 between the NCLAN-derived predictions and SoyFACE observations. 

9 Table 9-13 and Figure 9-18 present the results of comparisons between NCLAN-derived 

10 predictions and SoyFACE observations of yield, with the predictive function scaled to 

11 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.

Year 90-day 12-h W126 (ppm-h) observed at 

SoyFACE

Yield in Elevated O3 Relative to Ambient O3 (%)

Ambient Elevated Predicted by NCLANa Observed at SoyFACE

2007 4.39 46.23 75 76

2008 3.23 28.79 85 88

a

Two-parameter relative yield model.

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.

Year 90-day 12-h W126 (ppm-h) 

observed at SoyFACE

Yield predicted by NCLANa

(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

a

Three-parameter absolute yield model with intercept scaled to SoyFACE data.
Draft – Do Not Cite or Quote 9-139 June 2012

Note: Black dots are the median of 7 or 6 soybean genotypes in SoyFACE (2007, 2008); bars are Inter-Quartile Range for 

genotypes; dashed line is median composite model for 11 studies in NCLAN.

Source of data: Betzelberger et al. (2010); Morgan et al. (2006); Lee and Hogsett (1996).

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 9-140 June 2012

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 

2 Table 9-13 and Table 9-14, and in Table 9-18, the agreement between predictions based 

3 on NCLAN data and SoyFACE observations was notably close in single-year 

4 comparisons. Together with the very high agreement between median composite models 

5 for NCLAN and SoyFACE (Figure 9-19), it provides very strong mutual confirmation of 

6 those two projects’ results with respect to the response of yield of soybeans to O3

7 exposure. It is readily apparent from these results that the methodology described in 

8 Section 9.6.2 for obtaining predictions of yield or yield loss from NCLAN data is 

9 strongly validated by SoyFACE results. As described in Section 9.2, the exposure 

10 technologies used in the two projects were in sharp contrast, specifically with respect to 

11 the balance each achieved between control of potential interacting factors or confounders, 

12 and fidelity to natural conditions. The comparisons that EPA conducted therefore 

13 demonstrate that the methodology used in developing the composite functions is resistant 

14 to the influence of nuisance variables and that predictions are reliable. They may also 

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 9-141 June 2012

1 suggest that the aspects in which the two exposure technologies differ have less influence 

2 on exposure-response than initially supposed. These results are also in agreement with 

3 comparative studies reviewed in Section 9.2.6. 

9.6.3.2 Comparison of NHEERL/WED-Based Prediction of Tree 

Biomass Response and Aspen FACE Data

4 EPA also conducted two comparisons between prediction of above-ground biomass loss 

5 based on NHEERL/WED results and observations from Aspen FACE. The median 

6 composite function was developed from NHEERL/WED data for 11 studies that used 

7 wild-type seedlings of aspen as well as four clonally propagated genotypes. All plants 

8 were grown in OTCs for one growing season before being destructively harvested. Aspen 

9 FACE data were from clonally propagated trees of five genotypes grown from 1998 to 

10 2003, with above-ground biomass calculated using allometric equations derived from 

11 data for trees harvested destructively in 2000 and 2002 (King et al., 2005). 

12 The two parameter median composite function for relative biomass was used to predict 

13 biomass response under the observed elevated exposure, relative to its value under 

14 observed ambient exposure, for each separate year of Aspen FACE. EPA first compared 

15 Aspen FACE observations of the change in biomass between ambient and elevated 

16 exposure with the corresponding prediction at the same values of exposure. Comparisons 

17 between observed and predicted absolute biomass values were then conducted for each 

18 year by scaling the predictive function to yearly Aspen FACE data as described for 

19 soybean data in Section 9.6.3.1. In all cases, yearly 90 day 12-hour W126 values for 

20 Aspen FACE were computed as the cumulative average from the year of planting up to 

21 the year of interest. A comparison of composite functions between NHEERL/WED and 

22 Aspen FACE, similar to the one performed for NCLAN and SoyFACE, was not possible: 

23 as discussed in the introduction to Section 9.6, the pairing of 12 exposure values from 

24 separate years and 12 values of biomass cannot be the basis for a model of exposure25 response, because the trees continued growing for the six-year period of exposure. 

26 Because the same trees were used for the entire duration, and continued to grow, data 

27 could not be aggregated over years. Table 9-14 presents the results of ambient/elevated 

28 relative biomass comparisons between the NHEERL/WED-derived predictions and 

29 Aspen FACE observations. 

30 Table 9-15 and Figure 9-20 present the results of the comparison between 

31 NHEERL/WED-derived predictions and Aspen FACE observations for absolute biomass, 

32 using Aspen FACE data to scale the NHEERL/WED-derived composite function.
Draft – Do Not Cite or Quote 9-142 June 2012

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.

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.

Year 90 day 12-h W126 (ppm-h)

Cumulative Average 

observed at Aspen FACE

Biomass Predicted by 

NHEERL/WEDa (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 1468.4

2003 2.86 29.12 2640.1 1981.2 2695.2 1907.8

a

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

Above-Ground Biomass in

Elevated O3 relative To Ambient O3 (%)

Ambient Elevated Predicted by 

NHEERL/WEDa

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

a

Two-parameter relative biomass model
Draft – Do Not Cite or Quote 9-143 June 2012

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.

Source of data: King et al. (2005); Lee and Hogsett (1996).

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 9-144 June 2012

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 9-145 June 2012

Note: Line represents the three-parameter Weibull model.

Source: Modified with permission of 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.

1 As shown in Section 9.6.2, the median models available for trembling aspen and soybean 

2 have verifiable predictive ability for those particular species. This suggests that the 

3 corresponding NCLAN- and NHEERL/WED-based models for multiple crop and tree 

4 species can provide reliable estimates of losses for similar assortments of species. 

5 However, their predictive ability would likely be poor for individual species not tested. 

6 The cottonwood data of Gregg et al. (2003) show an extremely severe response to O3. 

7 They are consistent with the expectation that among species and genotypes, some are 

8 likely to be substantially more sensitive than a median measure, such as the estimate 

9 produced by NHEERL/WED (Figure 9-16), but the sensitivity of this particular species 

10 has not been studied elsewhere. 

11 An alternative hypothesis for the difference between the response of cottonwood in this 

12 experiment and deciduous tree species in NHEERL/WED, or the difference between the 

13 response of cottonwood and aspen in NHEERL/WED and Aspen FACE, could be the 

14 presence of confounding factors in the environments where the experiment was 

15 conducted. However, variability in temperature, moisture, soil fertility, and atmospheric 

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)
Draft – Do Not Cite or Quote 9-146 June 2012

1 deposition of N were all ruled out by Gregg et al. (2003) as contributing to the observed 

2 response to O3. In addition, this hypothesis would imply that the unrecognized 

3 confounder(s) were either absent from both OTC and FACE studies, or had the same 

4 value in both. This is not impossible, but the hypothesis that cottonwood is very sensitive 

5 to O3 exposure is more parsimonious, and sufficient.

9.6.3.4 Meta-analyses of growth and yield studies

6 Since the 2006 O3 AQCD, five studies have used meta-analytic methods to integrate 

7 results from experimental studies of crops or tree species relevant to the U.S. It is 

8 possible to obtain exposure-response data for growth and yield from those meta-analyses, 

9 but because all of them provided summary measurements of O3 exposure as hourly 

10 averages of various lengths of exposures, comparisons with exposure-response results 

11 where exposure is expressed as W126 are problematic. Table 9-16 summarizes the 

12 characteristics of the five meta-analyses. They all included studies conducted in the U.S. 

13 and other locations worldwide, and all of them expressed responses as comparative 

14 change between levels of exposure to O3, with carbon filtered air (CF) among those 

15 levels. Using hourly average concentration to summarize exposure, CF rarely equates 

16 with absence of O3, although it almost always near zero when exposure is summarized as 

17 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
Draft – Do Not Cite or Quote 9-147 June 2012

1 The only effect of O3 exposure on yield of rice reported in Ainsworth (2008) was a 

2 decrease of 14% with exposure increasing from CF to 62 ppb average concentration. 

3 Feng et al. (2008b) were able to separate exposure of wheat into four classes with average 

4 concentrations of 42, 69, 97, and 153 ppb, in data where O3 was the only treatment. Mean 

5 responses relative to CF were yield decreases of 17, 25, 49, and 61% respectively. Feng 

6 et al. (2008b) observed that wheat yield losses were smaller under conditions of drought, 

7 and that Spring wheat and Winter wheat appeared similarly affected. However, mean 

8 exposure in studies of Winter wheat was substantially higher than in studies of Spring 

9 wheat (86 versus 64 ppb), which suggests that the yield of Spring wheat was in fact more 

10 severely affected, since yield was approximately the same, even though Spring wheat was 

11 exposed to lower concentrations. Exposures of the six crops considered in Feng and 

12 Kobayashi (2009) were classified into two ranges, each compared to CF air. In the lower 

13 range of exposure (41-49 ppb), potato studies had the highest average exposure (45 ppb) 

14 and wheat and rice the lowest (41 ppb). In the higher range (51-75 ppb), wheat studies 

15 had the highest average exposure (65 ppb), and potato, barley and rice the lowest 

16 (63 ppb). In other words, across the studies included, all crops were exposed to very 

17 similar levels of O3. At approximately 42 ppb, the yield of potato, barley, wheat, rice, 

18 bean, and soybean declined by 5.3, 8.9, 9.7, 17.5, 19, and 7.7% respectively, relative to 

19 CF air. At approximately 64 ppb O3, declines were 11.9, 12.5, 21.1, 37.5, 41.4, and 

20 21.6%. Grantz et al. (2006) reported Relative Growth Rate (RGR) rather than growth, 

21 and did not report O3 exposure values in a way that would allow calculation of mean 

22 exposure for each of the three categories of plants for which RGR changes are reported. 

23 All studies used only two levels of exposure, with CF air as the lower one, and most used 

24 elevated exposure in the range of 40 to 70 ppb. Decline in RGR was 8.2% for the 34 

25 herbaceous dicots, 4.5% for the 21 herbaceous monocots, and 17.9% for the 5 tree 

26 species. Finally, Wittig et al. (2009) divided the studies analyzed into three classes of 

27 comparisons: CF versus ambient, CF versus elevated, and ambient versus elevated, but 

28 reported comparisons between three average levels of exposure besides CF: 40 ppb, 

29 64 ppb, and 97 ppb. Corresponding decreases in total biomass relative to CF were 7, 17, 

30 and 17%. 

31 These meta-analyses provide very strong confirmation of EPA’s conclusions from 

32 previous O3 AQCDs: compared to lower levels of ambient O3, current levels in many 

33 locations are having a substantial detrimental effect on the growth and yield of a wide 

34 variety of crops and natural vegetation. They also confirm strongly that decreases in 

35 growth and yield continue at exposure levels higher than current ambient levels. 

36 However, direct comparisons with the predictions of exposure-response models that use 

37 concentration-weighted cumulative metrics are difficult. 
Draft – Do Not Cite or Quote 9-148 June 2012

9.6.3.5 Additional exposure-response data 

1 The studies summarized in Table 9-17 and Table 9-18 contain growth or yield exposure2 response data at too few levels of exposure for exposure-response models, and/or used 

3 metrics other than W126. These tables update Tables AX9-16 through AX9-19 of the 

4 2006 O3 AQCD.

9.6.4 Summary

5 None of the information on effects of O3 on vegetation published since the 2006 O3

6 AQCD has modified the assessment of quantitative exposure-response relationships that 

7 was presented in that document. This assessment updates the 2006 exposure-response 

8 models by computing them using the W126 metric, cumulated over 90 days. Almost all 

9 of the experimental research on the effects of O3 on growth or yield of plants published 

10 since 2006 used only two levels of exposure. In addition, hourly O3 concentration data 

11 that would allow calculations of exposure using the W126 metric are generally 

12 unavailable. However, two long-term experiments, one with a crop species (soybean), 

13 one with a tree species (aspen), have produced data that can be used to validate the 

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

15 derive them. 

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

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

18 Network (NCLAN) and EPA National Health and Environmental Effects Research 

19 Laboratory, Western Ecology Division (NHEERL-WED) projects, using OTCs to expose 

20 crops and trees seedling to O3. In recent years, yield and growth results for two of the 

21 species that had provided extensive exposure-response information in those projects have 

22 become available from studies that used FACE technology, which is intended to provide 

23 conditions much closer to natural environments (Pregitzer et al., 2008; Morgan et al., 

24 2006; Morgan et al., 2004; Dickson et al., 2000). The robust methods that were used 

25 previously with exposure measured as SUM06 were applied to the NCLAN and 

26 NHEERL-WED data with exposure measured as W126, in order to derive single-species 

27 median models for soybean and aspen from studies involving different genotypes, years, 

28 and locations. The resulting models were used to predict the change in yield of soybean 

29 and biomass of aspen between the two levels of exposure reported in recent FACE 

30 experiments. Results from these new experiments were exceptionally close to predictions 

31 from the models. The accuracy of model predictions for two widely different plant 

32 species provides support for the validity of the corresponding multiple-species models for 

33 crops and trees in the NCLAN and NHEERL-WED projects. However, variability among 
Draft – Do Not Cite or Quote 9-149 June 2012

1 species in those projects indicates that the range of sensitivity is likely quite wide. This 

2 was confirmed by a recent experiment with cottonwood in a naturally occurring gradient 

3 of exposure (Gregg et al., 2006), which established the occurrence of species with 

4 responses substantially more severe under currently existing conditions than are predicted 

5 by the median model for multiple species.

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

7 yield of soybean, wheat, rice and other crops under broad categories of exposure, relative 

8 to charcoal-filtered air (Ainsworth, 2008; Feng et al., 2008b; Morgan et al., 2003). 

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

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

11 263 studies that reported effects on tree biomass. However, these analyses have proved 

12 difficult to compare with exposure-response models, especially given that exposure was 

13 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 aboveground biomassN/A (Highest 

treatment caused -

26% change)

Leitao et al. 

(2007a)
Draft – Do Not Cite or Quote 9-150 June 2012

Species

Facility

Location

Exposure 

Duration

O3 Exposure

(Additional Treatment)

Response 

Measured

Percent Change 

from CF

(Percent Change 

from Ambient)

Reference

Cotton cv. Pima

OTC; 9-L pots

San Joaquin 

Valley, CA

8 weeks 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 weeks 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)

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 and Fiscus 

(2008)
Draft – Do Not Cite or Quote 9-151 June 2012

Species

Facility

Location

Exposure 

Duration

O3 Exposure

(Additional Treatment)

Response 

Measured

Percent Change 

from CF

(Percent Change 

from Ambient)

Reference

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 Jaoudé et 

al. (2008b)

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. 

(2004b)

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. 

(2008b)

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)
Draft – Do Not Cite or Quote 9-152 June 2012

Species

Facility

Location

Exposure 

Duration

O3 Exposure

(Additional Treatment)

Response 

Measured

Percent Change 

from CF

(Percent Change 

from Ambient)

Reference

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 weeks.

12-h avg: CF (4 ppb); 

Ambient (58); 

Elevated (147)

(N/A)

Total biomass 

(g/plant)

-40 (-30) Grantz and Vu 

(2009)

Sweet Potato

Growth chambers

Bonn, Germany

4 weeks 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.) Dalstein and 

Vas (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

San Joaquin 

Valley, CA

8 weeks 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.
Draft – Do Not Cite or Quote 9-153 June 2012

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

San Joaquin Valley, 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. (2010a)

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-h: 

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)

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)
Draft – Do Not Cite or Quote 9-154 June 2012

Species

Facility

Location

Exposure 

Duration

O3 Exposure

(Additional 

Treatment)

Response 

Measured

Response Reference

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)
Draft – Do Not Cite or Quote 9-155 June 2012

Species

Facility

Location

Exposure 

Duration

O3 Exposure

(Additional 

Treatment)

Response 

Measured

Response Reference

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)

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.

9.7 Summary and Conclusions

1 Based on the evidence presented in Chapter 9 and summarized here, O3 is causally related 

2 or likely to be causally related to effects observed on vegetation and ecosystems. The 

3 evidence for these effects spans the entire continuum of biological organization, from the 

4 cellular and subcellular level to the whole plant, and up to ecosystem-level processes, and 

5 includes evidence for effects at lower levels of organization, leading to effects at higher 

6 levels. Given the current state of knowledge, exposure indices that cumulate and 

7 differentially weight the higher hourly average concentrations and also include the mid8 level values are the most appropriate for use in developing response functions and 

9 comparing studies. The framework for causal determinations (see Preamble) has been 

10 applied to the body of scientific evidence to examine effects attributed to O3 exposure 

11 collectively and the determinations are presented in Table 9-19.
Draft – Do Not Cite or Quote 9-156 June 2012

Table 9-19 Summary of ozone causal determinations for vegetation and 

ecosystem effects

Vegetation and 

Ecosystem Effects

Conclusions from 2006 O3 AQCD Conclusions from 

2011 3rd Draft ISA

Visible Foliar Injury Effects 

on Vegetation

Data published since the 1996 O3 AQCD strengthen previous 

conclusions that there is strong evidence that current ambient O3

concentrations cause impaired aesthetic quality of many native plants 

and trees by increasing foliar injury.

Causal Relationship

Reduced Vegetation Growth Data published since the 1996 O3 AQCD strengthen previous 

conclusions that there is strong evidence that current ambient O3

concentrations cause decreased growth and biomass accumulation in 

annual, perennial and woody plants, including agronomic crops, 

annuals, shrubs, grasses, and trees.

Causal Relationship

Reduced Productivity in 

Terrestrial Ecosystems

There is evidence that O3 is an important stressor of ecosystems and 

that the effects of O3 on individual plants and processes are scaled up 

through the ecosystem, affecting net primary productivity.

Causal Relationship

Reduced Carbon (C) 

Sequestration in Terrestrial 

Ecosystems

Limited studies from previous review Likely to be a Causal 

Relationship

Reduced Yield and Quality 

of Agricultural Crops

Data published since the 1996 O3 AQCD strengthen previous 

conclusions that there is strong evidence that current ambient O3

concentrations cause decreased yield and/or nutritive quality in a large 

number of agronomic and forage crops.

Causal Relationship

Alteration of Terrestrial 

Ecosystem Water Cycling

Ecosystem water quantity may be affected by O3 exposure at the 

landscape level.

Likely to be a Causal 

Relationship

Alteration of Below-ground 

Biogeochemical Cycles

Ozone-sensitive species have well known responses to O3 exposure, 

including altered C allocation to below-ground tissues, and altered rates 

of leaf and root production, turnover, and decomposition. These shifts 

can affect overall C and N loss from the ecosystem in terms of respired 

C, and leached aqueous dissolved organic and inorganic C and N.

Causal Relationship

Alteration of Terrestrial 

Community Composition

Ozone may be affecting above- and below -ground community 

composition through impacts on both growth and reproduction. 

Significant changes in plant community composition resulting directly 

from O3 exposure have been demonstrated.

Likely to be a Causal 

Relationship
Draft – Do Not Cite or Quote 9-157 June 2012

References

Agrell, J; Kopper, BJ; McDonald, EP; Lindroth, RL. (2005). CO2 and O3 effects on host plant preferences of 

the forest tent caterpillar (Malacosoma disstria). Global Change Biol 11: 588-599. 

http://dx.doi.org/10.1111/j.1365-2486.2005.00924.x

Ahlfors, R; Brosche, M; Kollist, H; Kangasjarvi, J. (2009). Nitric oxide modulates ozone-induced cell death, 

hormone biosynthesis and gene expression in Arabidopsis thaliana. Plant J 58: 1-12. 

http://dx.doi.org/10.1111/j.1365-313X.2008.03756.x

Ahsan, N; Nanjo, Y; Sawada, H; Kohno, Y; Komatsu, S. (2010). Ozone stress-induced proteomic changes in 

leaf total soluble and chloroplast proteins of soybean reveal that carbon allocation is involved in 

adaptation in the early developmental stage. Proteomics 10: 2605-2619. 

http://dx.doi.org/10.1002/pmic.201000180

Ainsworth, EA. (2008). Rice production in a changing climate: a meta-analysis of responses to elevated 

carbon dioxide and elevated ozone concentration. Global Change Biol 14: 1642-1650. 

http://dx.doi.org/10.1111/j.1365-2486.2008.01594.x

Ainsworth, EA; Long, SP. (2005). What have we learned from 15 years of free-air CO2 enrichment (FACE)? 

A meta-analytic review of the responses of photosynthesis, canopy properties and plant production to 

rising CO2 [Review]. New Phytol 165: 351-371. 

Ainsworth, EA; Rogers, A. (2007). The response of photosynthesis and stomatal conductance to rising 

[CO2]: Mechanisms and environmental interactions [Review]. Plant Cell Environ 30: 258-270. 

http://dx.doi.org/10.1111/j.1365-3040.2007.01641.x

Alexis, A; Garcia, A; Nystrom, M; Rosenkranz, K. (2001a). The 2001 California almanac of emissions and 

air quality. Sacremento, CA: California Air Resources Board. 

http://www.arb.ca.gov/aqd/almanac/almanac01/almanac01.htm

Allen, EB; Temple, PJ; Bytnerowicz, A; Arbaugh, MJ; Sirulnik, AG; Rao, LE. (2007). Patterns of understory 

diversity in mixed coniferous forests of southern California impacted by air pollution. 

ScientificWorldJournal 7: 247-263. http://dx.doi.org/10.1100/tsw.2007.72

Alonso, R; Bermejo, V; Sanz, J; Valls, B; Elvira, S; Gimeno, BS. (2007). Stomatal conductance of seminatural Mediterranean grasslands: Implications for the development of ozone critical levels. Environ 

Pollut 146: 692-698. http://dx.doi.org/10.1016/j.envpol.2006.06.009

Amthor, JS. (1988). Growth and maintenance respiration in leaves of bean (Phaseolus vulgaris L) exposed to 

ozone in open-top chambers in the field. New Phytol 110: 319-325. http://dx.doi.org/10.1111/j.1469-

8137.1988.tb00268.x

Andersen, CP. (2003). Source-sink balance and carbon allocation below ground in plants exposed to ozone. 

New Phytol 157: 213-228. 

Andersen, CP; Ritter, W; Gregg, J; Matyssek, R; Grams, TEE. (2010). Below-ground carbon allocation in 

mature beech and spruce trees following long-term, experimentally enhanced O3 exposure in southern 

Germany. Environ Pollut 158: 2604-2609. http://dx.doi.org/10.1016/j.envpol.2010.05.008

Andersen, CP; Wilson, R; Plocher, M; Hogsett, WE. (1997). Carry-over effects of ozone on root growth and 

carbohydrate concentrations of ponderosa pine seedlings. Tree Physiol 17: 805-811. 

Andrews, KM; Gibbons, JW; Jochimsen, DM. (2008). Ecological effects of roads on amphibians and 

reptiles: A literature review. In JC Mitchell; REJ Brown; B Bartholomew (Eds.), Urban Herpetology 

(pp. 121-143). Salt Lake City: Society for the Study of Amphibians and Reptiles. 

http://www.uga.edu/srel/Reprint/3091.htm

Aneja, MK; Sharma, S; Fleischmann, F; Stich, S; Heller, W; Bahnweg, G; Munch, JC; Schloter, M. (2007). 

Influence of ozone on litter quality and its subsequent effects on the initial structure of colonizing 

microbial communities. Microb Ecol 54: 151-160. http://dx.doi.org/10.1007/s00248-006-9183-0

Arbaugh, M; Bytnerowicz, A; Grulke, N; Fenn, M; Poth, M; Temple, P; Miller, P. (2003). Photochemical 

smog effects in mixed conifer forests along a natural gradient of ozone and nitrogen deposition in the 

San Bernardino Mountains. Environ Int 29: 401-406. http://dx.doi.org/10.1016/S0160-4120(02)00176-

9
Draft – Do Not Cite or Quote 9-158 June 2012

Arbaugh, MJ; Miller, PR; Carroll, JJ; Takemoto, BL; Proctor, T. (1998). Relationships of ozone exposure to 

pine injury in the Sierra Nevada and San Bernardino Mountains of California, USA. Environ Pollut 

101: 291-301. http://dx.doi.org/10.1016/S0269-7491(98)00027-X

Ariyaphanphitak, W; Chidthaisong, A; Sarobol, E; Bashkin, VN; Towprayoon, S. (2005). Effects of elevated 

ozone concentrations on Thai Jasmine rice cultivars (Oryza sativa L.). Water Air Soil Pollut 167: 179-

200. http://dx.doi.org/10.1007/s11270-005-8650-4

Ashmore, M; Emberson, L; Karlsson, PE; Pleijel, H. (2004a). Introduction for ozone deposition special issue. 

Atmos Environ 38: 2211-2212. 

Ashmore, M; Emberson, L; Karlsson, PE; Pleijel, H. (2004b). New directions: A new generation of ozone 

critical levels for the protection of vegetation in Europe (correspondence). Atmos Environ 38: 2213-

2214. 

Ashmore, MR. (2002). Effects of oxidants at the whole plant and community level. In JNB Bell; M Treshow 

(Eds.), Air pollution and plant life (pp. 89-118). London: Wiley. 

Ashmore, MR; Bell, JNB; Mimmack, A. (1988). Crop growth along a gradient of ambient air pollution. 

Environ Pollut 53: 99-121. http://dx.doi.org/10.1016/0269-7491(88)90028-0

Avnery, S; Mauzerall, DL; Liu, J; Horowitz, LW. (2011a). Global crop yield reductions due to surface ozone 

exposure: 1. Year 2000 crop production losses and economic damage. Atmos Environ 45: 2284-2296. 

http://dx.doi.org/10.1016/j.atmosenv.2010.11.045

Avnery, S; Mauzerall, DL; Liu, J; Horowitz, LW. (2011b). Global crop yield reductions due to surface ozone 

exposure: 2. Year 2030 potential crop production losses and economic damage under two scenarios of 

O3 pollution. Atmos Environ 45: 2297-2309. http://dx.doi.org/10.1016/j.atmosenv.2011.01.002

Awmack, CS; Harrington, R; Lindroth, RL. (2004). Aphid individual performance may not predict 

population responses to elevated CO2 or O3. Global Change Biol Biol.10: 1414-1423. 

Awmack, CS; Mondor, EB; Lindroth, RL. (2007). Forest understory clover populations in enriched CO2 and 

O-3 atmospheres: Interspecific, intraspecific, and indirect effects. Environ Exp Bot 59: 340-346. 

http://dx.doi.org/10.1016/j.envexpbot.2006.04.003

Bagard, M; Le Thiec, D; Delacote, E; Hasenfratz-Sauder, MP; Banvoy, J; Gerard, J; Dizengremel, P; Jolivet, 

Y. (2008). Ozone-induced changes in photosynthesis and photorespiration of hybrid poplar in relation 

to the developmental stage of the leaves. Physiol Plant 134: 559-574. http://dx.doi.org/10.1111/j.1399-

3054.2008.01160.x

Baier, M; Kandlbinder, A; Golldack, D; Dietz, K. (2005). Oxidative stress and ozone: Perception; signalling 

and response. Plant Cell Environ 28: 1012-1020. http://dx.doi.org/10.1111/j.1365-3040.2005.01326.x

Baldantoni, D; Fagnano, M; Alfani, A. (2011). Tropospheric ozone effects on chemical composition and 

decomposition rate of Quercus ilex L. leaves. Sci Total Environ 409: 979-984. 

http://dx.doi.org/10.1016/j.scitotenv.2010.11.022

Ball, GR; Palmer-Brown, D; Fuhrer, J; Skarby, L; Gimeno, BS; Mills, G. (2000). Identification of non-linear 

influences on the seasonal ozone dose-response of sensitive and resistant clover clones using artificial 

neural networks. Ecol Modell 129: 153-168. 

Balls, GR; Palmer-Brown, D; Sanders, GE. (1996). Investigating microclimatic influences on ozone injury in 

clover (Trifolium subterraneum) using artificial neural networks. New Phytol 132: 271-280. 

http://dx.doi.org/10.1111/j.1469-8137.1996.tb01846.x

Bandeff, JM; Pregitzer, KS; Loya, WM; Holmes, WE; Zak, DR. (2006). Overstory community composition 

and elevated atmospheric CO2 and O3 modify understory biomass production and nitrogen 

acquisition. Plant Soil 282: 251-259. http://dx.doi.org/10.1007/s11104-005-5930-0

Barnes, J; Zheng, Y; Lyons, T. (2002). Plant resistance to ozone: The role of ascorbate. In K Omasa; H Saji; 

S Youssefian; N Kondo (Eds.), Air pollution and plant biotechnology - Prospects for phytomonitoring 

and phytoremediation (pp. 235252). Tokyo: Springer-Verlag. 

Bassin, S; Volk, M; Fuhrer, J. (2007a). Factors affecting the ozone sensitivity of temperate European 

grasslands: An overview. Environ Pollut 146: 678-691. http://dx.doi.org/10.1016/j.envpol.2006.06.010

Bassin, S; Volk, M; Suter, M; Buchmann, N; Fuhrer, J. (2007b). Nitrogen deposition but not ozone affects 

productivity and community composition of subalpine grassland after 3 yr of treatment. New Phytol 

175: 523-534. http://dx.doi.org/10.1111/j.1469-8137.2007.02140.x

Bassin, S; Werner, RA; Sorgel, K; Volk, M; Buchmann, N; Fuhrer, J. (2009). Effects of combined ozone and 

nitrogen deposition on the in situ properties of eleven key plant species of a subalpine pasture. 

Oecologia 158: 747-756. http://dx.doi.org/10.1007/s00442-008-1191-y
Draft – Do Not Cite or Quote 9-159 June 2012

Bauer, MR; Hultman, NE; Panek, JA; Goldstein, AH. (2000). Ozone deposition to a ponderosa pine

plantation in the Sierra Nevada Mountains (CA): A comparison of two different climatic years. J 

Geophys Res 105: 22,123-122,136. http://dx.doi.org/10.1029/2000JD900168

Bender, J; Muntifering, RB; Lin, JC; Weigel, HJ. (2006). Growth and nutritive quality of Poa pratensis as 

influenced by ozone and competition. Environ Pollut 142: 109-115. 

http://dx.doi.org/10.1016/j.envpol.2005.09.012

Bender, J; Weigel, HJ. (2011). Changes in atmospheric chemistry and crop health: A review [Review]. Agron 

Sustain Dev 31: 81-89. http://dx.doi.org/10.1051/agro/2010013

Bennett, JP; Jepsen, EA; Roth, JA. (2006). Field responses of Prunus serotina and Asclepias syriaca to ozone 

around southern Lake Michigan. Environ Pollut 142: 354-366. 

http://dx.doi.org/10.1016/j.envpol.2005.09.024

Benoit, LF; Skelly, JM; Moore, LD; Dochinger, LS. (1982). Radial growth reductions of Pinus strobus L 

correlated with foliar ozone sensitivity as an indicator of ozone-induced losses in eastern forests. Can 

J For Res 12: 673-678. http://dx.doi.org/10.1139/x82-101

Bergweiler, CJ; Manning, WJ. (1999). Inhibition of flowering and reproductive success in spreading dogbane 

(Apocynum androsaemifolium) by exposure to ambient ozone. Environ Pollut 105: 333-339. 

http://dx.doi.org/10.1016/S0269-7491(99)00044-5

Bergweiler, CJ; Manning, WJ; Chevone, BI. (2008). Seasonal and diurnal gas exchange differences in ozonesensitive common milkweed (Asclepias syriaca L.) in relation to ozone uptake. Environ Pollut 152: 

403-415. http://dx.doi.org/10.1016/j.envpol.2007.06.019

Bernacchi, CJ; Leaky, ADB; Heady, LE; Morgan, PB; Dohleman, FG; McGrath, JM; Gillespie, KM; Wittig, 

VE; Rogers, A; Long, SP; Ort, DR. (2006). Hourly and seasonal variation in photosynthesis and 

stomatal conductance of soybean grown at future CO2 and ozone concentrations for 3 years under 

fully open-air field conditions. Plant Cell Environ 29: 2077-2090. http://dx.doi.org/10.1111/j.1365-

3040.2006.01581.x

Bernacchi, CJ; Morgan, PB; Ort, DR; Long, SP. (2005). The growth of soybean under free air CO2 

enrichment (FACE) stimulates photosynthesis while decreasing in vivo Rubisco capacity. Planta 220: 

434-446. http://dx.doi.org/10.1007/s00425-004-1320-8

Betzelberger, AM; Gillespie, KM; McGrath, JM; Koester, RP; Nelson, RL; Ainsworth, EA. (2010). Effects of 

chronic elevated ozone concentration on antioxidant capacity, photosynthesis and seed yield of 10 

soybean cultivars. Plant Cell Environ 33: 1569-1581. http://dx.doi.org/10.1111/j.1365-

3040.2010.02165.x

Bidart-Bouzat, MG; Imeh-Nathaniel, A. (2008). Global change effects on plant chemical defenses against 

insect herbivores. J Integr Plant Biol 50: 1339-1354. http://dx.doi.org/10.1111/j.1744-

7909.2008.00751.x

Billings, WD. (1978). Plants and the ecosystem. Belmont, CA: Wadsworth Publishing Company, Inc. 

Biswas, DK; Xu, H; Li, YG; Sun, JZ; Wang, XZ; Han, XG; Jiang, GM. (2008). Genotypic differences in leaf 

biochemical, physiological and growth responses to ozone in 20 winter wheat cultivars released over 

the past 60 years. Global Change Biol 14: 46-59. http://dx.doi.org/10.1111/j.1365-2486.2007.01477.x

Black, VJ; Black, CR; Roberts, JA; Stewart, CA. (2000). Impact of ozone on the reproductive development 

of plants. New Phytol 147: 421-447. 

Black, VJ; Stewart, CA; Roberts, JA; Black, CR. (2007). Ozone affects gas exchange, growth and 

reproductive development in Brassica campestris (Wisconsin Fast Plants). New Phytol 176: 150-163. 

http://dx.doi.org/10.1111/j.1469-8137.2007.02163.x

Black, VJ; Stewart, CA; Roberts, JA; Black, CR. (2010). Direct effects of ozone on reproductive 

development in Plantago major L. populations differing in sensitivity. Environ Exp Bot 69: 121-128. 

http://dx.doi.org/10.1016/j.envexpbot.2010.04.006

Blande, JD; Holopainen, JK; Li, T. (2010). Air pollution impedes plant-to-plant communication by volatiles. 

Ecol Lett 13: 1172-1181. http://dx.doi.org/10.1111/j.1461-0248.2010.01510.x

Bohler, S; Bagard, M; Oufir, M; Planchon, S; Hoffmann, L; Jolivet, Y; Hausman, JF; Dizengremel, P; 

Renaut, J. (2007). A DIGE analysis of developing poplar leaves subjected to ozone reveals major 

changes in carbon metabolism. Proteomics 7: 1584-1599. http://dx.doi.org/10.1002/pmic.200600822

Bohler, S; Sergeant, K; Lefèvre, I; Jolivet, Y; Hoffmann, L; Renaut, J; Dizengremel, P; Hausman, JF. (2010). 

Differential impact of chronic ozone exposure on expanding and fully expanded poplar leaves. Tree 

Physiol 30: 1415-1432. http://dx.doi.org/10.1093/treephys/tpq082
Draft – Do Not Cite or Quote 9-160 June 2012

Bonn, B; Von Kuhlmann, R; Lawrence, MG. (2004). High contribution of biogenic hydroperoxides to 

secondary organic aerosol formation. Geophys Res Lett 31: L10108. 

http://dx.doi.org/10.1029/2003GL019172

Booker, FL; Burkey, KO; Overmyer, K; Jones, AM. (2004a). Differential responses of G-protein Arabidopsis 

thaliana mutants to ozone. New Phytol 162: 633-641. 

Booker, FL; Burkey, KO; Pursley, WA; Heagle, AS. (2007). Elevated carbon dioxide and ozone effects on 

peanut: I. Gas-exchange, biomass, and leaf chemistry. Crop Sci 47: 1475-1487. 

http://dx.doi.org/10.2135/cropsci2006.08.0537

Booker, FL; Fiscus, EL. (2005). The role of ozone flux and antioxidants in the suppression of ozone injury by 

elevated CO2 in soybean. J Exp Bot 56: 2139-2151. http://dx.doi.org/10.1093/jxb/eri214

Booker, FL; Fiscus, EL; Miller, JE. (2004b). Combined effects of elevated atmospheric carbon dioxide and 

ozone on soybean whole-plant water use. Environ Manage 33: S355-S362. 

http://dx.doi.org/10.1007/s00267-003-9144-z

Booker, FL; Muntifering, R; McGrath, M; Burkey, K; Decoteau, D; Fiscus, E; Manning, W; Krupa, S; 

Chappelka, A; Grantz, D. (2009). The ozone component of global change: Potential effects on 

agricultural and horticultural plant yield, product quality and interactions with invasive species. J 

Integr Plant Biol 51: 337-351. http://dx.doi.org/10.1111/j.1744-7909.2008.00805.x

Booker, FL; Prior, SA; Torbert, HA; Fiscus, EL; Pursley, WA; Hu, S. (2005). Decomposition of soybean 

grown under elevated concentrations of CO2 and O3. Global Change Biol 11: 685-698. 

http://dx.doi.org/10.1111/j.1365-2486.2005.00939.x

Booker, FL; Reid, CD; Brunschon-Harti, S; Fiscus, EL; Miller, JE. (1997). Photosynthesis and 

photorespiration in soybean [Glycine max (L) Merr] chronically exposed to elevated carbon dioxide 

and ozone. J Exp Bot 48: 1843-1852. 

Borowiak, K; Rucinska-Sobkowiak, R; Rymer, K; Gwozdz, EA; Zbierska, J. (2009). Biochemical markers of 

tropospheric ozone: Experimentation with test-plants. Polish Journal of Ecology 57: 3-14. 

Bou Jaoudé, M; Katerji, N; Mastrorilli, M; Rana, G. (2008a). Analysis of the effect of ozone on soybean in 

the Mediterranean region I: The consequences on crop-water status. Eur J Agron 28: 508-518. 

http://dx.doi.org/10.1016/j.eja.2007.09.002

Bou Jaoudé, M; Katerji, N; Mastrorilli, M; Rana, G. (2008b). Analysis of the ozone effect on soybean in the 

Mediterranean region II. The consequences on growth, yield and water use efficiency. Eur J Agron 28: 

519-525. http://dx.doi.org/10.1016/j.eja.2007.09.001

Broadmeadow, MSJ; Jackson, SB. (2000). Growth responses of Quercus petraea, Fraxinus excelsior and 

Pinus sylvestris to elevated carbon dioxide, ozone and water supply. New Phytol 146: 437-451. 

http://dx.doi.org/10.1046/j.1469-8137.2000.00665.x

Brook, JR; DiGiovanni, F; Cakmak, S; Meyers, TP. (1997). Estimation of dry deposition velocity using 

inferential models and site-specific meteorology--uncertainty due to siting of meteorological towers. 

Atmos Environ 31: 3911-3919. 

Bulbovas, P; de Souza, SR; de Moraes, RM; Luizao, F; Artaxo, P. (2007). Soybean 'Tracaja' seedlings 

exposed to ozone under controlled conditions. Pesqui Agropecu Bras 42: 641-646. 

http://dx.doi.org/10.1590/S0100-204X2007000500005

Burkey, KO; Booker, FL; Pursley, WA; Heagle, AS. (2007). Elevated carbon dioxide and ozone effects on 

peanut: II. Seed yield and quality. Crop Sci 47: 1488-1497. 

http://dx.doi.org/10.2135/cropsci2006.08.0538

Burkey, KO; Eason, G; Fiscus, EL. (2003). Factors that affect leaf extracellular ascorbic acid content and 

redox status. Physiol Plant 117: 51-57. http://dx.doi.org/10.1034/j.1399-3054.2003.1170106.x

Burkey, KO; Neufeld, HS; Souza, L; Chappelka, AH; Davison, AW. (2006). Seasonal profiles of leaf ascorbic 

acid content and redox state in ozone-sensitive wildflowers. Environ Pollut 143: 427-434. 

http://dx.doi.org/10.1016/j.envpol.2005.12.009

Bytnerowicz, A; Arbaugh, M; Schilling, S; Fraczek, W; Alexander, D. (2008). Ozone distribution and 

phytotoxic potential in mixed conifer forests of the San Bernardino Mountains, Southern California. 

Environ Pollut 155: 398-408. http://dx.doi.org/10.1016/j.envpol.2008.01.046

Caird, MA; Richards, JH; Donovan, LA. (2007). Nighttime stomatal conductance and transpiration in C-3 

and C-4 plants. Plant Physiol 143: 4-10. http://dx.doi.org/10.1104/pp.106.092940

Cal EPA (California Environmental Protection Agency). (2010). Air quality data branch main page. Available 

online at http://www.arb.ca.gov/aqd/aqdpage.htm (accessed January 28, 2011).
Draft – Do Not Cite or Quote 9-161 June 2012

Calatayud, A; Alvarado, JW; Barreno, E. (2002). Similar effects of ozone on four cultivars of lettuce in open 

top chambers during winter. Photosynthetica 40: 195-200. 

http://dx.doi.org/10.1023/A:1021333305592

Calatayud, A; Pomares, F; Barreno, E. (2006). Interactions between nitrogen fertilization and ozone in 

watermelon cultivar Reina de Corazones in open-top chambers. Effects on chlorophyll alpha 

fluorescence, lipid peroxidation, and yield. Photosynthetica 44: 93-101. 

http://dx.doi.org/10.1007/s11099-005-0163-2

Calatayud, V; Cervero, J; Sanz, MJ. (2007a). Foliar, physiologial and growth responses of four maple species 

exposed to ozone. Water Air Soil Pollut 185: 239-254. http://dx.doi.org/10.1007/s11270-007-9446-5

Calatayud, V; Sanz, MJ; Calvo, E; Cervero, J; Ansel, W; Klumpp, A. (2007b). Ozone biomonitoring with 

Bel-W3 tobacco plants in the city of Valencia (Spain). Water Air Soil Pollut 183: 283-291. 

http://dx.doi.org/10.1007/s11270-007-9376-2

Campbell, SJ; Wanek, R; Coulston, JW. (2007). Ozone injury in west coast forests: 6 years of monitoring -

Introduction. Portland, OR: U.S. Department of Agriculture. 

Cannon, WN. (1990). Olfactory response of eastern spruce budworm larvae to red spruce needles exposed to 

acid rain and elevated levels of ozone. J Chem Ecol 16: 3255-3261. 

http://dx.doi.org/10.1007/BF00982096

Carde, RT; Haynes, KF. (2004). Stucture of the pheromone communication channel in moths. In Advances in 

insect chemical ecology. Cambridge: Cambridge University Press. 

http://dx.doi.org/10.1017/CBO9780511542664.009

Casteel, CL; O'Neill, BF; Zavala, JA; Bilgin, DD; Berenbaum, MR; DeLucia, EH. (2008). Transcriptional 

profiling reveals elevated CO2 and elevated O3 alter resistance of soybean (Glycine max) to Japanese 

beetles (Popillia japonica). Plant Cell Environ 31: 419-434. http://dx.doi.org/10.1111/j.1365-

3040.2008.01782.x

Chapman, JA; King, JS; Pregitzer, KS; Zak, DR. (2005). Effects of elevated concentrations of atmospheric 

CO2 and tropospheric O-3 on decomposition of fine roots. Tree Physiol 25: 1501-1510. 

Chappelka, A; Skelly, J; Somers, G; Renfro, J; Hildebrand, E. (1999a). Mature black cherry used as a 

bioindicator of ozone injury. Water Air Soil Pollut 116: 261-266. 

Chappelka, A; Somers, G; Renfro, J. (1999b). Visible ozone injury on forest trees in Great Smoky Mountains 

National Park, USA. Water Air Soil Pollut 116: 255-260. 

Chappelka, AH. (2002). Reproductive development of blackberry (Rubus cuneifolius) as influenced by 

ozone. New Phytol 155: 249-255. http://dx.doi.org/10.1046/j.1469-8137.2002.00464.x

Chappelka, AH; Neufeld, HS; Davison, AW; Somers, GL; Renfro, JR. (2003). Ozone injury on cutleaf 

coneflower (Rudbeckia laciniata) and crown-beard (Verbesina occidentalis) in Great Smoky 

Mountains National Park. Environ Pollut 125: 53-60. 

Chappelka, AH; Samuelson, LJ. (1998). Ambient ozone effects on forest trees of the eastern United States: A 

review [Review]. New Phytol 139: 91-108. http://dx.doi.org/10.1046/j.1469-8137.1998.00166.x

Chappelka, AH; Somers, GL; Renfro, JR. (2007). Temporal patterns of foliar ozone symptoms on tall 

milkweed (Asclepias exaltata L) in Great Smoky Mountains National Park. Environ Pollut 149: 358-

365. http://dx.doi.org/10.1016/j.envpol.2007.05.015

Chen, CW; Tsai, WT; Lucier, AA. (1998). A model of air-tree-soil system for ozone impact analysis. Ecol 

Modell 111: 207-222. 

Chen, Z; Gallie, DR. (2005). Increasing tolerance to ozone by elevating foliar ascorbic acid confers greater 

protection against ozone than increasing avoidance. Plant Physiol 138: 1673-1689. 

http://dx.doi.org/10.1104/pp.105.062000

Chen, Z; Wang, XK; Feng, ZZ; Xiao, Q; Duan, XN. (2009). Impact of elevated O3 on soil microbial 

community function under wheat crop. Water Air Soil Pollut 198: 189-198. 

http://dx.doi.org/10.1007/s11270-008-9838-1

Chen, Z; Wang, XK; Yao, FF; Zheng, FX; Feng, ZZ. (2010b). Elevated ozone changed soil microbial 

community in a rice paddy. Soil Sci Soc Am J 74: 829-837. http://dx.doi.org/10.2136/sssaj2009.0258

Cheng, FY; Burkey, KO; Robinson, JM; Booker, FL. (2007). Leaf extracellular ascorbate in relation to O3 

tolerance of two soybean cultivars. Environ Pollut 150: 355-362. 

http://dx.doi.org/10.1016/j.envpol.2007.01.022
Draft – Do Not Cite or Quote 9-162 June 2012

Cho, K; Shibato, J; Agrawal, GK; Jung, YH; Kubo, A; Jwa, NS; Tamogami, S; Satoh, K; Kikuchi, S; Higashi, 

T; Kimura, S; Saji, H; Tanaka, Y; Iwahashi, H; Masuo, Y; Rakwal, R. (2008). Integrated 

transcriptomics, proteomics, and metabolomics analyses to survey ozone responses in the leaves of 

rice seedling. J Proteome Res 7: 2980-2998. http://dx.doi.org/10.1021/pr800128q

Christman, MA; Donovan, LA; Richards, JH. (2009). Magnitude of nighttime transpiration does not affect 

plant growth or nutrition in well-watered Arabidopsis. Physiol Plant 136: 264-273. 

http://dx.doi.org/10.1111/j.1399-3054.2009.01216.x

Chung, HG; Zak, DR; Lilleskov, EA. (2006). Fungal community composition and metabolism under elevated 

CO2 and O3. Oecologia 147: 143-154. http://dx.doi.org/10.1007/s00442-005-0249-3

Colls, JJ; Unsworth, MH. (1992). Air pollution interactions with natural stressors. In JR Barker; DT Tingey 

(Eds.), Air pollution effects on biodiversity. New York, NY: Van Nostrand Reinhold. 

Cornelissen, T. (2011). Climate change and its effects on terrestrial insects and herbivory patterns. Neotrop 

Entomol 40: 155-163. http://dx.doi.org/10.1590/S1519-566X2011000200001

Costa, DL; Folinsbee, LJ; Raub, JA; Tilton, B; Tingey, DT. (1992). Summary of selected new information on 

effects of ozone on health and vegetation: Supplement to 1986 air quality criteria for ozone and other 

photochemical oxidants. (EPA/600/8-88/105F). Research Triangle Park, NC: U.S. Environmental 

Protection Agency. http://cfpub.epa.gov/ncea/isa/recordisplay.cfm?deid=31093

Coulston, JW; Smith, GC; Smith, WD. (2003). Regional assessment of ozone sensitive tree species using 

bioindicator plants. Environ Monit Assess 83: 113-127. 

Crous, KY; Vandermeiren, K; Ceulemans, R. (2006). Physiological responses to cumulative ozone uptake in 

two white clover (Trifolium repens L. cv. Regal) clones with different ozone sensitivity. Environ Exp 

Bot 58: 169-179. http://dx.doi.org/10.1016/j.envexpbot.2005.07.007

D'Haese, D; Vandermeiren, K; Asard, H; Horemans, N. (2005). Other factors than apoplastic ascorbate 

contribute to the differential ozone tolerance of two clones of Trifolium repens L. Plant Cell Environ 

28: 623-632. http://dx.doi.org/10.1111/j.1365-3040.2005.01308.x

Dalstein, L; Vas, N. (2005). Ozone concentrations and ozone-induced symptoms on coastal and alpine 

mediterranean pines in southern France. Water Air Soil Pollut 160: 181-195. 

Dammgen, U; Grunhage, L; Haenel, HD; Jager, HJ. (1993). Climate and stress in ecotoxicology: A coherent 

system of definitions and terms. J Appl Bot Food Qual 67: 157-162. 

Darbah, JNT; Kubiske, ME; Neilson, N; Oksanen, E; Vaapavuori, E; Karnosky, DF. (2007). Impacts of 

elevated atmospheric CO2 and O3 on paper birch (Betula papyrifera): Reproductive fitness. 

ScientificWorldJournal 7: 240-246. http://dx.doi.org/10.1100/tsw.2007.42

Darbah, JNT; Kubiske, ME; Nelson, N; Oksanen, E; Vapaavuori, E; Kamosky, DF. (2008). Effects of decadal 

exposure to interacting elevated CO2 and/or O-3 on paper birch (Betula papyrifera) reproduction. 

Environ Pollut 155: 446-452. http://dx.doi.org/10.1016/j.envpol.2008.01.033

Davidson, A. (1993). Update of ozone trends in California's South Coast Air Basin. Air Waste 43: 226-227. 

http://dx.doi.org/10.1080/1073161X.1993.10467130

Davis, DD. (2007a). Ozone-induced symptoms on vegetation within the Moosehorn National Wildlife 

Refuge in Maine. Northeast Nat 14: 403-414. http://dx.doi.org/10.1656/1092-

6194(2007)14[403:OSOVWT]2.0.CO;2

Davis, DD. (2007b). Ozone injury to plants within the Seney National Wildlife Refuge in northern Michigan. 

Northeast Nat 14: 415-424. 

Davis, DD. (2009). Ozone-induced stipple on plants in the Cape Romain National Wildlife Refuge, South 

Carolina. Southeastern Naturalist 8: 471-478. 

Davis, DD; Orendovici, T. (2006). Incidence of ozone symptoms on vegetation within a National Wildlife 

Refuge in New Jersey, USA. Environ Pollut 143: 555-564. 

http://dx.doi.org/10.1016/j.envpol.2005.10.051

Dawson, TE; Burgess, SS; Tu, KP; Oliveira, RS; Santiago, LS; Fisher, JB; Simonin, KA; Ambrose, AR.

(2007). Nighttime transpiration in woody plants from contrasting ecosystems. Tree Physiol 27: 561-

575. http://dx.doi.org/10.1093/treephys/27.4.561

de Lourdes de Bauer, M; Hernandez-Tejeda, T. (2007). A review of ozone-induced effects on the forests of 

central Mexico [Review]. Environ Pollut 147: 446-453. 

http://dx.doi.org/10.1016/j.envpol.2006.12.020

De Temmerman, L; Legrand, G; Vandermeiren, K. (2007). Effects of ozone on sugar beet grown in open-top 

chambers. Eur J Agron 26: 1-9. http://dx.doi.org/10.1016/j.eja.2006.08.001
Draft – Do Not Cite or Quote 9-163 June 2012

Degl'Innocenti, E; Guidi, L; Soldatini, GF. (2007). Effects of elevated ozone on chlorophyll a fluorescence in 

symptomatic and asymptomatic leaves of two tomato genotypes. Biol Plantarum 51: 313-321. 

http://dx.doi.org/10.1007/s10535-007-0061-5

Dermody, O; O'Neill, BF; Zangerl, AR; Berenbaum, MR; DeLucia, EH. (2008). Effects of elevated CO2 and 

O3 on leaf damage and insect abundance in a soybean agroecosystem. Arthropod-Plant Inte 2: 125-

135. 

DiBaccio, D; Castagna, A; Paoletti, E; Sebastiani, L; Ranieri, A. (2008). Could the differences in O3 

sensitivity between two poplar clones be related to a difference in antioxidant defense and secondary 

metabolic response to O3 influx? Tree Physiol 28: 1761-1772. 

Dickson, RE; Lewin, KF; Isebrands, JG; Coleman, MD; Heilman, WE; Riemenschneider, DE; Sober, J; Host, 

GE; Zak, DR; Hendrey, GR; Pregitzer, KS; Karnosky, DF. (2000). Forest Atmosphere Carbon Transfer 

and Storage (FACTS-II) the Aspen Free-Air CO2 and O3 Enrichment (FACE) project: An overview. 

(General Technical Report NC-214). St. Paul, MN: U.S. Dept. of Agriculture, Forest Service. 

http://nrs.fs.fed.us/pubs/278

Ditchkoff, SS; Lewis, JS; Lin, JC; Muntifering, RB; Chappelka, AH. (2009). Nutritive quality of highbush 

blackberry (Rubus argutus) exposed to tropospheric ozone. Rangeland Ecol Manag 62: 364-370. 

Dizengremel, P; Le Thiec, D; Bagard, M; Jolivet, Y. (2008). Ozone risk assessment for plants: Central role of 

metabolism-dependent changes in reducing power. Environ Pollut 156: 11-15. 

http://dx.doi.org/10.1016/j.envpol.2007.12.024

Dizengremel, P; Le Thiec, D; Hasenfratz-Sauder, MP; Vaultier, MN; Bagard, M; Jolivet, Y. (2009). 

Metabolic-dependent changes in plant cell redox power after ozone exposure. Plant Biol (Stuttg) 11: 

35-42. http://dx.doi.org/10.1111/j.1438-8677.2009.00261.x

Dizengremel, P; Sasek, T; Brown, K; Richardson, C. (1994). Ozone-induced changes in primary carbon 

metabolism enzymes of loblolly pine needles. J Plant Physiol 144: 300-306. 

Dobson, HEM. (1994). Floral volatiles in insect biology. In EA Bernays (Ed.), Insect-plant interactions: Vol 5 

(pp. 47-82). Boca Raton, FL: CRC Press. 

Dohm, MR; Mautz, WJ; Andrade, JA; Gellert, KS; Salas-Ferguson, LJ; Nicolaisen, N; Fujie, N. (2005). 

Effects of ozone exposure on nonspecific phagocytic capacity of pulmonary macrophages from an 

amphibian, Bufo marinus. Environ Toxicol Chem 24: 205-210. 

Dohm, MR; Mautz, WJ; Doratt, RE; Stevens, JR. (2008). Ozone exposure affects feeding and locomotor 

behavior of adult Bufo marinus. Environ Toxicol Chem 27: 1209-1216. http://dx.doi.org/10.1897/07-

388.1

Dohm, MR; Mautz, WJ; Looby, PG; Gellert, KS; Andrade, JA. (2001). Effects of ozone on evaporative water 

loss and thermoregulatory behavior of marine toads (Bufo marinus). Environ Res 86: 274-286. 

Dohrmann, AB; Tebbe, CC. (2005). Effect of elevated tropospheric ozone on the structure of bacterial 

communities inhabiting the rhizosphere of herbaceous plants native to Germany. Appl Environ 

Microbiol 71: 7750-7758. http://dx.doi.org/10.1128/AEM.71.12.7750-7758.2005

Drogoudi, PD; Ashmore, M. (2001). 14C-allocation of flowering and deblossomed strawberry in response to 

elevated ozone. New Phytol 152: 455-461. http://dx.doi.org/10.1046/j.0028-646X.2001.00270.x

Drogoudi, PD; Ashmore, MR. (2000). Does elevated ozone have differing effects in flowering and 

deblossomed strawberry? New Phytol 147: 561-569. http://dx.doi.org/10.1046/j.1469-

8137.2000.00718.x

Dudareva, N; Negre, F; Nagegowda, DA; Orlova, I. (2006). Plant volatiles: Recent advances and future 

perspectives [Review]. Crit Rev Plant Sci 25: 417-440.

Ederli, L; Morettini, R; Borgogni, A; Wasternack, C; Miersch, O; Reale, L; Ferranti, F; Tosti, N; Pasqualini, 

S. (2006). Interaction between nitric oxide and ethylene in the induction of alternative oxidase in 

ozone-treated tobacco plants. Plant Physiol 142: 595-608. http://dx.doi.org/10.1104/pp.106.085472

Edwards, IP; Zak, DR. (2011). Fungal community composition and function after long-term exposure of 

northern forests to elevated atmospheric CO2 and tropospheric O3. Global Change Biol 17: 2184-

2195. http://dx.doi.org/10.1111/j.1365-2486.2010.02376.x

Ellenson, JL; Amundson, RG. (1982). Delayed light imaging for the early detection of plant stress. Science 

215: 1104-1106. http://dx.doi.org/10.1126/science.215.4536.1104

Ellsworth, DS; Reich, PB; Naumburg, ES; Koch, GW; Kubiske, ME; Smith, SD. (2004). Photosynthesis, 

carboxylation and leaf nitrogen responses of 16 species to elevated pCO2 across four free-air CO2

enrichment experiments in forest, grassland and desert. Global Change Biol 10: 2121-2138. 

http://dx.doi.org/10.1111/j.1365-2486.2004.00867.x
Draft – Do Not Cite or Quote 9-164 June 2012

Eltayeb, AE; Kawano, N; Badawi, GH; Kaminaka, H; Sanekata, T; Morishima, I; Shibahara, T; Inanaga, S; 

Tanaka, K. (2006). Enhanced tolerance to ozone and drought stresses in transgenic tobacco 

overexpressing dehydroascorbate reductase in cytosol. Physiol Plant 127: 57-65. 

http://dx.doi.org/10.1111/j.1399-3054.2005.00624.x

Eltayeb, AE; Kawano, N; Badawi, GH; Kaminaka, H; Sanekata, T; Shibahara, T; Inanaga, S; Tanaka, K.

(2007). Overexpression of monodehydroascorbate reductase in transgenic tobacco confers enhanced 

tolerance to ozone, salt and polyethylene glycol stresses. Planta 225: 1255-1264. 

http://dx.doi.org/10.1007/s00425-006-0417-7

Emberson, L; Ashmore, MR; Cambridge, HM; Simpson, D; Tuovinen, JP. (2000a). Modelling stomatal ozone 

flux across Europe. Environ Pollut 109: 403-413. http://dx.doi.org/10.1016/S0269-7491(00)00043-9

Emberson, LD; Wieser, G; Ashmore, MR. (2000b). Modelling of stomatal conductance and ozone flux of 

Norway spruce: Comparison with field data. Environ Pollut 109: 393-402. 

http://dx.doi.org/10.1016/S0269-7491(00)00042-7

Enders, G. (1992). Deposition of ozone to a mature spruce forest: Measurements and comparison to models. 

Environ Pollut 75: 61-67. http://dx.doi.org/10.1016/0269-7491(92)90057-H

Esperschutz, J; Pritsch, K; Gattinger, A; Welzl, G; Haesler, F; Buegger, F; Winkler, JB; Munch, JC; Schloter, 

M. (2009). Influence of chronic ozone stress on carbon translocation pattern into rhizosphere 

microbial communities of beech trees (Fagus sylvatica L.) during a growing season. Plant Soil 323: 

85-95. http://dx.doi.org/10.1007/s11104-009-0090-2

Fares, S; Barta, C; Brilli, F; Centritto, M; Ederli, L; Ferranti, F; Pasqualini, S; Reale, L; Tricoli, D; Loreto, F.

(2006). Impact of high ozone on isoprene emission, photosynthesis and histology of developing 

Populus alba leaves directly or indirectly exposed to the pollutant. Physiol Plant 128: 456-465. 

http://dx.doi.org/10.1111/j.1399-3054.2006.00750.x

Fares, S; Gentner, DR; Park, JH; Ormeno, E; Karlik, J; Goldstein, AH. (2011). Biogenic emissions from 

Citrus species in California. Atmos Environ 45: 4557-4568. 

http://dx.doi.org/10.1016/j.atmosenv.2011.05.066

Fares, S; Loreto, F; Kleist, E; Wildt, J. (2008). Stomatal uptake and stomatal deposition of ozone in isoprene 

and monoterpene emitting plants. Plant Biol (Stuttg) 10: 44-54. http://dx.doi.org/10.1055/s-2007-

965257

Fares, S; McKay, M; Holzinger, R; Goldstein, AH. (2010a). Ozone fluxes in a Pinus ponderosa ecosystem are 

dominated by non-stomatal processes: Evidence from long-term continuous measurements. Agr Forest 

Meteorol 150: 420-431. http://dx.doi.org/10.1016/j.agrformet.2010.01.007

Fares, S; Oksanen, E; Lannenpaa, M; Julkunen-Tiitto, R; Loreto, F. (2010b). Volatile emissions and phenolic 

compound concentrations along a vertical profile of Populus nigra leaves exposed to realistic ozone 

concentrations. Photosynth Res 104: 61-74. http://dx.doi.org/10.1007/s11120-010-9549-5

Felzer, B; Kicklighter, D; Melillo, J; Wang, C; Xhuang, Q; Prinn, R. (2004). Effects of ozone on net primary 

production and carbon sequestration in the conterminous United States using a biogeochemistry 

model. Tellus B Chem Phys Meteorol 56: 230-248. http://dx.doi.org/10.1111/j.1600-

0889.2004.00097.x

Felzer, B; Reilly, J; Melillo, J; Kicklighter, D; Sarofim, M; Wang, C; Prinn, R; Zhuang, Q. (2005). Future 

effects of ozone on carbon sequestration and climate change policy using a global biogeochemical 

model. Clim Change 73: 345-373. http://dx.doi.org/10.1007/s10584-005-6776-4

Felzer, BS; Cronin, TW; Melillo, JM; Kicklighter, DW; Schlosser, CA. (2009). Importance of carbonnitrogen interactions and ozone on ecosystem hydrology during the 21st century. J Geophys Res 114: 

G01020. 

Feng, YW; Komatsu, S; Furukawa, T; Koshiba, T; Kohno, Y. (2008a). Proteome analysis of proteins 

responsive to ambient and elevated ozone in rice seedlings. Agric Ecosyst Environ 125: 255-265. 

http://dx.doi.org/10.1016/j.agee.2008.01.018

Feng, Z; Pang, J; Nouchi, I; Kobayashi, K; Yamakawa, T; Zhu, J. (2010). Apoplastic ascorbate contributes to 

the differential ozone sensitivity in two varieties of winter wheat under fully open-air field conditions. 

Environ Pollut 158: 3539-3545. http://dx.doi.org/10.1016/j.envpol.2010.08.019

Feng, ZZ; Kobayashi, K. (2009). Assessing the impacts of current and future concentrations of surface ozone 

on crop yield with meta-analysis. Atmos Environ 43: 1510-1519. 

http://dx.doi.org/10.1016/j.atmosenv.2008.11.033
Draft – Do Not Cite or Quote 9-165 June 2012

Feng, ZZ; Kobayashi, K; Ainsworth, EA. (2008b). Impact of elevated ozone concentration on growth, 

physiology, and yield of wheat (Triticum aestivum L.): A meta-analysis. Global Change Biol 14: 2696-

2708. http://dx.doi.org/10.1111/j.1365-2486.2008.01673.x

Fenn, ME; de Bauer, LI; Hernández-Tejeda, T. (2002). Summary of air pollution impacts on forests in the 

Mexico City air basin. In Urban air pollution and forests. New York, NY: Springer-Verlag. 

Fenn, ME; Poth, MA; Johnson, DW. (1996). Evidence for nitrogen saturation in the San Bernardino 

Mountains in southern California. For Ecol Manage 82: 211-230. http://dx.doi.org/10.1016/0378-

1127(95)03668-7

Findley, DA; Keever, GJ; Chappelka, AH; Eakes, DJ; Gillian, DJ. (1997). Differential responses of buddleia 

(Buddleia davidii Franch) to ozone. Environ Pollut 98: 105-111. 

Finkelstein, PL; Ellestad, TG; Clarke, JF; Meyers, TP; Schwede, DB; Hebert, EO; Neal, JA. (2000). Ozone 

and sulfur dioxide dry deposition to forests: Observations and model evaluation. J Geophys Res 105: 

15365-15377. 

Finnan, JM; Burke, JL; Jones, MB. (1997). An evaluation of indices that describe the impact of ozone on the 

yield of spring wheat (Triticum aestivum L). Atmos Environ 31: 2685-2693. 

http://dx.doi.org/10.1016/S1352-2310(97)00105-2

Finnan, JM; Jones, MB; Burke, JI. (1996). A time-concentration study on the effects of ozone on spring 

wheat (Triticum aestivum L): 2. A comparison of indices. Agric Ecosyst Environ 57: 169-177. 

http://dx.doi.org/10.1016/0167-8809(95)01004-1

Fiscus, EL; Booker, FL; Burkey, KO. (2005). Crop responses to ozone: Uptake, modes of action, carbon 

assimilation and partitioning. Plant Cell Environ 28: 997-1011. 

Fiscus, EL; Philbeck, R; Britt, AM; Booker, FL. (1999). Growth of Arabidopsis flavonoid mutants under 

solar radiation and UV filters. Environ Exp Bot 41: 231-245. http://dx.doi.org/10.1016/S0098-

8472(99)00011-8

Fishman, J; Bowman, KW; Burrows, JP; Richter, A; Chance, KV; Edwards, DP; Martin, RV; Morris, GA; 

Pierce, RB; Ziemke, JR; Al-Saadi, JA; Creilson, JK; Schaack, TK; Thompson, AM. (2008). Remote 

sensing of tropospheric pollution from space. Bull Am Meteorol Soc 89: 805-821. 

http://dx.doi.org/10.1175/2008BAMS2526.1

Fishman, J; Creilson, JK; Parker, PA; Ainsworth, EA; Vining, GG; Szarka, J; Booker, FL; Xu, XJ. (2010). An 

investigation of widespread ozone damage to the soybean crop in the upper Midwest determined from 

ground-based and satellite measurements. Atmos Environ 44: 2248-2256. 

http://dx.doi.org/10.1016/j.atmosenv.2010.01.015

Flagler, RB. (1998). Recognition of air pollution injury to vegetation: A pictorial atlas (2nd ed.). Pittsburgh, 

PA: Air & Waste Management Association. 

Flowers, MD; Fiscus, EL; Burkey, KO; Booker, FL; Dubois, JJB. (2007). Photosynthesis, chlorophyll 

fluorescence, and yield of snap bean (Phaseolus vulgaris L.) genotypes differing in sensitivity to 

ozone. Environ Exp Bot 61: 190-198. http://dx.doi.org/10.1016/j.envexpbot.2007.05.009

Fontan, JA; Minga, A; Lopez, A; Druilhet, A. (1992). Vertical ozone profiles in a pine forest. Atmos Environ 

26: 863-869. http://dx.doi.org/10.1016/0960-1686(92)90245-G

Foyer, CH; Noctor, G. (2005a). Oxidant and antioxidant signalling in plants: A re-evaluation of the concept 

of oxidative stress in a physiological context. Plant Cell Environ 28: 1056-1071. 

Foyer, CH; Noctor, G. (2005b). Redox homeostasis and antioxidant signaling: A metabolic interface between 

stress perception and physiological responses. Plant Cell 17: 1866-1875. 

http://dx.doi.org/10.1105/tpc.105.033589

Fredericksen, TS; Joyce, BJ; Skelly, JM; Steiner, KC; Kolb, TE; Kouterick, KB; Savage, JE; Snyder, KR.

(1995). Physiology, morphology, and ozone uptake of leaves of black cherry seedlings, saplings, and 

canopy trees. Environ Pollut 89: 273-283. http://dx.doi.org/10.1016/0269-7491(94)00077-Q

Fredericksen, TS; Kolb, TE; Skelly, JM; Steiner, KC; Joyce, BJ; Savage, JE. (1996). Light environment alters 

ozone uptake per net photosynthetic rate in black cherry trees. Tree Physiol 16: 485-490. 

Freiwald, V; Haikio, E; Julkunen-Tiitto, R; Holopainen, JK; Oksanen, E. (2008). Elevated ozone modifies 

the feeding behaviour of the common leaf weevil on hybrid aspen through shifts in developmental, 

chemical, and structural properties of leaves. Entomol Exp Appl 128: 66-72. 

http://dx.doi.org/10.1111/j.1570-7458.2008.00677.x

Fuentes, JD; Gillespie, TJ; den Hartog, G; Neumann, HH. (1992). Ozone deposition onto a deciduous forest 

during dry and wet conditions. Agr Forest Meteorol 62: 1-18. http://dx.doi.org/10.1016/0168-

1923(92)90002-L
Draft – Do Not Cite or Quote 9-166 June 2012

Fuhrer, J. (1994). Effects of ozone on managed pasture: 1. Effects of open-top chambers on microclimate, 

ozone flux, and plant growth. Environ Pollut 86: 297-305. 

Fuhrer, J; Skarby, L; Ashmore, MR. (1997). Critical levels for ozone effects on vegetation in Europe. Environ 

Pollut 97: 91-106. http://dx.doi.org/10.1016/S0269-7491(97)00067-5

Gate, IM; McNeill, S; Ashmore, MR. (1995). Effects of air pollution on the searching behaviour of an insect 

parasitoid. Water Air Soil Pollut 85: 1425-1430. http://dx.doi.org/10.1007/BF00477181

Geiser, LH; Neitlich, PN. (2007). Pollution and climate gradients in western Oregon and Washington 

indicated by epiphytic macrolichens. Environ Pollut 145: 203-218. 

http://dx.doi.org/10.1016/j.envpol.2006.03.024

Gerosa, G; Marzuoli, R; Rossini, M; Panigada, C; Meroni, M; Colombo, R; Faoro, F; Iriti, M. (2009). A fluxbased assessment of the effects of ozone on foliar injury, photosynthesis, and yield of bean (Phaseolus 

vulgaris L. cv. Borlotto Nano Lingua di Fuoco) in open-top chambers. Environ Pollut 157: 1727-1736. 

http://dx.doi.org/10.1016/j.envpol.2008.06.028

Gielen, B; Vandermeiren, K; Horemans, N; D'Haese, D; Serneels, R; Valcke, R. (2006). Chlorophyll a 

fluorescence imaging of ozone-stressed Brassica napus L. plants differing in glucosinolate 

concentrations. Plant Biol (Stuttg) 8: 698-705. http://dx.doi.org/10.1055/s-2006-924150

Gitay, H; Brown, S; Easterling, W; Jallow, B. (2001). Ecosystems and their goods and services. In Climate 

change 2001: Impacts, adaptation and vulnerability: Contribution of Working Group II to the third 

assessment report of the Intergovernmental Panel on Climate Change. Cambridge, United Kingdom: 

Cambridge University Press. 

Gombert, S; Asta, J; Seaward, MRD. (2006). Lichens and tobacco plants as complementary biomonitors of 

air pollution in the Grenoble area (Isere, southeast France). Ecol Indicat 6: 429-443. 

http://dx.doi.org/10.1016/j.ecolind.2005.06.001

Gonzalez-Fernandez, I; Kaminska, A; Dodmani, M; Goumenaki, E; Quarrie, S; Barnes, JD. (2010). 

Establishing ozone flux-response relationships for winter wheat: Analysis of uncertainties based on 

data for UK and Polish genotypes. Atmos Environ 44: 621-630. 

http://dx.doi.org/10.1016/j.atmosenv.2009.11.021

Goumenaki, E; Taybi, T; Borland, A; Barnes, J. (2010). Mechanisms underlying the impacts of ozone on 

photosynthetic performance. Environ Exp Bot 69: 259-266. 

http://dx.doi.org/10.1016/j.envexpbot.2010.04.011

Grantz, DA; Gunn, S; Vu, HB. (2006). O3 impacts on plant development: A meta-analysis of root/shoot 

allocation and growth. Plant Cell Environ 29: 1193-1209. http://dx.doi.org/10.1111/j.1365-

3040.2006.01521.x

Grantz, DA; Shrestha, A. (2006). Tropospheric ozone and interspecific competition between yellow nutsedge 

and pima cotton. Crop Sci 46: 1879-1889. http://dx.doi.org/10.2135/cropsci2005.06.0167

Grantz, DA; Shrestha, A; Vu, HB. (2008). Early vigor and ozone response in horseweed (Conyza canadensis) 

biotypes differing in glyphosate resistance. Weed Sci 56: 224-230. http://dx.doi.org/10.1614/ws-07-

130.1

Grantz, DA; Shrestha, A; Vu, HB. (2010a). Ozone impacts on assimilation and allocation to reproductive 

sinks in the vegetatively propagated C-4 weed, yellow nutsedge. Crop Sci 50: 246-252. 

http://dx.doi.org/10.2135/cropsci2009.03.0127

Grantz, DA; Vu, HB. (2009). O3 sensitivity in a potential C4 bioenergy crop: Sugarcane in California. Crop 

Sci 49: 643-650. 

Grantz, DA; Vu, HB; Aguilar, C; Rea, MA. (2010b). No interaction between methyl jasmonate and ozone in 

Pima cotton: Growth and allocation respond independently to both. Plant Cell Environ 33: 717-728. 

http://dx.doi.org/10.1111/j.1365-3040.2009.02096.x

Grantz, DA; Zhang, XJ; Massman, W; Delany, A; Pederson, R. (1997). Ozone deposition to a cotton 

(Gossypium hirsutum L) field: Stomatal and surface wetness effects during the California Ozone 

Deposition experiment. Agr Forest Meteorol 85: 19-31. http://dx.doi.org/10.1016/S0168-

1923(96)02396-9

Grantz, DA; Zhang, XJ; Massman, WJ; Den Hartog, G; Neumann, HH; Pederson, JR. (1995). Effects of 

stomatal conductance and surface wetness on ozone deposition in field-grown grape. Atmos Environ 

29: 3189-3198. http://dx.doi.org/10.1016/1352-2310(95)00129-M

Grebenc, T; Kraigher, H. (2007). Changes in the community of ectomycorrhizal fungi and increased fine root 

number under adult beech trees chronically fumigated with double ambient ozone concentration. Plant 

Biol (Stuttg) 9: 279-287. http://dx.doi.org/10.1055/s-2006-924489
Draft – Do Not Cite or Quote 9-167 June 2012

Gregg, JW; Jones, CG; Dawson, TE. (2003). Urbanization effects on tree growth in the vicinity of New York 

City [Letter]. Nature 424: 183-187. http://dx.doi.org/10.1038/nature01728

Gregg, JW; Jones, CG; Dawson, TE. (2006). Physiological and developmental effects of O3 on cottonwood 

growth in urban and rural sites. Ecol Appl 16: 2368-2381. http://dx.doi.org/10.1890/1051-

0761(2006)016[2368:PADEOO]2.0.CO;2

Grennfelt, P. (2004). New directions: Recent research findings may change ozone control policies. Atmos 

Environ 38: 2215-2216. 

Groppa, MD; Benavides, MP. (2008). Polyamines and abiotic stress: Recent advances. Amino Acids 34: 35-

45. http://dx.doi.org/10.1007/s00726-007-0501-8

Grulke, N; Neufeld, H; Davison, A; Roberts, M; Chappelka, A. (2007a). Stomatal behavior of ozonesensitive and -insensitive coneflowers (Rudbeckia laciniata var. digitata) in Great Smoky Mountains 

National Park. New Phytol 173: 100-109. http://dx.doi.org/10.1111/j.1469-8137.2006.01872.x

Grulke, NE. (1999). Physiological responses of ponderosa pine to gradients of environmental stressors. In PR 

Miller; JR McBride (Eds.), Oxidant air pollution impacts in the montane forests of Southern 

California (pp. 126-163). New York, NY: Springer-Verlag. 

Grulke, NE; Alonso, R; Nguyen, T; Cascio, C; Dobrowolski, W. (2004). Stomata open at night in pole-sized 

and mature ponderosa pine: Implications for O3 exposure metrics. Tree Physiol 24: 1001-1010. 

Grulke, NE; Dobrowolski, W; Mingus, P; Fenn, ME. (2005). California black oak response to nitrogen 

amendment at a high O3, nitrogen-saturated site. Environ Pollut 137: 536-545. 

http://dx.doi.org/10.1016/j.envpol.2005.01.039

Grulke, NE; Johnson, R; Esperanza, A; Jones, D; Nguyen, T; Posch, S; Tausz, M. (2003a). Canopy 

transpiration of Jeffrey pine in mesic and xeric microsites: O3 uptake and injury response. Trees Struct 

Funct 17: 292-298. 

Grulke, NE; Johnson, R; Monschein, S; Nikolova, P; Tausz, M. (2003b). Variation in morphological and 

biochemical O3 injury attributes of mature Jeffrey pine within canopies and between microsites. Tree 

Physiol 23: 923-929. 

Grulke, NE; Lee, EH. (1997). Assessing visible ozone-induced injury in ponderosa pine. Can J For Res 27: 

1658-1668. 

Grulke, NE; Paoletti, E; Heath, RL. (2007b). Chronic vs. short-term acute O3 exposure effects on nocturnal 

transpiration in two Californian oaks. ScientificWorldJournal 7: 134-140. 

http://dx.doi.org/10.1100/tsw.2007.33

Grulke, NE; Paoletti, E; Heath, RL. (2007c). Comparison of calculated and measured foliar O3 flux in crop 

and forest species. Environ Pollut 146: 640-647. http://dx.doi.org/10.1016/j.envpol.2006.04.014

Grulke, NE; Preisler, HK; Rose, C; Kirsch, J; Balduman, L. (2002). O3 uptake and drought stress effects on 

carbon acquisition of ponderosa pine in natural stands. New Phytol 154: 621-631. 

http://dx.doi.org/10.1046/j.1469-8137.2002.00403.x

Grunhage, L; Haenel, HD. (1997). PLATIN (PLant-ATmosphere-INteraction) I: A model of plant-atmosphere 

interaction for estimating absorbed doses of gaseous air pollutants. Environ Pollut 98: 37-50. 

http://dx.doi.org/10.1016/S0269-7491(97)00114-0

Grunhage, L; Jager, HJ. (2003). From critical levels to critical loads for ozone: a discussion of a new 

experimental and modelling approach for establishing flux-response relationships for agricultural 

crops and native plant species. Environ Pollut 125: 99-110. 

Grunhage, L; Krupa, SV; Legge, AH; Jager, HJ. (2004). Ambient flux-based critical values of ozone for 

protecting vegetation: Differing spatial scales and uncertainties in risk assessment. Atmos Environ 38: 

2433-2437. http://dx.doi.org/10.1016/j.atmosenv.2003.12.039

Guenther, A; Karl, T; Harley, P; Wiedinmyer, C; Palmer, PI; Geron, C. (2006). Estimates of global terrestrial 

isoprene emissions using MEGAN (Model of Emissions of Gases and Aerosols from Nature). Atmos 

Chem Phys 6: 3181-3210. http://dx.doi.org/10.5194/acp-6-3181-2006

Guidi, L; Degl'Innocenti, E; Martinelli, F; Piras, M. (2009). Ozone effects on carbon metabolism in sensitive 

and insensitive Phaseolus cultivars. Environ Exp Bot 66: 117-125. 

http://dx.doi.org/10.1016/j.envexpbot.2008.12.005

Guidi, L; Degl'lnnocenti, E. (2008). Ozone effects on high light-induced photoinhibition in Phaseolus 

vulgaris. Plant Sci 174: 590-596. http://dx.doi.org/10.1016/j.plantsci.2008.03.003

Gül, H; Gaga, EO; Döğeroğlu, T; Ozden, O; Ayvaz, O; Ozel, S; Güngör, G. (2011). Respiratory health 

symptoms among students exposed to different levels of air pollution in a Turkish city. Int J Environ 

Res Public Health 8: 1110-1125. http://dx.doi.org/10.3390/ijerph8041110
Draft – Do Not Cite or Quote 9-168 June 2012

Gumpertz, ML; Rawlings, JO. (1992). Nonlinear regression with variance components: Modeling effects of 

ozone on crop yield. Crop Sci 32: 219-224. 

http://dx.doi.org/10.2135/cropsci1992.0011183X003200010045x

Gunderson, CA; Sholtis, JD; Wullschleger, SD; Tissue, DT; Hanson, PJ; Norby, RJ. (2002). Environmental 

and stomatal control of photosynthetic enhancement in the canopy of a sweetgum (Liquidambar 

styraciflua L) plantation during 3 years of CO2 enrichment. Plant Cell Environ 25: 379-393. 

http://dx.doi.org/10.1046/j.0016-8025.2001.00816.x

Haberer, K; Herbinger, K; Alexou, M; Rennenberg, H; Tausz, M. (2008). Effects of drought and canopy 

ozone exposure on antioxidants in fine roots of mature European beech (Fagus sylvatica). Tree Physiol 

28: 713-719. http://dx.doi.org/10.1093/treephys/28.5.713

Hamel, LP; Miles, GP; Samuel, MA; Ellis, BE; Seguin, A; Beaudoin, N. (2005). Activation of stressresponsive mitogen-activated protein kinase pathways in hybrid poplar (Populus trichocarpa x Populus 

deltoides). Tree Physiol 25: 277-288. http://dx.doi.org/10.1093/treephys/25.3.277

Hamilton, JG; Dermody, O; Aldea, M; Zangerl, AR; Rogers, A; Berenbaum, MR; DeLucia, EH. (2005). 

Anthropogenic changes in tropospheric composition increase susceptibility of soybean to insect 

herbivory. Environ Entomol 34: 479-485. 

Handley, T; Grulke, NE. (2008). Interactive effects of O3 exposure on California black oak (Quercus 

kelloggii Newb.) seedlings with and without N amendment. Environ Pollut 156: 53-60. 

http://dx.doi.org/10.1016/j.envpol.2008.01.002

Hanson, PJ; Wullschleger, SD; Norby, RJ; Tschaplinski, TJ; Gunderson, CA. (2005). Importance of changing 

CO2, temperature, precipitation, and ozone on carbon and water cycles of an upland-oak forest: 

incorporating experimental results into model simulations. Global Change Biol 11: 1402-1423. 

http://dx.doi.org/10.1111/j.1365-2486.2005.00991.x

Harward, M; Treshow, M. (1975). Impact of ozone on the growth and reproduction of understorey plants in 

the Aspen zone of western USA. Environ Conserv 2: 17-23. 

http://dx.doi.org/10.1017/S0376892900000564

Hassan, R; Scholes, R; Ash, N. (2005). Ecosystems and human well-being: Current state and trends. 

Washington, DC: Island Press. 

Hayes, F; Jones, MLM; Mills, G; Ashmore, M. (2007). Meta-analysis of the relative sensitivity of seminatural vegetation species to ozone. Environ Pollut 146: 754-762. 

http://dx.doi.org/10.1016/j.envpol.2006.06.011

Hayes, F; Mills, G; Ashmore, M. (2009). Effects of ozone on inter- and intra-species competition and 

photosynthesis in mesocosms of Lolium perenne and Trifolium repens. Environ Pollut 157: 208-214. 

http://dx.doi.org/10.1016/j.envpol.2008.07.002

Hayes, F; Mills, G; Ashmore, M. (2010). How Much Does the Presence of a Competitor Modify the WithinCanopy Distribution of Ozone-Induced Senescence and Visible Injury? Water Air Soil Pollut 210: 

265-276. http://dx.doi.org/10.1007/s11270-009-0248-9

He, XY; Fu, SL; Chen, W; Zhao, TH; Xu, S; Tuba, Z. (2007). Changes in effects of ozone exposure on 

growth, photosynthesis, and respiration of Ginkgo biloba in Shenyang urban area. Photosynthetica 45: 

555-561. http://dx.doi.org/10.1007/s11099-007-0095-0

He, XY; Ruan, YN; Chen, W; Lu, T. (2006). Responses of anti-oxidative system in leaves of Ginkgo biloba 

to elevated ozone concentration in urban area. Botanical Studies 47: 409-416. 

Heagle, AS. (1979). Effects of growth media, fertiliser rate and hour and season of exposure on sensitivity of 

four soybean cultivars to ozone. Environ Pollut 18: 313-322. 

Heagle, AS. (1989). Ozone and crop yield*. Annu Rev Phytopathol 27: 397-423. 

http://dx.doi.org/10.1146/annurev.py.27.090189.002145

Heagle, AS; Body, DE; Heck, WW. (1973). An open-top field chamber to assess the impact of air pollution 

on plants. J Environ Qual 2: 365-368. 

Heagle, AS; Brandenburg, RL; Burns, JC; Miller, JE. (1994a). Ozone and carbon dioxide effects on spider 

mites in white clover and peanut. J Environ Qual 23: 1168-1176. 

http://dx.doi.org/10.2134/jeq1994.00472425002300060006x

Heagle, AS; Heck, WW; Lesser, VM; Rawlings, JO. (1987). Effects of daily ozone exposure duration and 

concentration fluctuation on yield of tobacco. Phytopathology 77: 856-862. 

http://dx.doi.org/10.1094/Phyto-77-856
Draft – Do Not Cite or Quote 9-169 June 2012

Heagle, AS; Kress, LW; Temple, PJ; Kohut, RJ; Miller, JE; Heggestad, HE. (1988). Factors influencing 

ozone dose-yield response relationships in open-top field chamber studies. In WW Heck; OC Taylor; 

DT Tingey (Eds.), Assessment of crop loss from air pollutants: Proceedings of an international 

conference (pp. 141-179). New York, NY: Elsevier Applied Science. 

Heagle, AS; Letchworth, MB; Mitchell, CA. (1983). Effects of growth medium and fertilizer rate on the yield 

response of soybeans exposed to chronic doses of ozone. Phytopathology 73: 134-139. 

http://dx.doi.org/10.1094/Phyto-73-134

Heagle, AS; Miller, JE; Rawlings, JO; Vozzo, SF. (1991). Effect of growth stage on soybean response to 

chronic ozone exposure. J Environ Qual 20: 562-570. 

http://dx.doi.org/10.2134/jeq1991.00472425002000030010x

Heagle, AS; Miller, JE; Sherrill, DE. (1994b). A white clover system to estimate effects of tropospheric 

ozone on plants. J Environ Qual 23: 613-621. 

http://dx.doi.org/10.2134/jeq1994.00472425002300030030x

Heagle, AS; Reinert, RA; Miller, JE. (1996). Response of white clover to ozone in different environments. J 

Environ Qual 25: 273-278. http://dx.doi.org/10.2134/jeq1996.00472425002500020010x

Heath, RL. (2008). Modification of the biochemical pathways of plants induced by ozone: What are the 

varied routes to change? Environ Pollut 155: 453-463. http://dx.doi.org/10.1016/j.envpol.2008.03.010

Heath, RL; Lefohn, AS; Musselman, RC. (2009). Temporal processes that contribute to nonlinearity in 

vegetation responses to ozone exposure and dose. Atmos Environ 43: 2919-2928. 

http://dx.doi.org/10.1016/j.atmosenv.2009.03.011

Heck, WW; Cowling, EB. (1997). The need for a long term cumulative secondary ozone standard - An 

ecological perspective. EM January: 23-33. 

Heck, WW; Cure, WW; Rawlings, JO; Zaragoza, LJ; Heagle, AS; Heggestad, HE; Kohut, RJ; Kress, LW; 

Temple, PJ. (1984). Assessing impacts of ozone on agricultural crops: II. Crop yield functions and 

alternative exposure statistics. J Air Pollut Control Assoc 34: 810-817. 

Heck, WW; Heagle, AS; Miller, JE; Rawlings, JO. (1991). A national program (NCLAN) to assess the impact 

of ozone on agricultural resources. In RL Berglund; DR Lawson; DJ McKee (Eds.), Tropospheric 

ozone and the environment: papers from an international conference; March 1990; Los Angeles, CA 

(pp. 225-254). Pittsburgh, PA: Air & Waste Management Association. 

Heck, WW; Philbeck, RB; Dunning, JA. (1978). A continuous stirred tank reactor (CSTR) system for 

exposing plants to gaseous air contaminants: Principles, specifications, construction, and operation. 

Washington, DC: U.S. Government Printing Office. 

Heck, WW; Taylor, OC; Adams, R; Bingham, G; Miller, J; Preston, E; Weinstein, L. (1982). Assessment of 

crop loss from ozone. J Air Pollut Control Assoc 32: 353-361. 

Heggestad, HE. (1991). Origin of Bel-W3, Bel-C and Bel-B tobacco varieties and their use as indicators of 

ozone. Environ Pollut 74: 263-291. http://dx.doi.org/10.1016/0269-7491(91)90076-9

Heidenreich, B; Haberer, G; Mayer, K; Sandermann, H; Ernst, D. (2005). CDNA array-analysis of mercuryand ozone-induced genes in Arabidopsis thaliana. Acta Physiologiae Plantarum 27: 45-51. 

http://dx.doi.org/10.1007/s11738-005-0035-1

Hendrey, GR; Ellsworth, DS; Lewin, KF; Nagy, J. (1999). A free-air enrichment system for exposing tall 

forest vegetation to elevated atmospheric CO2. Global Change Biol 5: 293-309. 

http://dx.doi.org/10.1046/j.1365-2486.1999.00228.x

Hendrey, GR; Kimball, BA. (1994). The FACE program. Agr Forest Meteorol 70: 3-14. 

http://dx.doi.org/10.1016/0168-1923(94)90044-2

Hildebrand, E; Skelly, JM; Fredericksen, TS. (1996). Foliar response of ozone-sensitive hardwood tree 

species from 1991 to 1993 in the Shenandoah National Park, Virginia. Can J For Res 26: 658-669. 

Hillstrom, M; Meehan, TD; Kelly, K; Lindroth, RL. (2010a). Soil carbon and nitrogen mineralization 

following deposition of insect frass and greenfall from forests under elevated CO2 and O3. Plant Soil 

336: 75-85. http://dx.doi.org/10.1007/s11104-010-0449-4

Hillstrom, ML; Lindroth, RL. (2008). Elevated atmospheric carbon dioxide and ozone alter forest insect 

abundance and community composition. Insect Conservation and Diversity 1: 233-241. 

http://dx.doi.org/10.1111/j.1752-4598.2008.00031.x

Hillstrom, ML; Vigue, LM; Coyle, DR; Raffa, KF; Lindroth, RL. (2010b). Performance of the invasive 

weevil Polydrusus sericeus is influenced by atmospheric CO2 and host species. Agr Forest Entomol 

12: 285-292. http://dx.doi.org/10.1111/j.1461-9563.2010.00474.x
Draft – Do Not Cite or Quote 9-170 June 2012

Himanen, SJ; Nerg, AM; Holopainen, JK. (2009a). Degree of herbivore feeding damage as an important 

contributor to multitrophic plant-parasitoid signaling under climate change. Plant Signalling & 

Behavior 4: 249-251. 

Himanen, SJ; Nerg, AM; Nissinen, A; Pinto, DM; Stewart, CN; Poppy, GM; Holopainen, JK. (2009b). 

Effects of elevated carbon dioxide and ozone on volatile terpenoid emissions and multitrophic 

communication of transgenic insecticidal oilseed rape (Brassica napus). New Phytol 181: 174-186. 

http://dx.doi.org/10.1111/j.1469-8137.2008.02646.x

Himanen, SJ; Nerg, AM; Nissinen, A; Stewart, CN; Poppy, GM; Holopainen, JK. (2009c). Elevated 

atmospheric ozone increases concentration of insecticidal Bacillus thuringiensis (Bt) Cry1Ac protein 

in Bt Brassica napus and reduces feeding of a Bt target herbivore on the non-transgenic parent. 

Environ Pollut 157: 181-185. http://dx.doi.org/10.1016/j.envpol.2008.07.006

Hogg, A; Uddling, J; Ellsworth, D; Carroll, MA; Pressley, S; Lamb, B; Vogel, C. (2007). Stomatal and nonstomatal fluxes of ozone to a northern mixed hardwood forest. Tellus B Chem Phys Meteorol 59: 514-

525. http://dx.doi.org/10.1111/j.1600-0889.2007.00269.x

Hogsett, WE; Olszyk, D; Ormrod, DP; Taylor, GE, Jr; Tingey, DT. (1987a). Air pollution exposure systems 

and experimental protocols: Volume 2: Description of facilities. (EPA/600/3-87/037b). Corvallis, OR: 

U.S. Environmental Protection Agency, Environmental Research Laboratory. 

http://nepis.epa.gov/Exe/ZyPURL.cgi?Dockey=30000KQH.txt

Hogsett, WE; Olszyk, D; Ormrod, DP; Taylor, GE, Jr; Tingey, DT. (1987b). Air pollution exposure systems 

and experimental protocols: Volume I: A review and evaluation of performance. (EPA/600/3-87/037a). 

Corvallis, OR: U.S. Environmental Protection Agency. 

Hogsett, WE; Tingey, DT; Holman, SR. (1985). A programmable exposure control system for determination 

of the effects of pollutant exposure regimes on plant growth. Atmos Environ 19: 1135-1145. 

http://dx.doi.org/10.1016/0004-6981(85)90198-2

Hogsett, WE; Tingey, DT; Lee, EH. (1988). Ozone exposure indices: Concepts for development and 

evaluation of their use. In Assessment of crop loss from air pollutants: Proceedings of an international 

conference. New York: Elsevier Applied Science. 

Hogsett, WE; Tingey, DT; Lee, EH; Beedlow, PA; Andersen, CP. (2008). An approach for evaluating the 

effectiveness of various ozone Air Quality Standards for protecting trees. Environ Manage 41: 937-

948. http://dx.doi.org/10.1007/s00267-007-9057-3

Hogsett, WE; Weber, JE; Tingey, D; Herstrom, A; Lee, EH; Laurence, JA. (1997). Environmental auditing: 

An approach for characterizing tropospheric ozone risk to forests. J Environ Manage 21: 105-120. 

http://dx.doi.org/10.1007/s002679900010

Holmes, WE; Zak, DR; Pregitzer, KS; King, JS. (2006). Elevated CO2 and O3 alter soil nitrogen 

transformations beneath trembling aspen, paper birch, and sugar maple. Ecosystems 9: 1354-1363. 

http://dx.doi.org/10.1007/s10021-006-0163-5

Hong, B; Weinstein, D; Swaney, D. (2006). Assessment of ozone effects on nitrate export from Hubbard 

Brook Watershed 6. Environ Pollut 141: 8-21. http://dx.doi.org/10.1016/j.envpol.2005.08.030

Horvath, L; Nagy, Z; Weidinger, T; Artz, R; Luke, WT; Valigura, R; Pinto, JP; Womack, J. (1995). 

Measurement of fluxes of trace gases (O3, NOX, SO2, CO2, HNO3), particulate sulfate and nitrate, 

water vapour over short vegetation by gradient and eddy correlation techniques in Hungary. Ann 

Geophys 13: C490. 

Howard, AR; van Iersel, MW; Richards, JH; Donovan, LA. (2009). Night-time transpiration can decrease 

hydraulic redistribution. Plant Cell Environ 32: 1060-1070. http://dx.doi.org/10.1111/j.1365-

3040.2009.01988.x

Hui, D; Sims, DA; Johnson, DW; Chang, W; Luo, Y. (2002). Effects of gradual versus step increases in 

carbon dioxide on Plantago photosynthesis and growth in a microcosm study. Environ Exp Bot 47: 51-

66. 

ICP M&M (International Cooperative Programme on Modelling and Mapping). (2004). Mapping critical 

levels for vegetation. In Manual on methodologies and criteria for modelling and mapping critical 

loads and levels, and air pollution effects, risks and trends. 

http://www.rivm.nl/en/themasites/icpmm/manual-and-downloads/manual-english/index.html

Innes, JL; Skelly, JM; Schaub, M. (2001). Ozon, Laubholz- und Krautpflanzen: Ein Fuhrer zum Bestimmen 

von Ozonsymptomen. Bern, Switzerland: Paul Haupt Publishers. 

IPCC (Intergovernmental Panel on Climate Change). (2007a). Climate change 2007: Impacts, adaptation and 

vulnerability. Cambridge, UK: Cambridge University Press. 
Draft – Do Not Cite or Quote 9-171 June 2012

Iriti, M; Di Maro, A; Bernasconi, S; Burlini, N; Simonetti, P; Picchi, V; Panigada, C; Gerosa, G; Parente, A; 

Faoro, F. (2009). Nutritional traits of bean (Phaseolus vulgaris) seeds from plants chronically exposed 

to ozone pollution. J Agric Food Chem 57: 201-208. http://dx.doi.org/10.1021/jf802819m

Iriti, M; Faoro, F. (2009). Chemical diversity and defence metabolism: How plants cope with pathogens and 

ozone pollution. International Journal of Molecular Sciences 10: 3371-3399. 

http://dx.doi.org/10.3390/ijms10083371

IRRI (International Rice Research Institute). (2002). IRRI annual report 2001-2002. Los Baños, Laguna in 

the Philippines. http://www.scribd.com/collections/2641758/Annual-Reports

Isebrands, JG; Dickson, RE; Rebbeck, J; Karnosky, DF. (2000). Interacting effects of multiple stresses on 

growth and physiological processes in northern forest trees. In RA Mickler; RA Birsdey; J Hom 

(Eds.), Responses of northern US forests to environmental change (pp. 149-180). New York, NY: 

Springer-Verlag. 

Isebrands, JG; McDonald, EP; Kruger, E; Hendrey, G; Percy, K; Pregitzer, K; Sober, J; Karnosky, DF.

(2001). Growth responses of Populus tremuloides clones to interacting carbon dioxide and 

tropospheric ozone. Environ Pollut 115: 359-371. 

Jackson, DM; Heagle, AS; Eckel, RVW. (1999). Ovipositional response of tobacco hornworm moths 

(Lepidoptera: Sphyngidae) to tobacco plants grown under elevated levels of ozone. Environ Entomol 

28: 566-571. 

Janzik, I; Preiskowski, S; Kneifel, H. (2005). Ozone has dramatic effects on the regulation of the 

prechorismate pathway in tobacco (Nicotiana tabacum L. cv. Bel W3). Planta 223: 20-27. 

http://dx.doi.org/10.1007/s00425-005-0060-8

Johnson, RM; Pregitzer, KS. (2007). Concentration of sugars, phenolic acids, and amino acids in forest soils 

exposed to elevated atmospheric CO2 and O3. Soil Biol Biochem 39: 3159-3166. 

http://dx.doi.org/10.1016/j.soilbio.2007.07.010

Jones, ME; Paine, TD. (2006). Detecting changes in insect herbivore communities along a pollution gradient. 

Environ Pollut 143: 377-387. http://dx.doi.org/10.1016/j.envpol.2005.12.013

Jones, MLM; Hodges, G; Mills, G. (2010). Nitrogen mediates above-ground effects of ozone but not belowground effects in a rhizomatous sedge. Environ Pollut 158: 559-565. 

http://dx.doi.org/10.1016/j.envpol.2009.08.002

Jones, TG; Freeman, C; Lloyd, A; Mills, G. (2009). Impacts of elevated atmospheric ozone on peatland 

below-ground doc characteristics. Ecol Eng 35: 971-977. 

http://dx.doi.org/10.1016/j.ecoleng.2008.08.009

Joo, JH; Wang, SY; Chen, JG; Jones, AM; Fedoroff, NV. (2005). Different signaling and cell death roles of 

heterotrimeric G protein alpha and beta subunits in the arabidopsis oxidative stress response to ozone. 

Plant Cell 17: 957-970. http://dx.doi.org/10.1105/tpc.104.029603

Joss, U; Graber, WK. (1996). Profiles and simulated exchange of H2O, O3, NO2 between the atmosphere 

and the HartX Scots pine plantation. Theor Appl Climatol 53: 157-172. 

Jovan, S; McCune, B. (2006). Using epiphytic macrolichen communities for biomonitoring ammonia in 

forests of the greater Sierra Nevada, California. Water Air Soil Pollut 170: 69-93. 

Kanerva, T; Palojarvi, A; Ramo, K; Manninen, S. (2008). Changes in soil microbial community structure 

under elevated tropospheric O3 and CO2. Soil Biol Biochem 40: 2502-2510. 

http://dx.doi.org/10.1016/j.soilbio.2008.06.007

Kanerva, T; Palojarvi, A; Ramo, K; Ojanpera, K; Esala, M; Manninen, S. (2006). A 3-year exposure to CO2 

and O3 induced minor changes in soil N cycling in a meadow ecosystem. Plant Soil 286: 61-73. 

http://dx.doi.org/10.1007/s11104-006-9026-2

Kanerva, T; Regina, K; Ramo, K; Ojanpera, K; Manninen, S. (2007). Fluxes of N2O, CH4 and CO2 in a 

meadow ecosystem exposed to elevated ozone and carbon dioxide for three years. Environ Pollut 145: 

818-828. http://dx.doi.org/10.1016/j.envpol.2006.03.055

Kangasjarvi, J; Jaspers, P; Kollist, H. (2005). Signalling and cell death in ozone-exposed plants. Plant Cell 

Environ 28: 1021-1036. 

Karlsson, PE; Sellden, G; Plaijel, H. (2003). Establishing ozone critical levels II: UNECE workshop report. 

Gothenburg, Sweden: IVL Swedish Environmental Institute. 

Karlsson, PE; Uddling, J; Braun, S; Broadmeadow, M; Elvira, S; Gimeno, BS; Le Thiec, D; Okansen, E; 

Vandermeiren, K; Wilkinson, M; Emberson, L. (2004). New critical levels for ozone effects on young 

trees based on AOT40 and simulated cumulative leaf uptake of ozone. Atmos Environ 38: 2283-2294. 
Draft – Do Not Cite or Quote 9-172 June 2012

Karnosky, DF; Gagnon, ZE; Dickson, RE; Coleman, MD; Lee, EH; Isebrands, JG. (1996). Changes in 

growth, leaf abscission, biomass associated with seasonal tropospheric ozone exposures of Populus 

tremuloides clones and seedlings. Can J For Res 26: 23-37. 

Karnosky, DF; Mankovska, B; Percy, K; Dickson, RE; Podila, GK; Sober, J; Noormets, A; Hendrey, G; 

Coleman, MD; Kubiske, M; Pregitzer, KS; Isebrands, JG. (1999). Effects of tropospheric ozone on 

trembling aspen and interaction with CO2: Results from an O3-gradient and a FACE experiment. 

Water Air Soil Pollut 116: 311-322. 

Karnosky, DF; Pregitzer, KS; Zak, DR; Kubiske, ME; Hendrey, GR; Weinstein, D; Nosal, M; Percy, KE.

(2005). Scaling ozone responses of forest trees to the ecosystem level in a changing climate. Plant Cell 

Environ 28: 965-981. http://dx.doi.org/10.1111/j.1365-3040.2005.01362.x

Karnosky, DF; Zak, DR; Pregitzer, KS; Awmack, CS; Bockheim, JG; Dickson, RE; Hendrey, GR; Host, GE; 

King, JS; Kopper, BJ; Kruger, EL; Kubiske, ME; Lindroth, RL; Mattson, WJ; McDonald, EP; 

Noormets, A; Oksanen, E; Parsons, WFJ; Percy, KE; Podila, GK; Riemenschneider, DE; Sharma, P; 

Thakur, R; Sober, A; Sober, J; Jones, WS; Anttonen, S; Vapaavuori, E; Mankovska, B; Heilman, W; 

Isebrands, JG. (2003). Tropospheric O3 moderates responses of temperate hardwood forests to 

elevated CO2: A synthesis of molecular to ecosystem results from the Aspen FACE project. Funct 

Ecol 17: 289-304. 

Kasurinen, A; Keinanen, MM; Kaipainen, S; Nilsson, LO; Vapaavuori, E; Kontro, MH; Holopainen, T.

(2005). Below-ground responses of silver birch trees exposed to elevated CO2 and O3 levels during 

three growing seasons. Global Change Biol 11: 1167-1179. http://dx.doi.org/10.1111/j.1365-

2486.2005.00970.x

Kasurinen, A; Peltonen, PA; Julkunen-Tiitto, R; Vapaavuori, E; Nuutinen, V; Holopainen, T; Holopainen, JK.

(2007). Effects of elevated CO2 and O3 on leaf litter phenolics and subsequent performance of litterfeeding soil macrofauna. Plant Soil 292: 25-43. http://dx.doi.org/10.1007/s11104-007-9199-3

Kasurinen, A; Riikonen, J; Oksanen, E; Vapaavuori, E; Holopainen, T. (2006). Chemical composition and 

decomposition of silver birch leaf litter produced under elevated CO2 and O3. Plant Soil 282: 261-

280. http://dx.doi.org/10.1007/s11104-005-6026-6

Kats, G; Olszyk, DM; Thompson, CR. (1985). Open top experimental chambers for trees. J Air Waste Manag 

Assoc 35: 1298-1301. 

Kats, G; Thompson, CR; Kuby, WC. (1976). Improved ventilation of open top greenhouses. J Air Pollut 

Control Assoc 26: 1089-1090. 

Keller, T; Häsler, R. (1984). The influence of a fall fumigation with ozone on the stomatal behavior of spruce 

and fir. Oecologia 64: 284-286. http://dx.doi.org/10.1007/BF00376884

Kellomaki, S; Wang, KY. (1997). Effects of elevated O3 and CO2 concentrations on photosynthesis and 

stomatal conductance in Scots pine. Plant Cell Environ 20: 995-1006. http://dx.doi.org/10.1111/j.1365-

3040.1997.tb00676.x

Kerner, R; Winkler, J; Dupuy, J; Jürgensen, M; Lindermayr, C; Ernst, D; Müller-starck, G. (2011). Changes 

in the proteome of juvenile European beech following three years exposure to free-air elevated ozone. 

iForest 4: 69-76. http://dx.doi.org/10.3832/ifor0570-004

Keutgen, AJ; Noga, G; Pawelzik, E. (2005). Cultivar-specific impairment of strawberry growth, 

photosynthesis, carbohydrate and nitrogen accumulation by ozone. Environ Exp Bot 53: 271-280. 

http://dx.doi.org/10.1016/j.envexpbot.2004.04.003

Keutgen, N; Keutgen, AJ; Janssens, MJJ. (2008). Sweet potato [Ipomoea batatas (L.) Lam.] cultivated as 

tuber or leafy vegetable supplier as affected by elevated tropospheric ozone. J Agric Food Chem 56: 

6686-6690. http://dx.doi.org/10.1021/jf8006272

King, JS; Kubiske, ME; Pregitzer, KS; Hendrey, GR; McDonald, EP; Giardina, CP; Quinn, VS; Karnosky, 

DF. (2005). Tropospheric O3 compromises net primary production in young stands of trembling

aspen, paper birch and sugar maple in response to elevated atmospheric CO2. New Phytol 168: 623-

635. http://dx.doi.org/10.1111/j.1469-8137.2005.01557.x

King, JS; Pregitzer, KS; Zak, DR; Sober, J; Isebrands, JG; Dickson, RE; Hendrey, GR; Karnosky, DF. (2001). 

Fine-root biomass and fluxes of soil carbon in young stands of paper birch and trembling aspen as 

affected by elevated atmospheric CO2 and tropospheric O3. Oecologia 128: 237-250. 

Kitao, M; Low, M; Heerdt, C; Grams, TEE; Haberle, KH; Matyssek, R. (2009). Effects of chronic elevated 

ozone exposure on gas exchange responses of adult beech trees (Fagus sylvatica) as related to the 

within-canopy light gradient. Environ Pollut 157: 537-544. 

http://dx.doi.org/10.1016/j.envpol.2008.09.016
Draft – Do Not Cite or Quote 9-173 June 2012

Kline, LJ; Davis, DD; Skelly, JM; Decoteau, DR. (2009). Variation in ozone sensitivity within Indian hemp 

and common milkweed selections from the Midwest. Northeast Nat 16: 307-313. 

http://dx.doi.org/10.1656/045.016.0210

Kline, LJ; Davis, DD; Skelly, JM; Savage, JE; Ferdinand, J. (2008). Ozone sensitivity of 28 plant selections 

exposed to ozone under controlled conditions. Northeast Nat 15: 57-66. 

http://dx.doi.org/10.1656/1092-6194(2008)15[57:OSOPSE]2.0.CO;2

Kohut, R. (2007). Assessing the risk of foliar injury from ozone on vegetation in parks in the US National 

Park Service's Vital Signs Network. Environ Pollut 149: 348-357. 

http://dx.doi.org/10.1016/j.envpol.2007.04.022

Kolb, TE; Fredericksen, TS; Steiner, KC; Skelly, JM. (1997). Issues in scaling tree size and age responses to 

ozone: A review [Review]. Environ Pollut 98: 195-208. http://dx.doi.org/10.1016/S0269-

7491(97)00132-2

Kollist, T; Moldau, H; Rasulov, B; Oja, V; Ramma, H; Huve, K; Jaspers, P; Kangasjarvi, J; Kollist, H.

(2007). A novel device detects a rapid ozone-induced transient stomatal closure in intact Arabidopsis 

and its absence in abi2 mutant. Physiol Plant 129: 796-803. http://dx.doi.org/10.1111/j.1399-

3054.2006.00851.x

Kostka-Rick, R; Hahn, HU. (2005). Biomonitoring using tobacco Bel W3 provides supplemental information 

for risk assessment of vegetation injury due to ozone. Gefahrstoffe Reinhaltung Der Luft 65: 485-491. 

Kozovits, AR; Matyssek, R; Blaschke, H; Gottlein, A; Grams, TEE. (2005). Competition increasingly 

dominates the responsiveness of juvenile beech and spruce to elevated CO2 and/or O3 concentrations 

throughout two subsequent growing seasons. Global Change Biol 11: 1387-1401. 

http://dx.doi.org/10.1111/j.1365-2486.2005.00993.x

Krupa, SV; Grunhage, L; Jager, HJ; Nosal, M; Manning, WJ; Legge, AH; Hanewald, K. (1995). Ambient 

ozone (O3) and adverse crop response: A unified view of cause and effect. Environ Pollut 87: 119-

126. http://dx.doi.org/10.1016/S0269-7491(99)80014-1

Krupa, SV; Nosal, M; Peterson, DL. (2001). Use of passive ozone O3 samplers in vegetation effects 

assessment. Environ Pollut 112: 303-309. 

Kubiske, ME; Quinn, VS; Heilman, WE; McDonald, EP; Marquardt, PE; Teclaw, RM; Friend, AL; 

Karnoskey, DF. (2006). Interannual climatic variation mediates elevated CO2 and O3 effects on forest 

growth. Global Change Biol 12: 1054-1068. http://dx.doi.org/10.1111/j.1365-2486.2006.01152.x

Kubiske, ME; Quinn, VS; Marquardt, PE; Karnosky, DF. (2007). Effects of elevated atmospheric CO2 and/or 

O3 on intra- and interspecific competitive ability of aspen. Plant Biol (Stuttg) 9: 342-355. 

http://dx.doi.org/10.1055/s-2006-924760

Laffray, X; Rose, C; Garrec, JP. (2007). Estimation of ozone concentration in a valley of the alps mountains 

based on bel-w3 tobacco leaf injury. Water Air Soil Pollut 186: 29-42. 

http://dx.doi.org/10.1007/s11270-007-9460-7

Lambers, H; Chapin, FS, III; Pons, TL. (1998). Plant physiological ecology. In H Lambers; FS Chapin, III; 

TL Pons (Eds.). New York: Springer. 

Langebartels, C; Kerner, K; Leonardi, S; Schraudner, M; Trost, M; Heller, W; Sandermann, H, Jr. (1991). 

Biochemical plant responses to ozone: I. Differential induction of polyamine and ethylene 

biosynthesis in tobacco. J Plant Physiol 95: 882-889. http://dx.doi.org/10.1104/pp.95.3.882

Larson, JL; Zak, DR; Sinsabaugh, RL. (2002). Extracellular enzyme activity beneath temperate trees growing 

under elevated carbon dioxide and ozone. Soil Sci Soc Am J 66: 1848-1856. 

Lawlor, DW. (1998). Plant responses to global change: Temperature and drought stress. In LJ De Kok; I 

Stulen (Eds.), Responses of plant metabolism to air pollution and global change. Leiden, The 

Netherlands: Backhuys Publishers. 

Leakey, ADB; Bernacchi, CJ; Ort, DR; Long, SP. (2006). Long-term growth of soybean at elevated CO2 does 

not cause acclimation of stomatal conductance under fully open-air conditions. Plant Cell Environ 29: 

1794-1800. http://dx.doi.org/10.1111/j.1365-3040.2006.01556.x

Lee, EH; Hogsett, WE. (1996). Methodology for calculating inputs for ozone secondary standard benefits 

anaylsis: Part II. Research Triangle Park, NC: U.S. Environmental Protection Agency. 

Lee, EH; Hogsett, WE. (1999). Role of concentrations and time of day in developing ozone exposure indices 

for a secondary standard. J Air Waste Manag Assoc 49: 669-681. 

Lee, EH; Hogsett, WE; Tingey, DT. (1994). Attainment and effects issues regarding alternative secondary 

ozone air quality standards. J Environ Qual 23: 1129-1140. 

http://dx.doi.org/10.2134/jeq1994.00472425002300060002x
Draft – Do Not Cite or Quote 9-174 June 2012

Lee, EH; Tingey, DT; Hogsett, WE. (1987). Selection of the best exposure-response model using various 7-

hour ozone exposure statistics. Research Triangle Park, NC: U.S. Environmental Protection Agency. 

Lee, EH; Tingey, DT; Hogsett, WE. (1988a). Evaluation of ozone-exposure indices for relating exposure to 

plant production and for estimating agricultural losses. (EPA/600/3-88/039). Washington, DC: U.S. 

Environmental Protection Agency. 

Lee, EH; Tingey, DT; Hogsett, WE. (1988b). Evaluation of ozone exposure indices in exposure-response 

modeling. Environ Pollut 53: 43-62. http://dx.doi.org/10.1016/0269-7491(88)90024-3

Lee, EH; Tingey, DT; Hogsett, WE. (1989). Interrelation of experimental exposure and ambient air quality 

data for comparison of ozone exposure indices and estimating agricultural losses. (EPA/600/3-89/047). 

Corvallis, OR: U.S. Environmental Protection Agency. 

Lee, EH; Tingey, DT; Hogsett, WE; Laurence, JA. (2003a). History of tropospheric ozone for the San 

Bernardino Mountains of southern California, 1963-1999. Atmos Environ 37: 2705-2717. 

http://dx.doi.org/10.1016/S1352-2310(03)00203-6

Lee, EH; Tingey, DT; Waschmann, RS; Phillips, DL; Olszyk, DM; Johnson, MG; Hogsett, WE. (2009a). 

Seasonal and long-term effects of CO2 and O3 on water loss in ponderosa pine and their interaction 

with climate and soil moisture. Tree Physiol 29: 1381-1393. http://dx.doi.org/10.1093/treephys/tpp071

Lee, S; Yun, SC. (2006). The ozone stress transcriptome of pepper (Capsicum annuum L.). Molecules and 

Cells 21: 197-205. 

Lee, WS; Chevone, BI; Seiler, JR. (1990). Growth and gas exchange of loblolly pine seedlings as influenced 

by drought and air pollutants. Water Air Soil Pollut 51: 105-116. 

http://dx.doi.org/10.1007/BF00211508

Lefohn, AS; Benedict, HM. (1982). Development of a mathematical index that describes ozone 

concentration, frequency and duration. Atmos Environ 16: 2529-2532. http://dx.doi.org/10.1016/0004-

6981(82)90145-7

Lefohn, AS; Jackson, W; Shadwick, DS; Knudsen, HP. (1997). Effect of surface ozone exposures on 

vegetation grown in the southern Appalachian Mountains: Identification of possible areas of concern. 

Atmos Environ 31: 1695-1708. http://dx.doi.org/10.1016/S1352-2310(96)00258-0

Lefohn, AS; Laurence, JA; Kohut, RJ. (1988). A comparison of indices that describe the relationship between 

exposure to ozone and reduction in the yield of agricultural crops. Atmos Environ 22: 1229-1240. 

http://dx.doi.org/10.1016/0004-6981(88)90353-8

Lefohn, AS; Runeckles, VC. (1987). Establishing standards to protect vegetation - ozone exposure/dose 

considerations. Atmos Environ 21: 561-568. http://dx.doi.org/10.1016/0004-6981(87)90038-2

Lefohn, AS; Shadwick, DS. (2000). Differences in trending estimates in the United States using several 

ozone metrics. In Proceedings of the 93rd Air & Waste Management Association Annual Conference 

and Exhibition. Pittsburgh, PA: Air & Waste Management Association. 

Legge, AH; Grunhage, L; Nosal, M; Jager, HJ; Krupa, SV. (1995). Ambient ozone and adverse crop 

response: An evaluation of North American and European data as they relate to exposure indices and 

critical levels. J Appl Bot Food Qual 69: 192-205. 

Leitao, L; Bethenod, O; Biolley, JP. (2007a). The impact of ozone on juvenile maize (Zea mays L.) plant 

photosynthesis: Effects on vegetative biomass, pigmentation, and carboxylases (PEPc and Rubisco). 

Plant Biol (Stuttg) 9: 478-488. http://dx.doi.org/10.1055/s-2007-964942

Leitao, L; Delacote, E; Dizengremel, P; Le Thiec, D; Biolley, JP. (2007b). Assessment of the impact of 

increasing concentrations of ozone on photosynthetic components of maize (Zea mays L.), a C-4 

plant. Environ Pollut 146: 5-8. http://dx.doi.org/10.1016/j.envpol.2006.05.019

Leitao, L; Maoret, JJ; Biolley, JP. (2007c). Changes in PEP carboxylase, rubisco and rubisco activase mRNA 

levels from maize (Zea mays) exposed to a chronic ozone stress. Biol Res 40: 137-153. 

http://dx.doi.org/10.4067/S0716-97602007000200005

Lesser, VM; Rawlings, JO; Spruill, SE; Somerville, MC. (1990). Ozone effects on agricultural crops: 

Statistical methodologies and estimated dose-response relationships. Crop Sci 30: 148-155. 

Leuning, R; Unsworth, MH; Neumann, HN; King, KM. (1979). Ozone fluxes to tobacco and soil under field 

conditions. Atmos Environ 13: 1155-1163. http://dx.doi.org/10.1016/0004-6981(79)90039-8

Levine, JS; Pinto, JP. (1998). The production of CO by biomass burning. In MAK Khalil; JP Pinto; MJ 

Shearer (Eds.), Atmospheric carbon monoxide and its environmental effects: Proceedings of the 

international conference; December 1997; Portland, Oregon (pp. 251-256). Research Triangle Park, 

NC: U.S. Environmental Protection Agency, Office of Research and Development. 
Draft – Do Not Cite or Quote 9-175 June 2012

Lewis, JS; Ditchkoff, SS; Lin, JC; Muntifering, RB; Chappelka, AH. (2006). Nutritive quality of big 

bluestem (Andropogon gerardii) and eastern gamagrass (Tripsacum dactyloides) exposed to 

tropospheric ozone. Rangeland Ecol Manag 59: 267-274. 

Li, PH; Mane, SP; Sioson, AA; Robinet, CV; Heath, LS; Bohnert, HJ; Grene, R. (2006b). Effects of chronic 

ozone exposure on gene expression in Arabidopsis thaliana ecotypes and in Thellungielia halophila. 

Plant Cell Environ 29: 854-868. http://dx.doi.org/10.1111/j.1365-3040.2005.01465.x

Lin, JC; Nosal, M; Muntifering, RB; Krupa, SV. (2007). Alfalfa nutritive quality for ruminant livestock as 

influenced by ambient air quality in west-central Alberta. Environ Pollut 149: 99-103. 

http://dx.doi.org/10.1016/j.envpol.2006.12.009

Lindroth, RL. (2010). Impacts of elevated atmospheric CO2 and O3 on forests: Phytochemistry, trophic 

interactions, and ecosystem dynamics. J Chem Ecol 36: 21-Feb. http://dx.doi.org/10.1007/s10886-

009-9731-4

Liu, L; King, J; Giardina, C. (2005). Effects of elevated concentrations of atmospheric CO2 and tropospheric 

O3 on leaf litter production and chemistry in trembling aspen and paper birch communities. Tree 

Physiol 25: 1511-1522. 

Liu, L; King, JS; Giardina, CP. (2007a). Effects of elevated atmospheric CO2 and tropospheric O3 on 

nutrient dynamics: Decomposition of leaf litter in termbling aspen and paper birch communities. Plant 

Soil 299: 65-82. 

Liu, LL; King, JS; Giardina, CP; Booker, FL. (2009b). The influence of chemistry, production and 

community composition on leaf litter decomposition under elevated atmospheric CO2 and 

tropospheric O-3 in a northern hardwood ecosystem. Ecosystems 12: 401-416. 

http://dx.doi.org/10.1007/s10021-009-9231-y

Loats, KV; Rebbeck, J. (1999). Interactive effects of ozone and elevated carbon dioxide on the growth and 

physiology of black cherry, green ash, and yellow poplar seedlings. Environ Pollut 106: 237-248. 

http://dx.doi.org/10.1016/S0269-7491(99)00069-X

Long, SP. (1991). Modification of the response of photosynthetic productivity to rising temperature by 

atmospheric CO2 concentrations: Has its importance been underestimated? Plant Cell Environ 14: 

729-739. http://dx.doi.org/10.1111/j.1365-3040.1991.tb01439.x

Loranger, GI; Pregitzer, KS; King, JS. (2004). Elevated CO2 and O3t concentrations differentially affect 

selected groups of the fauna in temperate forest soils. Soil Biol Biochem 36: 1521-1524. 

Lorenzini, G; Nali, C. (1995). Analysis of vertical ozone and nitrogen oxides profiles in a Prunus cerasifera 

canopy. Int J Biometeorol 39: 1-4. http://dx.doi.org/10.1007/BF01320885

Loreto, F; Fares, S. (2007). Is ozone flux inside leaves only a damage indicator? Clues from volatile 

isoprenoid studies. Plant Physiol 143: 1096-1100. http://dx.doi.org/10.1104/pp.106.091892

Loreto, F; Velikova, V. (2001). Isoprene produced by leaves protects the photosynthetic apparatus against 

ozone damage, quenches ozone products, and reduces lipid peroxidation of cellular membranes. Plant 

Physiol 127: 1781-1787. http://dx.doi.org/10.1104/pp.010497

Low, M; Herbinger, K; Nunn, AJ; Haberle, KH; Leuchner, M; Heerdt, C; Werner, H; Wipfler, P; Pretzsch, H; 

Tausz, M; Matyssek, R. (2006). Extraordinary drought of 2003 overrules ozone impact on adult beech 

trees (Fagus sylvatica). Trees Struct Funct 20: 539-548. http://dx.doi.org/10.1007/s00468-006-0069-z

Loya, WM; Pregitzer, KS; Karberg, NJ; King, JS; Giardina, CP. (2003). Reduction of soil carbon formation 

by tropospheric ozone under elevated carbon dioxide. Nature 425: 705-707. 

Ludwikow, A; Gallois, P; Sadowski, J. (2004). Ozone-induced oxidative stress response in Arabidopsis: 

Transcription profiling by microarray approach. Cell Mol Biol Lett 9: 829-842. 

Ludwikow, A; Kierzek, D; Gallois, P; Zeef, L; Sadowski, J. (2009). Gene expression profiling of ozonetreated Arabidopsis abi1td insertional mutant: Protein phosphatase 2C ABI1 modulates biosynthesis 

ratio of ABA and ethylene. Planta 230: 1003-1017. http://dx.doi.org/10.1007/s00425-009-1001-8

Ludwikow, A; Sadowski, J. (2008). Gene networks in plant ozone stress response and tolerance. J Integr 

Plant Biol 50: 1256-1267. http://dx.doi.org/10.1111/j.1744-7909.2008.00738.x

Luo, Y. (2001). Transient ecosystem response to free-air CO2 enrichment (FACE): Experimental evidence 

and methods of analysis. New Phytol 152: 3-8. 

Luo, Y; Reynolds, JF. (1999). Validity of extrapolating field CO2 experiments to predict carbon sequestration 

in natural ecosystems. Ecology 80: 1568-1583. 

Lyons, TM; Barnes, JD. (1998). Influence of plant age on ozone resistance in Plantago major. New Phytol 

138: 83-89. http://dx.doi.org/10.1046/j.1469-8137.1998.00879.x
Draft – Do Not Cite or Quote 9-176 June 2012

Maggio, A; Chiaranda, FQ; Cefariello, R; Fagnano, M. (2009). Responses to ozone pollution of alfalfa 

exposed to increasing salinity levels. Environ Pollut 157: 1445-1452. 

http://dx.doi.org/10.1016/j.envpol.2008.09.013

Mahalingam, R; Jambunathan, N; Gunjan, SK; Faustin, E; Weng, H; Ayoubi, P. (2006). Analysis of oxidative 

signalling induced by ozone in Arabidopsis thaliana. Plant Cell Environ 29: 1357-1371. 

http://dx.doi.org/10.1111/j.1365-3040.2006.01516.x

Mahalingam, R; Shah, N; Scrymgeour, A; Fedoroff, N. (2005). Temporal evolution of the Arabidopsis 

oxidative stress response. Plant Mol Biol 57: 709-730. http://dx.doi.org/10.1007/s11103-005-2860-4

Maier-Maercker, U. (1998). Predisposition of trees to drought stress by ozone. Tree Physiol 19: 71-78. 

Mandl, RH; Laurence, JA; Kohut, RJ. (1989). Development and testing of open-top chambers for exposing 

large, perennial plants to air pollutants. J Environ Qual 18: 534-540. 

http://dx.doi.org/10.2134/jeq1989.00472425001800040026x

Mandl, RH; Weinstein, LH; McCune, DC; Keveny, M. (1973). A cylindrical, open-top chamber for the 

exposure of plants to air pollutants in the field. J Environ Qual 2: 371-376. 

Maňkovská, B; Percy, KE; Karnosky, DF. (2005). Impacts of greenhouse gases on epicuticular waxes of 

Populus tremuloides Michx.: Results from an open-air exposure and a natural O3 gradient. Environ 

Pollut 137: 580-586. http://dx.doi.org/10.1016/j.envpol.2005.01.043

Manning, WJ. (2003). Detecting plant effects is necessary to give biological significance to ambient ozone 

monitoring data and predictive ozone standards. Environ Pollut 126: 375-379. 

Manning, WJ; Krupa, SV. (1992). Experimental methodology for studying the effects of ozone on crops and 

trees. In AS Lefohn (Ed.), Surface level ozone exposures and their effects on vegetation (pp. 93-156). 

Chelsea, MI: Lewis Publishers. 

Martin, MJ; Host, GE; Lenz, KE; Isebrands, JG. (2001). Simulating the growth response of aspen to elevated 

ozone: A mechanistic approach to scaling a leaf-level model of ozone effects on photosynthesis to a 

complex canopy architecture. Environ Pollut 115: 425-436. 

Massman, WJ. (2004). Toward an ozone standard to protect vegetation based on effective dose: A review of 

deposition resistances and a possible metric [Review]. Atmos Environ 38: 2323-2337. 

Massman, WJ; Grantz, DA. (1995). Estimating canopy conductance to ozone uptake from observations of 

evapotranspiration at the canopy scale and at the leaf scale. Global Change Biol 1: 183-198. 

http://dx.doi.org/10.1111/j.1365-2486.1995.tb00020.x

Massman, WJ; Musselman, RC; Lefohn, AS. (2000). A conceptual ozone dose-response model to develop a 

standard to protect vegetation. Atmos Environ 34: 745-759. http://dx.doi.org/10.1016/S1352-

2310(99)00395-7

Matyssek, R; Gunthardt-Goerg, MS; Maurer, S; Keller, T. (1995). Nighttime exposure to ozone reduces 

whole-plant production in Betula pendula. Tree Physiol 15: 159-165. 

Matyssek, R; Le Thiec, D; Low, M; Dizengremel, P; Nunn, AJ; Haberle, KH. (2006). Interactions between 

drought and O3 stress in forest trees. Plant Biol (Stuttg) 8: 11-17. http://dx.doi.org/10.1055/s-2005-

873025

Matyssek, R; Sandermann, H; Wieser, G; Booker, F; Cieslik, S; Musselman, R; Ernst, D. (2008). The 

challenge of making ozone risk assessment for forest trees more mechanistic. Environ Pollut 156: 567-

582. http://dx.doi.org/10.1016/j.envpol.2008.04.017

Matyssek, R; Wieser, G; Ceulemans, R; Rennenberg, H; Pretzsch, H; Haberer, K; Low, M; Nunn, AJ; Werner, 

H; Wipfler, P; Obwald, W; Nikolova, P; Hanke, DE; Kraigher, H; Tausz, M; Bahnweg, G; Kitao, M; 

Dieler, J; Sandermann, H; Herbinger, K; Grebenc, T; Blumenrother, M; Deckmyn, G; Grams, TEE; 

Heerdt, C; Leuchner, M; Fabian, P; Haberle, KH. (2010). Enhanced ozone strongly reduces carbon 

sink strength of adult beech (Fagus sylvatica): Resume from the free-air fumigation study at 

Kranzberg Forest. Environ Pollut 158: 2527-2532. http://dx.doi.org/10.1016/j.envpol.2010.05.009

Mautz, WJ; Dohm, MR. (2004). Respiratory and behavioral effects of ozone on a lizard and a frog. Comp 

Biochem Physiol A Mol Integr Physiol 139: 371-377. 

McAinsh, MR; Evans, NH; Montgomery, LT; North, KA. (2002). Calcium signalling in stomatal responses 

to pollutants. New Phytol 153: 441-447. 

McBride, JR; Laven, RD. (1999). Impact of oxidant air pollutants on forest succession in the mixed conifer 

forests of the San Bernardino Mountains. In PR Miller; JR McBride (Eds.), Oxidant air pollution 

impacts in the montane forests of southern California: A case study of the San Bernardino Mountains 

(pp. 338-352). New York, NY: Springer-Verlag. 
Draft – Do Not Cite or Quote 9-177 June 2012

McCarthy, HR; Oren, R; Johnsen, KH; Gallet-Budynek, A; Pritchard, SG; Cook, CW; LaDeau, SL; Jackson, 

RB; Finzi, AC. (2009). Re-assessment of plant carbon dynamics at the Duke free-air CO2 enrichment 

site: Interactions of atmospheric [CO2] with nitrogen and water availability over stand development. 

New Phytol 185: 514-528. http://dx.doi.org/10.1111/j.1469-8137.2009.03078.x

McCool, PM; Musselman, RC; Younglove, T; Teso, RR. (1988). Response of kidney bean to sequential

ozone exposures. Environ Exp Bot 28: 307-313. 

McFrederick, QS; Fuentes, JD; Roulston, T; Kathilankal, JC; Lerdau, M. (2009). Effects of air pollution on 

biogenic volatiles and ecological interactions. Oecologia 160: 411-420. 

http://dx.doi.org/10.1007/s00442-009-1318-9

McFrederick, QS; Kathilankal, JC; Fuentes, JD. (2008). Air pollution modifies floral scent trails. Atmos 

Environ 42: 2336-2348. http://dx.doi.org/10.1016/j.atmosenv.2007.12.033

McLaughlin, SB; Nosal, M; Wullschleger, SD; Sun, G. (2007a). Interactive effects of ozone and climate on 

tree growth and water use in a southern Appalachian forest in the USA. New Phytol 174: 109-124. 

http://dx.doi.org/10.1111/j.1469-8137.2007.02018.x

McLaughlin, SB; Wullschleger, SD; Sun, G; Nosal, M. (2007b). Interactive effects of ozone and climate on 

water use, soil moisture content and streamflow in a southern Appalachian forest in the USA. New 

Phytol 174: 125-136. http://dx.doi.org/10.1111/j.1469-8137.2007.01970.x

McLeod, AR; Long, SP. (1999). Free-air carbon dioxide enrichment (FACE) in global change research: A 

review. Adv Ecol Res 28: 1-56. http://dx.doi.org/10.1016/S0065-2504(08)60028-8

Medlyn, BE; Barton, CVM; Broadmeadow, MSJ; Ceulemans, R; De Angelis, P; Forstreuter, M; Freeman, M; 

Jackson, SB; Kellomaki, S; Laitat, E; Rey, A; Roberntz, P; Sigurdsson, BD; Strassemeyer, J; Wang, K; 

Curtis, PS; Jarvis, PG. (2001). Stomatal conductance of forest species after long-term exposure to 

elevated CO2 concentration: A synthesis. New Phytol 149: 247-264. http://dx.doi.org/10.1046/j.1469-

8137.2001.00028.x

Meehan, TD; Crossley, MS; Lindroth, RL. (2010). Impacts of elevated CO2 and O3 on aspen leaf litter 

chemistry and earthworm and springtail productivity. Soil Biol Biochem 42: 1132-1137. 

http://dx.doi.org/10.1016/j.soilbio.2010.03.019

Menendez, AI; Romero, AM; Folcia, AM; Martinez-Ghersa, MA. (2009). Getting the interactions right: Will 

higher O3 levels interfere with induced defenses to aphid feeding? Basic Appl Ecol 10: 255-264. 

http://dx.doi.org/10.1016/j.baae.2008.03.010

Menendez, AI; Romero, AM; Folcia, AM; Martinez-Ghersa, MA. (2010). Aphid and episodic O3 injury in 

arugula plants (Eruca sativa Mill) grown in open-top field chambers. Agric Ecosyst Environ 135: 10-

14. http://dx.doi.org/10.1016/j.agee.2009.08.005

Mereu, S; Gerosa, G; Finco, A; Fusaro, L; Muys, B; Manes, F. (2009). Improved sapflow methodology 

reveals considerable night-time ozone uptake by Mediterranean species. Biogeosciences 6: 3151-

3162. http://dx.doi.org/10.5194/bg-6-3151-2009

Miles, GP; Samuel, MA; Zhang, YL; Ellis, BE. (2005). RNA interference-based (RNAi) suppression of 

AtMPK6, an Arabidopsis mitogen-activated protein kinase, results in hypersensitivity to ozone and 

misregulation of AtMPK3. Environ Pollut 138: 230-237. 

http://dx.doi.org/10.1016/j.envpol.2005.04.017

Miller, PL. (1973). Oxidant-induced community change in a mixed conifer forest. In JA Naegele (Ed.), Air 

pollution damage to vegetation (pp. 101-117). Washington, DC: American Chemical Society. 

Miller, PR; McCutchan, MH; Ryan, BC. (1972). Influence of climate and topography on oxidant air pollution 

concentrations that damage conifer forests in southern California. Mitt Forstl Bundesversuchsanst 

Wien 97: 585-607. 

Miller, PR; Parmeter, JR, Jr; Taylor, OC; Cardiff, EA. (1963). Ozone injury to the foliage of Pinus ponderosa. 

Phytopathology 53: 1072-1076. 

Miller, PR; Rechel, J. (1999). Temporal changes in crown condition indices, needle litterfall, and collateral 

needle injuries of Ponderosa and Jeffrey pines. In PR Miller; JR McBride (Eds.), Oxidant air pollution 

impacts in the montane forests of southern California: A case study of the San Bernardino Mountains 

(pp. 164-178). New York, NY: Springer. 

Mills, G. (2002). Modification of plant response by environmental conditions. In JNB Bell; M Treshow 

(Eds.), Air pollution and plant life (2nd ed., pp. 343-358). Chichester, United Kingdom: John Wiley & 

Sons. 
Draft – Do Not Cite or Quote 9-178 June 2012

Mills, G; Ball, G; Hayes, F; Fuhrer, J; Skarby, L; Gimeno, B; De Temmerman, L. (2000). Development of a 

multi-factor model for predicting the effects of ambient ozone on the biomass of white clover. Environ 

Pollut 109: 533-542. http://dx.doi.org/10.1016/S0269-7491(00)00057-9

Mills, G; Buse, A; Gimeno, B; Bermejo, V; Holland, M; Emberson, L; Pleijel, H. (2007a). A synthesis of 

AOT40-based response functions and critical levels of ozone for agricultural and horticultural crops. 

Atmos Environ 41: 2630-2643. http://dx.doi.org/10.1016/j.atmosenv.2006.11.016

Mills, G; Hayes, F; Jones, MLM; Cinderby, S. (2007b). Identifying ozone-sensitive communities of (semi-

)natural vegetation suitable for mapping exceedance of critical levels. Environ Pollut 146: 736-743. 

http://dx.doi.org/10.1016/j.envpol.2006.04.005

Mills, G; Hayes, F; Wilkinson, S; Davies, WJ. (2009). Chronic exposure to increasing background ozone 

impairs stomatal functioning in grassland species. Global Change Biol 15: 1522-1533. 

http://dx.doi.org/10.1111/j.1365-2486.2008.01798.x

Mondor, EB; Awmack, CS; Lindroth, RL. (2010). Individual growth rates do not predict aphid population 

densities under altered atmospheric conditions. Agr Forest Entomol 12: 293-299. 

http://dx.doi.org/10.1111/j.1461-9563.2010.00478.x

Mondor, EB; Tremblay, MN; Awmack, CS; Lindroth, RL. (2004). Divergent pheromone-mediated insect 

behaviour under global atmospheric change. Global Change Biol 10: 1820-1824. 

Mondor, EB; Tremblay, MN; Awmack, CS; Lindroth, RL. (2005). Altered genotypic and phenotypic 

frequencies of aphid populations under enriched CO2 and O3 atmospheres. Global Change Biol 11: 

1990-1996. http://dx.doi.org/10.1111/j.1365-2486.2005.01054.x

Morgan, PB; Ainsworth, EA; Long, SP. (2003). How does elevated ozone impact soybean? A meta-analysis 

of photosynthesis, growth and yield. Plant Cell Environ 26: 1317-1328. 

Morgan, PB; Bernacchi, CJ; Ort, DR; Long, SP. (2004). An in vivo analysis of the effect of season-long 

open-air elevation of ozone to anticipated 2050 levels on photosynthesis in soybean. J Plant Physiol 

135: 2348-2357. 

Morgan, PB; Mies, TA; Bollero, GA; Nelson, RL; Long, SP. (2006). Season-long elevation of ozone 

concentration to projected 2050 levels under fully open-air conditions substantially decreases the 

growth and production of soybean. New Phytol 170: 333-343. http://dx.doi.org/10.1111/j.1469-

8137.2006.01679.x

Morison, JIL; Lawlor, DW. (1999). Interactions between increasing CO2 concentration and temperature on 

plant growth. Plant Cell Environ 22: 659-682. http://dx.doi.org/10.1046/j.1365-3040.1999.00443.x

Morsky, SK; Haapala, JK; Rinnan, R; Tiiva, P; Saarnio, S; Silvola, J; Holopainen, T; Martikainen, PJ.

(2008). Long-term ozone effects on vegetation, microbial community and methane dynamics of boreal 

peatland microcosms in open-field conditions. Global Change Biol 14: 1891-1903. 

http://dx.doi.org/10.1111/j.1365-2486.2008.01615.x

Mudd, JB. (1996). Biochemical basis for the toxicity of ozone. In M Yunus; M Iqbal (Eds.), Plant response to 

air pollution (pp. 267-283). New York, NY: John Wiley & Sons. 

Muntifering, RB; Chappelka, AH; Lin, JC; Karnosky, DF; Somers, GL. (2006a). Chemical composition and 

digestibility of Trifolium exposed to elevated ozone and carbon dioxide in a free-air (FACE) 

fumigation system. Funct Ecol 20: 269-275. http://dx.doi.org/10.1111/j.1365-2435.2006.01093.x

Muntifering, RB; Manning, WJ; Lin, JC; Robinson, GB. (2006b). Short-term exposure to ozone altered the 

relative feed value of an alfalfa cultivar. Environ Pollut 140: 1-3. 

Musselman, RC; Lefohn, AS; Massman, WJ; Heath, RL. (2006). A critical review and analysis of the use of 

exposure- and flux-based ozone indices for predicting vegetation effects [Review]. Atmos Environ 40: 

1869-1888. http://dx.doi.org/10.1016/j.atmosenv.2005.10.064

Musselman, RC; Massman, WJ. (1999). Ozone flux to vegetation and its relationship to plant response and 

ambient air quality standards. Atmos Environ 33: 65-73. 

Musselman, RC; McCool, PM; Younglove, T. (1988). Selecting ozone exposure statistics for determining 

crop yield loss from air pollutants. Environ Pollut 53: 63-78. http://dx.doi.org/10.1016/0269-

7491(88)90025-5

Musselman, RC; Minnick, TJ. (2000). Nocturnal stomatal conductance and ambient air quality standards for 

ozone. Atmos Environ 34: 719-733. http://dx.doi.org/10.1016/S1352-2310(99)00355-6

Nali, C; Balducci, E; Frati, L; Paoli, L; Loppi, S; Lorenzini, G. (2007). Integrated biomonitoring of air 

quality with plants and lichens: A case study on ambient ozone from central Italy. Chemosphere 67: 

2169-2176. http://dx.doi.org/10.1016/j.chemosphere.2006.12.036
Draft – Do Not Cite or Quote 9-179 June 2012

NAPAP (National Acid Precipitation Assessment Program). (1987). Diagnosing injury to Eastern forest trees: 

A manual for identifying damage caused by air pollution, pathogens, insects, and abiotic stresses. 

University Park, PA: Pennsylvania State University. 

Nash, TH, III. (2008). Lichen sensitivity to air pollution. In Lichen Biology. Cambridge, United Kingdom: 

Cambridge University Press. 

Neufeld, HS; Lee, EH; Renfro, JR; Hacker, WD. (2000). Seedling insensitivity to ozone for three conifer 

species native to Great Smoky Mountains National Park. Environ Pollut 108: 141-151. 

http://dx.doi.org/10.1016/S0269-7491(99)00247-X

Neufeld, HS; Lee, EH; Renfro, JR; Hacker, WD; Yu, BH. (1995). Sensitivity of seedlings of black cherry 

(Prunus serotina Ehrh) to ozone in Great Smoky Mountains National Park I Exposure-response curves 

for biomass. New Phytol 130: 447-459. http://dx.doi.org/10.1111/j.1469-8137.1995.tb01839.x

Neufeld, HS; Renfro, JR; Hacker, WD; Silsbee, D. (1992). Ozone in Great Smoky Mountains National Park: 

Dynamics and effects on plants. In RL Berglund (Ed.), Tropospheric ozone and the environment II: 

Effects, modeling and control (pp. 594-617). Pittsburgh, PA: Air & Waste Management Association. 

Nikolova, PS; Andersen, CP; Blaschke, H; Matyssek, R; Häberle, KH. (2010). Belowground effects of 

enhanced tropospheric ozone and drought in a beech/spruce forest (Fagus sylvatica L./Picea abies [L.] 

Karst). Environ Pollut 158: 1071-1078. http://dx.doi.org/10.1016/j.envpol.2009.07.036

Noctor, G; Foyer, CH. (1998). Ascorbate and glutathione: Keeping active oxygen under control. 49: 249-

279. 

Norby, RJ; DeLucia, EH; Gielen, B; Calfapietra, C; Giardina, CP; King, JS; Ledford, J; McCarthy, HR; 

Moore, DJP; Ceulemans, R; De Angelis, P; Finzi, AC; Karnosky, DF; Kubiske, ME; Lukac, M; 

Pregitzer, KS; Scarascia-Mugnozza, GE; Schlesinger, WH; Oren, R. (2005). Forest response to 

elevated CO2 is conserved across a broad range of productivity. PNAS 102: 18052-18056. 

http://dx.doi.org/10.1073/pnas.0509478102

Novak, K; Cherubini, P; Saurer, M; Fuhrer, J; Skelly, JM; Kräuchi, N; Schaub, M. (2007). Ozone air 

pollution effects on tree-ring growth, delta(13)C, visible foliar injury and leaf gas exchange in three 

ozone-sensitive woody plant species. Tree Physiol 27: 941-949. 

NPS (U.S. National Park Service). (2006). Ozone bioindicators. Washington, DC. 

http://www.nature.nps.gov/air/Pubs/bioindicators/index.cfm

NPS (U.S. National Park Service). (2007). Ozone effects studies. Washington, DC: U.S. Department of the 

Interior, National Park Service. http://www.nature.nps.gov/air/studies/ecoOzone.cfm

Nunn, AJ; Reiter, IM; Haberle, KH; Werner, H; Langebartels, C; Sandermann, H; Heerdt, C; Fabian, P; 

Matyssek, R. (2002). Free-air ozone canopy fumigation in an old-growth mixed forest: concept and 

observations in beech. Phyton (Buenos Aires) 42: 105-119. 

Nussbaum, S; Geissmann, M; Fuhrer, J. (1995). Ozone exposure-response relationships for mixtures of 

perennial ryegrass and white clover depend on ozone exposure patterns. Atmos Environ 29: 989-995. 

http://dx.doi.org/10.1016/1352-2310(94)00368-U

O'Gara, PJ. (1922). Sulfur dioxide and fume problems and their solution. J Ind Eng Chem 14: 744-745. 

O'Neill, BF; Zangerl, AR; Delucia, EH; Berenbaum, MR. (2008). Longevity and fecundity of Japanese beetle 

(Popillia japonica) on foliage grown under elevated carbon dioxide. Environ Entomol 37: 601-607. 

O'Neill, BF; Zangerl, AR; Dermody, O; Bilgin, DD; Casteel, CL; Zavala, JA; DeLucia, EH; Berenbaum, MR.

(2010). Impact of elevated levels of atmospheric CO2 and herbivory on flavonoids of soybean 

(Glycine max Linnaeus). J Chem Ecol 36: 35-45. http://dx.doi.org/10.1007/s10886-009-9727-0

Ogawa, D; Nakajima, N; Sano, T; Tamaoki, M; Aono, M; Kubo, A; Kanna, M; Ioki, M; Kamada, H; Saji, H.

(2005). Salicylic acid accumulation under O-3 exposure is regulated by ethylene in tobacco plants. 

Plant Cell Physiol 46: 1062-1072. http://dx.doi.org/10.1093/pcp/pci118

Oksanen, E. (2003). Responses of selected birch (Betula pendula Roth) clones to ozone change over time. 

Plant Cell Environ 26: 875-886. 

Oksanen, E; Holopainen, T. (2001). Responses of two birch (Betula pendula Roth) clones to different ozone 

profiles with similar AOT40 exposure. Atmos Environ 35: 5245-5254. 

http://dx.doi.org/10.1016/S1352-2310(01)00346-6

Olbrich, M; Betz, G; Gerstner, E; Langebartels, C; Sandermann, H; Ernst, D. (2005). Transcriptome analysis 

of ozone-responsive genes in leaves of European beech (Fagus sylvatica L.). Plant Biol (Stuttg) 7: 

670-676. http://dx.doi.org/10.1055/s-2005-873001
Draft – Do Not Cite or Quote 9-180 June 2012

Olbrich, M; Gerstner, E; Bahnweg, G; Haberle, KH; Matyssek, R; Welzl, G; Heller, W; Ernst, D. (2010). 

Transcriptional signatures in leaves of adult European beech trees (Fagus sylvatica L.) in an 

experimentally enhanced free air ozone setting. Environ Pollut 158: 977-982.

http://dx.doi.org/10.1016/j.envpol.2009.08.001

Olbrich, M; Gerstner, E; Welzl, G; Winkler, JB; Ernst, D. (2009). Transcript responses in leaves of ozonetreated beech saplings seasons at an outdoor free air model fumigation site over two growing seasons. 

Plant Soil 323: 61-74. http://dx.doi.org/10.1007/s11104-009-0129-4

Ollinger, SV; Aber, JD; Reich, PB. (1997a). Simulating ozone effects on forest productivity: Interactions 

among leaf- and stand-level processes. Ecol Appl 123: 351-358. 

Ollinger, SV; Aber, JD; Reich, PB. (1997b). Simulating ozone effects on forest productivity: Interactions 

among leaf-, canopy-, and stand-level processes. Ecol Appl 7: 1237-1251. 

Ollinger, SV; Aber, JD; Reich, PB; Freuder, RJ. (2002). Interactive effects of nitrogen deposition, 

tropospheric ozone, elevated CO2 and land use history on the carbon dynamics of northern hardwood 

forests. Global Change Biol 8: 545-562. http://dx.doi.org/10.1046/j.1365-2486.2002.00482.x

Olszyk, DM; Kats, G; Dawson, PJ; Bytnerowicz, A; Wolf, J; Thompson, CR. (1986). Characteristics of air 

exclusion systems vs chambers for field air pollution studies. J Environ Qual 15: 326-334. 

Omasa, K; Shimazaki, KI; Aiga, I; Larcher, W; Onoe, M. (1987). Image analysis of chlorophyll fluorescence 

transients for diagnosing the photosynthetic system of attached leaves. Plant Physiol 84: 748-752. 

http://dx.doi.org/10.1104/pp.84.3.748

Orendovici-Best, T; Skelly, JM; Davis, DD; Ferdinand, JA; Savage, JE; Stevenson, RE. (2008). Ozone 

uptake (flux) as it relates to ozone-induced foliar symptoms of Prunus serotina and Populus 

maximowizii x trichocarpa. Environ Pollut 151: 79-92. http://dx.doi.org/10.1016/j.envpol.2007.03.003

Orendovici, T; Skelly, JM; Ferdinand, JA; Savage, JE; Sanz, MJ; Smith, GC. (2003). Response of native 

plants of northeastern United States and southern Spain to ozone exposures; determining 

exposure/response relationships. Environ Pollut 125: 31-40. 

Oshima, RJ; Braegelmann, PK; Baldwin, DW; Van Way, V; Taylor, OC. (1977). Reduction of tomato fruit 

size and yield by ozone. J Am Soc Hortic Sci 102: 289-293. 

Oshima, RJ; Poe, MP; Braegelmann, PK; Baldwin, DW; Van Way, V. (1976). Ozone dosage-crop loss 

function for alfalfa: A standardized method for assessing crop losses from air pollutants. J Air Pollut 

Control Assoc 26: 861-865. 

Overmyer, K; Brosche, M; Pellinen, R; Kuittinen, T; Tuominen, H; Ahlfors, R; Keinanen, M; Saarma, M; 

Scheel, D; Kangasjarvi, J. (2005). Ozone-induced programmed cell death in the Arabidopsis radicalinduced cell death1 mutant. Plant Physiol 137: 1092-1104. http://dx.doi.org/10.1104/pp.104.055681

Overmyer, K; Kollist, H; Tuominen, H; Betz, C; Langebartels, C; Wingsle, G; Kangasjarvi, S; Brader, G; 

Mullineaux, P; Kangasjarvi, J. (2008). Complex phenotypic profiles leading to ozone sensitivity in 

Arabidopsis thaliana mutants. Plant Cell Environ 31: 1237-1249. http://dx.doi.org/10.1111/j.1365-

3040.2008.01837.x

Overmyer, K; Tuominen, H; Kettunen, R; Betz, C; Langebartels, C; Sandermann, H, Jr; Kangasjarvi, J.

(2000). Ozone-sensitive Arabidopsis rcd1 mutant reveals opposite roles for ethylene and jasmonate 

signaling pathways in regulating superoxide-dependent cell death. Plant Cell 12: 1849-1862. 

http://dx.doi.org/10.1105/tpc.12.10.1849

Pan, YD; Birdsey, R; Hom, J; McCullough, K. (2009). Separating effects of changes in atmospheric 

composition, climate and land-use on carbon sequestration of US Mid-Atlantic temperate forests. For 

Ecol Manage 259: 151-164. http://dx.doi.org/10.1016/j.foreco.2009.09.049

Panek, J; Kurpius, MR; Goldstein, AH. (2002). An evaluation of ozone exposure metrics for a seasonally 

drought-stressed ponderosa pine ecosystem. Environ Pollut 117: 93-100. 

http://dx.doi.org/10.1016/S0269-7491(01)00155-5

Panek, JA. (2004). Ozone uptake, water loss and carbon exchange dynamics in annually drought-stressed 

Pinus ponderosa forests: Measured trends and parameters for uptake modeling. Tree Physiol 24: 277-

290. 

Panek, JA; Goldstein, AH. (2001). Responses of stomatal conductance to drought in ponderosa pine: 

Implications for carbon and ozone uptake. Tree Physiol 21: 337-344. 

Paolacci, AR; Miraldi, C; Tanzarella, OA; Badiani, M; Porceddu, E; Nali, C; Lorenzini, G; Ciaffi, M. (2007). 

Gene expression induced by chronic ozone in the Mediterranean shrub Phillyrea latifolia: Analysis by 

cDNA-AFLP. Tree Physiol 27: 1541-1550. http://dx.doi.org/10.1093/treephys/27.11.1541
Draft – Do Not Cite or Quote 9-181 June 2012

Paoletti, E; Grulke, NE. (2005). Does living in elevated CO2 ameliorate tree response to ozone? A review on 

stomatal responses [Review]. Environ Pollut 137: 483-493. 

http://dx.doi.org/10.1016/j.envpol.2005.01.035

Paoletti, E; Grulke, NE. (2010). Ozone exposure and stomatal sluggishness in different plant physiognomic 

classes. Environ Pollut 158: 2664-2671. http://dx.doi.org/10.1016/j.envpol.2010.04.024

Paoletti, E; Manning, WJ. (2007). Toward a biologically significant and usable standard for ozone that will 

also protect plants. Environ Pollut 150: 85-95. http://dx.doi.org/10.1016/j.envpol.2007.06.037

Paoletti, E; Seufert, G; Della Rocca, G; Thomsen, H. (2007). Photosynthetic responses to elevated CO2 and 

O-3 in Quercus ilex leaves at a natural CO2 spring. Environ Pollut 147: 516-524. 

http://dx.doi.org/10.1016/j.envpol.2006.08.039

Parsons, WFJ; Bockheim, JG; Lindroth, RL. (2008). Independent, interactive, and species-specific responses 

of leaf litter decomposition to elevated CO2 and O3 in a northern hardwood forest. Ecosystems 11: 

505-519. 

Pearson, M; Mansfield, TA. (1993). Interacting effects of ozone and water stress on the stomatal resistance of 

beech (Fagus sylvatica L). New Phytol 123: 351-358. http://dx.doi.org/10.1111/j.1469-

8137.1993.tb03745.x

Pearson, S; Davison, AW; Reiling, K; Ashenden, T; Ollerenshaw, JH. (1996). The effects of different ozone 

exposures on three contrasting populations of Plantago major. New Phytol 132: 493-502. 

http://dx.doi.org/10.1111/j.1469-8137.1996.tb01869.x

Pei, ZM; Murata, Y; Benning, G; Thomine, S; Klüsener, B; Allen, GJ; Grill, E; Schroeder, JI. (2000). 

Calcium channels activated by hydrogen peroxide mediate abscisic acid signalling in guard cells. 

Nature 406: 731-734. http://dx.doi.org/10.1038/35021067

Pell, EJ; Sinn, JP; Brendley, BW; Samuelson, L; Vinten-Johansen, C; Tien, M; Skillman, J. (1999). 

Differential response of four tree species to ozone-induced acceleration of foliar senescence. Plant 

Cell Environ 22: 779-790. http://dx.doi.org/10.1046/j.1365-3040.1999.00449.x

Peltonen, PA; Julkunen-Tiitto, R; Vapaavuori, E; Holopainen, JK. (2006). Effects of elevated carbon dioxide 

and ozone on aphid oviposition preference and birch bud exudate phenolics. Global Change Biol 12: 

1670-1679. http://dx.doi.org/10.1111/j.1365-2486.2006.01226.x

Peltonen, PA; Vapaavuori, E; Heinonen, J; Julkunen-Tiitto, R; Holopainen, JK. (2010). Do elevated 

atmospheric CO2 and O-3 affect food quality and performance of folivorous insects on silver birch? 

Global Change Biol 16: 918-935. http://dx.doi.org/10.1111/j.1365-2486.2009.02073.x

Percy, KE; Nosal, M; Heilman, W; Dann, T; Sober, J; Legge, AH; Karnosky, DF. (2007). New exposurebased metric approach for evaluating O3 risk to North American aspen forests. Environ Pollut 147: 

554-566. http://dx.doi.org/10.1016/j.envpol.2006.10.009

Peterson, DL; Arbaugh, MJ; Wakefield, VA; Miller, PR. (1987). Evidence of growth reduction in ozoneinjured Jeffrey pine (Pinus jeffreyi Grev and Balf) in Sequoia and Kings Canyon National Parks. J Air 

Waste Manag Assoc 37: 906-912. 

Pfleeger, TG; Plocher, M; Bichel, P. (2010). Response of pioneer plant communities to elevated ozone 

exposure. Agric Ecosyst Environ 138: 116-126. http://dx.doi.org/10.1016/j.agee.2010.04.009

Phillips, DL; Johnson, MG; Tingey, DT; Storm, MJ. (2009). Elevated CO2 and O3 effects on fine-root 

survivorship in ponderosa pine mesocosms. Oecologia 160: 827-837. 

http://dx.doi.org/10.1007/s00442-009-1339-4

Phillips, RL; Zak, DR; Holmes, WE; White, DC. (2002). Microbial community composition and function 

beneath temperate trees exposed to elevated atmospheric carbon dioxide and ozone. Oecologia 131: 

236-244. 

Piikki, K; De Temmerman, L; Hogy, P; Pleijel, H. (2008a). The open-top chamber impact on vapour pressure 

deficit and its consequences for stomatal ozone uptake. Atmos Environ 42: 6513-6522. 

http://dx.doi.org/10.1016/j.atmosenv.2008.04.014

Piikki, K; De Temmerman, L; Ojanpera, K; Danielsson, H; Pleijel, H. (2008b). The grain quality of spring 

wheat (Triticum aestivum L.) in relation to elevated ozone uptake and carbon dioxide exposure. Eur J 

Agron 28: 245-254. http://dx.doi.org/10.1016/j.eja.2007.07.004

Piikki, K; Vorne, V; Ojanpera, K; Pleijel, H. (2007). Impact of elevated O3 and CO2 exposure on potato 

(Solanum tuberosum L. cv. Bintje) tuber macronutrients (N, P, K, Mg, Ca). Agric Ecosyst Environ 

118: 55-64. http://dx.doi.org/10.1016/j.agee.2006.04.012
Draft – Do Not Cite or Quote 9-182 June 2012

Pinto, DM; Blande, JD; Nykanen, R; Dong, WX; Nerg, AM; Holopainen, JK. (2007a). Ozone degrades 

common herbivore-induced plant volatiles: Does this affect herbivore prey location by predators and 

parasitoids? J Chem Ecol 33: 683-694. http://dx.doi.org/10.1007/s10886-007-9255-8

Pinto, DM; Blande, JD; Souza, SR; Nerg, AM; Holopainen, JK. (2010). Plant volatile organic compounds 

(VOCs) in ozone (O3) polluted atmospheres: The ecological effects. J Chem Ecol 36: 22-34. 

http://dx.doi.org/10.1007/s10886-009-9732-3

Pinto, DM; Himanen, SJ; Nissinen, A; Nerg, AM; Holopainen, JK. (2008). Host location behavior of Cotesia 

plutellae Kurdjumov (Hymenoptera: Braconidae) in ambient and moderately elevated ozone in field 

conditions. Environ Pollut 156: 227-231. http://dx.doi.org/10.1016/j.envpol.2007.12.009

Pinto, DM; Nerg, AM; Holopainen, JK. (2007b). The role of ozone-reactive compounds, terpenes, and green 

leaf volatiles (GLVs), in the orientation of Cotesia plutellae. J Chem Ecol 33: 2218-2228. 

http://dx.doi.org/10.1007/s10886-007-9376-0

Pinto, J. (2009). Wyoming winter smog. Nat Geosci 2: 88-90. http://dx.doi.org/10.1038/ngeo430

Pleijel, H; Danielsson, H; Gelang, J; Sild, E; Sellden, G. (1998). Growth stage dependence of the grain yield 

response to ozone in spring wheat (Triticum aestivum L). Agric Ecosyst Environ 70: 61-68. 

http://dx.doi.org/10.1016/S0167-8809(97)00167-9

Pleijel, H; Danielsson, H; Ojanpera, K; De Temmerman, L; Hogy, P; Badiani, M; Karlsson, PE. (2004a). 

Relationships between ozone exposure and yield loss in European wheat and potato--a comparison of 

concentration- and flux-based exposure indices. Atmos Environ 38: 2259-2269. 

Pleijel, H; h, D; Ojanpera, K; De Temmerman, L; Hogy, P. (2004b). Relationships between ozone exposure 

and yield loss in wheat and potato - Suggestions of critical levels for ozone effects on crops. Atmos 

Environ 38: 2259-2269. http://dx.doi.org/10.1016/j.atmosenv.2003.09.076

Pleijel, H; Ojanpera, K; Mortensen, L. (1997). Effects of tropospheric ozone on the yield and grain protein 

content of spring wheat (Triticum aestivum L) in the nordic countries. Acta Agric Scand B Soil Plant 

Sci 47: 20-25. http://dx.doi.org/10.1080/09064719709362434

Plessl, M; Elstner, EF; Rennenberg, H; Habermeyer, J; Heiser, I. (2007). Influence of elevated CO2 and 

ozone concentrations on late blight resistance and growth of potato plants. Environ Exp Bot 60: 447-

457. http://dx.doi.org/10.1016/j.envexpbot.2007.01.003

Plochl, M; Lyons, T; Ollerenshaw, J; Barnes, J. (2000). Simulating ozone detoxification in the leaf apoplast 

through the direct reaction with ascorbate. Planta 210: 454-467. 

http://dx.doi.org/10.1007/PL00008153

Pollastrini, M; Desotgiu, R; Cascio, C; Bussotti, F; Cherubini, P; Saurer, M; Gerosa, G; Marzuoli, R. (2010). 

Growth and physiological responses to ozone and mild drought stress of tree species with different 

ecological requirements. Trees Struct Funct 24: 695-704. http://dx.doi.org/10.1007/s00468-010-0439-

4

Polle, A; Pell, EJ. (1999). Role of carbon dioxide in modifying the plant response to ozone. In Y Luo; HA 

Mooney (Eds.), Carbon dioxide and environmental stress (pp. 193-213). San Diego, CA: Academic 

Press. 

Pregitzer, K; Loya, W; Kubiske, M; Zak, D. (2006). Soil respiration in northern forests exposed to elevated 

atmospheric carbon dioxide and ozone. Oecologia 148: 503-516. http://dx.doi.org/10.1007/s00442-

006-0381-8

Pregitzer, KS; Burton, AJ; King, JS; Zak, DR. (2008). Soil respiration, root biomass, and root turnover 

following long-term exposure of northern forests to elevated atmospheric Co-2 and tropospheric O-3. 

New Phytol 180: 153-161. http://dx.doi.org/10.1111/j.1469-8137.2008.02564.x

Pretzsch, H; Dieler, J; Matyssek, R; Wipfler, P. (2010). Tree and stand growth of mature Norway spruce and 

European beech under long-term ozone fumigation. Environ Pollut 158: 1061-1070. 

http://dx.doi.org/10.1016/j.envpol.2009.07.035

Pritsch, K; Esperschuetz, J; Haesler, F; Raidl, S; Winkler, B; Schloter, M. (2009). Structure and activities of 

ectomycorrhizal and microbial communities in the rhizosphere of Fagus sylvatica under ozone and 

pathogen stress in a lysimeter study. Plant Soil 323: 97-109. http://dx.doi.org/10.1007/s11104-009-

9972-6

Puckette, MC; Tang, YH; Mahalingam, R. (2008). Transcriptomic changes induced by acute ozone in 

resistant and sensitive Medicago truncatula accessions. BMC Plant Biol 8: 46. 

http://dx.doi.org/10.1186/1471-2229-8-46
Draft – Do Not Cite or Quote 9-183 June 2012

Pujol Pereira, EI; Chung, H; Scow, K; Sadowsky, MJ; van Kessel, C; Six, J. (2011). Soil nitrogen 

transformations under elevated atmospheric CO and O during the soybean growing season. Environ 

Pollut 159: 401-407. http://dx.doi.org/10.1016/j.envpol.2010.10.033

Ramo, K; Kanerva, T; Ojanpera, K; Manninen, S. (2007). Growth onset, senescence, and reproductive 

development of meadow species in mesocosms exposed to elevated O3 and CO2. Environ Pollut 145: 

850-860. http://dx.doi.org/10.1016/j.envpol.2006.03.054

Rao, MV; Hale, BA; Ormrod, DP. (1995). Amelioration of ozone-induced oxidative damage in wheat plants 

grown under high carbon dioxide: Role of antioxidant enzymes. J Plant Physiol 109: 421-432. 

Rawlings, JO; Cure, WW. (1985). The Weibull function as a dose-response model to describe ozone effects 

on crop yields. Crop Sci 25: 807-814. 

Reich, PB. (1987). Quantifying plant response to ozone: A unifying theory. Tree Physiol 3: 63-91. 

http://dx.doi.org/10.1093/treephys/3.1.63

Reich, PB; Lassoie, JP. (1984). Effects of low level O3 exposure on leaf diffusive conductance and water-use 

efficiency in hybrid poplar. Plant Cell Environ 7: 661-668. http://dx.doi.org/10.1111/1365-

3040.ep11571645

Reid, CD; Fiscus, EL. (2008). Ozone and density affect the response of biomass and seed yield to elevated 

CO2 in rice. Global Change Biol 14: 60-76. http://dx.doi.org/10.1111/j.1365-2486.2007.01472.x

Reiling, K; Davison, AW. (1992). Effects of a short ozone exposure given at different stages in the 

development of Plantago major L. New Phytol 121: 643-647. http://dx.doi.org/10.1111/j.1469-

8137.1992.tb01135.x

Reiling, K; Davison, AW. (1994). Effects of exposure to ozone at different stages in the development of 

Plantago major L on chlorophyll fluorescence and gas exchange. New Phytol 128: 509-514. 

http://dx.doi.org/10.1111/j.1469-8137.1994.tb02998.x

Reinert, RA; Eason, G; Barton, J. (1997). Growth and fruiting of tomato as influenced by elevated carbon 

dioxide and ozone. New Phytol 137: 411-420. http://dx.doi.org/10.1046/j.1469-8137.1997.00846.x

Reinert, RA; Ho, MC. (1995). Vegetative growth of soybean as affected by elevated carbon dioxide and 

ozone. Environ Pollut 89: 89-96. http://dx.doi.org/10.1016/0269-7491(94)00039-G

Ren, W; Tian, H; Chen, G; Liu, M; Zhang, C; Chappelka, AH; Pan, S. (2007a). Influence of ozone pollution 

and climate variability on net primary productivity and carbon storage in China's grassland ecosystems 

from 1961 to 2000. Environ Pollut 149: 327-335. http://dx.doi.org/10.1016/j.envpol.2007.05.029

Ren, W; Tian, H; Tao, B; Chappelka, A; Sun, G; Lu, C; Liu, M; Chen, G; Xu, X. (2011). Impacts of 

tropospheric ozone and climate change on net primary productivity and net carbon exchange of 

China's forest ecosystems. Global Ecology and Biogeography 20: 391-406. 

http://dx.doi.org/10.1111/j.1466-8238.2010.00606.x

Ren, W; Tian, HQ; Liu, ML; Zhang, C; Chen, GS; Pan, SF; Felzer, B; Xu, XF. (2007b). Effects of 

tropospheric ozone pollution on net primary productivity and carbon storage in terrestrial ecosystems 

of China. J Geophys Res 112: D22S09. http://dx.doi.org/10.1029/2007jd008521

Rhea, L; King, J; Kubiske, M; Saliendra, N; Teclaw, R. (2010). Effects of elevated atmospheric CO2 and 

tropospheric O3 on tree branch growth and implications for hydrologic budgeting. Environ Pollut 158: 

1079-1087. http://dx.doi.org/10.1016/j.envpol.2009.08.038

Riddell, J; Padgett, PE; Nash, TH, III. (2010). Responses of the lichen Ramalina menziesii Tayl. to ozone 

fumigations. In TH Nash, III (Ed.), Biology of lichens: Symbiosis, ecology, environmental 

monitoring, systematics, cyber applications (pp. 113-123). Stuttgart: J. Cramer in der Gebrüder 

Borntraeger Verlagsbuchhandlung. 

Riikonen, J; Kets, K; Darbah, J; Oksanen, E; Sober, A; Vapaavuori, E; Kubiske, ME; Nelson, N; Karnosky, 

DF. (2008). Carbon gain and bud physiology in Populus tremuloides and Betula papyrifera grown 

under long-term exposure to elevated concentrations of CO2 and O3. Tree Physiol 28: 243-254. 

http://dx.doi.org/10.1093/treephys/28.2.243

Riikonen, J; Maenpaa, M; Alavillamo, M; Silfver, T; Oksanen, E. (2009). Interactive effect of elevated 

temperature and O3 on antioxidant capacity and gas exchange in Betula pendula saplings. Planta 230: 

419-427. http://dx.doi.org/10.1007/s00425-009-0957-8

Rizzo, M; Bernardi, R; Salvini, M; Nali, C; Lorenzini, G; Durante, M. (2007). Identification of differentially 

expressed genes induced by ozone stress in sensitive and tolerant poplar hybrids. J Plant Physiol 164: 

945-949. http://dx.doi.org/10.1016/j.jplph.2006.07.012
Draft – Do Not Cite or Quote 9-184 June 2012

Rodenkirchen, H; Gottlein, A; Kozovits, AR; Matyssek, R; Grams, TEE. (2009). Nutrient contents and 

efficiencies of beech and spruce saplings as influenced by competition and O3/CO2 regime. European 

Journal of Forest Research 128: 117-128. http://dx.doi.org/10.1007/s10342-008-0221-y

Rogers, A; Allen, DJ; Davey, PA; Morgan, PB; Ainsworth, EA; Bernacchi, CJ; Cornic, G; Dermody, OC; 

Dohleman, FG; Heaton, EA; Mahoney, J; Zhu, XG; Delucia, EH; Ort, DR; Long, SP. (2004). Leaf 

photosynthesis and carbohydrate dynamics of soybean grown throughout their life-cycle under free-air 

carbon dioxide enrichment. Plant Cell Environ 27: 449-458. 

Rowland-Bamford, AJ. (2000). Plant responses to changing carbon dioxide and temperature. In SN Singh 

(Ed.), Trace gas emissions and plants (pp. 63-74). Dordecht, The Netherlands: Kluwer Academic 

Publishers. 

Ryan, A; Cojocariu, C; Possell, M; Davies, WJ; Hewitt, CN. (2009). Defining hybrid poplar (Populus 

deltoides x Populus trichocarpa) tolerance to ozone: Identifying key parameters. Plant Cell Environ 

32: 31-45. http://dx.doi.org/10.1111/j.1365-3040.2008.01897.x

Ryang, SZ; Woo, SY; Kwon, SY; Kim, SH; Lee, SH; Kim, KN; Lee, DK. (2009). Changes of net 

photosynthesis, antioxidant enzyme activities, and antioxidant contents of Liriodendron tulipifera 

under elevated ozone. Photosynthetica 47: 19-25. http://dx.doi.org/10.1007/s11099-009-0005-8

Samuel, MA; Ellis, BE. (2002). Double jeopardy: Both overexpression and suppression of a redox-activated 

plant mitogen-activated protein kinase render tobacco plants ozone sensitive. Plant Cell 14: 2059-

2069. http://dx.doi.org/10.1105/tpc.002337

Samuel, MA; Miles, GP; Ellis, BE. (2000). Ozone treatment rapidly activates MAP kinase signalling in 

plants. Plant J 22: 367-376. http://dx.doi.org/10.1046/j.1365-313x.2000.00741.x

Samuel, MA; Walia, A; Mansfield, SD; Ellis, BE. (2005). Overexpression of SIPK in tobacco enhances 

ozone-induced ethylene formation and blocks ozone-induced SA accumulation. J Exp Bot 56: 2195-

2201. http://dx.doi.org/10.1093/jxb/eri219

Samuelson, LJ; Kelly, JM. (1997). Ozone uptake in Prunus serotina, Acer rubrum and Quercus rubra forest 

trees of different sizes. New Phytol 136: 255-264. http://dx.doi.org/10.1046/j.1469-8137.1997.00734.x

Sánchez, MJS; Peña, GS; Lorente, VC; Gallego, TM; Albert, JC. (2001). La contaminación atmosférica en 

los bosques: Guía para la identificacíon de daños visibles causados por Ozono. Madrid, Spain: 

Ministerio de Medio Ambiente. http://www.ceam.es/VentaLibros/guia_O3/index.htm

Sandermann, H. (2008). Ecotoxicology of ozone: Bioactivation of extracellular ascorbate. Biochem Biophys 

Res Commun 366: 271-274. http://dx.doi.org/10.1016/j.bbrc.2007.12.018

Sanz, J; Bermejo, V; Gimeno, BS; Elvira, S; Alonso, R. (2007). Ozone sensitivity of the Mediterranean 

terophyte trifolium striatum is modulated by soil nitrogen content. Atmos Environ 41: 8952-8962. 

http://dx.doi.org/10.1016/j.atmosenv.2007.08.016

Sanz, J; Muntifering, RB; Bermejo, V; Gimeno, BS; Elvira, S. (2005). Ozone and increased nitrogen supply 

effects on the yield and nutritive quality of Trifolium subterraneum. Atmos Environ 39: 5899-5907. 

http://dx.doi.org/10.1016/j.atmosenv.2005.06.022

Sarkar, A; Rakwal, R; Agrawal, SB; Shibato, J; Ogawa, Y; Yoshida, Y; Agrawal, GK; Agrawal, M. (2010). 

Investigating the impact of elevated levels of ozone on tropical wheat using integrated phenotypical, 

physiological, biochemical, and proteomics approaches. J Proteome Res 9: 4565-4584. 

http://dx.doi.org/10.1021/Pr1002824

Saviranta, NMM; Julkunen-Tiitto, R; Oksanen, E; Karjalainen, RO. (2010). Leaf phenolic compounds in red 

clover (Trifolium pratense L.) induced by exposure to moderately elevated ozone. Environ Pollut 158: 

440-446. http://dx.doi.org/10.1016/j.envpol.2009.08.029

Sawada, H; Kohno, Y. (2009). Differential ozone sensitivity of rice cultivars as indicated by visible injury 

and grain yield. Plant Biol (Stuttg) 11: 70-75. http://dx.doi.org/10.1111/j.1438-8677.2009.00233.x

Scebba, F; Giuntini, D; Castagna, A; Soldatini, G; Ranieri, A. (2006). Analysing the impact of ozone on 

biochemical and physiological variables in plant species belonging to natural ecosystems. Environ 

Exp Bot 57: 89-97. http://dx.doi.org/10.1016/j.envexpbot.2005.04.005

Schaub, M; Skelly, JM; Zhang, JW; Ferdinand, JA; Savage, JE; Stevenson, RE; Davis, DD; Steiner, KC.

(2005). Physiological and foliar symptom response in the crowns of Prunus serotina, Fraxinus 

americana and Acer rubrum canopy trees to ambient ozone under forest conditions. Environ Pollut 

133: 553-567. http://dx.doi.org/10.1016/j.envpol.2004.06.012

Schraudner, M; Moeder, W; Wiese, C; Van Camp, W; Inze, D; Langebartels, C; Sandermann, H, Jr. (1998). 

Ozone-induced oxidative burst in the ozone biomonitor plant, tobacco Bel W3. Plant J 16: 235-245. 

http://dx.doi.org/10.1046/j.1365-313x.1998.00294.x
Draft – Do Not Cite or Quote 9-185 June 2012

Severino, JF; Stich, K; Soja, G. (2007). Ozone stress and antioxidant substances in Trifolium repens and 

Centaurea jacea leaves. Environ Pollut 146: 707-714. http://dx.doi.org/10.1016/j.envpol.2006.04.006

Sharkey, TD; Wiberley, AE; Donohue, AR. (2008). Isoprene emission from plants: Why and how. Ann Bot 

101: 5-18. http://dx.doi.org/10.1093/aob/mcm240

Singh, E; Tiwari, S; Agrawal, M. (2009). Effects of elevated ozone on photosynthesis and stomatal 

conductance of two soybean varieties: A case study to assess impacts of one component of predicted 

global climate change. Plant Biol (Stuttg) 11: 101-108. http://dx.doi.org/10.1111/j.1438-

8677.2009.00263.x

Singh, E; Tiwari, S; Agrawal, M. (2010a). Variability in antioxidant and metabolite levels, growth and yield 

of two soybean varieties: An assessment of anticipated yield losses under projected elevation of ozone. 

Agric Ecosyst Environ 135: 168-177. http://dx.doi.org/10.1016/j.agee.2009.09.004

Sitch, S; Cox, PM; Collins, WJ; Huntingford, C. (2007). Indirect radiative forcing of climate change through 

ozone effects on the land-carbon sink. Nature 448: 791-794. http://dx.doi.org/10.1038/nature06059

Skarby, L; Ottosson, S; Karlsson, PE; Wallina, G; Sellden, G; Medina, EL; Pleijel, H. (2004). Growth of 

Norway spruce (Picea abies) in relation to different ozone exposure indices: A synthesis. Atmos 

Environ 38: 2225-2236. 

Skarby, L; Troeng, E; Bostrom, CA. (1987). Ozone uptake and effects on transpiration, net photosynthesis, 

and dark respiration in Scots pine. Forest Sci 33: 801-808. 

Smith, G; Coulston, J; Jepsen, E; Prichard, T. (2003). A national ozone biomonitoring program: Results from 

field surveys of ozone sensitive plants in northeastern forests (1994-2000). Environ Monit Assess 87: 

271-291. 

Soja, G; Barnes, JD; Posch, M; Vandermeiren, K; Pleijel, H; Mills, G. (2000). Phenological weighting of 

ozone exposures in the calculation of critical levels for wheat, bean and plantain. Environ Pollut 109: 

517-524. http://dx.doi.org/10.1016/S0269-7491(00)00055-5

Soja, G; Reichenauer, TG; Eid, M; Soja, AM; Schaber, R; Gangl, H. (2004). Long-term ozone exposure and 

ozone uptake of grapevines in open-top chambers. Atmos Environ 38: 2313-2321. 

Somers, GL; Chappelka, AH; Rosseau, P; Renfro, JR. (1998). Empirical evidence of growth decline related 

to visible ozone injury. For Ecol Manage 104: 129-137. http://dx.doi.org/10.1016/S0378-

1127(97)00252-1

Souza, L; Neufeld, HS; Chappelka, AH; Burkey, KO; Davison, AW. (2006). Seasonal development of ozoneinduced foliar injury on tall milkweed (Asclepias exaltata) in Great Smoky Mountains National Park. 

Environ Pollut 141: 175-183. http://dx.doi.org/10.1016/j.envpol.2005.07.022

Stampfli, A; Fuhrer, J. (2010). Spatial heterogeneity confounded ozone-exposure experiment in semi-natural 

grassland. Oecologia 162: 515-522. http://dx.doi.org/10.1007/s00442-009-1462-2

Stewart, CA. (1998) Impact of ozone on the reproductive biology of Brassica campestris L and Plantago 

major L. (Doctoral Dissertation). Loughborough University of Technology, England. Retrieved from 

http://ethos.bl.uk/OrderDetails.do?did=1&uin=uk.bl.ethos.299673

Stewart, CA; Black, VJ; Black, CR; Roberts, JA. (1996). Direct effects of ozone on the reproductive 

development of Brassica species. J Plant Physiol 148: 172-178. 

Stoelken, G; Pritsch, K; Simon, J; Mueller, CW; Grams, TEE; Esperschuetz, J; Gayler, S; Buegger, F; 

Brueggemann, N; Meier, R; Zeller, B; Winkler, JB; Rennenberg, H. (2010). Enhanced ozone exposure 

of European beech (Fagus sylvatica) stimulates nitrogen mobilization from leaf litter and nitrogen 

accumulation in the soil. Plant Biosystems 144: 537-546. 

http://dx.doi.org/10.1080/11263500903429346

Street, NR; James, TM; James, T; Mikael, B; Jaakko, K; Mark, B; Taylor, G. (2011). The physiological, 

transcriptional and genetic responses of an ozone-sensitive and an ozone tolerant poplar and selected 

extremes of their F2 progeny. Environ Pollut 159: 45-54. 

http://dx.doi.org/10.1016/j.envpol.2010.09.027

Summers, CG; Retzlaff, WA; Stephenson, S. (1994). The effect of ozone on the mean relative growth-rate of 

Diuraphis-noxia (Mordvilko) (Homoptera, Aphididae). J Agr Entomol 11: 181-187. 

Talhelm, AF; Pregitzer, KS; Zak, DR. (2009). Species-specific responses to atmospheric carbon dioxide and 

tropospheric ozone mediate changes in soil carbon. Ecol Lett 12: 1219-1228. 

http://dx.doi.org/10.1111/j.1461-0248.2009.01380.x
Draft – Do Not Cite or Quote 9-186 June 2012

Tamaoki, M; Nakajima, N; Kubo, A; Aono, M; Matsuyama, T; Saji, H. (2003). Transcriptome analysis of 

O3-exposed Arabidopsis reveals that multiple signal pathways act mutually antagonistically to induce 

gene expression. Plant Mol Biol 53: 443-456. 

http://dx.doi.org/10.1023/B:PLAN.0000019064.55734.52

Temple, PJ; Kupper, RS; Lennox, RW; Rohr, K. (1988). Injury and yield responses of differentially irrigated 

cotton to ozone. Agron J 80: 751-755. http://dx.doi.org/10.2134/agronj1988.00021962008000050011x

Temple, PJ; Riechers, GH; Miller, PR. (1992). Foliar injury responses of ponderosa pine seedlings to ozone, 

wet and dry acidic deposition, and drought. Environ Exp Bot 32: 101-113. 

http://dx.doi.org/10.1016/0098-8472(92)90035-Z

Theis, N; Raguso, RA. (2005). The effect of pollination on floral fragrance in thistles. J Chem Ecol 31: 2581-

2600. 

Thomas, VFD; Braun, S; Fluckiger, W. (2005). Effects of simultaneous ozone exposure and nitrogen loads on 

carbohydrate concentrations, biomass, and growth of young spruce trees (Picea abies). Environ Pollut 

137: 507-516. http://dx.doi.org/10.1016/j.envpol.2005.02.002

Thomas, VFD; Braun, S; Fluckiger, W. (2006). Effects of simultaneous ozone exposure and nitrogen loads on 

carbohydrate concentrations, biomass, growth, and nutrient concentrations of young beech trees 

(Fagus sylvatica). Environ Pollut 143: 341-354. http://dx.doi.org/10.1016/j.envpol.2005.11.036

Tian, H; Melillo, J; Lu, C; Kicklighter, D; Liu, M; Ren, W; Xu, X; Chen, G; Zhang, C; Pan, S; Liu, J; 

Running, S. (2011). China's terrestrial carbon balance: Contributions from multiple global change 

factors. Global Biogeochem Cycles 25: GB1007. http://dx.doi.org/10.1029/2010GB003838

Tingey, DT; Hogsett, WE; Lee, EH; Herstrom, AA; Azevedo, SH. (1991). An evaluation of various 

alternative ambient ozone standards based on crop yield loss data. In RL Berglund; DR Lawson; DJ 

McKee (Eds.), Tropospheric Ozone and the Environment (pp. 272-288). Pittsburgh, PA: Air & Waste 

Management Association. 

Tingey, DT; Hogsett, WE; Lee, EH; Laurence, JA. (2004). Stricter ozone ambient air quality standard has 

beneficial effect on ponderosa pine in California. J Environ Manage 34: 397-405. 

Tingey, DT; Johnson, MG; Lee, EH; Wise, C; Waschmann, R; Olszyk, DM; Watrud, LS; Donegan, KK.

(2006). Effects of elevated CO2 and O3 on soil respiration under ponderosa pine. Soil Biol Biochem 

38: 1764-1778. http://dx.doi.org/10.1016/j.soilbio.2005.12.003

Tingey, DT; McVeety, BD; Waschmann, R; Johnson, MG; Phillips, DL; Rygiewicz, PT; Olszyk, DM. (1996). 

A versatile sun-lit controlled-environment facility for studying plant and soil processes. J Environ 

Qual 25: 614-625. 

Tingey, DT; Rodecap, KD; Lee, EH; Hogsett, WE; Gregg, JW. (2002). Pod development increases the ozone 

sensitivity of Phaseolus vulgaris. Water Air Soil Pollut 139: 325-341. 

Tissue, DT; Griffin, KL; Ball, T. (1999). Photosynthetic adjustment in field-grown ponderosa pine trees after 

six years of exposure to elevated CO2. Tree Physiol 19: 221-228. 

Tissue, DT; Thomas, RB; Strain, BR. (1997). Atmospheric CO2 enrichment increases growth and 

photosynthesis of Pinus taeda: A 4 year experiment in the field. Plant Cell Environ 20: 1123-1134. 

http://dx.doi.org/10.1046/j.1365-3040.1997.d01-140.x

Tjoelker, MG; Volin, JC; Oleksyn, J; Reich, PB. (1995). Interaction of ozone pollution and light effects on 

photosynthesis in a forest canopy experiment. Plant Cell Environ 18: 895-905. 

http://dx.doi.org/10.1111/j.1365-3040.1995.tb00598.x

Tobiessen, P. (1982). Dark opening of stomata in successional trees. Oecologia 52: 356-359. 

http://dx.doi.org/10.1007/BF00367959

Toet, S; Ineson, P; Peacock, S; Ashmore, M. (2011). Elevated ozone reduces methane emissions from 

peatland mesocosms. Global Change Biol 17: 288-296. http://dx.doi.org/10.1111/j.1365-

2486.2010.02267.x

Tong, D; Mathur, R; Schere, K; Kang, D; Yu, S. (2007). The use of air quality forecasts to assess impacts of 

air pollution on crops: Methodology and case study. Atmos Environ 41: 8772-8784. 

http://dx.doi.org/10.1016/j.atmosenv.2007.07.060

Tong, DQ; Mauzerall, DL. (2008). Summertime state-level source-receptor relationships between nitrogen 

oxides emissions and surface ozone concentrations over the continental United States. Environ Sci 

Technol 42: 7976-7984. http://dx.doi.org/10.1021/es7027636

Topa, MA; Vanderklein, DW; Corbin, A. (2001). Effects of elevated ozone and low light on diurnal and 

seasonal carbon gain in sugar maple. Plant Cell Environ 24: 663-677. 
Draft – Do Not Cite or Quote 9-187 June 2012

Torsethaugen, G; Pell, EJ; Assmann, SM. (1999). Ozone inhibits guard cell K+ channels implicated in 

stomatal opening. PNAS 96: 13577-13582. 

Tosti, N; Pasqualini, S; Borgogni, A; Ederli, L; Falistocco, E; Crispi, S; Paolocci, F. (2006). Gene expression 

profiles of O3-treated Arabidopsis plants. Plant Cell Environ 29: 1686-1702. 

http://dx.doi.org/10.1111/j.1365-3040.2006.01542.x

Turnipseed, AA; Burns, SP; Moore, DJP; Hu, J; Guenther, AB; Monson, RK. (2009). Controls over ozone 

deposition to a high elevation subalpine forest. Agr Forest Meteorol 149: 1447-1459. 

http://dx.doi.org/10.1016/j.agrformet.2009.04.001

U.S. EPA (U.S. Environmental Protection Agency). (1977). Photochemical oxidant air pollutant effects on a 

mixed conifer forest ecosystem: A progress report [EPA Report]. (EPA-600/3-77-104). Corvallis, OR. 

http://nepis.epa.gov/Exe/ZyPURL.cgi?Dockey=20015IYP.txt

U.S. EPA (U.S. Environmental Protection Agency). (1978a). Air quality criteria for ozone and other 

photochemical oxidants [EPA Report]. (EPA/600/8-78/004). Washington, DC. 

U.S. EPA (U.S. Environmental Protection Agency). (1984). Air quality criteria for ozone and other 

photochemical oxidants, volume III of V (review draft) [EPA Report]. (EPA-600/8-84-020A3). 

Research Triangle Park, NC. http://www.ntis.gov/search/product.aspx?ABBR=PB85126050

U.S. EPA (U.S. Environmental Protection Agency). (1986). Air quality criteria for ozone and other 

photochemical oxidants [EPA Report]. (EPA-600/8-84-020aF - EPA-600/8-84-020eF). Research 

Triangle Park, NC. http://www.ntis.gov/search/product.aspx?ABBR=PB87142949

U.S. EPA (U.S. Environmental Protection Agency). (1996b). Air quality criteria for ozone and related 

photochemical oxidants, Vol. II of III [EPA Report]. (EPA/600/P-93/004BF). Research Triangle Park, 

NC. 

U.S. EPA (U.S. Environmental Protection Agency). (1996c). Air quality criteria for ozone and related 

photochemical oxidants, Vol. III of III [EPA Report]. (EPA/600/P-93/004cF). Research Triangle Park, 

NC. http://www.ntis.gov/search/product.aspx?ABBR=PB96185608

U.S. EPA (U.S. Environmental Protection Agency). (1996e). Review of national ambient air quality 

standards for ozone: Assessment of scientific and technical information: OAQPS staff paper [EPA 

Report]. (EPA/452/R-96/007). Research Triangle Park, NC. 

http://www.ntis.gov/search/product.aspx?ABBR=PB96203435

U.S. EPA (U.S. Environmental Protection Agency). (2006b). Air quality criteria for ozone and related 

photochemical oxidants [EPA Report]. (EPA/600/R-05/004AF). Research Triangle Park, NC. 

http://cfpub.epa.gov/ncea/cfm/recordisplay.cfm?deid=149923

U.S. EPA (U.S. Environmental Protection Agency). (2007b). Review of the national ambient air quality 

standards for ozone: Policy assessment of scientific and technical information: OAQPS staff paper 

[EPA Report]. (EPA/452/R-07/003). Research Triangle Park, NC. 

http://www.epa.gov/ttn/naaqs/standards/ozone/data/2007_01_ozone_staff_paper

U.S. Forest Service, National Park Service,, and U.S. Fish and Wildlife Service, . (2000). Federal land 

managers air quality related values work group (FLAG): Phase I report. Denver, CO: U.S. Forest 

Service. http://www.nature.nps.gov/air/permits/flag/index.cfm

Uddling, J; Hogg, AJ; Teclaw, RM; Carroll, MA; Ellsworth, DS. (2010). Stomatal uptake of O3 in aspen and 

aspen-birch forests under free-air CO2 and O3 enrichment. Environ Pollut 158: 2023-2031. 

http://dx.doi.org/10.1016/j.envpol.2009.12.001

Uddling, J; Teclaw, RM; Kubiske, ME; Pregitzer, KS; Ellsworth, DS. (2008). Sap flux in pure aspen and 

mixed aspen-birch forests exposed to elevated concentrations of carbon dioxide and ozone. Tree 

Physiol 28: 1231-1243. 

Uddling, J; Teclaw, RM; Pregitzer, KS; Ellsworth, DS. (2009). Leaf and canopy conductance in aspen and 

aspen-birch forests under free-air enrichment of carbon dioxide and ozone. Tree Physiol 29: 1367-

1380. http://dx.doi.org/10.1093/treephys/tpp070

UIUC (University of Illinois at Urbana-Champain). (2010). SoyFACE. Available online at 

http://soyface.illinois.edu/ (accessed December 8, 2010).

UNECE (United Nations Economic Commission for Europe). (1988). ECE critical levels workshop; March; 

Bad Harzburg, Germany. Geneva, Switzerland. 

UNEP (United Nations Environment Programme). (2003). Millennium Ecosystem Assessment: Ecosystems 

and human well-being: A framework for assessment. Washington, DC: Island Press. 
Draft – Do Not Cite or Quote 9-188 June 2012

Unsworth, MH; Heagle, AS; Heck, WW. (1984a). Gas exchange in open-top field chambers: I. Measurement 

and analysis of atmospheric resistances to gas exchange. Atmos Environ 18: 373-380. 

http://dx.doi.org/10.1016/0004-6981(84)90111-2

Unsworth, MH; Heagle, AS; Heck, WW. (1984b). Gas exchange in open-top field chambers: II. Resistances 

to ozone uptake by soybeans. Atmos Environ 18: 381-385. http://dx.doi.org/10.1016/0004-

6981(84)90112-4

USDA (U.S. Department of Agriculture). (2011). Ozone biomonitoring program. Available online at 

http://www.nrs.fs.fed.us/fia/topics/ozone/ (accessed January 28, 2011).

Vahisalu, T; Kollist, H; Wang, YF; Nishimura, N; Chan, WY; Valerio, G; Lamminmäki, A; Brosché, M; 

Moldau, H; Desikan, R; Schroeder, JI; Kangasjärvi, J. (2008). SLAC1 is required for plant guard cell 

S-type anion channel function in stomatal signalling. Nature 452: 487-491. 

http://dx.doi.org/10.1038/nature06608

Valkama, E; Koricheva, J; Oksanen, E. (2007). Effects of elevated O3, alone and in combination with 

elevated CO2, on tree leaf chemistry and insect herbivore performance: A meta-analysis. Global 

Change Biol 13: 184-201. http://dx.doi.org/10.1111/j.1365-2486.01284.x

Van Buuren, ML; Guidi, L; Fornale, S; Ghetti, F; Franceschetti, M; Soldatini, GF; Bagni, N. (2002). Ozoneresponse mechanisms in tobacco: Implications of polyamine metabolism. New Phytol 156: 389-398. 

http://dx.doi.org/10.1046/j.1469-8137.2002.00539.x

Van Dingenen, R; Dentener, FJ; Raes, F; Krol, MC; Emberson, L; Cofala, J. (2009). The global impact of 

ozone on agricultural crop yields under current and future air quality legislation. Atmos Environ 43: 

604-618. http://dx.doi.org/10.1016/j.atmosenv.2008.10.033

Vandermeiren, K; Black, C; Pleijel, H; de Temmerman, L. (2005). Impact of rising tropospheric ozone on 

potato: Effects on photosynthesis, growth, productivity and yield quality. Plant Cell Environ 28: 982-

996. http://dx.doi.org/10.1111/j.1365-3040.2005.01316.x

VanDingenen, R; Dentener, FJ; Raes, F; Krol, MC; Emberson, L; Cofala, J. (2009). The global impact of 

ozone on agricultural crop yields under current and future air quality legislation. Atmos Environ 43: 

604-618. http://dx.doi.org/10.1016/j.atmosenv.2008.10.033

Velikova, V; Pinelli, P; Pasqualini, S; Reale, L; Ferranti, F; Loreto, F. (2005). Isoprene decreases the 

concentration of nitric oxide in leaves exposed to elevated ozone. New Phytol 166: 419-426. 

http://dx.doi.org/10.1111/j.1469-8137.2005.01409.x

Vickers, CE; Possell, M; Cojocariu, CI; Velikova, VB; Laothawornkitkul, J; Ryan, A; Mullineaux, PM; 

Hewitt, CN. (2009). Isoprene synthesis protects transgenic tobacco plants from oxidative stress. Plant 

Cell Environ 32: 520-531. http://dx.doi.org/10.1111/j.1365-3040.2009.01946.x

Vigue, LM; Lindroth, RL. (2010). Effects of genotype, elevated CO2 and elevated O3 on aspen 

phytochemistry and aspen leaf beetle Chrysomela crotchi performance. Agr Forest Entomol 12: 267-

276. http://dx.doi.org/10.1111/j.1461-9563.2010.00475.x

Volk, M; Bungener, P; Contat, F; Montani, M; Fuhrer, J. (2006). Grassland yield declined by a quarter in 5 

years of free-air ozone fumigation. Global Change Biol 12: 74-83. http://dx.doi.org/10.1111/j.1365-

2486.2005.01083.x

Volk, M; Geissmann, M; Blatter, A; Contat, F; Fuhrer, J. (2003). Design and performance of a free-air 

exposure system to study long-term effects of ozone on grasslands. Atmos Environ 37: 1341-1350. 

Volk, M; Obrist, D; Novak, K; Giger, R; Bassin, S; Fuhrer, J. (2011). Subalpine grassland carbon dioxide 

fluxes indicate substantial carbon losses under increased nitrogen deposition, but not at elevated ozone 

concentration. Global Change Biol 17: 366-376. http://dx.doi.org/10.1111/j.1365-2486.2010.02228.x

Vollenweider, P; Woodcock, H; Kelty, MJ; Hofer, RM. (2003). Reduction of stem growth and site 

dependency of leaf injury in Massachusetts black cherries exhibiting ozone symptoms. Environ Pollut 

125: 467-480. 

Vollsnes, AV; Kruse, OMO; Eriksen, AB; Oxaal, U; Futsaether, CM. (2010). In vivo root growth dynamics of 

ozone exposed Trifolium subterraneum. Environ Exp Bot 69: 183-188. 

http://dx.doi.org/10.1016/j.envexpbot.2010.03.007

Vuorinen, T; Nerg, AM; Holopainen, JK. (2004). Ozone exposure triggers the emission of herbivore-induced 

plant volatiles, but does not disturb tritrophic signalling. Environ Pollut 131: 305-311. 

http://dx.doi.org/10.1016/j.envpol.2004.02.027

Wallin, G; Skärby, L. (1992). The influence of ozone on the stomatal and non-stomatal limitation of 

photosynthesis in Norway spruce, Picea abies (L.) Karst, exposed to soil moisture deficit. Trees Struct 

Funct 6: 128-136. http://dx.doi.org/10.1007/BF00202428
Draft – Do Not Cite or Quote 9-189 June 2012

Wang, L; He, X; Chen, W. (2009b). Effects of elevated ozone on photosynthetic CO2 exchange and 

chlorophyll a fluorescence in leaves of Quercus mongolica grown in urban area. Bull Environ Contam 

Toxicol 82: 478-481. http://dx.doi.org/10.1007/s00128-008-9606-3

Wang, X; Mauzerall, DL. (2004). Characterizing distributions of surface ozone and its impact on grain 

production in China, Japan and South Korea: 1990 and 2020. Atmos Environ 38: 4383-4402. 

http://dx.doi.org/10.1016/j.atmosenv.2004.03.067

Wang, X; Taub, DR. (2010). Interactive effects of elevated carbon dioxide and environmental stresses on root 

mass fraction in plants: A meta-analytical synthesis using pairwise techniques. Oecologia 163: 1-11. 

http://dx.doi.org/10.1007/s00442-010-1572-x

Wang, X; Zheng, Q; Feng, Z; Xie, J; Feng, Z; Ouyang, Z; Manning, WJ. (2008). Comparison of a diurnal vs 

steady-state ozone exposure profile on growth and yield of oilseed rape (Brassica napus L.) in opentop chambers in the Yangtze Delta, China. Environ Pollut 156: 449-453. 

http://dx.doi.org/10.1016/j.envpol.2008.01.027

Watanabe, M; Yamaguchi, M; Tabe, C; Iwasaki, M; Yamashita, R; Funada, R; Fukami, M; Matsumura, H; 

Kohno, Y; Izuta, T. (2007). Influences of nitrogen load on the growth and photosynthetic responses of 

Quercus serrata seedlings to O3. Trees Struct Funct 21: 421-432. http://dx.doi.org/10.1007/s00468-

007-0134-2

Weinstein, DA; Gollands, B; Retzlaff, WA. (2001). The effects of ozone on a lower slope forest of the Great 

Smoky Mountain National Park: Simulations linking an individual tree model to a stand model. Forest 

Sci 47: 29-42. 

Weinstein, DA; Laurence, JA; Retzlaff, WA; Kern, JS; Lee, EH; Hogsett, WE; Weber, J. (2005). Predicting 

the effects of tropospheric ozone on regional productivity of ponderosa pine and white fir. For Ecol 

Manage 205: 73-89. http://dx.doi.org/10.1016/j.foreco.2004.10.007

Werner, H; Fabian, P. (2002). Free-air fumigation of mature trees: A novel system for controlled ozone 

enrichment in grown-up beech and spruce canopies. Environ Sci Pollut Res Int 9: 117-121. 

Wesely, ML; Hicks, BB. (2000). A review of the current status of knowledge on dry deposition [Review]. 

Atmos Environ 34: 2261-2282. http://dx.doi.org/10.1016/S1352-2310(99)00467-7

Whitfield, CP; Davison, AW; Ashenden, TW. (1996). Interactive effects of ozone and soil volume on 

Plantago major. New Phytol 134: 287-294. http://dx.doi.org/10.1111/j.1469-8137.1996.tb04633.x

Whitfield, CP; Davison, AW; Ashenden, TW. (1997). Artificial selection and heritability of ozone resistance 

in two populations of Plantago major. New Phytol 137: 645-655. 

Wieser, G; Havranek, WM. (1995). Environmental control of ozone uptake in Larix decidua Mill: A 

comparison between different altitudes. Tree Physiol 15: 253-258. 

Wieser, G; Manning, WJ; Tausz, M; Bytnerowicz, A. (2006). Evidence for potential impacts of ozone on 

Pinus cembra L. at mountain sites in Europe: An overview. Environ Pollut 139: 53-58. 

http://dx.doi.org/10.1016/j.envpol.2005.04.037

Wilkinson, S; Davies, WJ. (2010). Drought, ozone, ABA and ethylene: New insights from cell to plant to 

community. Plant Cell Environ 33: 510-525. http://dx.doi.org/10.1111/j.1365-3040.2009.02052.x

Will, RE; Ceulemans, R. (1997). Effects of elevated CO2 concentration on photosynthesis, respiration and 

carbohydrate status of coppice Populus hybrids. Physiol Plant 100: 933-939. 

http://dx.doi.org/10.1111/j.1399-3054.1997.tb00020.x

Winner, WE; Lefohn, AS; Cotter, IS; Greitner, CS; Nellessen, J; McEvoy, LR, Jr; Olson, RL; Atkinson, CJ; 

Moore, LD. (1989). Plant responses to elevational gradients of O3 exposures in Virginia. PNAS 86: 

8828-8832. 

Wittig, VE; Ainsworth, EA; Long, SP. (2007). To what extent do current and projected increases in surface 

ozone affect photosynthesis and stomatal conductance of trees? A meta-analytic review of the last 3 

decades of experiments [Review]. Plant Cell Environ 30: 1150-1162. http://dx.doi.org/10.1111/j.1365-

3040.2007.01717.x

Wittig, VE; Ainsworth, EA; Naidu, SL; Karnosky, DF; Long, SP. (2009). Quantifying the impact of current 

and future tropospheric ozone on tree biomass, growth, physiology and biochemistry: A quantitative 

meta-analysis. Global Change Biol 15: 396-424. http://dx.doi.org/10.1111/j.1365-2486.2008.01774.x

Woo, SY; Hinckley, TM. (2005). The effects of ozone on growth and stomatal response in the F-2 generation 

of hybrid poplar (Populus trichocarpa x Populus deltoides). Biol Plantarum 49: 395-404. 

http://dx.doi.org/10.1007/s10535-005-0014-9
Draft – Do Not Cite or Quote 9-190 June 2012

Wright, GA; Lutmerding, A; Dudareva, N; Smith, BH. (2005). Intensity and the ratios of compounds in the 

scent of snapdragon flowers affect scent discrimination by honeybees (Apis mellifera). J Comp 

Physiol A Neuroethol Sens Neural Behav Physiol 191: 105-114. 

Yamaguchi, M; Watanabe, M; Iwasaki, M; Tabe, C; Matsumura, H; Kohno, Y; Izuta, T. (2007). Growth and 

photosynthetic responses of Fagus crenata seedlings to O3 under different nitrogen loads. Trees Struct 

Funct 21: 707-718. http://dx.doi.org/10.1007/s00468-007-0163-x

Yan, K; Chen, W; He, XY; Zhang, GY; Xu, S; Wang, LL. (2010). Responses of photosynthesis, lipid 

peroxidation and antioxidant system in leaves of Quercus mongolica to elevated O3. Environ Exp Bot 

69: 198-204. http://dx.doi.org/10.1016/j.envexpbot.2010.03.008

Yoshida, S; Tamaoki, M; Ioki, M; Ogawa, D; Sato, Y; Aono, M; Kubo, A; Saji, S; Saji, H; Satoh, S; 

Nakajima, N. (2009). Ethylene and salicylic acid control glutathione biosynthesis in ozone-exposed 

Arabidopsis thaliana. Physiol Plant 136: 284-298. http://dx.doi.org/10.1111/j.1399-3054.2009.01220.x

Younglove, T; McCool, PM; Musselman, RC; KahI, ME. (1994). Growth-stage dependent crop yield 

response to ozone exposure. Environ Pollut 86: 287-295. http://dx.doi.org/10.1016/0269-

7491(94)90169-4

Yuan, JS; Himanen, SJ; Holopainen, JK; Chen, F; Stewart, CN, Jr. (2009). Smelling global climate change: 

Mitigation of function for plant volatile organic compounds. Trends Ecol Evol 24: 323-331. 

http://dx.doi.org/10.1016/j.tree.2009.01.012

Yun, SC; Laurence, JA. (1999). The response of sensitive and tolerant clones of Populus tremuloides to 

dynamic ozone exposure under controlled environmental conditions. New Phytol 143: 305-313. 

Zak, DR; Holmes, WE; Pregitzer, KS. (2007). Atmospheric CO2 and O3 alter the flow of N15 in developing 

forest ecosystems. Ecology 88: 2630-2639. 

Zhang, C; Tian, HQ; Chappelka, AH; Ren, W; Chen, H; Pan, SF; Liu, ML; Styers, DM; Chen, GS; Wang, 

YH. (2007a). Impacts of climatic and atmospheric changes on carbon dynamics in the Great Smoky 

Mountains National Park. Environ Pollut 149: 336-347. 

http://dx.doi.org/10.1016/j.envpol.2007.05.028

Zhang, J; Schaub, M; Ferdinand, JA; Skelly, JM; Steiner, KC; Savage, JE. (2010a). Leaf age affects the 

responses of foliar injury and gas exchange to tropospheric ozone in Prunus serotina seedlings. 

Environ Pollut 158: 2627-2634. http://dx.doi.org/10.1016/j.envpol.2010.05.003

Zheng, F; Wang, X; Lu, F; Hou, P; Zhang, W; Duan, X; Zhou, X; Ai, Y; Zheng, H; Ouyang, Z; Feng, Z.

(2011). Effects of elevated ozone concentration on methane emission from a rice paddy in Yangtze 

River Delta, China. Global Change Biol 17: 898-910. http://dx.doi.org/10.1111/j.1365-

2486.2010.02258.x