From: Lydia Wegman/RTP/USEPA/US

To: "Frey, Nathan J." <Nathan_J._Frey@omb.eop.gov>, "Lee, Amanda I."

<Amanda_I._Lee@omb.eop.gov> Cc: Joseph

- J Dougherty/DC/USEPA/US@EPA, Amy Lamson/RTP/USEPA/US@EPA, Darryl

Weatherhead/RTP/USEPA/US@EPA Date: 12/12/2012 05:00 PM Subject: Fw: Revised Chapter 6 of PM RIA

Please find attached a revised version of Chapter 6 in track changes from the version previously sent to you. This version clarifies some of the language regarding visibility co

- benefits, including the Chestnut and Rowe study. Please let us know if you have any questions. Thanks.** * E.O. 12866 Review

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- 2 E.O. 12866_PM NAAQS_2060

- AO47_FRN_Welfare_Co

- Benefits_20121120 CHAPTER 6 WELFARE CO

- BENEFITS OF THE PRIMARY STAND ARD

6

. 1 Synopsis
T he emission reductions to meet the revised primary standard will have “welfare” co

benefits in addition to human health benefits, including changes in visibility, materials damage, ecological effects from PM deposition, ecological effects from nitrogen and sulfur emissions, vegetation effects from ozone exposure, ecological effects from mercury deposition, and climate effects. 1 Despite our goal to quantify and monetize as many of the benefits as possible for the revised primary standard, the welfare co

- benefits of the revised primary standard remain unquantified and nonmonetized in this RIA due to data, methodology, and resource limitations. Specifically, we do not have air quality model runs for the regulatory baseline and the alternative standard levels that would allow us to calculate the visibili ty co

- benefits of attaining the revised primary standard even though we have a complete methodology for estimating these co

- benefits. However, using the approach described in this chapter, we provide the results of an illustrative analysis in Appendix 6.B using the 2020 base case and 2020 control case simulation s that were used to develop the air quality ratios.

6

. 2 Introduction to Welfare Benefits
Illustrative e mission reduction strategies to attain the revised and alternative annual primary standard have numerous documented effects on environmental quality that affect human welfare. The Clean Air Act define s welfare effects to include any non

- health effects, including direct damages to property, either through impacts on material structures or by soi ling of surfaces, direct economic damages in the form of lost productivity of crops and trees, indirect damages through alteration of ecosystem functions, and indirect economic damages through the loss in value of recreational experiences or the existence value of important

resources. The EPA ’ s Integrated Science Assessments for Particulate Matter (hereafter, “ PM ISA ” ) (U.S. EPA, 2009b ) and ) , NO X /SO X — Ecological Criteria (U.S. EPA, 2008), as well as the Criteria Document for ozone and Ozone (U.S. EPA, 2006 201 2a ) identify numerous physical and ecological effects known to be causally linked to these pollutants. This chapter describes these individual effects and how we would quantify and monetize them if there is enough data to do

so. These welfare effects inclu de changes in visibility, materials damage, ecological effects from

1 While we understand that “welfare” can include health and non

- health effects in the economic sense, we use the term “welfare” in this RIA in the same context as the definitions in the Clean Air Act for NAAQS. In practice, welfare benefits represent non

- health effects.** * E.O. 12866 Review

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- 3 PM deposition, ecological effects from nitrogen and sulfur emissions, vegetation effects from ozone exposure, ecological effects from mercury deposition, and climate effects.

These welfare co
- benefits are associated with reductions in emissions of specific pollutants resulting from illustrative e mission reduction strategies to attain the revised and alternative annual primary standard , not the form or intent of any specific standard. Even though the primary standards are designed to protect against adverse effects to human health, the emission reductions have welfare co

- benefits in addition to the direct human health benefits.

The impacts of the illustrative emission reduction strategi es can be grouped into four categories: directly emitted PM (e.g., metals, organic compounds, dust), reductions of PM 2.5 precursors (e.g., NO x , SO x , VOCs), other ancillary reductions from the illustrative emission reduction strategies (e.g., mercury and CO 2 ) , and secondary co

- pollutant formation from PM precursors (e.g., ozone from NO x and VOCs). Regardless of the category, these emission changes are anticipated to affect ambient concentrations and deposition, and consequently affect public welfare. It is t herefore appropriate and reasonable to include all the benefits and

co
- benefits associated with these emission reductions to provide a comprehensive understanding of the likely public impacts of attaining the revised or alternative annual standards. Table 6

- 1 shows the welfare effects associated with the various pollutants (either directly or as a precursor to secondary formation of PM or ozone) that would be reduced by the illustrative emission reduction strategies to attain the revised and alternative ann ual standard.

Based on the EPA’s previous analyses, we believe the welfare co
- benefits associated with these non

- health benefit categories could be significant (U.S. EPA, 2011b). Despite our goal to quantify and monetize as many of the benefits and co

- bene fits as possible, w elfare co
benefits of the revised primary standard remain unquantified and nonmonetized in this RIA due to data, methodology, and resource limitations. Therefore, the total benefits would be larger than we have estimated

. The monetized v alue of these unquantified effects is represented by adding an unknown “ B, ” which includes both unmonetized health benefits and welfare co
benefits , to the aggregate total for the cost

- benefit comparison. These unquantified benefits and co

- benefits may be substantial, although the magnitude is highly uncertain. W e include a qualitative description of the anticipated welfare effects in this chapter to characterize the type and potential extent of those co
- benefits , as identified i n Table 6

- 2.** * E.O. 12866 Review

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- 4 Table 6

- 1. Welfare Effects by Pollutants Potentially Affected by Attainment of the PM

NAAQS

Pollutant Atmospheric Effects Atmospheric and Deposition Effects Deposition Effects

Visibility Impairment Vegetation

Injury (SO 2 ) Vegetation

Injury (Ozone) Materials

Damage Climate Ecosystem

Effects

— (Organics & Metals) Acidification

(freshwater) Nitrogen

Enrichment Mercury

Methylation

Direct PM 2.5    

NO x      

SO 2      

VOCs    

PM 10    Hg  

CO 2 

 = Welfare category affected by this pollutant.
The remainder of this chapter is organized as follows: Section 6.3 provides a qualitative discussion of the visibility co

- benefits and describes the our approach to estimat e those visibility

co
- benefits . E ven though if we do not have had the data to quantify the visibility co
- benefits using this approach for the revised primary standard , we include this information as a reference for future analyses do so . Sections 6.4 through 6.6 provide qualitative co
- benefits for the unquantified benefits categories of materials damage, climate, and ecosystem effects

. References are provided in Section 6.7. Additional information regarding technical details of the visibility co
- benefits appr oach i s provided in Appendix 6A

. The illustrative visibility co

- benefits results for the specific modeled scenario (not the revised standard scenario) are provided in Appendix 6.B

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- 5 (continued) Table 6

- 2

. Quantified and Unquantified Welfare Co

- Benefits

Benefits Categ ory Specific Effect Effect Has Been Quantified Effect Has

Been Monetized More

Information Improved Environment Reduced visibility impairment Visibility in Class 1 areas in SE, SW, and CA regions a a Section 6.3 , Appendix 6.B Visibility in Class 1 areas in other regions — a Section 6.3 , Appendix 6.B Visibility in 8 cities

— a Section 6.3 , Appendix 6.B Visibility in other residential areas

— a Section 6.3 , Appendix 6.B Reduced climate effects Global climate impacts from CO 2 — — Section 6.5, SCC TSD b Climate impacts from ozone and PM — — Section 6.5, Ozone ISA, PM ISA c Other climate impacts (e.g., other GHGs, other impacts) — — Section 6.5, IPCC c

Reduced effects on materials Household soiling

— — Section 6.4, PM ISA c Materials damage (e.g., corrosion, increased wear) — — Section 6.4, PM ISA c Reduced effects from PM deposition (metals and organics) Effects on Individual organisms and ecosystems — — Section 6.6.1, PM ISA c

Reduced vegetation and ecosystem effects from exposure to ozone Visible foliar injury on vegetation

— — Section 6.6.4, Ozone ISA c Reduced vegetation growth and reproduction — — Section 6.6.4, Ozone ISA b Yield and quality of commercial forest products and crops — — Section 6.6.4, Ozone ISA b,d Damage to urban ornamental plants — — Section 6.6.4, Ozone ISA c Carbon sequestration in terrestrial ecosystems — — Ozone ISA c

Recreational demand associated with forest aesthetics — — Ozone ISA c

Other non

- use effects Ozone ISA c Ecosystem functions (e.g., water cycling, biogeochemical cycles, net primary productivity, leaf

- gas exchange, community composition) — — Ozone ISA 2** * E.O. 12866 Review

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- 6 Table 6

- 2. Quantified and Unquantified Welfare Co

- Benefits (Cont.)

Benefits Category Specific Effect Effect Has Been Quantified Effect Has

Been Monetized More

Information Improved Environment (Cont.) Reduced effects from acid deposition Recreational fishing

— — Section 6.6.2, NO x SO x ISA b Tree mortality and decline

— — Section 6.6.2, NO x SO x ISA c Commercial fishing and forestry effects — — Section 6.6.2, NO x SO x ISA c Recreational demand in terrestrial and aquatic ecosystems — — Section 6.6.2, NO x SO x ISA c Other non

- use effects Section 6.6.2, NO x SO x ISA c Ecosystem functions (e.g., biogeochemical cycles) — — Section 6.6.2, NO x SO x ISA c Reduced effects from nutrient enrichment Species composition and biodiversity in terrestrial and estuarine ecosystems — — Section 6.6.2, NO x SO x ISA c

Coastal eutrophication

— — Section 6.6.2, NO x SO x ISA c Recreational demand in terrestrial and estuarine ecosystems — — Section 6.6.2, NO x SO x ISA c Other non

- use effects Section 6.6.2, NO x SO x ISA c Ecosystem functions (e.g., biogeochemical cycles, fire regulation) — — Section 6.6.2, NO x SO x ISA c

Reduced vegetation effects from ambient exposure to SO 2 and NO x Injury to vegetation from SO 2 exposure — — Section 6.6.2, NO x SO x ISA c Injury to vegetation from NO x exposure — — Section 6.6.2, NO x SO x ISA c Reduced ecosystem effects from exposure to methylmercury (through the role of sulfate in methylation) Effects on fish, birds, and mammals (e.g., reproductive effects) — — Section 6.2 and 6.6.3, Mercury Study RTC c,d Commercial, subsistence and recreational fishing — — Section 6.2 and 6.6.3, Mercury Study RTC c

a We quantify these co
- benefits in a n illustrative analysis using the methods discussed in this chapter for the specific modeled scenario. These results are provided in Appendix 6.B , but these results of that illustrative scenario are not an estimate of the co
- benefits for the revised primary standard

. b We assess these co
- benefits qualitative ly due to time and resource limitations for this RIA

. c We assess these co
- benefits qualitatively because we do not have sufficient confidence in available data or methods. d We assess these co
- benefits qualitatively because current evidence is only suggestive of causality or there are other significant concerns over the strength of the association.** * E.O. 12866 Review

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- 7 6

. 3 Visibility Co

- Benefits Approach

6

. 3 .1 Visibility and Light Extinction Background
The illustrative emission reduction strategies designed to attain the revised and alternative annual standards would reduce emissions of directly emitted PM 2.5 as well as precursor emissions such as NO x and SO 2 for an alternative annual standard at 11 μg/m 3

. These emission reductions would improve the level of visibility because these suspended particles and gases impair visibility by scattering and absorbing light (U.S. EPA, 2009b). Visibility is also referred to as visual air qua lity (VAQ) , 2 and it directly affects people ’ s enjoyment of a variety of daily activities (U.S. EPA, 2009b). Good visibility increases quality of life where individuals live and work, and where they travel for recreational activities, including sites of uni que public value, such as the Great Smoky Mountains National Park (U. S. EPA, 2009b). This section discusses the economic co
- benefits associated with improved visibility as a result of emission reductions associated with the revised and alternative annual standards.
Air pollution affects light extinction, a measure of how much the components of the atmosphere scatter and absorb light. More light extinction means that the clarity of visual images and visual range is reduced, all else held constant

. Light extinction is the optical characteristic of the atmosphere that occurs when light is either scattered or absorbed, which converts the light to heat. Particulate matter and gases can both scatter and absorb light. Fine particles with significant light

- extin ction efficiencies include sulfates, nitrates, organic carbon, elemental carbon, and soil (Sisler, 1996). The extent to which any amount of light extinction affects a person ’ s ability to view a scene depends on both scene and light characteristics. For

exa mple, the appearance of a nearby object (e.g., a building) is generally less sensitive to a change in light extinction than the appearance of a similar object at a greater distance. See Figure 6

- 1 for an illustration of the important factors affecting visi bility.
According to the PM ISA, there is strong and consistent evidence that PM is the overwhelming source of visibility impairment in both urban and remote areas (U.S. EPA, 2009b). After reviewing all of the evidence, the PM ISA concluded that the eviden ce was sufficient to conclude that a causal relationship exists between PM and visibility impairment.

2 We use the term VAQ to refer to the visibility effects caused solely by air quality conditions, excluding fog.** * E.O. 12866 Review

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- 8 Figure 6

- 1. Important Factors Involved in Seeing a Scenic Vista (Malm, 1999)
Visibility is commonly measured as either light extinction (β ext ), which is defined as the loss of light per unit of distance in terms of inverse megameters (Mm - 1 ), or using the deciview (dv) metric, which is a logarithmic function of extinction (Pitchford and Malm, 1994). Deciviews, a unitless measure of visibility, are standardized for a reference distance in such a way that one deciview corresponds to a change of about 10% in available light. 3 Pitchford and Malm (1994) characterize a change of one deciview as “ a small but perceptible scenic change under many circumstanc es. ” 4 Extinction and deciviews are both physical measures of the amount of visibility impairment (e.g., the amount of “ haze ” ), with both extinction and deciview increasing as the amount of haze increases. Using the relationships derived by Pitchford and Ma lm (1994),

3 Note that deciviews are inver sely related to visual range, such that a decrease in deciviews implies an increase in visual range (i.e., improved visibility). Conversely, an increase in deciviews implies a decrease in visual range (i.e., decreased visibility). Deciview, in effect, is a measure of the lack of visibility. 4 An instantaneous change of less than 1 deciview (i.e., less than 10% in the light extinction budget) represents a measurable improvement in visibility but may not be perceptible to the eye. The visibility co
- benefits analysis approach described in this chapter reflects annual average changes in visibility, which are likely made up of periods with changes less than one deciview and periods with changes exceeding one deciview. Annual averages appear to more closely corres pond to the economic literature relied upon for valuation of visibility changes in this analysis. T he secondary PM NAAQS uses a different averaging time than the benefits analysis (see Chapter 13)

.** * E.O. 12866 Review

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- 9 (6.1)
where VR denotes visual range (in kilometers) and β ext denotes light extinction (in Mm - 1 ). 5
Annual average visibility conditions (reflecting light extinction due to both anthropogenic and non

- anthropogenic sources) vary regionally across the U.S. and by season (U.S. EPA, 2009b). Particulate sulfate is the dominant source of regional haze in the eastern U.S

. (> 50% of the particulate light extinction) and an important contributor to haze elsewhere in the country (>20% of particulate light extinction) (U.S. EPA, 2009b). Particulate nitrate is an important contributor to light extinction in California and the upper Midwestern U.S., particularly during winter (U.S. EPA, 2009b). Smoke plumes from large wildfires dominate many of the worst haze periods in the western U.S., while Asian dust only caused a few of the worst haze episodes, primarily in the more northerly re gions of the west (U.S. EPA, 2009b). Higher visibility impairment levels in the East are due to generally higher concentrations of fine particles, particularly sulfates, and higher average relative humidity levels (U.S. EPA, 2009b). Humidity increases visi bility impairment because some particles such as ammonium sulfate and ammonium nitrate absorb water and form droplets that become larger when relative humidity increases, thus resulting in increased light scattering (U.S. EPA, 2009b).
Reductions in air pol lution from implementation of various programs associated with the Clean Air Act Amendments of 1990 (CAAA) provisions have resulted in substantial improvements in visibility, and will continue to do so in the future. Because trends in haze are closely asso ciated with trends in particulate sulfate and nitrate due to the simple relationship between their concentration and light extinction, visibility trends have improved as emissions of SO 2 and NO x have decreased over time due to air pollution regulations suc h as the Acid Rain Program (U.S. EPA, 2009b). For example, Figure 6

- 2 shows that visual range increased nearly 50% in the eastern U.S. since 1992. 6 The EPA’s r ecent regulations such as the Cross

- State Air Pollution Rule (U.S. EPA, 2011 c ) and the Mercury an d Air Toxics Standard (U.S. EPA, 2011 d ) are anticipated to reduce SO 2 emissions down to 2 million tons nationally, which would lead to substantial further improve ment in visibility levels in the Eastern U.S

. Calculated from light extinction efficiencies fr om Trijonis et al. (1987, 1988), annual average visual range under natural conditions in the East is estimated to be 150 km ± 45 km (i.e., 65 to 120 miles) and 230 km ± 35 km (i.e., 120 to 165 miles) in the West (Irving, 1991). Figure 6

- 2 reflects the aver age

5 It has been noted that, for a given deciview value, there can be many different visual ranges, depending on the other factors that affect visual range

— such as light angle and altitude. See Appendix 6a for more detail. 6 In Figure 6

- 2, the “ best days ” are defined as the best 20% of days, the “ mid

- range days ” are defined as the middle 20%, and the “ worst days ” are defined as the worst 20% of days (IMPROVE, 2010).** * E.O. 12866 Review

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- 10 trends in visual ranges at select monitors in the eastern and western areas of the U.S. since 1992 using data from the IMPROVE monitoring network (U.S. EPA (2008) updated; IMPROVE (2010)). As an illustration of the improvements in visibility att ributab le to the CAAA, Figure 6

- 3 depicts the modeled improvements in visibility associated with all the CAAA provisions in 2020 compared to a counterfactual scenario without the CAAA (U.S. EPA, 2011b). While visibility trends have improved in most National Parks , the recent data show that these areas continue to suffer from visibility impairment beyond natural background levels (U.S. EPA, 2009b).

Figure 6

- 2. Visibility in Selected National Parks and Wilderness Areas in the U.S., 1992

–

20 08 a,b (Source: U.S. EPA (2008) updated, IMPROVE (2010))** * E.O. 12866 Review

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- 11 Figure 6

- 3. Estimated Improvement in Annual Average Visibility Levels Associated with the

CAAA Provisions in 2020 Source: U.S. EPA, 2011b 7

7 It is important to note that visibility levels shown in these maps were modeled differently than the modeling conducted for this analysis using an earlier method that we would currently use , including coarser grid resolution (i.e., 36 km instead of 12 km). Please In addition, please note that these maps present annual average visibility levels , which are different than the short

- term averages being co nsidered for the secondary standard

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- 12 6. 3

. 3 Quantifying Light Extinction for Assessing Visibility Co

- benefits For this RIA, we do not have air quality model runs for the regulatory baseline and the alternative standard levels that would allow us to calculate the visibility co

- benefits of attaining the revised primary standard. However, we provide an illustrative a nalysis in Appendix 6.B using the 2020 base case and 2020 control case simulations that were used to develop the air quality ratios. 8 In our approach , we generate light extinction estimates using the CMAQ model in conjunction with the IMPROVE (Interagency Monitoring of Protected Visual Environments) algorithm that estimates light extinction as a function of PM concentrations and relative humidity levels (U.S. EPA, 2009b). 9 The procedure for calculating light extinction associated with the revised and alternative annual s tandards is described in detail in Chapter 3 of th is RIA. In addition, Appendix 6.A describes how the spatial resolution of the light extinction estimates would be adjusted in our approach

. It is important to note that the light extinct ion estimates used in our approach represent annual averages, which is different from the averaging times currently being considered for the secondary PM NAAQS. While the annual averages are influenced by days with extremely impaired visibility, the light extinction data is not sufficient to provide higher temporal resolution than quarterly averages. While we suspect that the most impaired days would have disproportionately improved visibility from the emission reductions associated with attaining the revis ed or alternative primary standards, we are not able to quantify those impacts. These data gaps result in an underestimate of visibility co
- benefits associated with extreme days. We recognize that recent advice from the Science Advisory Board’s Advisory Co uncil on Clean Air Compliance Analysis ( SAB

- Council ) recommends estimating visibility co
- benefits considering daytime visibility on days with severe impairment (U.S. EPA

- SAB, 2010 a ), but the available data and valuation studies do not allow such fine tempo ral resolution.

8 These simulations are described in Chapter 3. 9 According to the PM ISA, the IMPROVE algorithm performs reasonably well despite its simplicity (U.S. EPA, 2009b).** * E.O. 12866 Review

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- 13 While our approach is a substantial improvement in the methods to estimat e light extinction nationally, we are still developing a method to estimate coarse particle concentrations for the entire continental U.S. for estimating light extinction. As an interim solution, our approach includes sensitivity analyses to show the potent ial impact of omitting coarse particles from the light extinction estimates for recreational and residential visibility . For the s e sensitivity analys e s, we selected the levels of coarse particles to represent the full range of possible annual concentration s from a recent report on the IMPROVE monitoring network (Debell et al., 2006). Specifically, for these sensitivity analyses, we assume four levels of coarse particles: no coarse particles, 5 μg/m 3 nationwide, 15 μg/m 3 in the Southwest with 5 μg/m 3 in the rest of the country, and 15 μg/m 3 in the Southwest with 8 μg/m 3 in the rest of the country. 10 In Table 6

- 3, we provide a qualitative assessment of how key assumptions in the estimation of light extinction would affect the visibility co
- benefits.

Table 6

- 3

. Key Assumptions in the Light Extinction Estimates Affecting the Visibility Co

Benefits Approach a

Key Assumption Direction of Bias Magnitude of Effect The light extinction estimates are annual averages to correspond with the valuation studies. People may value large changes to the haziest days differently than small changes to many days. We assume that annual average light extinction is the most appropriate temporal scale for estimating visibility benefits. Potential Underestimate Medium

Coarse particles are a component of light extinction, but we were unable to include coarse particles in the light extinction estimates. We provide sensitivity analyses with up to 15 μg/m 3 in the Southwest and 8 μg/m 3 in the rest of the country. Potential Overestimate Very Low

a A description of the classifications for magnitude of effects can be found in Appendix 5.B of this RIA

.

10 We define “ Southwest ” for this sensitivity analysis to be the states of California, Nevada, Utah, Arizona, New Mexico, Colorado, and Texas.** * E.O. 12866 Review

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- 14 6

. 3

. 3 Visibility Valuation Overview In the Clean Air Act Amendments of 1977, the U.S. Government recognized visibility ’ s value to society by establishing a national goal to protect national parks and wilderness areas from visibility impairment caused by manmade pollution. 11 Air pollution impairs visibility in both residential and recreational settings, and an individual ’ s willingness to pay (WTP ) to improve visibility differs in these two settings. Benefits of residential visibility relate to the impact of visibility changes on an individual ’ s daily life (e.g., at home, at work, and while engaged in routine recreational activities). Benefits of r ecreational visibility relate to the impact of visibility changes manifested at parks and wilderness areas that are expected to be experienced by recreational visitors.
Both recreational and residential visibility benefits consist of use values and nonuse values. Use values include the aesthetic benefits of better visibility, improved road and air safety, and enhanced recreation in activities like hunting and birdwatching. Nonuse values are based on a belief that the environment ought to exist free of human

- induced haze. This includes the value of better visibility for use by others now and in the future (bequest value). Nonuse values may be more important for recreational areas, particularly national parks and
monuments.
The relationship between a household ’ s WTP and changes in visibility can be derived from a number of contingent valuation (CV) studies published in the peer

- reviewed economics literature. The studies used in the approach to estimate the residential and recreational visibility co
- benefits ass ociated with the revised and alternative annual standards are described in the following sections. In addition to CV studies, hedonic valuation studies ( Beron et al., 2001, 2004) also demonstrate that visibility has value, but we are unable to apply these valuation estimates in the context of estimating the visibility co
- benefits associated with national regulations that reduce air pollution (Leggett and Neumann, 2004)

. 6. 3

. 3.1 Visibility Valuation Approach
In our approach , we assume that individuals value visibility for aesthetic reasons rather than viewing visibility as a proxy for other impacts associated with air pollution, such as health or ecological improvements. Some studies in the literature indicate that individuals may have difficulty distinguishi ng visibility from other aspects of air pollution (e.g., McClelland et al., 1993; Chestnut and Rowe, 1990c; Carson, Mitchell, and Rudd, 1990). Because visual air quality is inherently multi

- attribute, it is a challenge for all visibility valuation studies to isolate the

11 See Section 169(a) of the Clean Air Act.** * E.O. 12866 Review

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- 15 value of visibility from the collection of intertwined benefits. Each study used in our approach attempts to isolate visibility from other effect categories, but the different studies take different approaches (U.S. EPA, 2009b). 12 Because we believe that residual potential for double

counting visibility and health effects is relatively minimal, we would not further adjust the visibility co

- benefits to account for potentially embedded health effects beyond what the studies have already done. Ho wever, t he degree to which the studies were successful in convincing respondents to focus solely on visibility is unclear
Similarly, it is important to try to distinguish residential visibility from recreational visibility co
- benefits, specifically whether these can these can be treated as distinct and additive benefit categories based on the available literature. In our approach , we assume that residential and recreational visibility co
- benefits are distinct and separable. It is conceivable that respondents to the recreational visibility survey may have partially included values for their own residential visibility when evaluating changes at national parks and wilderness areas located in their region of the country. In our app roach, we take care to minimize the number of overlapping areas and their contributions. Specifically, we believe that the potential for double

counting recreational and residential visibility is minimal for several reasons. First, in our approach, we only include a subset of areas in the primary estimates of recreational and residential visibility co
- benefits, which overlap in only a few places. 13 Second, a number of the overlapping counties are wilderness areas, which would contribute little to the overall monetized visibility co

- benefits due to low visitation rates, rather than highly visited national parks. For example, Los Angeles County is home to the San Gabriel Wilderness Area, which has 10 thousand annual visitors (NPS, 2008). If we were to exclude t he residential visibility co
benefits that accrue to 10 million residents in Los Angeles County and only include the very small recreational visibility co
- benefits for the wilderness area, we would be substantially biasing the overall estimates downward. F or these reasons, we believe that the potential for double

- counting is minimal.
In the next sections, we describe the methodology and limitations of the recreational and residential visibility approach . Consistent with the health benefits analysis, the mon etized visibility co
- benefits would be adjusted for inflation and income growth. These co
- benefits would be specific to the analysis year , and as population and income increase over time, these

12 See Leggett and Neumann (2004) for a more detailed discussion of this issue. 13 As describ ed in detail in Sections 6.3.3 and 6.3.4, our approach includes only a subset of visibility co
- benefits in the primary main benefits estimates, while providing the rest of the visibility co
- benefits in sensitivity analyses.** * E.O. 12866 Review

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- 16 co
- benefits would be expected to increase each year for the sa me incremental change in light extinction.

6

. 3

. 4 Recreational Visibility

6

. 3

. 4 .1 Methodology
Our approach for estimating recreational visibility co
- benefits is well

- establis hed and has been used in numerous analyses by the EPA (U.S. EPA, 1999; 2005 ; 20 06 ; 2010; 2011b). In our approach , recreational visibility co
- benefits apply to Class 1 areas, such as National Parks and Wilderness Areas. 14 Although other recreational settings such as National Forests, state parks, or even hiking trails or roadside areas hav e important scenic vistas, a lack of suitable economic valuation literature to identify these other areas and/or a lack of visitation data prevents us from generating estimates for those recreational vista areas.
Under the 1999 Regional Haze Rule (64 FR 35 714), states are required to set goals develop long

- term strategies to improve visibility in Class 1 areas, with the goal of achieving natural background visibility levels by 2064. In conjunction with the U.S. National Park Service (NPS), the U.S. Forest S ervice (USFS), other Federal land managers, and State organizations in the U.S., the EPA has supported visibility monitoring in national parks and wilderness areas since 1988. The monitoring network known as IMPROVE includes 156 sites that represent the Cl ass 1 areas across the country (U.S. EPA, 2009b). 15 The IMPROVE monitoring network measures fine particles, coarse particles, and key PM 2.5 constituents that affect visibility, such as sulfate, nitrate, organic and elemental carbon, soil dust, and several other elements. Figure 6

- 4 identifies where each of these parks are located in the U.S.

14 Hereafter referred to as Class 1 areas, which are defined as areas of the country such as national parks, national wilderness areas, and national monuments that have been set aside under Section 169(a) of the Clean Air Act to receive the most stringent degree of air quality protection. Mandatory Class 1 federal lands fall under the jurisdiction of three federal agencies, the National Park Service, the Fish and Wildlife Service, and the Forest Service. EPA has designated 156 areas as mandatory Class 1 federal areas for visibility protecti on, including national parks that exceed 6,000 acres and wilderness areas that exceed 5,000 acres (40 CFR § 81.400 ). 15 The formula used to estimate light extinction from concentrations of PM constituents and relative humidity is referred to as the IMPROVE algorithm.** * E.O. 12866 Review

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- 17 Figure 6

- 4. Mandatory Class 1 Areas in the U.S.

F or recreational visibility, the EPA has determined that only one existing study provides adequate monetary estimates of the value of changes in recreational visibility: relies upon a contingent valuation (CV) survey conducted in 1988 ( Chestnut and Rowe , 199 0a; 1990b ). ) to estimate the recreational visibility co

- benefits

. Although there are several other studies in the literature on recreational visibility valuation, they are even older and use less robust methods. In the EPA ’ s judgment , despite the inherent limitations in the survey , the Chestnut and Rowe study contains many of the elements of a valid CV study and is sufficiently reliable to serve as the basis for monetary estimates of the co
- benefits of visibility changes in recreational areas

. e E ven though this specific report has not been published in a peer

- reviewed journal

. , these data were subjected to peer review as part of the National Acid Precipitation Assessment Program (NAPAP) and have been applied to estimate visibility benefits in several peer

- re viewed journal articles (Chestnut and Dennis, 1997; Burtraw et al., 1998; Delucchi et al, 2002; Chestnut and Mills, 2005) , as well as several peer reviewed in reports ( NAPAP, 1991; SAMI, 2002 ; U.S.** * E.O. 12866 Review

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- 18 EPA, 1999b; 2011b )

. 16 This study serves as an essential input to our approach for estimating the co
- benefits from improving recreational visibility.

In our approach , we assume that the household WTP is higher if the Class 1 recreational area is located close to the person ’ s home (i.e., in the same region of the country). People appear to be willing to pay more for visibility improvements at parks and wilderness areas that are in the same region as their household than at those that are not in the same region as their househol d (Chestnut and Rowe, 1990a, 1990b). This is plausible, because people are more likely to visit, be familiar with, and care about parks and wilderness areas in their own part of the country. However, studies have also found many people who had never visite d and never planned to visit the parks still had positive values for visibility improvements in those locations (Chestnut and Rowe, 1990b).
The Chestnut and Rowe survey measured the demand for visibility in Class 1 areas managed by the NPS in three broad r egions of the country: California, the Colorado Plateau (Southwest), and the Southeast. 17 Respondents in five states were asked about their WTP to protect national parks or NPS

- managed wilderness areas within a particular region. The survey used photographs reflecting different visibility levels in the specified recreational areas. The authors used the survey data to estimate household WTP values for improved visibility in each region.

16 In 1999 , as part of The Benefits and Costs of the Clean Air Act, 1990

– 2010 (U.S. EPA, 1999b), the EPA asked the SAB

- Council whether they concurred with the EPA’s SAB stated, “ proposed use of the WTP values from Chestnut and Rowe (1990a,b) to support the main (“primary”) benefit estimate. The SAB

- Council responded, “[t]he Council believes that EPA should rely on the peer

- reviewed literature as the basis for its regulatory analyses. Although many members of the Council believe that the Chestnu t and Rowe study is the best available [ study for , we believe that the general principle should be adhered to.” ” (U.S. EPA

- SAB, 1999). In response to this comment, the EPA clarified to the SAB

- Council that the visibility valuation functions and the sour ces of estimates for the coefficients for the functions were originally developed as part of NAPAP and were subjected to peer review as part of that program (Chestnut and Rowe, 1990c). The Integrated Assessment Report , including the data contained therein, received several rounds of public and peer review (NAPAP, 1991). After this clarification, the SAB

- Council identified no concerns in their final review letter with the EPA using the Chestnut and Rowe data to quantify recreational visibility valuation ]” ben efits in the main benefits estimate (U.S. EPA

- SAB, 1999 b ). In July 2010 , the SAB as part of The Benefits and Costs of the Clean Air Act 1990 to 2020 (U.S. EPA, 2011b), the SAB

Council stated that the studies were dated, but the EPA “used what the Council u nderstands to be the only relevant studies” (U.S. EPA

- SAB, 2010). 2010a). In addition, the SAB

- Council identified no concerns with quantifying recreational visibility benefits using these data in the main benefits estimate (U.S. EPA

- SAB, 2010a). 17 The Colo rado Plateau (Southwest) region is defined as the states of Colorado, New Mexico, Arizona, and Utah. The Southeast region is defined as the states of West Virginia, Virginia, North Carolina, South Carolina, Georgia, Florida, Alabama, Mississippi, Louisiana , Tennessee, and Kentucky. The California region includes the state of California and one wilderness area in Nevada.** * E.O. 12866 Review

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- 19 The separate regions were developed to capture differences in household W TP values based on proximity to recreational areas. Chestnut (1997) also concluded that, for a given region, a substantial proportion of the WTP is attributable to one specific park within the region. This so called “ indicator park ” is the most well

- known and frequently visited park within a particular region. The indicator parks for the three studied park regions are Yosemite National Park for the California region, the Grand Canyon National Park for the Southwest region, and Shenandoah National Park for t he Southeast region. In accordance with the methodology in Chestnut (1997), our approach calculates the benefits from households for a particular region for a given change in visibility at a particular Class 1 area. In theory, summing benefits from househo lds in all regions would yield the total monetary benefits associated with a given visibility improvement at a particular park, which could then be summed with other parks and regions to estimate national benefits. Because recreational visibility benefits may reflect the value an individual places on visibility improvements regardless of whether the person plans to visit the park, all households in the U.S. are assumed to derive some benefit from improvements to Class 1 areas.

To value recreational visibili ty improvements associated with its rulemakings, the EPA developed a valuation WTP equation function based on the baseline of visibility, the magnitude of the visibility improvement, and household income. This function requires light extinction estimates measured as visual range. The behavioral parameters of this equation were taken from an analysis of the survey described in Chestnut and Rowe (1990a, 1990b). These parameters were used to calibrate WTP for the visibility changes resulting from this rule. 18 As an example, household WTP for a visibility improvement at a park in its region takes the following form:

(6.2)

where: i indexes region, k indexes park,
m = household income, ρ = shape parameter (0. 1),

18 The parameters for each region are available in Appendix 6a of this RIA.** * E.O. 12866 Review

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- 20 γ = parameter corresponding to the visibility at in

- region parks,
Q 0 = starting visibility, and
Q 1 = visibility after change.
As discussed in more detail in Appendix 6 .A of this RIA, our approach to valuing recreational visibility changes is an application of the Constant Elasticity of Substitution (CES) utility function approach and is based on the preference calibration method developed by Smith, Van Houtven, and Pattanayak (2002). 19 Ava ilable evidence indicates that households are willing to pay more for a given visibility improvement as their income increases (Chestnut, 1997). Using the income elasticity calculated by Chestnut (1997), the recreational visibility benefits assume a 1% inc rease in income is associated with a 0.9% increase in WTP for a given change in visibility. WTP responses reported in Chestnut and Rowe (1990a, 1990b) were also region

- specific, rather than park

- specific. As visibility improvements are not constant across all parks in a region, we must infer park

- specific visibility parameters in order to calculate WTP for projected visibility changes. As the quantity and quality of parks differs between regions, we apportion the regional WTP parameters based on relative vi sitation rates at the different parks, because this statistic likely captures both park quality (more people visit parks with more desirable attributes, so collective WTP is likely higher) and quantity (more people visit parks in a region if the parks are more numerous, so collective WTP is likely higher). 20 We also adjust the

co
- benefits for inflation and growth in real income.
Recreational visibility co
- benefits can be calculated as the sum of the household WTPs for changes in light extinction. We assume t hat each household is valuing the first or only visibility change that occurs in a particular area. The co
- benefits at particular areas can be calculated by assuming that the subset of visibility changes of interest is the first or the only set of changes being valued by households. Estimating benefit components in this way will yield slightly upwardly biased estimates of co
- benefits , because disposable in come is not reduced by

19 The Constant Elasticity of Substitution utility function has been chosen for use in this analysis due to its flexibility when illustrating the degree of substitutability present in various economic relationships (in this case, the tradeoff between income and improvements in visibility). 20 We use 2008 park visitation data from the National Park Service Statistical Abstracts (NPS, 2008), as this is the most current data available. Where the data for a particular park was not representative of normal visitation rates at that park (for example due to fire damage that occ urred during that year), we substitute data from the prior year. We use 1997 visitation data for those wilderness areas not included in the National Park Service Statistical Abstracts, as more current data is not readily available. As visitation rates for Wilderness Areas are small compared to visitation rates in National Parks, the inaccuracies generated by using 1997 data are likely to also be small.** * E.O. 12866 Review

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- 21 the WTPs for any prior visibility improvements. The upward bias should be extremely small, however, because all of the WTPs for visibility changes are very small relative to income.
In our approach, t he primary estimate for recreatio nal visibility only includes co
- benefits for 86 Class 1 areas in the original study regions (i.e., California, the Southwest, and the Southeast). 21 These co
- benefits reflect the value to households living in the same region as the Class 1 area as well as va lues for all households in the United States living outside the state containing the Class 1 area. The Chestnut and Rowe study did not measure values for visibility improvement in Class 1 areas in the Northwest, Northern Rockies, and Rest of U.S. regions. 22 In order to obtain estimates of WTP for visibility changes for the 70 additional Class 1 areas in these non
- studied regions, we have to transfer the WTP values from the studied regions. 23 This co
- benefits transfer approach introduces additional uncertainty into the estimates. However, we have taken steps to adjust the WTP values to account for the possibility that a visibility improvement in parks within one region may not necessarily represent the same vi sibility improvement at parks within a different region in terms of environmental improvement. This may be due to differences in the scenic vistas at different parks, uniqueness of the parks, or other factors, such as public familiarity with the park resou rce. To account for this potential difference, we adjusted the transferred WTP being transferred by the ratio of visitor days in the two regions. 24 A complete description of the co
- benefits transfer method used to infer values for visibility changes in Clas s 1 areas outside the study regions is provided in Appendix 6a of this RIA.

Table 6

- 4 indicates which studied park regions we used to estimate the value in the non

studied park regions in our approach

. Figure 6

- 5 shows how the visitation rates vary across Class 1 areas and regions and indicates whether each Class 1 area is located within one of the studied regions.

Table 6

- 4. WTP for Visibility Improvements in Class 1 Areas in Non

- Studied Park Regions

Park Region Source of WTP Estimate

21 The 86 Class 1 areas in the three studied park regions represented 68% of the total visitor days to Cla ss 1 areas in 2008 (NPS, 2008). 22 The Northwest region is defined as the states of Washington and Oregon. The Northern Rockies region include s the states of Idaho, Montana, Wyoming, North Dakota, and South Dakota. The Rest of the U.S. region includes all
o ther states not included in the other 5 regions. 23 The 70 additional Class 1 areas represented 32% of the total visitor days to Class 1 areas in 2008 (NPS, 2008). 24 For example, if total park visitation in a transfer region was less than visitation in a s tudy region, transferred WTP would be adjusted downward by the ratio of the two.** * E.O. 12866 Review

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- 22 1. Northwest Benefits transfer from California
2. Northern Rockies Benefits transfer from Colorado Plateau
3. Rest of U.S. Benefits transfer from Southeast

Figure 6

- 5. Visitation Rates and Park Regions for Class 1 Areas a

a The a The colors in this map correspond to the park regions used in the valuation study and the extrapolation to parks in other regions. Red = California, light red = Northwest (extrapolated from California), blue = Colorado Plateau, light blue = Northern Rockies ( extrapolated from Colorado Plateau), green = Southeast, light green = Rest of U.S. (extrapolated from Southeast).
In a more recent study, Smith et al. (2005) conducted a contingent valuation survey that updated and expanded a portion of the 1988 survey by Chestnut and Rowe (1990). Specifically, the Smith et al. (2005) survey relied on a demographically representative national panel maintained by Knowledge Networks with 2,020 participants completing the survey. Similar to the Chestnut and Rowe survey, the Smith et al. survey assessed WTP for changes in summertime visibility using the base photograph of Shenandoah National Park. Unlike the Chestnut and Rowe survey, t he Smith et al. survey only assessed the Shenandoah National Park, did not estimate in

- region estimates of WTP, and evaluated several options for incorporating budgetary constraints into the survey. The authors concluded that WTP for recreational visibilit y is skewed and sensitive to information about budgetary constraints. We are still ys** * E.O. 12866 Review

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- 23 evaluating the potential error identified by Smith et al. (2005) regarding the visibility levels in the photographs for Shenandoah National Park in the Chestnut and Rowe sur vey (1990a,b).
Even though this survey represents several advantages over the older survey (e.g., more recent, national, demographically representative, larger sample, etc

. ), we are unable incorporate the results generated by this survey into our existing method for calculating recreational visibility co
- benefits because the survey did not account for the differences between WTP for in

- region parks and out

- of

- region parks. This omission precludes us from combining this new survey for only 1 region of the co untry with the WTP for the other regions of the country from Chestnut and Rowe (1990). Furthermore, Smith et al. (2005) provide a variety of WTP estimates reflecting different versions of the survey and different methods of summarizing the typical response , which makes it difficult to select estimates to incorporate into the recreational visibility benefits calculation.

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. 3

. 4 .2 Recreational Visibility Limitations, Caveats, and Uncertainties Our approach relies upon several data sources as inputs, including emission inventories, air quality data from models (with their associated parameters and inputs), relative humidity measurements, park information, economic data and assumptions for monetizing co
- benefits

. Each of these inputs may contain uncertainty that would affect the recreational visibility co
benefits estimates. Though we are unable to quantify the cumulative effect of all of these uncertainties in our approach , we do provide information on uncertainty based on the available data, including model eva luation 25 and sensitivity analyses to characterize major omissions ( i.e., benefits from parks in non

- studied park regions and inclusion of coarse particles). Although we strive to incorporate as many quantitative assessments of uncertainty as possible, we a re severely limited by the available data, and there are several aspects that we are only able to address qualitatively. A summary of the key assumptions including direction and magnitude of bias is provided in Table 6

- 5

. One major source of uncertainty for the estimation of recreational visibility co
- benefits is the benefits transfer process. Choices regarding the functional form and key parameters of the estimating equation for WTP for the affected population could have s ignificant effects on the magnitude of the estimates. Assumptions about how individuals respond to changes in visibility that are either very small or outside the range covered in the Chestnut and Rowe study could also affect the estimates.

25 See Chapter 4 for more information on model evaluation.** * E.O. 12866 Review

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- 24 Table 6

- 5

. Summary of Key Assumptions in Estimating R ecreational Visibility Co

- benefits a

Key Assumption Direction of Bias Potential Magnitude of Effect Chestnut and Rowe study covers parks in three regions: California, Southwest, and Southeast. Benefits to other reg ions in the U.S. are not included in the primary benefits estimate. Underestimate Medium

Benefits to other recreational settings, such as National Forests and state parks, are not included in our approach

. Underestimate Medium

- Low

Chestnut and Rowe study conducted on populations in five states. These results are applied to the entire U.S. population. Unclear Unclear

Individuals have a greater WTP for visibility changes in parks within their region. Unclear Unclear

WTP values reflect only visibili ty improvements and not overall air quality improvements. Potential Overestimate Unclear

We assume that there are 2.68 people per household. Because this estimate has been decreasing over time, this may underestimate the number of households. Potential Underestimate Medium

- Low

a A a A description of the classifications for magnitude of effects can be found in Appendix 5.B of this RIA

. Since the valuation of recreational visibility co
- benefits relies upon one study (Chestnut and Rowe, 1990a; 1990b), all of the uncertainties within that study also pertain to any analysis that uses it

. In general, the survey design and implementation reflect the period in which the Chestnut and Rowe study was con ducted. Since that time, many improvements to the design of stated preference surveys have been developed (e.g., Arrow, 1993) , but we are currently unaware of newer studies that we could incorporate into our visibility co
- benefits methodology

. Although Che stnut and Rowe still offers the best available WTP estimates, the study has a number of limitations, including:

 The vintage of the survey (late 1980s) invites questions whether the values would still be valid for current populations, or more importantly fo r our approach , future populations in 2020.

 The survey focused on visibility improvements in and around specific national parks and wilderness areas. Given that national parks and wilderness areas exhibit unique characteristics, it is not clear whether the WTP estimate obtained from this survey can be transferred to other national parks and wilderness areas, even other parks within the studied park regions, without introducing additional uncertainty.** * E.O. 12866 Review

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- 25  The survey focused only on populations in five states, so the application of the estimated values to populations outside those states requires that preferences of populations in the five surveyed states be similar to those of non

- surveyed states.

 There is an inherent difficulty in separating values expressed for visibility improvements from an overall value for improved air quality. The survey attempted to control for this by informing respondents that “ other households are being asked about visibility, human health, and vegetation protections in urban areas and at national parks in other regions. ” However, most of the respondents did not feel that they were able to segregate recreational visibility at national parks entirely from residential visibility and health effects.

 It is not clear exactly what visibility i mprovements the respondents to the survey were valuing. The WTP question asked about changes in average visibility, but the survey respondents were shown photographs of only daytime, summer conditions, when visibility is generally at its worst. It is possi ble that the respondents believed those visibility conditions held year

- round, in which case they would have been valuing much larger overall improvements in visibility than what otherwise would be the case. In our approach , the EPA assumed that respondent s provided values for changes in annual average visibility. Because most policies would result in a shift in the distribution of visibility (usually affecting the worst days more than the best days), the annual average may not be the most relevant metric f or policy analysis.

 The survey did not include reminders of possible substitutes (e.g., visibility at other parks) or budget constraints. These reminders are considered to be best practice for stated preference surveys.

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. 3

. 5 Residential Visibility

6

. 3

. 5 .1 Methodology Residential visibility co
- benefits are those that occur from visibility changes in urban, suburban, and rural areas where people live. These co
- benefits are important because s ome people living in certain urban areas may place a high value on unique scenic resources in or near these areas that are outside of Class 1 areas. For example, the State of Colorado established a local visibility standard for the Denver metropolitan area in 1990 (Ely et al., 1991). In our approach , residential visibilit y improvements are defined as those that occur specifically in Metropo litan Statistical Areas (MSAs).

In the Urban

-

focused Visibility Assessment (U.S. EPA, 2010b) and the Policy Assessment for the Review of the PM NAAQS (U.S. EPA, 2011a), several preference studies provide the** * E.O. 12866 Review

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- 26 foundation for the secondary PM NAAQS. 26 The three completed survey studies (all in the west) included Denver, Colorado (Ely et al., 1991), one in the lower Fraser River valley near Vancouver, Briti sh Columbia (BC), Canada (Pryor, 1996), and one in Phoenix, Arizona (BBC Research & Consulting, 2003). A pilot focus group study was conducted in Washington, DC on behalf of the EPA to inform the 2006 PM NAAQS review (Abt Associates Inc., 2001). Although
t hese studies indicate that some individuals considered the visual air quality associated with ambient levels of air pollution in urban areas to be unacceptable, the se studies do not provide sufficient information on which to develop monetized co
- benefits e stimates. Specifically, the public perception studies do not provide preferences expressed in dollar values, even though they do suggest that the co
- benefits associated with improving residential visibility are positive

. Studies in the peer

- reviewed litera ture support a non

- zero value for residential visibility ( e.g., Brookshire et al., 1982; Loehman et al., 1994). Furthermore, Chestnut and Rowe (1990c) conclude that residential visibility co
- benefits are likely to be at least as high as recreational visibi lity co
- benefits because of the quantity of time most people spend in and near their homes and the substantial number of people affected. In previous assessments, the EPA used a study on residential visibility valuation conducted in 1990 (McClelland et al., 1993). Consistent with advice from SAB

- Council , the EPA designated the McClelland et al. study as significantly less reliable for regulatory benefit

- cost analysis, alth ough it does provide useful estimates on the order of magnitude of residential visibility co
- benefits (U.S. EPA

- SAB, 1999). 27 In our approach for estimating residential visibility co
- benefits , we replaced the previous methodology with a new benefits transfe r approach and incorporated additional valuation studies. This new approach was developed for The Benefits and Costs of the Clean Air Act 1990 to 2020: EPA Report to Congress (U.S. EPA, 2011) 28 and reviewed by the SAB

- Council (U. S. EPA

- SAB, 2004, 2010 a, 20 10b ).
To value residential visibility improvements, the new approach draws upon information from reported in the Brookshire et al. (1979), Loehman et al. (1985) and Tolley et al. (198 4 6 )

26 For more detail about these preference studies, including information about study designs and sampling protocols, pl ease see Section 2 of the Particulate Matter Urban

- Focused Visibility Assessment (U.S. EPA, 2010b). 27 EPA ’ s Advisory Council on Clean Air Compliance Analysis noted that the McClelland et al. (1993) study may not incorporate two potentially important adjust ments. First, their study does not account for the “ warm glow ” effect, in which respondents may provide higher willingness to pay estimates simply because they favor “ good
causes ” such as environmental improvement. Second, while the study accounts for non
- response bias, it may not employ the best available methods. As a result of these concerns, the Council recommended that reside ntial visibility be omitted from the overall primary benefits estimate. (U. S. EPA

- SAB, 1999) 28 This report is also known as t he Second Prospective 812 analysis.** * E.O. 12866 Review

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- 27 studies. 29 Each of the studies provides estimates of household WTP to improve visibility conditions. While uncertainty exists regarding the precision of these older, stated

- preference residential valuation studies, we believe their results support the argument that individuals have a non

- zero value for residential visibility improvements. These studies provide primary visibility values for Atlanta, Boston, Chicago, Denver, Los Angeles, Mobile, San Francisco, and Washington D.C. 30
In accordance with Chestnut and Rowe (1990c), we utilize the WTP estimates and the associated chan ge in visual range from each study to estimate the β parameter for the eight study areas. The β parameter represents the WTP for a specific improvement in visibility in a specific location. Where studies provide multiple estimates for visual range improvem ents, we estimate β by regressing the natural log of the ratio of visual range following and prior to improvement against WTP. To express these value estimates in comparable terms across study locations, we express household WTP for a change in visual rang e in a specific MSA using the following function:
(6.3)

where:
VR 0 = mean annual visual range in miles before the improvement,

VR 1 = mean annual visual range in miles after the improvement, and β = parameter.

29 Loehman et al. ( (1994) and Brookshire et al. (1982) published results in peer

- reviewed journals based on the same underlying data we obtained from Loehman et al. ( 1985) and Brookshire et al. (1979 ) were subsequently published in peer

- reviewed journals (see Loehman et al. (1994) and Brookshire et al. (1982 ). While t he specific details need to compute visibility benefits using Tolley et al. ( 198 4 6 ) were not included in the published study (likely due to length restricti ons in the in a peer

- reviewed journal ) , , the overall work including study and survey design was subject to peer review during study development

. (see Leggett et al and Neumann , 2004 and Patterson et al., 2005 ) ). In addition, Tolley et al. subsequently published a book (Tolley and Fabian, 1988) based on this research, which notes in the preface that the methods were critiqued throughout by various external economists (Tolley et al., 1988). . The EPA does not claim that this external critique necessarily c onstituted a formal peer review process, but we provide this information for transparency regarding the review of this work. The use of these studies as the only available information to estimate residential visibility co

- benefits in the main estimate
was supported by the SAB

- Council (U.S. EPA

- SAB, 1999a, 2010a). 30 Recognizing potential fundamental issues associated with data collected in Cincinnati and Miami (e.g., see Chestnut et al. (1986) and Chestnut and Rowe (1990c), we do not include values for these cities in our analysis. The 8 MSAs where the valuation studi es were conducted represent 15% of the total US population in 2020 (U.S. Census).** * E.O. 12866 Review

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- 28 Total residential visi bility co
- benefits within a particular MSA are driven by visibility improvements, population density, and the WTP value applied. Only those people living within in an MSA are assumed to receive co
- benefits from improved residential visibility. In other

wor ds, unlike recreational visibility, we do not assume a non

- use value by people who live outside the MSA for residential visibility. Table 6

- 6 provides a summary of these valuation estimates for each study location, as well as an illustrative implied WTP value for a 10% improvement in visual range. As shown, the implied annual per

- household WTP estimates for a hypothetical 10% improvement ranges from $21 to $220, depending on the study area. It is not surprising that such a range of values exists, as these study areas all feature different landscapes and vistas, populations and prevailing visibility conditions.

Table 6

- 6

. Summary of Residential V isibility Valuation Estimates

City Study β Estimate Implied WTP for 10% Improvement in Visual Range (1990$, 1990 income) Implied WTP for 10%

Improvement in Visual Range (2006$, 2020 income)
Atlanta Tolley et al. ( 198 4 6 ) 321 $31 $72

Boston Tolley et al. ( 198 4 6 ) 398 $38 $89 Chicago Tolley et al. ( 198 4 6 ) 310 $30 $69

Denver Tolley et al. ( 198 4 6 ) 696 $66 $155 Los Angeles Brookshire et al. (1979) 94 $9 $21
Mobile Tolley et al. ( 198 4 6 ) 313 $30 $70
San Francisco Loehman et al. (1985) 989 $94 $220 Washington, DC Tolley et al. ( 198 4 6 ) 614 $59 $137

Similar to recreational visibility co
- benefits , we then incorporate preference calibration using the method developed by Smith, Van Houtven, and Pattanayak (2002), which is discussed in more detail in Appendix 6a of this RIA. This preference calibration is a change since The Benefits and Costs of the Clean Air Act 1990 to 2020: EPA Report to Congress (U.S. EPA, 2011 ) intended to address the SAB

- Council’s concern (U.S. EPA

- SAB, 2010a) regarding the inconsistency regarding household income in the estimation of recreational and residential visibility. To express these “ preference

- calibrated ” value estimates across study locations, we express household WTP for a change in visual range in a specific MSA using the following function:

(6

- 4)** * E.O. 12866 Review

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- 29 where:
m = household income, ρ = shape parameter (0.1), θ = WTP parameter corresponding to the visibility at MSA,
VR 0 = starting visibility, and
VR 1 = visibility after change.
While the primary estimate for residential visibility includes co
- benefits in only the eight MSAs included in the valuation studies, people living in other urban areas also have non

- zero values for residential visibility. For th is reason, our approach includes a sensitivity analysis for the extrapolated residential visibility in the 351 additional MSAs. 31 Because there is considerable uncertainty about the validity of this benefit transfer approach, these extrapolated co
- benefits are included in a sensitivity analysis only. This is an important distinction between the approach used in The Benefits and Costs of the Clean Air Act 1990 to 2020: EPA Report to Congress (U.S. EPA, 2011 ), where all cities were included in the total benefi ts approach. We believe that it is appropriate to deviate from the previous approach in order to be consistent with the approach used to estimate recreational visibility co
- benefits and to recognize the uncertainty associated with extrapolating beyond the studied cities. Figure 6

- 6 indicates the study cities as well as the assignment of the other MSAs to the study cities.
The degree to which the three studies were successful in convincing respondents to focus solely on visibility is unclear, as none of the three studies includes follow

- up questions necessary to investigate the issue. Furthermore, no other residential visibi lity CV studies provide evidence regarding the degree to which health effects are embedded in visibility values. Although the McClelland et al. (1991) study has a follow

- up question designed to allocate WTP across several categories, the CV question in the McClelland et al. study was focused on air pollution generally rather than visibility. As a result, we do not adjust the results from these studies to account for potentially embedded health effects.
There are many factors that could influence WTP for res idential visibility, and these factors vary across urban areas. In our approach , we utilize the benefit transfer approach developed for The Benefits and Costs of the Clean Air Act 1990 to 2020: EPA Report to Congress (U.S. EPA, 2011 ) report , but we recogni ze that there are alternative methods that we could

31 The 351 additional MSAs plus the 8 study area MSAs represent 84% of the total US population in 2020 (U.S. Census).** * E.O. 12866 Review

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- 30 have used. We assigned a valuation study area to each MSA based on two factors: geographic proximity to one of the eight study cities and elevation. Any MSA with a county elevation above 1 , 500 meters was assigned the Denver valuation instead of the nearest study area. 32 Because residents of Denver have a dramatic view of the Rocky Mountains that is rarely obstructed by trees, it is plausible that they might have a greater interest in protecting visibility t han a city without nearby mountains. The geographic proximity factor is constrained in two areas. The San Francisco valuation study is only assigned to the six counties in the San Francisco Bay area MSAs because the study is unique among the three regardin g the temporal description of visibility conditions, landscape/vistas, and prevailing weather conditions. In addition, the Los Angeles valuation was assigned to the Riverside MSA despite exceeding the elevation threshold. 33

32 Elevation data represent the county

- level maximum, whic h were calculated using the ArcGIS Spatial Analyst tool

“ Zonal Statistics ” using the geographic database HYDRO1K for North America (U.S. Geological Survey, 1997). This dataset and associated documentation are available on the Internet at DEMhttp://eros.usg s.gov/#/Find_Data/Products_and_Data_Available/gtopo30/hydro/namerica

. 33 Riverside MSA is assigned to the Los Angeles study area because a significant portion of Riverside County itself is located in the South Coast Air Quality Management District, which ca n be considered by to be part of the same regulated airshed as Los Angeles. The geographic assignment is preserved despite exceeding the elevation threshold because Riverside is adjacent to one of the study cities and this region has a particular set of lo cation

specific characteristics that set it apart from Denver.** * E.O. 12866 Review

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- 31 Figure 6

- 6

. Residential Visibility Study City Assignment

6. 3

. 5 .2 Residential Visibility Limitations, Caveats, and Uncertainties Similar to recreational visibility co
- benefits , there are many data inputs into the residential visibility co
- benefits that contribute to overall uncert ainty. Our approach includes sensitivity analyses to characterize major omissions (i.e., co
- benefits in other MSAs and coarse particles). A summary of the key assumptions including direction and magnitude of bias in our approach is provided in Table 6

- 7.

T he valuation studies relied upon for the residential visibility co
- benefits , although representing the best available estimates, have a number of limitations. These include the following:

 The survey design and implementation reflects the period in which th e surveys were conducted. Since that time, many improvements to the stated preference methods have been developed.

 The vintage of the surveys (1970s and 1980s) invites questions whether the values are still valid for current populations, or more importantl y for our approach , future populations in 2020.** * E.O. 12866 Review

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- 32  The survey focused only on populations in eight cities, so the transfer of the WTP estimates values to populations outside those cities requires that their preferences be similar to those in non

- surveyed citi es, as well as the visibility attributes be similar across study and transfer MSAs.

 There is an inherent difficulty in separating values expressed for visibility improvements from an overall value for improved air quality. The studies attempted to control for this, but most of the respondents did not feel that they were able to segregate residential visibility entirely from recreational visibility and health effects.

Table 6

- 7

. Summary of Key Assumptions in the Residential Visibility Co

- benefits a

Key Assumption Direction of Bias Magnitude of Effect Residential and recreational visibility benefits are distinct and separable. Potential Overestimate Medium

- Low

Estimates residential visibility benefits are limited to populations within the boundaries of MSAs. Areas outside of an MSA are not included in our approach

. Underestimate Low

WTP values reflect only visibility improvements and not overall air quality improvements. Potential Overestimate Medium

- Low

WTP values from studies in Atlanta, Boston, Chicago, Denver, Los Angeles, Mobile, San Francisco, and Washington D.C. can be accurately transferred to MSAs across the U.S. based on proximity and elevation Unclear Unclear

We assume that there are 2.68 peopl e per household. Because this estimate has been decreasing over time, this may underestimate the number of households. Potential Underestimate Medium

- Low

a A description of the classifications for magnitude of effects can be found in Appendix 5.B of this RIA

. 6. 3

. 5.3 Using Hedonic Economic Literature to E stimate Visibility Co

- benefits
The hedonic model assumes that consumers d o not value the consumption of a good directly, but rather value the characteristics contained within a good. In the contex t of property values, the consumer values both the physical attributes of the property (i.e., number of rooms, square footage, etc.) as well as geographic and environmental attributes (e.g., proximity to parks, visibility, etc.). Following the technique de veloped by Rosen (1974), property characteristics are regressed on the observed price of the properties within a given housing market to estimate the contribution of each characteristic to the overall price.

Numerous studies have applied hedonic methods t o estimate the willingness to pay (WTP) for air quality changes (see Smith and Huang (1995) and Boyle and Kiel (2001) for** * E.O. 12866 Review

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- 33 literature reviews), but fewer researchers have focused specifically on visibility. Studies that have estimated the WTP to improve vis ibility have focused on specific housing markets like Los Angeles (Murdoch and Thayer, 1998; Beron et al., 2001) or San Francisco (Graves et al., 1998), and it is unclear if these results would be more broadly applicable to the rest of the county. While th e literature demonstrates a link between pollutant concentrations and home

- buying behavior, it is difficult to partition the WTP for changes in pollution between health and aesthetic concerns. Murdoch and Thayer (1998) use visibility as a surr ogate for env ironmental quality and Beron et al. (2001) acknowledge that their parameter estimates likely reflect a combination of visual aesthetics and an absence of health effects. Delucchi et al. (2002) deals with this issue by partitioning WTP estimates from a hedo nic model into health and visibility components using results from previous contingent valuation studies, and find that the estimate of visibility co
- benefits is similar to estimates based on contingent valuation alone

. In 2004, the Advisory Council on Clean Air Compliance Analysis ( SAB

- Council ) recommended that the EPA evaluate the available studies addressing residential visibility and consider the possibility of using hedonic property models to estimate residential visibility co
be nefits (U.S. EPA

- SAB, 2004). In response to this recommendation, the EPA evaluated the existing economic literature, and determined that there were substantial limitations that precluded the Agency from using these studies to make inferences regarding indi viduals ’ WTP for improved visibility (Leggett and Neumann, 2004). Specifically, the literature did not provide support for the assumption that market participants are aware of the spatial variation in visibility, and consider this variation when purchasing a home, and can successfully separate visibility effects from health effects (Leggett and Neumann, 2004). This conclusion is also supported by Delucchi et al

. (2004), which found that hedonic price analysis does not capture all of the health effects of ai r pollution because homebuyers may not be fully informed about these effects .
Research since 2004 has attempted to address limitations of the hedonic method through the use of U.S. Census microdata (Bayer et al., 2009), spatial statistic al methods (Anseli n and Le Gallo, 2006; Anselin and L ozana
- Gracia , 2009; Beron et al., 2004; Kim et al., 2010 ) and more complete air quality data and information about nonattainment status (Chay and Greenstone, 2005). However, none of these studies specifically address visi bility, and they are therefore of limited use at this time. While the current state of the literature does not provide a basis for using hedonics

- based approaches , continued innovations in methodology and the further development of national, micro

- level ho using and demographic datasets may open possibilities for national

- scale hedonics

- based benefit analysis in the future.** * E.O. 12866 Review

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- 34 Regardless of whether we use hedonic models or stated preference surveys to estimate

co
- benefits arising from improved visibility, it i s important to emphasize that estimates of WTP for residential or recreational visibility improvements are not substitutes for health benefits. As previously mentioned, people often have difficulty separating their health concerns from their aesthetic conc erns when evaluating preferences for visibility, which could overestimate visibility co
- benefits if not properly controlled. However, because we use a damage

- function approach to estimate health benefits (see Chapter 5 of this RIA), the health benefits est imates are unaffected by any potential confounding with visibility preferences.

6

. 3

. 6 Discussion of Visibility Co

- benefits As described in the previous sections of this chapter, the estimation of visibility co
benefits is complex and suffers from unavoida ble limitations. While we are confident that the underlying scientific literature supports a non

- zero estimate for visibility co
- benefits attributable to emission reductions, we are less confident in the magnitude of those co
- benefits outside of previously studied locations. While acknowledging these limitations, it is important
to note that this general approach was peer included in The Benefits and Costs of the Clean Air Act 1990 to 2020: EPA Report to Congress (U.S. EPA, 2011 ) , whic h was reviewed by the SAB

Council (U.S. EPA

- SAB, 2010 a, 2010b ) , and we )

. Although the SAB

- Council highlighted concerns with the visibility approach used in the study, it did not recommend that visibility benefits be excluded. W e have addressed the SAB

- Council’s concern regarding inconsistency between estimation of residential and recreational visibility highlighted by the SAB

- Council in our approach

. However, we do not have the data to address the SAB

- Council’s concern regarding inclu sion of night

- time benefits of visibility improvements in our annual average , which may lead to an underestimation of visibility benefits

. To minimize uncertainties related to extrapolation and geographical double counting, our approach only include s a subset of monetized vi sibility co
- benefits in the core monetized visibility co

- benefits estimate to correspond with our higher level of confidence in recreational co
- benefits within the study regions and residential co
- benefits within the study cities. Although we are confident that visibility co
- benefits extend beyond these studied areas, we are less confident about the magnitude of those co
- benefits

. However, it is unclear the degree to which the visibility valuation surveys were successful in contro l l ing for potential double counting embedded health benefits.

Consistent with the approach described in the proposal RIA , we have described a revised approach for estimating visibility co
- benefits , including light extinction estimation methods, visitation data for Class 1 areas (used in extrapolating co
- benefits ) , valuation studies** * E.O. 12866 Review

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- 35 for residential visibility co
- benefits , and the benefit transfer technique for residential co
benefits

. Including residential visibility co
- benefits in the core visibility co

- benefits estimates reflec ts an evolution in our understanding of the nature and importance of the effect on public welfare from visibility impairment to a more multifaceted approach that includes non

- Class 1 areas, such as urban areas. This evolution has occurred in conjunction wi th the expansion of available PM data and information from associated studies of public perception, valuation and personal comfort and well

- being. While visibility preference studies ( Abt Associates Inc., 2001, Ely et al., 1991, Pryor, 1996, BBC Research & Consulting, 2003 ) also provide support for a non

zero benefits estimate, these surveys did not include questions that would enable monetization of those preferences.
Despite these improvements, we are limited by the available peer

- reviewed studies on visi bility co
- benefits , which have not undergone a similar expansion as the health literature. Each of these valuation studies has limitations, which are identified in the section s 6.3.4.2 and 6.3.5.2. When the SAB

- Council reviewed the visibility benefits anal ysis for The Benefits and Costs of the Clean Air Act 1990 to 2020: EPA Report to Congress (U.S. EPA, 2011 ) , they also lamented on the need for additional research to improve methods and estimates (U.S. EPA

SAB, 2010 a, 2010b ). Because of time and resource constraints, performing original research for regulatory analyses of specific policy actions is infeasible. Most importantly, we are interested in recently published national

- scale visibility valuation studies that incorporate current CV best practices, as the existing studies are limited to specific subset of geographic areas. Other important research questions that remain unresolved include identifying factors that affect valuation preferences in order to facilitate benefits transfer from the original stu dies to transfer sites across localities, disentangling health and ecosystem valuation from visibility valuation, usefulness of preference calibration, and potential role of hedonic valuation approaches. Many of these same research needs were identified by Cropper (2000), but they have yet to be addressed by the research community.
In Appendix 6.B, w e provide the results of an illustrative analysis of the visibility co
benefits associated with the 2020 base case 2020 control case simulation described in Cha pter 3 that were used to develop the air quality ratios; however, we do not have air quality model runs for the regulatory baseline and the alternative standard levels that would allow us to calculate the visibility co

- benefits of attaining the revised pri mary standard.** * E.O. 12866 Review

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- 36 6

. 4 Materials Damage Co

- benefits Building materials including metals, stones, cements, and paints undergo natural weathering processes from exposure to environmental elements (e.g., wind, moisture, temperature fluctuations, sunlight, etc. ). Pollution can worsen and accelerate these effects. Deposition of PM is associated with both physical damage (materials damage effects) and impaired aesthetic qualities (soiling effects). Wet and dry deposition of PM can physically affect materials, adding to the effects of natural weathering processes, by potentially promoting or accelerating the corrosion of metals, by degrading paints and by deteriorating building materials such as stone, concrete and marble (U.S. EPA, 2009b). The effects of PM are

exacer bated by the presence of acidic gases and can be additive or synergistic due to the complex mixture of pollutants in the air and surface characteristics of the material. Acidic deposition has been shown to have an effect on materials including zinc/galvani zed steel and other metal, carbonate stone (as monuments and building facings), and surface coatings (paints) (Irving, 1991). The effects on historic buildings and outdoor works of art are of particular concern because of the uniqueness and irreplaceabilit y of many of these objects.
The PM ISA concludes that evidence is sufficient to support a causal relationship between PM and effects on materials (U.S. EPA, 2009b). Considerable research has been conducted on the effects of air pollutants on metal surfaces due to the economic importance of these materials, especially steel, zinc, aluminum, and copper. Moisture is the single greatest factor promoting metal corrosion; however, deposited PM can have additive, antagonistic or synergistic effects. In general, SO 2 is more corrosive than NO x although mixtures of NO x , SO 2 and other particulate matter corrode some metals at a faster rate than either pollutant alone (U.S. EPA, 2008). Metal structures are usually coated by alkaline corrosion product layers and thus are subject to increased corrosion by acidic deposition. In addition, research has demonstrated that iron, copper, and aluminum

- based products are subject to increased corrosion due to pollution (Irving, 1991). Information from both the PM ISA (U.S. EPA, 2009 b) and NO x /SO x ISA (U.S. EPA, 2008) suggest that the extent of damage to metals due to ambient PM is variable and dependent upon the type of metal, prevailing environmental conditions, rate of natural weathering and presence or absence of other pollutants
In addition, the deposition of PM can cause soiling, which is the accumulation of dirt, dust, and ash on exposed surfaces such as metal, glass, stone and paint. Particles consisting primarily of carbonaceous compounds can cause soiling of commonly used bui lding materials and culturally important items such as statues and works of art. Soiling occurs when PM accumulates on an object and alters the optical characteristics (appearance). The reflectivity of** * E.O. 12866 Review

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- 37 a surface may be changed or presence of particulates m ay alter light transmission. These effects can reduce the aesthetic value of a structure or result in reversible or irreversible damage to statues, artwork and architecturally or culturally significant buildings. Due to soiling of building surfaces by PM, the frequency and duration of cleaning or repainting may be increased. In addition to natural factors, exposure to PM may give painted surfaces a dirty appearance. Pigments in works of art can be degraded or discolored by atmospheric pollutants, especially sulfates (U.S. EPA, 2008). Previous assessments estimated household soiling co
benefits based on the Manuel et al. (1982) study of consumer expenditures on cleaning and household maintenance. However, the data used to estimate household soiling damages in the Manuel et al. study is from a 1972 consumer expenditure survey and as such may not accurately represent consumer preferences in the future. In light of this significant limitation, we believe that this study cannot provide reliable estimates of the li kely magnitude of the co
benefits of reduced PM household soiling.

In order to estimate the monetized co
- benefits associated with reducing materials damage and household soiling, quantitative relationships are needed between particle size, concentration, c hemical concentrations and frequency of maintenance and repair. Such an analysis would require three steps:

1. Develop a national inventory of sensitive materials;

2. Derive concentration

- response functions that relate material damage to change in pollution concentration or deposition; and,

3. Estimate the value of lost materials and/or repair of damage.
Due to data limitations and uncertainties inherent in each of these steps, we have chosen not to include a monetized estimate of materials damage a nd household soiling in this analysis. The PM ISA concluded that there is considerable uncertainty with regard to interaction of co

- pollutants in regards to materials damage and soiling processes (U.S. EPA, 2009b). Previous benefits analyses by the EPA hav e provided quantitative estimates of materials damage (U.S. EPA, 2011b) and household soiling damage (U.S. EPA, 1999). Consistent with SAB advice (U.S. EPA, 1998), we determined that the existing data are not sufficient to calculate a reliable estimate of future year household soiling damages (U.S. EPA, 1998). These previous analyses have shown that materials damage co
- benefits are significantly smaller than the health benefits associated with reduced exposure to PM 2.5 and ozone, or even visibility co
benef its

. However, studies of materials damage to historic buildings and outdoor artwork in** * E.O. 12866 Review

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- 38 Sweden (Grosclaude and Soguel, 1994) indicate that these co
- benefits could be an order of magnitude larger than household soiling co
- benefits

. In the absence of quantifi ed co
- benefits , we provide a qualitative description of the avoided damage associated with reducing PM and PM precursor pollutants. Table 6

- 8 shows the effect of various PM 2.5 precursor pollutants and other co

- pollutants on various materials.
Table 6

- 8

. Ma terials Damaged by Pollutants Affected by this Rule (U.S. EPA, 2011b)

Pollutant Unquantified Effects / Damage to:
Sulfur oxides Infrastructural materials

— galvanized and painted carbon steel Commercial buildings

— carbonate stone, metal, and painted wood surfaces Residential buildings

— carbonate stone, metal, and painted wood surfaces Monuments

— carbonate stone and metal Structural aesthetics Automotive finishes

— painted metal Hydrogen ion and nitrogen oxides Infrastructural materials

— galvanized and painted carbon steel Zinc

- based metal products, such as galvanized steel Commercial and residential buildings

— carbonate stone, metal, and wood surfaces Monuments

— carbonate stone and metal Structural aesthetics Automotive finishes

— painted metal
Carbon dioxide Zinc

- based metal products, such as galvanized steel Formaldehyde Zinc

- based metal products, such as galvanized steel
Particulate matter Household cleanliness (i.e., household soiling)
Ozone Rubber products (e.g., tires)

6. 5 Climate Co

- benefits Actions taken by s tate and l ocal governments to implement the revised annual primary standard are likely to have implications for climate change because emission reductions ultimately implemented to meet the standard may have impacts on emissions of long

- l ived greenhouse gas (GHG) such as carbon dioxide (CO 2 ), short

- lived climate forcers such as black carbon (BC), and cooling aerosols like organic carbon (OC). Our ability to quantify the climate effects of these revised standard is limited due to lack of av ailable information on the co

controlled GHG emission reductions, the energy and associated climate gas implications of control technologies assumed in the illustrative regulatory alternative s, and remaining** * E.O. 12866 Review

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- 39 uncertainties regarding the impact of long

- lived and short

- lived climate forcer impacts on climate change. For this RIA, we discuss qualitatively the implications of potential emission reductions in warming and cooling aerosols and changes in long

- lived GHG emissions such as CO 2 for the regulatory alter native s. Implementation strategies undertaken by s tate and l ocal governments to comply with the standard s may differ from the illustrative emission reduction strategies in this RIA

. It is important to note that the net climate forcing depends on the specif ic combinations of emission reductions chosen to meet the revised standard because of the differences in warming and cooling potential of the difference pollutants

. 6

. 5 .1 Climate Effects of Short Lived Climate Forcers Pollutants that affect the energy balance of the earth are referred to as climate forcers. A pollutant that increases the amount of energy in the Earth ’ s climate system is said to exert

“ positive radiative forcing, ” which leads to warming and climate chang e. In contrast, a pollutant that exerts negative radiative forcing reduces the amount of energy in the Earth ’ s system and leads to cooling.
Long

- lived gases such as CO 2 differ from short

- lived pollutants such as BC in the length of time they remain in the atmosphere affecting the earth ’ s energy balance. Long

- lived gases remain in the atmosphere for hundreds to thousands of years. Short

- lived climate forcers (SLCFs), in contrast, remain in the atmosphere for short periods of time ranging from days to weeks. The potential to affect near

- term climate change and the rate of climate change with policies to address these emissions is gaining attention nationally and internationally ( e.g., Black Carbon Report to Congress ( U.S. EPA, 2012b ), Arctic Council Task Force , Global Methane Initiative, and Convention on Long

- Range Trans

- boundary Air Pollution of the United Nations Economic Commission for Europe). A recent United Nations Environmental Programme (UNEP) study provides the most comprehensive analysis to date of t he co
- benefits of measures to reduce SLCFs including methane, ozone, and black carbon assessing the health, climate, and agricultural co
- benefits of a suite of mitigation technologies. The report concludes that the climate is changing now, and these change s have the potential to “ trigger abrupt transitions such as the release of carbon from thawing permafrost and biodiversity loss. ” While reducing long

- lived GHGs such as CO 2 is necessary to protect against long

- term climate change, reducing SLCF gases inclu ding BC and ozone is beneficial and will slow the rate of climate change within the first half of this century (UNEP , 2011).** * E.O. 12866 Review

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- 40 6

. 5 .1.1 Climate Effects of Black Carbon Black carbon is the most strongly light

‐ absorbing component of PM 2.5 , and is formed by incomplete combustion of fossil fuels, biofuels, and biomass. The short atmospheric lifetime of BC lasting from days to weeks and the mechanisms by which BC affects climate distinguish it from long

‐ lived GHGs like CO 2 . This means that ac tions taken to reduce the BC constituents in direct PM 2.5 will have almost immediate effects on climate change. Emissions sources and ambient concentrations of BC vary geographically and temporally resulting in climate effects that are more regionally and seasonally dependent than the effects of long‐lived, well‐mixed GHGs. Likewise, mitigation actions for BC will produce different climate impacts depending on the region, season, and emission source category affected.
BC influences climate in multiple ways: directly absorbing light , reducing the reflectivity ( “ albedo ” ) of snow and ice through deposition, and interacting with clouds. BC affects climate directly by absorbing both incoming and outgoing radiation of all wavelengths. In contrast, GHGs mainly trap outgoing infrared radiation from the earth ’ s surface. Per unit of mass in the atmosphere, BC can absorb a million times more energy than CO 2 (Bond and Sun 2005). This strong absorptive capacity is the property most relevant to its potential to affect the Earth ’ s climate. W hen BC is deposited on snow and ice, it darkens the surface and decreases albedo , thereby increasing absorption and accelerating melting. Finally, BC also affects climate indirectly by altering the properties of clouds, affecting cloud reflectivity, precipitation, and surface dimming. These indirect impacts of BC are associated with all ambient particles and may lead to cooling, but are not associated with l ong
- lived well mixed GHGs.
Regional climate impacts of BC are highly variable, and sensitive regions such as the Arctic and the Himalayas are particularly vulnerable to the warming and melting effects of BC. Snow and ice cover in the Western U.S. has also been affected by BC. Specifically, deposition of BC on mountain glaciers and snow packs produces a positive snow and ice albedo effect, contributing to the melting of snowpack earlier in the spring and reducing the amount of snowmelt that normally would o ccur later in the spring and summer (Hadley et al. 2010). This has implications for freshwater resources in regions of the U.S. dependent on snow‐fed or glacier‐fed water systems. In the Sierra Nevada mountain range, Hadley et al. (2010) found BC at differ ent depths in the snowpack, deposited over the winter months by snowfall. In the spring, the continuous uncovering of the BC contributed to the early melt. A model capturing the effects of soot on snow in the western U.S. shows significant decreases in sno wpack between December and May (Qian et al., 2009). Snow water equivalent (the amount of water that would be produced by melting all the snow) is reduced 2‐50 millimeters (mm) in** * E.O. 12866 Review

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- 41 mountainous areas, particularly over the Central Rockies, Sierra Nevadas, and western Canada. A study found that biomass burning emissions in Alaska and the Rocky Mountain region during the summer can enhance snowmelt. Dust deposition on snow, at high concentrations, can have similar effects to BC (Koch et al., 2007). Similarly, a study done by Painter et al. (2007) in the San Juan Mountains in Colorado indicated a decrease in snow cover duration of 18‐35 days as a result of dust transported from non‐local desert sources.

The illustrative emission reduction strategies evaluated for this rule include reductions in BC emissions that will tend to have a beneficial cooling effect on the atmosphere. BC and elemental carbon ( EC ) (or particulate elemental carbon (PEC)) are used interchangeably in this report because the EPA t raditionally estimates EC emissions rather than BC and for the purpose of this analysis these measures are essentially equivalent.

6

. 5 .1.2 Climate Effects of Nitrates, Sulfate, and Organic Carbon (excluding BC) The composition of the total emissions mixtu re is also relevant as to whether emissions are warming or cooling to the atmosphere . P ollutants such as SO 2 , NO X , and most OC particles tend to produce a cooling influence on climate

. Exceptions include OC deposition on snow and ice , which leads to increased melting.
In addition, it is important to account for the indirect effects of all PM constituents on climate: all aerosols (including BC) affect climate indirectly by changing the reflectivity and lifetime of clouds. The net indirect effect of all aerosols is very uncertain but is thought to be a net cooling influence.

6

. 5 .1.3 Climate Effects of Ozone Ozone changes due to this revised annual standard are not estimated for this analysis but may occur due to the NO x reductions estimated. Ozone is a w ell

- known SLCF (U.S. EPA, 2006). Stratospheric ozone (the upper ozone layer) is beneficial because it protects life on Earth from the sun ’ s harmful ultraviolet (UV) radiation. In contrast, tropospheric ozone (ozone in the lower atmosphere) is a harmful air pollutant that adversely affects human health and the environment and contributes significantly to regional and global climate change. Due to its short atmospheric lifetime, tropospheric ozone concentrations exhibit large spatial and temporal variability (U.S. EPA, 2009). The discernible influence of ground level ozone on climate leads to increases in global surface temperature and changes in hydrological cycles. While reducing long

- lived GHGs such as CO 2 is necessary to protect against long

- term climate c hange, reducing SLCF gases including ozone is beneficial and will slow the rate of climate change within the first half of this century (UNEP , 2011).** * E.O. 12866 Review

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- 42 6

. 5 .1.4 SLCFs Summary and Conclusions Assessing the net climate impact of SLCFs for the illustrative emiss ion reduction strategies is outside the scope of this regulatory analysis and requires climate atmospheric modeling not undertaken due to time and resource constraints. Information about the amount of BC relative to non

- BC constituents emitted from a sourc e is important. In general, these non

BC constituents are emitted in greater volume than BC, counteracting the warming influence of BC. Qualitatively, it seems likely that BC emission reductions associated with direct ly emitted PM 2.5 reductions will be beneficial for the climate in terms of reduced radiative forcing and deposition on snow and ice. Reductions in OC, sulfates and nitrates are likely to produce warming in the atmosphere. The indirect impacts of aerosols on clouds and precipitation remain the subject of great uncertainty making it more difficult to estimate the quantitative impact of aerosol reductions on climate.

6

. 5 .2 Climate Effects of Long

- Lived Greenhouse Gases
The EPA Administrator found in 2009 that elevated concentrations of the six major GHGs, including CO 2 , endanger the public health and public welfare of current and future generations (FR 77 66496). While addressing short

- lived climate forcers can result in near

- term (and sometimes regionally specific) co
- benefits as well a s reductions in the rate of warming, reductions of long

- term warming would require mitigation of long

- lived GHGs. We are unable to quantify the impact of the illustrative emission reduction strategies for this rulemaking on long

- lived climate gases due lac k of available data

. However, State and Local governments may want to consider human health, welfare, and climate implications of regulatory strategies undertaken to implement the promulgated PM standards.

6

. 6 Ecosystem Co

- benefits and Services
The effects of air pollution on the health and stability of ecosystems are potentially very important. At present, it is difficult to measure the impact of reducing air pollution in a national scale analysis across different types of ecosystems and different pollutan t effects. Previous science assessments by the EPA (U.S. EPA, 2006a; 2008c; 2009b) have determined that air pollution can be directly linked to aquatic and terrestrial acidification, nutrient enrichment, vegetation injury, and metal bioaccumulation in animals. Ecosystem services are a useful conceptual framework for analyzing the impact of ecosystem changes on public welfare.
Ecosystem service s can be generally defined as the benefits that individuals and organizations obtain from ecosystems. The EPA has defined ecological goods and services as the “ outputs of ecological functions or processes that directly or indirectly contribute to social** * E.O. 12866 Review

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- 43 we lfare or have the potential to do so in the future. Some outputs may be bought and sold, but most are not marketed ” (U.S. EPA, 2006c). Figure 6

- 7 provides the Millennium Ecosystem Assessment ’ s schematic demonstrating the connections between the categories of ecosystem services and human well

- being. The interrelatedness of these categories means that any one ecosystem may provide multiple services. Changes in these services can affect human well

being by affecting security, health, social relationships, and access to basic material goods (MEA, 2005).

Figure 6

- 7

. Linkages between Categories of Ecosystem Se rvices and Components of Human

Well

- Being from Millennium Ecosystem Assessment (MEA, 2005)

In the Millennium Ecosystem Assessment (MEA, 2005), ecosystem s ervices are classified into four main categories:

1. Provisioning: Products obtained from ecosystems, such as the production of food and water
2. Regulating: Benefits obtained from the regulation of ecosystem processes, such as the control of climate and disease
3. Cultural: Nonmaterial benefits that people obtain from ecosystems through spiritual enrichment, cognitive development, reflection, recreation, and aesthetic experiences
4. Supporting: Services necessary for the production of all other ecosystem services, such as nutrient cycles and crop pollination** * E.O. 12866 Review

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- 44 The monetization of ecosystem services generally involves estimating the value of ecological goods and services based on what people are willing to pay (WTP) to increase ecological services or by what people are willing to accept (WTA) in compensation for reductions in them (U.S. EPA, 2006c). There are three primary approaches for estimating the monetary va lue of ecosystem services: market

- based approaches, revealed preference methods, and stated preference methods (U.S. EPA, 2006c). Because economic valuation of ecosystem services can be difficult, nonmonetary valuation using biophysical measurements and co ncepts also can be used. An example of a nonmonetary valuation method is the use of relative

- value indicators (e.g., a flow chart indicating uses of a water body, such as boatable, fishable, swimmable, etc.). It is necessary to recognize that in the analys is of the environmental responses associated with any particular policy or environmental management action, only a subset of the ecosystem services likely to be affected are readily identified. Of those ecosystem services that are identified, only a subset of the changes can be quantified. Within those services whose changes can be quantified, only a few will likely be monetized, and many will remain nonmonetized. The stepwise concept leading up to the valuation of ecosystems services is graphically depicte d in Figure 6

- 8

.

Figure 6

- 8

. Schematic of the Benefits Assessment Pro cess (U.S. EPA, 2006c)

6

. 6 .1 Ecosystem Co

- benefits for Metallic and Organic Constituents of PM Several significant ecological effects are associated with deposition of chemical constituents of ambient PM such as metals and organics (U.S. EPA, 2009b). The trace metal constituents of PM include cadmium, copper, chromium, mercury, nickel, zinc, and lead. The** * E.O. 12866 Review

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- 45 organics include persistent organic pollutants (POPs), polyaromatic hydroca rbons (PAHs) and polybromiated diphenyl ethers (PBDEs). Exposure to PM for direct effects occur via deposition ( e.g., wet, dry or occult) to vegetation surfaces, while indirect effects occur via deposition to ecosystem soils or surface waters where the dep osited constituents of PM then interacts with biological organisms. While both fine and coarse

- mode particles may affect plants and other organisms, more often the chemical constituents drive the ecosystem response to PM (Grantz et al., 2003). Ecological e ffects of PM include direct effects to metabolic processes of plant foliage; contribution to total metal loading resulting in alteration of soil biogeochemistry and microbiology, plant and animal growth and reproduction; and contribution to total organics loading resulting in bioaccumulation and biomagnification across trophic levels.
The PM ISA concludes that a causal relationship is likely to exist between deposition of PM and a variety of effects on individual organisms and ecosystems (U.S. EPA 2009b). M ost direct ecosystem effects associated with particulate pollution occur in severely polluted areas near industrial point sources (quarries, cement kilns, metal smelting) (U.S. EPA, 2009b). However the PM ISA also finds, in many cases, it is difficult to c haracterize the nature and magnitude of effects and to quantify relationships between ambient concentrations of PM and ecosystem response due to significant data gaps and uncertainties as well as considerable variability that exists in the components of PM and their various ecological effects (U.S. EPA, 2009b).
Particulate matter can adversely impact plants and ecosystem services provided by plants by deposition to vegetative surfaces (U.S. EPA, 2009b). Particulates deposited on the surfaces of leaves and n eedles can block light, altering the radiation received by the plant. PM deposition near sources of heavy deposition can obstruct stomata limiting gas exchange, damage leaf cuticles and increase plant temperatures (U.S. EPA, 2009b). Plants growing on roads ides exhibit impact damage from near

- road PM deposition, having higher levels of organics and heavy metals, and accumulate salt from road de

- icing during winter months (U.S. EPA, 2009b). In addition, atmospheric PM can convert direct solar radiation to dif fuse radiation, which is more uniformly distributed in a tree canopy, allowing radiation to reach lower leaves (U.S. EPA, 2009b). Decreases in crop yields (a provisioning service) due to reductions in solar radiation have been attributed to regional scale air pollution in other counties with especially severe regional haze (Chameides et al., 1999).
In addition to damage to plant surfaces, deposited PM can be taken up by plants from soil or foliage. Copper, zinc, and nickel have been shown to be directly tox ic to vegetation under field conditions (U.S. EPA, 2009b). The ability of vegetation to take up heavy metals is** * E.O. 12866 Review

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- 46 dependent upon the amount, solubility and chemical composition of the deposited PM. Uptake of PM by plants from soils and vegetative surfaces ca n disrupt photosynthesis, alter pigments and mineral content, reduce plant vigor, decrease frost hardiness and impair root development.

Particulate matter can also contain organic air toxic pollutants, including PAHs, which are a class of polycyclic organi c matter (POM). PAHs can accumulate in sediments and bioaccumulate in freshwater, flora and fauna. The uptake of organics depends on the plant species, site of deposition, physical and chemical properties of the organic compound and prevailing environmenta l conditions (U.S. EPA, 2009b). Different species can have different uptake rates of PAHs. For example, zucchini ( Cucurbita pepo ) accumulated significantly more PAHs than related plant species (Parrish et al., 2006). PAHs can accumulate to high enough

conc entrations in some coastal environments to pose an environmental health threat that includes cancer in fish populations, toxicity to organisms living in the sediment and risks to those (e.g., migratory birds) that consume these organisms (Simcik et al., 19 96; Simcik et al., 1999). Atmospheric deposition of particles is thought to be the major source of PAHs to the sediments of Lake Michigan, Chesapeake Bay, Tampa Bay and other coastal areas of the U.S. (Arz avus, Dickhut, and Canuel, 2001 ).

Contamination of plant leaves by heavy metals can lead to elevated concentrations in the soil. Trace metals absorbed into the plant, frequently bind to the leaf tissue, and then are lost when the leaf drops. As the fallen leaves decompose, the heavy metals are transferred into the soil (Cotrufo et al

. , 1995; Niklinska et al., 1998). Many of the major indirect plant responses to PM deposition are chiefly soil

- mediated and depend on the chemical composition of individual components of deposited PM. Upon entering the soil envi ronment, PM pollutants can alter ecological processes of energy flow and nutrient cycling, inhibit nutrient uptake to plants, change microbial community structure and, affect biodiversity. Accumulation of heavy metals in soils depends on factors such as lo cal soil characteristics, geologic origin of parent soils, and metal bioavailability. Heavy metals, such as zinc, copper, and cadmium, and some pesticides can interfere with microorganisms that are responsible for decomposition of soil litter, an important regulating ecosystem service that serves as a source of soil nutrients (U.S. EPA, 2009b). Surface litter decomposition is reduced in soils having high metal concentrations. Soil communities have associated bacteria, fungi, and invertebrates that are essen tial to soil nutrient cycling processes. Changes to the relative species abundance and community composition are associated with deposited PM to soil biota (U.S. EPA, 2009b).
Atmospheric deposition can be the primary source of some organics and metals to watersheds. Deposition of PM to surfaces in urban settings increases the metal and organic** * E.O. 12866 Review

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- 47 component of storm water runoff (U.S. EPA, 2009b). This atmospherically

- associated pollutant burden can then be toxic to aquatic biota. The contribution of atmospher ically deposited PAHs to aquatic food webs was demonstrated in high elevation mountain lakes with no other anthropogenic contaminant sources (U.S. EPA, 2009b). Metals associated with PM deposition limit phytoplankton growth, affecting aquatic trophic struc ture. Long

- range atmospheric transport of 47 pesticides and degradation products to the snowpack in seven national parks in the Western U.S. was recently quantified indicating PM

- associated contaminant inputs to receiving waters during spring snowmelt (Hag eman et al., 2006).
The recently completed Western Airborne Contaminants Assessment Project (WACAP) is the most comprehensive database on contaminant transport and PM depositional effects on sensitive ecosystems in the Western U.S. (Landers et al., 2008). In this project, the transport, fate, and ecological impacts of anthropogenic contaminants from atmospheric sources were assessed from 2002 to 2007 in seven ecosystem components (air, snow, water, sediment, lichen, conifer needles and fish) in eight core n ational parks. The study concluded that bioaccumulation of semi

- volatile organic compounds occurred throughout park ecosystems, an elevational gradient in PM deposition exists with greater accumulation in higher altitude areas, and contaminants accumulate in proximity to individual agriculture and industry sources, which is counter to the original working hypothesis that most of the contaminants would originate from Eastern Europe and Asia.
Although there is considerable data on impacts of PM on ecological receptors, few studies link ambient PM levels to observed effect. This is due, in part, to the nature, deposition, transport and fate of PM in ecosystems. Some of the difficulties in quantifying the ecosystem

co
- benefits associated with reduced PM depositi on include the following:

 PM is not a single pollutant, but a heterogeneous mixture of particles differing in size, origin and chemical composition. Since vegetation and other ecosystem components are affected more by particulate chemistry than size fracti on, exposure to a given mass concentration of airborne PM may lead to widely differing plant or ecosystem responses, depending on the particular mix of deposited particles.

 Composition of ambient PM varies in time and space and the particulate mixture may have synergistic, antagonistic or additive effects on ecological receptors depending upon the chemical species present.

 Presence of co

- pollutants makes it difficult to attribute observed effects to ecological receptors to PM alone or one component of depos ited PM.** * E.O. 12866 Review

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- 48  Ecosystem effects linked to PM are difficult to determine because the changes may not be observed until pollutant deposition has occurred for many decades. Furthermore, many PM components bioaccumulate over time in organisms or plants, making corr elations to ambient levels of PM difficult.

 Multiple ecological stressors can confound attempts to link specific ecosystem responses to PM deposition. These stressors can be anthropogenic ( e.g., habitat destruction, eutrophication, other pollutants) or nat ural ( e.g., drought, fire, disease). Deposited PM interacts with other stressors to affect ecosystem patterns and processes.

 Each ecosystem has a unique topography, underlying bedrock, soils, climate, meteorology, hydrologic regime, natural and land use hi story, and species composition. Sensitivity of ecosystem response can be highly variable in space and time. Because of this variety and lack of data for most ecosystems, extrapolating these effects from one ecosystem to another is highly uncertain.

6

. 6 .2 E cosystem Co

- benefits from Reductions in Nitrogen and Sulfur Emissions Emissions of the PM precursors, such as n itrogen and sulfur oxides occur over large regions of North America. Once these pollutants are lofted to the middle and upper troposphere, they t ypically have a much longer lifetime and, with the generally stronger winds at these altitudes, can be transported long distances from their source regions. The length scale of this transport is highly variable owing to differing chemical and meteorologica l conditions encountered along the transport path (U.S. EPA, 2008c). Secondary particles are formed from NO X and SO 2 gaseous emissions and associated chemical reactions in the atmosphere. Deposition can occur in either a wet (i.e., rain, snow, sleet, hail, clouds, or fog) or dry form (i.e., gases or particles). Together these emissions are deposited onto terrestrial and aquatic ecosystems across the U.S., contributing to the problems of acidification, nutrient enrichment, and methylmercury production as rep resented in Figures 6

- 9 and 6

- 1 0

. Although there is some evidence that nitrogen deposition may have positive effects on agricultural and forest output through passive fertilization, it is likely that the overall value is very small relative to other health and welfare effects. In addition to deposition effects, SO 2 can affect vegetation at ambient levels near pollution sources.** * E.O. 12866 Review

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- 49 NO X Atmospheric Fate and Transport

Deposition Process

Acidification

Aquatic Ecosystems Terrestrial

Ecosystems Nutrient Enrichment

Aquatic Ecosystems Terrestrial

Ecosystems** * E.O. 12866 Review

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- 50 SO 2 Atmospheric Fate and Transport

Deposition Process

Acidification

Aquatic Ecosystems Terrestrial Ecosystems Methylmercury Production

NO X Atmospheric Fate and Transport

Deposition Process

Acidification

Aquatic Ecosystems Terrestrial

Ecosystems Nutrient Enrichment

Aquatic Ecosystems Terrestrial

Ecosystems** * E.O. 12866 Review

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- 51 Figure 6

- 9

. Schematics of Ecological Effects of Nitrogen and Sulfur Deposition

Figure 6

- 1 0

. Nitrogen and Sulfur Cycling , and Interactions in the Environment Source: U.S. EPA, 2008c SO 2 Atmospheric Fate and Transport

Deposition Process

Acidification

Aquatic Ecosystems Terrestrial Ecosystems Methylmercury Production** * E.O. 12866 Review

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- 52 The atmospheric lifetimes of particles vary with particle size. Accumulation

- mode particles such as sulfates are kept in suspension by normal air motions and have a lower deposition velocity tha n coarse
- mode particles; they can be transported thousands of kilometers and remain in the atmosphere for a number of days. They are removed from the atmosphere primarily by cloud processes. Particulates affect acid deposition by serving as cloud condensat ion nuclei and contribute directly to the acidification of rain. In addition, the gas

phase species that lead to the dry deposition of acidity are also precursors of particles. Therefore, reductions in NO X and SO 2 emissions will decrease both acid depositi on and PM concentrations, but not necessarily in a linear fashion (U.S. EPA, 2008c). Sulfuric acid is also deposited on surfaces by dry deposition and can contribute to environmental effects (U.S. EPA, 2008c).

6

. 6 .2.1 Ecological Effects of Acidification
De position of nitrogen and sulfur can cause acidification, which alters biogeochemistry and affects animal and plant life in terrestrial and aquatic ecosystems across the U.S. Soil acidification is a natural process, but is often accelerated by acidifying de position, which can decrease concentrations of exchangeable base cations in soils (U.S. EPA, 2008c). Major terrestrial effects include a decline in sensitive tree species, such as red spruce ( Picea rubens ) and sugar maple ( Acer saccharum ) (U.S. EPA, 2008c)

. Biological effects of acidification in terrestrial ecosystems are generally linked to aluminum toxicity and decreased ability of plant roots to take up base cations (U.S. EPA, 2008c). Decreases in the acid neutralizing capacity and increases in inorganic aluminum concentration contribute to declines in zooplankton, macro invertebrates, and fish species richness in aquatic ecosystems (U.S. EPA, 2008c).
Geology (particularly surficial geology) is the principal factor governing the sensitivity of terrestrial and aquatic ecosystems to acidification from nitrogen and sulfur deposition (U.S. EPA, 2008c). Geologic formations having low base cation supply generally underlie the watersheds of acid

- sensitive lakes and streams. Other factors contribute to the sensiti vity of soils and surface waters to acidifying deposition, including topography, soil chemistry, land use, and hydrologic flow path (U.S. EPA, 2008c).
Aquatic Acidification. Aquatic effects of acidification have been well studied in the U.S. and elsewhere at various trophic levels. These studies indicate that aquatic biota have been affected by acidification at virtually all levels of the food web in acid sensitive aquatic ecosystems. The ISA for NO x /SO x — Ecological Criteria concluded that the evidence is sufficient to infer a causal relationship between acidifying deposition and effects on biogeochemistry** * E.O. 12866 Review

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- 53 related to aquatic ecosystems and biota in aquatic ecosystems (U.S. EPA, 2008c). Effects have been mos t clearly documented for fish, aquatic insects, other invertebrates, and algae. Biological effects are primarily attributable to a combination of low pH and high inorganic aluminum concentrations. Such conditions occur more frequently during rainfall and s nowmelt that cause high flows of water and less commonly during low

- flow conditions, except where chronic acidity conditions are severe. Biological effects of episodes include reduced fish condition factor 34 , changes in species composition and declines in a quatic species richness across multiple taxa, ecosystems and regions. These conditions may also result in direct fish mortality (Van Sickle et al., 1996). Biological effects in aquatic ecosystems can be divided into two major categories: effects on health, vigor, and reproductive success; and effects on biodiversity. Surface water with ANC values greater than 50 μeq/L generally provides moderate protection for most fish (i.e., brook trout, others) and other aquatic organisms (U.S. EPA, 2009c). Table 6

- 9 pro vides a summary of the biological effects experienced at various ANC levels.
Table 6

- 9

. Aquatic Status Categories
Category Label ANC Levels Expected Ecological Effects

Acute Concern <0 micro

equivalent per Liter (μeq/L) Near complete loss of fish populations is expected. Planktonic communities have

extremely low diversity and are dominated by acidophilic forms. The number of individuals in plankton species that are present is greatly reduced.
Severe Concern 0

– 20 μeq/L Highly sensitive to episodic acidification. During episodes of high acidifying deposition, brook trout populations may experience lethal effects. Diversity and distribution of zooplankton communities decline sharply.
Elevated Concern 20

– 50 μeq/L Fish species richness is greatly reduced (i.e., more than half of expected species can be missing). On average, brook trout populations experience sublethal effects, including loss of health, reproduction capacity, and fitness. Diversity and distribution of zooplankton communities decline

. Moderate Concern 50

– 100 μeq/L Fish species richness begins to decline (i.e., sensitive species are lost from lakes). Brook trout populations are sensitive and variable, with possible sublethal effects. Diversity and distribution of zooplankton communiti es also begin to decline as species that are sensitive to acidifying deposition are affected.
Low Concern >100 μeq/L Fish species richness may be unaffected. Reproducing brook trout populations are expected where habitat is suitable. Zooplankton communities are unaffected and exhibit expected diversity and distribution.

34 Condition factor is an index that describes the relationship between fish weight and length, and is one measure of sublethal acidification stress that has been used to quantify effects of ac idification on an individual fish (U.S. EPA, 2008f).** * E.O. 12866 Review

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- 54 A number of national and regional assessments have been conducted to estimate the distribution and extent of surface water acidity in the U.S. (U.S. EPA, 2008c). As a result, sev eral regions of the U.S. have been identified as containing a large number of lakes and streams that are seriously impacted by acidification. Figure 6

- 11 illustrates those areas of the U.S. where aquatic ecosystems are at risk from acidification.

Because a cidification primarily affects the diversity and abundance of aquatic biota, it also affects the ecosystem services that are derived from the fish and other aquatic life found in these surface waters.
While acidification is unlikely to have serious negativ e effects on, for example, water supplies, it can limit the productivity of surface waters as a source of food (i.e., fish). In the northeastern United States, the surface waters affected by acidification are not a major source of commercially raised or ca ught fish; however, they are a source of food for some recreational and subsistence fishermen and for other consumers. For example, there is evidence that certain population subgroups in the northeastern United States, such as the Hmong and Chippewa ethnic groups, have particularly high rates of self

- caught fish consumption (Hutchison and Kraft, 1994; Peterson et al., 1994). However, it is not known if and how their consumption patterns are affected by the reductions in available fish populations caused by surface water acidification.** * E.O. 12866 Review

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- 55 Figure 6

- 11. Areas Potentially Sensitive to Aquatic Acidification Source: U.S. EPA, 2008c
Inland surface waters support several cultural services, including aesthetic and educational services and recreational fishing. Recreational fishing in lakes and streams is among the most popular outdoor recreational activities in the northeastern United States. Based on studies conducted in the northeastern United States, Kaval and Loomis (2003) estimated average consumer surplus values per day of $36 for recreational fishing (in 2007 dollars); therefore, the implied total annual value of freshwater fishing in the northeastern United States was $5.1 billion in 2006. 35 For recreation days, consumer surplus value is most commonly

meas ured using recreation demand, travel cost models.
Another estimate of the overarching ecological co
- benefits associated with reducing lake acidification levels in Adirondacks National Park can be derived from the contingent valuation (CV) survey (Banzhaf e t al., 2006), which elicited values for specific improvements in acidification

- related water quality and ecological conditions in Adirondack lakes. The survey described a base version with minor improvements said to result from the program, and a

35 These estimates reflect the total value of the service, not the marginal change in the value of the service as a result of the emission reductions achieved by this rule.** * E.O. 12866 Review

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- 56 scope ver sion with large improvements due to the program and a gradually worsening status quo. After adapting and transferring the results of this study and converting the 10

- year annual payments to permanent annual payments using discount rates of 3% and 5%, the W TP estimates ranged from $48 to $107 per year per household (in 2004 dollars) for the base version and $54 to $154 for the scope version. Using these estimates, the aggregate annual benefits of eliminating all anthropogenic sources of NO x and SO x emissions were estimated to range from $291 million to $829 million (U.S. EPA, 2009c). 36
In addition, inland surface waters provide a number of regulating services associated with hydrological and climate regulation by providing environments that sustain a quatic food webs. These services are disrupted by the toxic effects of acidification on fish and other aquatic life. Although it is difficult to quantify these services and how they are affected by acidification, some of these services may be captured thro ugh measures of provisioning and cultural services.

Terrestrial Acidification. Acidifying deposition has altered major biogeochemical processes in the U.S. by increasing the nitrogen and sulfur content of soils, accelerating nitrate and sulfate leaching fr om soil to drainage waters, depleting base cations (especially calcium and magnesium) from soils, and increasing the mobility of aluminum. Inorganic aluminum is toxic to
some tree roots. Plants affected by high levels of aluminum from the soil often have r educed root growth, which restricts the ability of the plant to take up water and nutrients, especially calcium (U. S. EPA, 2008c). These direct effects can, in turn, influence the response of these plants to climatic stresses such as droughts and cold tem peratures. They can also influence the sensitivity of plants to other stresses, including insect pests and disease (Joslin et al., 1992) leading to increased mortality of canopy trees. In the U.S., terrestrial effects of acidification are best described fo r forested ecosystems (especially red spruce and sugar maple ecosystems) with additional information on other plant communities, including shrubs and lichen (U.S. EPA, 2008c). The ISA for NO x /SO x — Ecological Criteria concluded that the evidence is sufficien t to infer a causal relationship between acidifying deposition and effects on biogeochemistry related to terrestrial ecosystems and biota in terrestrial ecosystems (U.S. EPA, 2008c).
Certain ecosystems in the continental U.S. are potentially sensitive to t errestrial acidification, which is the greatest concern regarding nitrogen and sulfur deposition U.S. EPA (2008c). Figure 6

- 1 2 depicts the areas across the U.S. that are potentially sensitive to terrestrial acidification.

36 These estimates reflect the total value of the service, not the marginal change in the value of the service as a result of the emission reductions achieved by this rule.** * E.O. 12866 Review

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- 57 Figure 6

- 1 2

. Areas Potentially Sensitive to Terrestrial Acidification
Source: U.S. EPA, 2008c
Both coniferous and deciduous forests throughout the eastern U.S. are experiencing gradual losses of base cation nutrients from the soil due to accelerated leaching from acidifying deposition. This change in nutrient availability may reduce the quality of forest nutrition over the long term. Evidence suggests that red spruce and sugar maple in some areas in the eastern U.S. have experienced declining health because of this deposition. For red sp ruce, ( Picea rubens ) dieback or decline has been observed across high elevation landscapes of the northeastern U.S., and to a lesser extent, the southeastern U.S., and acidifying deposition has been implicated as a causal factor (DeHayes et al., 1999). Fig ure 6

- 1 3 shows the distribution of red spruce (brown) and sugar maple (green) in the eastern U.S.
Terrestrial acidification affects several important ecological endpoints, including declines in habitat for threatened and endangered species (cultural), decl ines in forest aesthetics (cultural), declines in forest productivity (provisioning), and increases in forest soil erosion and reductions in water retention (cultural and regulating).** * E.O. 12866 Review

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- 58 Figure 6

- 1 3

. Distribution of Red Spruce (pink) and Sugar Maple (green) in the Eastern U.S. Source: U.S. EPA, 2008c
Forests in the northeastern United States provide several important and valuable provisioning services in the form of tree products. Sugar maples are a particularly important commercial hardwood tree species, pr oviding timber and maple syrup. In the United States, sugar maple saw timber was nearly 900 million board feet in 2006 (USFS, 2006), and annual production of maple syrup was nearly 1.4 million gallons, accounting for approximately 19% of worldwide producti on. The total annual value of U.S. production in these years was approximately $160 million (NASS, 2008). 37 Red spruce is also used in a variety of products including lumber, pulpwood, poles, plywood, and musical instruments. The total removal of red spruce saw timber from timberland in the United States was over 300 million board feet in 2006 (USFS, 2006).
Forests in the northeastern United States are also an important source of cultural ecosystem services

— nonuse (i.e., existence value for threatened and en dangered species),

37 These estimates reflect the total value of the service, not the marginal change in the value of the service as a result of the emission reductions achieved by this rule.** * E.O. 12866 Review

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- 59 recreational, and aesthetic services. Red spruce forests are home to two federally listed species and one delisted species:

1. Spruce

- fir moss spider ( Microhexura montivaga )

— endangered

2. Rock gnome lichen ( Gymnoderma lineare )

— endangered

3. Virginia northern flying squirrel ( Glaucomys sabrinus fuscus )

— delisted, but important
Forestlands support a wide variety of outdoor recreational activities, including fishing, hiking, camping, off

- road driving, hunting, and wildlife viewing. Regional s tatistics on recreational activities that are specifically forest based are not available; however, more general data on outdoor recreation provide some insights into the overall level of recreational services provided by forests. More than 30% of the U.S. adult population visited a wilderness or primitive area during the previous year and engaged in day hiking (Cordell et al., 2008). From 1999 to 2004, 16% of adults in the northeastern United States participated in off

- road vehicle recreation, for an avera ge of 27 days per year (Cordell et al., 2005). The average consumer surplus value per day of off

- road driving in the United States was $25 (in 2007 dollars), and the implied total annual value of off

- road driving recreation in the northeastern United State s was more than $9 billion (Kaval and Loomis, 2003). More than 5% of adults in the northeastern United States participated in nearly 84 million hunting days (U.S. FWS and U.S. Census Bureau, 2007). Ten percent of adults in northeastern states participated in wildlife viewing away from home on 122 million days in 2006. For these recreational activities in the northeastern United States, Kaval and Loomis (2003) estimated average consumer surplus values per day of $52 for hunting and $34 for wildlife viewing ( in 2007 dollars). The implied total annual value of hunting and wildlife viewing in the northeastern United States was, therefore, $4.4 billion and $4.2 billion, respectively, in 2006 (U.S. EPA, 2009c). 38
As previously mentioned, it is d ifficult to estimate the portion of these recreational services that are specifically attributabl e to forests and to the health of specific tree species. However, one recreational activity that is directly dependent on forest conditions is fall color viewing. Sugar maple tree s, in particular, are known for their bright colors and are, therefore, an essential aesthetic component of most fall color landscapes. A survey of residents in the Great Lakes area found that roughly 30% of residents reported at least one trip in the prev ious year involving fall color viewing (Spencer and Holecek, 2007). In a separate study conducted in

38 These estimates reflect the total value of the service, not the marginal change in the value of the service as a result of the emission reductions achieved by this rule.** * E.O. 12866 Review

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- 60 Vermont, Brown (2002) reported that more than 22% of households visiting Vermont in 2001 made the trip primarily for viewing fall colors.

Two studies estim ated values for protecting high

- elevation spruce forests in the southern Appalachian Mountains. Kramer et al. (2003) conducted a contingent valuation study estimating households ’ WTP for programs to protect remaining high

- elevation spruce forests from dama ges associated with air pollution and insect infestation. Median household WTP was estimated to be roughly $29 (in 2007 dollars) for a smaller program, and $44 for the more extensive program. Jenkins et al. (2002) conducted a very similar study in seven So uthern Appalachian states on a potential program to maintain forest conditions at status quo levels. The overall mean annual WTP for the forest protection programs was $208 (in 2007 dollars). Multiplying the average WTP estimate from these studies by the t otal number of households in the seven

- state Appalachian region results in an aggregate annual range of $470 million to $3.4 billion for avoiding a significant decline in the health of high

- elevation spruce forests in the Southern Appalachian region (U.S. EPA, 2009c). 39
Forests in the northeastern United States also support and provide a wide variety of valuable regulating services, including soil stabilization and erosion control, water regulation, and climate regulation. The total value of these ecosystem services is very difficult to quantify in a meaningful way, as is the reduction in the value of these services associated with total nitrogen and sulfur deposition. As terrestrial acidification contributes to root damages, reduced biomass growth, and tree mortality, all of these services are likely to be affected; however, the magnitude of these impacts is currently very uncertain.

6

. 6 .2.2 Ecological Effects from Nitrogen Enrichment Aquatic Enrichment

. The ISA for NO x /SO x — Ecological Criteria concluded that the evidence is sufficient to infer a causal relationship between nitrogen deposition and the alteration of species richness, species composition, and biodiversity in wetland, freshwater aquatic and coastal marine ecosystems (U.S. EPA, 2008c).

One of the m ain adverse ecological effects resulting from nitrogen deposition, particularly in the Mid

- Atlantic region of the United States, is the effect associated with nutrient enrichment in estuarine waters. A recent assessment of 141 estuaries nationwide by the N ational Oceanic and Atmospheric Administration (NOAA) concluded that 19 estuaries (13%) suffered from moderately high or high levels of eutrophication due to excessive inputs of both

39 These estimates reflect the marginal value of the service for the hypothetical program described in the survey, not the marginal change in the value of the service as a result of the emission reductions achieved by this rule.** * E.O. 12866 Review

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- 61 N and phosphorus, and a majority of these estuaries are located in the co astal area from North Carolina to Massachusetts (NOAA, 2007). For estuaries in the Mid

- Atlantic region, the contribution of atmospheric distribution to total N loads is estimated to range between 10% and 58% (Valigura et al., 2001).
Eutrophication in estua ries is associated with a range of adverse ecological effects. The conceptual framework developed by NOAA emphasizes four main types of eutrophication effects

— low dissolved oxygen (DO), harmful algal blooms (HABs), loss of submerged aquatic vegetation (SAV ), and low water clarity. Low DO disrupts aquatic habitats, causing stress to fish and shellfish, which, in the short

- term, can lead to episodic fish kills and, in the long

- term, can damage overall growth in fish and shellfish populations. Low DO also degr ades the aesthetic qualities of surface water. In addition to often being toxic to fish and shellfish, and leading to fish kills and aesthetic impairments of estuaries, HABs can, in some instances, also be harmful to human health. SAV provides critical hab itat for many aquatic species in estuaries and, in some instances, can also protect shorelines by reducing wave strength; therefore, declines in SAV due to nutrient enrichment are an important source of concern. Low water clarity is in part the result of accumulations of both algae and sediments in estuarine waters. In addition to contributing to declines in SAV, high levels of turbidity also degrade the aesthetic qualities of the estuarine environment.
Estuaries in the eastern United States are an importa nt source of food production, in particular fish and shellfish production. The estuaries are capable of supporting large stocks of resident commercial species, and they serve as the breeding grounds and interim habitat for several migratory species. To pro vide an indication of the magnitude of provisioning services associated with coastal fisheries, from 2005 to 2007, the average value of total catch was $1.5 billion per year. It is not known, however, what percentage of this value is directly attributable to or dependent upon the estuaries in these states.
In addition to affecting provisioning services through commercial fish harvests, eutrophication in estuaries may also affect the demand for seafood. For example, a well

publicized toxic pfiesteria bloom i n the Maryland Eastern Shore in 1997, which involved thousands of dead and lesioned fish, led to an estimated $56 million (in 2007 dollars) in lost seafood sales for 360 seafood firms in Maryland in the months following the outbreak (Lipton, 1999).

Estuari es in the United States also provide an important and substantial variety of cultural ecosystem services, including water

- based recreational and aesthetic services. The** * E.O. 12866 Review

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- 62 water quality in the estuary directly affects the quality of these experiences. For exa mple, there were 26 million days of saltwater fishing coastal states from North Carolina to Massachusetts in 2006 (FWA and Census, 2007). Assuming an average consumer surplus value for a fishing day at $36 (in 2007 dollars) in the Northeast and $87 in the Southeast (Kaval and Loomis, 2003), the aggregate value was approximately $1.3 billion (in 2007 dollars) (U.S. EPA, 2009c). 40 In addition, almost 6 million adults participated in motorboating in coastal states from North Carolina to Massachusetts, for a to tal of nearly 63 million days annually during 1999

– 2000 (Leeworthy and Wiley, 2001). Using a national daily value estimate of $32 (in 2007 dollars) for motorboating (Kaval and Loomis (2003), the aggregate value of these coastal motorboating outings was $2 billion per year (U.S. EPA, 2009c). 41 Almost 7 million participated in birdwatching for 175 million days per year, and more than 3 million participated in visits to non

- beach coastal waterside

areas.
Estuaries and marshes have the potential to support a wi de range of regulating services, including climate, biological, and water regulation; pollution detoxification; erosion prevention; and protection against natural hazards from declines in SAV (MEA, 2005). SAV can help reduce wave energy levels and thus pro tect shorelines against excessive erosion, which increases the risks of episodic flooding and associated damages to near

- shore properties or public infrastructure or even contribute to shoreline retreat.
We are unable to provide an estimate of the aquatic enrichment co
- benefits associated with the revised or alternative annual standard s due to data, time, and resource limitations.

Terrestrial Enrichment

. Terrestrial enrichment occurs when terrestrial ecosystems receive N loadings in excess of natural background levels, through either atmospheric deposition or direct application. Evidence presented in the ISA for NO x /SO x (U.S. EPA, 2008c) supports a causal relationship between atmospheric N deposition and biogeochemical cycling and fluxes of N and carbo n in terrestrial systems. Furthermore, evidence summarized in the report supports a causal link between atmospheric N deposition and changes in the types and number of species and biodiversity in terrestrial systems. Nitrogen enrichment occurs over a long time period; as a result, it may take as much as 50 years or more to see changes in ecosystem conditions and indicators. This long time scale also affects the timing of the ecosystem service changes. The ISA for NO x /SO x — Ecological Criteria concluded that t he

40 These estimates reflect the total value of the service, not the marginal change in the value of the service as a result of the emission reductions achieved by this rule. 41 These estimates reflect the total value of the service, not the marginal change in the value of the service as a result of the emission reductions achieved by this rule.** * E.O. 12866 Review

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- 63 evidence is sufficient to infer a causal relationship between nitrogen deposition and the alteration of species richness, species composition, and biodiversity in terrestrial ecosystems (U.S. EPA, 2008c).
One of the main provisioning services potentiall y affected by N deposition is grazing opportunities offered by grasslands for livestock production in the Central U.S. Although N deposition on these grasslands can offer supplementary nutritive value and promote overall grass production, there are concern s that fertilization may favor invasive grasses and shift the species composition away from native grasses. This process may ultimately reduce the productivity of grasslands for livestock production. Losses due to invasive grasses can be significant; for e xample, based on a bioeconomic model of cattle grazing in the upper Great Plains, Leitch, Leistritz, and Bangsund (1996) and Leistritz, Bangsund, and Hodur (2004) estimated $130 million in losses due to a leafy spurge infestation in the Dakotas, Montana, a nd Wyoming. 42 However, the contribution of N deposition to these losses is still uncertain.
Terrestrial nutrient enrichment also affects cultural and regulating services. For example, in California, Coastal Sage Scrub (CSS) habitat concerns focus on a decli ne in CSS and an increase in nonnative grasses and other species, impacts on the viability of threatened and endangered species associated with CSS, and an increase in fire frequency. Changes in Mixed Conifer Forest (MCF) include changes in habitat suitabi lity and increased tree mortality, increased fire intensity, and a change in the forest ’ s nutrient cycling that may affect surface water quality through nitrate leaching (U.S. EPA, 2008c). CSS and MCF are an integral part of the California landscape, and t ogether the ranges of these habitats include the densely populated and valuable coastline and the mountain areas. Numerous threatened and endangered species at both the state and federal levels reside in CSS and MCF. The value that California residents and the U.S. population as a whole place on CSS and MCF habitats is reflected in the various federal, state, and local government measures that have been put in place to protect these habitats, including the Endangered Species Act, conservation planning progr ams, and private and local land trusts. CSS and MCF habitat are showcased in many popular recreation areas in California, including several national parks and monuments. In addition, millions of individuals are involved in fishing, hunting, and wildlife vi ewing in California every year (DOI, 2007). The quality of these trips depends in part on the health of the ecosystems and their ability to support the diversity of plants and animals found in important habitats found in CSS or MCF ecosystems and the parks associated with those ecosystems.

42 These estimates reflect the total value of the service, not the marginal change in the value of the service as a result of the emission reductions achieved by this rule.** * E.O. 12866 Review

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- 64 Based on analyses in the NO x SO x REA average values of the total benefits in 2006 from fishing, hunting, and wildlife viewing away from home in California were approximately $950 million, $170 million, and $3.6 billion, r espectively (U.S. EPA, 2009c). 43 In addition, data from California State Parks (2003) indicate that in 2002, 69% of adult residents participated in trail hiking for an average of 24 days per year. The aggregate annual benefit for California residents from trail hiking in 2007 was $11 bil lion (U.S. EPA, 2009c). 44 It is not currently possible to quantify the loss in value of services due to nitrogen deposition as those losses are already reflected in the estimates of the contemporaneous total value of these recreational activities. Restorati on of services through decreases in nitrogen deposition would likely increase the total value of recreational services.
Fire regulation is also an important regulating service that could be affected by nutrient enrichment of the CSS and MCF ecosystems by e ncouraging growth of more flammable grasses, increasing fuel loads, and altering the fire cycle. Over the 5

- year period from 2004 to 2008, Southern California experienced, on average, over 4,000 fires per year burning, on average, over 400,000 acres per ye ar (National Association of State Foresters [NASF], 2009). It is not possible at this time to quantify the contribution of nitrogen deposition, among many other factors, to increased fire risk.
We are unable to provide an estimate of the terrestrial nutrie nt enrichment co
- benefits associated with the revised or alternative annual st andards due to data, time, and resource limitations. Methods are not yet available to allow estimation of changes in ecosystem services due to nitrogen deposition.

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. 7 .2.3 Vegeta tion Effects Associated with Gaseous Sulfur Dioxide Uptake of gaseous sulfur dioxide in a plant canopy is a complex process involving adsorption to surfaces (leaves, stems, and soil) and absorption into leaves. SO 2 penetrates into leaves through to the sto mata, although there is evidence for limited pathways via the cuticle (U.S. EPA, 2008c). Pollutants must be transported from the bulk air to the leaf boundary layer in order to get to the stomata. When the stomata are closed, as occurs under dark or drough t conditions, resistance to gas uptake is very high and the plant has a very low degree of susceptibility to injury. In contrast, mosses and lichens do not have a protective cuticle barrier to gaseous pollutants or stomates and are generally more sensitive to gaseous sulfur and

43 These estimates refle ct the total value of the service, not the marginal change in the value of the service as a result of the emission reductions achieved by this rule. 44 These estimates reflect the total value of the service, not the marginal change in the value of the serv ice as a result of the emission reductions achieved by this rule.** * E.O. 12866 Review

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- 65 nitrogen than vascular plants (U.S. EPA, 2008c). Acute foliar injury usually happens within hours of exposure, involves a rapid absorption of a toxic dose, and involves collapse or necrosis of plant tissues. Another type of visible i njury is termed chronic injury and is usually a result of variable SO 2 exposures over the growing season. Besides foliar injury, chronic exposure to low SO 2 concentrations can result in reduced photosynthesis, growth, and yield of plants (U.S. EPA, 2008c). These effects are cumulative over the season and are often not associated with visible foliar injury. As with foliar injury, these effects vary among species and growing environment. SO 2 is also considered the primary factor causing the death of lichens i n many urban and industrial areas (Hutchinson et al., 1996). The ISA for NO x /SO x — Ecological Criteria concluded that the evidence is sufficient to infer a causal relationship between SO 2 injury to vegetation (U.S. EPA, 2008c).

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. 6 .2.4 Mercury

- Related Co

- benefits Associated with the Role of Sulfate in Mercury Methylation Mercury is a persistent, bioaccumulative toxic metal that is emitted from in three forms: gaseous elemental Hg (Hg 0 ), oxidized Hg compounds (Hg +2 ), and particle

- bound Hg (Hg P ). Methylme rcury (MeHg) is formed by microbial action in the top layers of sediment and soils, after Hg has precipitated from the air and deposited into waterbodies or land. Once formed, MeHg is taken up by aquatic organisms and bioaccumulates up the aquatic food web

. Larger predatory fish may have MeHg concentrations many times, typically on the order of one million times, that of the concentrations in the freshwater body in which they live.

The NO x SO x ISA

— Ecological Criteria concluded that evidence is sufficient to infer a causal relationship between sulfur deposition and increased mercury methylation in wetlands and aquatic environments (U.S. EPA, 2008c). Specifically, there appears to be a relationship between SO 4 2

- deposition and mercury methylation; however, the rate of mercury methylation varies according to several spatial and biogeochemical factors whose influence has not been fully quantified (see Figure 6

- 1 4 ). Therefore, the correlation between SO 4 2

- deposition and MeHg could not be quantified for the purpose of interpolating the association across waterbodies or regions. Nevertheless, because changes in MeHg in ecosystems represent changes in significant human and ecological health risks, the association between sulfur and mercury cannot be neglected ( U.S. EPA, 2008c).** * E.O. 12866 Review

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- 66 Figure 6

- 1 4

. Spatial and Biogeochemical Factors Influencing MeHg Production

As research evolves and the computational capacity of models expands to meet the complexity of mercury methylation processes in ecosystems, the role of interac ting factors may be better parsed out to identify ecosystems or regions that are more likely to generate higher concentrations of MeHg. Figure 6

- 1 5 illustrates the type of current and forward

- looking research being developed by the U.S. Geological Survey ( USGS) to synthesize the contributing factors of mercury and to develop a map of sensitive watersheds. The mercury score referenced in Figure 6

- 1 5 is based on SO 4 2

- concentrations, acid neutralizing capacity (ANC), levels of dissolved organic carbon and pH, mercury species concentrations, and soil types to gauge the methylation sensitivity (Myers et al., 2007).** * E.O. 12866 Review

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- 67 Figure 6

- 1 5

. Preliminary USGS Map of Mercury Methylation

– Sensitive Watersheds Source: Myers et al., 2007
Interdependent biogeochemical factors preclude the existence of simple sulfate

- related mercury methylation models. It is clear that decreasing sulfate deposition is likely to result in decreased MeHg concentrations. Future research may allow for the charac terization of a usable sulfate

- MeHg response curve; however, no regional or classification calculation scale can be created at this time because of the number of confounding factors.

Decreases in SO 4 2

- deposition have already shown promising reductions in MeHg. Observed decreases in MeHg fish tissue concentrations have been linked to decreased acidification and declining SO 4 2

- and mercury deposition in Little Rock Lake, WI (Hrabik and Watras, 2002), and to decreased SO 4 2

- deposition in Isle Royale in Lake S uperior, MI (Drevnick
et al., 2007). Although the possibility exists that reductions in SO 4 2

- emissions could generate a pulse in MeHg production because of decreased sulfide inhibition in sulfate

- saturated waters, this effect would likely involve a limite d number of U.S. waters (Harmon et al., 2007). Also, because of the diffusion and outward flow of both mercury

- sulfide complexes and SO 4 2

- , increased mercury methylation downstream may still occur in sulfate

- enriched ecosystems with increased organic matte r and/or downstream transport capabilities.** * E.O. 12866 Review

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- 68 Remediation of sediments heavily contaminated with mercury has yielded significant reductions of MeHg in biotic tissues. Establishing quantitative relations in biotic responses to MeHg levels as a result of chang es in atmospheric mercury deposition, however, presents difficulties because direct associations can be confounded by all of the factors discussed in this section. Current research does suggest that the levels of MeHg and total mercury in ecosystems are po sitively correlated, so that reductions in mercury deposited into ecosystems would also eventually lead to reductions in MeHg in biotic tissues. Ultimately, an integrated approach that involves the reduction of both sulfur and mercury emissions may be most efficient because of the variability in ecosystem responses. Reducing SO X emissions could have a beneficial effect on levels of MeHg in many waters of the United States.
MeHg is the only form of mercury that biomagnifies in the food web. Concentrations of MeHg in fish are generally on the order of a million times the MeHg concentration in water. In addition to mercury deposition, key factors affecting MeHg production and accumulation in fish include the amount and forms of sulfur and carbon species present in a given waterbody. Thus, two adjoining water bodies receiving the same deposition can have significantly different fish mercury concentrations.
Methylmercury builds up more in some types of fish and shellfish than in others. The levels of methylmercury in high and shellfish vary widely depending on what they eat, how long they live, and how high they are in the food chain. Most fish, including ocean species and local freshwater fish, contain some methylmercury. In general, higher mercury concentrations are expected in top predators, which are often large fish relative to other species in a waterbody.
The ecosystem service most directly affected by sulfate

- mediated mercury methylation is the provision of fish for consumption as a food source. This service is of particular importance to groups engaged in subsistence fishing, pregnant women and young children.

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. 6 .3 Ecosystem Co

- benefits from Reductions in Mercury Emissions Deposition of mercury to waterbodies can also have an impact on ecosystems and wildli fe. Mercury contamination is present in all environmental media with aquatic systems experiencing the greatest exposures due to bioaccumulation. Bioaccumulation refers to the net uptake of a contaminant from all possible pathways and includes the accumulat ion that may occur by direct exposure to contaminated media as well as uptake from food.
Atmospheric mercury enters freshwater ecosystems by direct deposition and through runoff from terrestrial watersheds. Once mercury deposits, it may be converted to org anic** * E.O. 12866 Review

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- 69 methylmercury mediated primarily by sulfate

- reducing bacteria. Methylation is enhanced in anaerobic and acidic environments, greatly increasing mercury toxicity and potential to bioaccumulate in aquatic foodwebs. A number of key biogeochemical control s influence the production of methylmercury in aquatic ecosystems. These include sulfur, pH, organic matter, iron, mercury “ aging , ” and bacteria type and activity (Munthe et al., 2007).
Wet and dry deposition of oxidized mercury is a dominant pathway for bringing mercury to terrestrial surfaces. In forest ecosystems, elemental mercury may also be absorbed by plants stomatally, incorporated by foliar tissues and released in litterfall (Eri cksen et al., 2003). Mercury in throughfall, direct deposition in precipitation, and uptake of dissolved mercury by roots (Rea et al., 2002) are also important in mercury accumulation in terrestrial ecosystems.
Soils have significant capacity to store larg e quantities of atmospherically deposited mercury where it can leach into groundwater and surface waters. The risk of mercury exposure extends to insectivorous terrestrial species such as songbirds, bats, spiders, and amphibians that receive mercury deposi tion or from aquatic systems near the forest areas they inhabit (Bergeron et al., 2010a, b; Cristol et al., 2008; Rimmer et al., 2005; Wada et al., 2009 & 2010).
Numerous studies have generated field data on the levels of mercury in a variety of wild speci es. Many of the data from these environmental studies are anecdotal in nature rather than representative or statistically designed studies. The body of work examining the effects of these exposures is growing but still incomplete given the complexities of the natural world. A large portion of the adverse effect research conducted to date has been carried out in the laboratory setting rather than in the wild; thus, conclusions about overarching ecosystem health and population effects are difficult to make at this time. In the sections that follow numerous effects have been identified at d iffering exposure levels.

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. 6 .3.1 Mercury Effects on Fish A review of the literature on effects of mercury on fish (Crump and Trudeau, 2009) reports results for numerous spec ies including trout, bass (large and smallmouth), northern pike, carp, walleye, salmon and others from laboratory and field studies. The effects studied are reproductive and include deficits in sperm and egg formation, histopathological changes in
testes a nd ovaries, and disruption of reproductive hormone synthesis. These studies were conducted in areas from New York to Washington and while many were conducted by adding MeHg to water or diet many were conducted at current environmental levels. While we cann ot determine at this time whether these reproductive deficits are affecting fish populations across the United States it should be noted that it is possible that over time reproductive deficits could** * E.O. 12866 Review

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- 70 have an effect on populations. Lower fish populations wo uld conceivably impact the ecosystem services like recreational fishing derived from having healthy aquatic ecosystems quite apart from the effects of consumption advisories due to the human health effects of mercury.

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. 6 .3.2 Mercury Effects on Birds
In ad dition to effects on fish, mercury also affects avian species. In previous reports (U.S. EPA , 1997; 2005), much of the focus has been on large piscivorous species in particular the common loon. The loon is most visible to the public during the summer breed ing season on northern lakes and they have become an important symbol of wilderness in these areas (McIntyre and Barr, 1997). A multitude of loon watch, preservation, and protection groups have formed over the past few decades and have been instrumental in promoting conservation, education, monitoring, and research of breeding loons (McIntyre and Evers, 2000; Evers, 2006). Significant adverse effects on breeding loons from mercury have been found to occur including behavioral (reduced nest

- sitting), physiol ogical (flight feather asymmetry) and reproductive (chicks fledged/territorial pair) effects (Evers, 2008). Additionally Evers, et al (2008) report that they believe that the weight of evidence indicates that population

- level effects occur in parts of Main e and New Hampshire, and potentially in broad areas of the loon ’ s range.
Recently attention has turned to other piscivorous species such as the white ibis, and great snowy egret. While considered to be fish

- eating generally these wading birds have a very

w ide diet including crayfish, crabs, snails, insects and frogs. These species are experiencing a range of adverse effects due to exposure to mercury. The white ibis has been observed to have decreased foraging efficiency (Adams and Frederick, 2008). Additio nally ibises have been shown to exhibit decreased reproductive success and altered pair behavior (Frederick and Jayasena, 2010). These effects include significantly more unproductive nests, male/male pairing, reduced courtship behavior (head bobbing and pa ir bowing) and lower nestling production by exposed males. In this study, a worst

- case scenario suggested by the results could involve up to a 50% reduction in fledglings due to MeHg in diet. These estimates may be conservative if male/male pairing in the wild it could result in a shortage of partners for females and the effect of homosexual breeding would be magnified. In egrets, mercury has been implicated in the decline of the species in south Florida (Sepulveda, et al., 1999) and Hoffman (2010) has show n that egrets show liver and possibly kidney effects. While ibises and egrets are most abundant in coastal areas and these studies were conducted in south Florida and Nevada the ranges of ibises and egrets extend to a large portion of the United States. Ib is territory can range inland to Oklahoma, Arkansas and Tennessee. Egret range covers virtually the entire United States except the mountain west.** * E.O. 12866 Review

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- 71 Insectivorous birds have also been shown to suffer adverse effects due to mercury exposure. These songbirds s uch as Bicknell ’ s thrush, tree swallows and the great tit have shown reduced reproduction, survival, and changes in singing behavior. Exposed tree swallows produced fewer fledglings (Brasso, 2008), lower survival (Hallinger, 2010) and had compromised immun e competence (Hawley, 2009). The great tit has exhibited reduced singing behavior and smaller song repertoire in an area of high contamination in the vicinity of a metallurgic smelter in Flanders (Gorissen, 2005).

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. 6 .3.3 Mercury Effects on Mammals In mammals, adverse effects have been observed in mink and river otter, both fish eating species. For otter from Maine and Vermont maximum concentrations on Hg in fur nearly equal or exceed a concentration associated with mortality and concentration in liv er for mink in Massachusetts/Connecticut and the levels in fur from mink in Maine exceed concentrations associated with acute mortality (Yates, 2005). Adverse sublethal effects may be associated with lower Hg concentrations and consequently be more widespr ead than potential acute effects. These effects may include increased activity, poorer maze performance, abnormal startle reflex, and impaired escape and avoidance behavior (Scheuhammer et al., 2007).

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. 6 .3.4 Mercury Ecological Conclusions
The studies cite d here provide a glimpse of the scope of mercury effects on wildlife particularly reproductive and survival effects. These effects range across species from fish to mammals and spatially across a wide area of the United States. The literature is far from

c omplete however. Much more research is required to establish a link between the ecological effects on wildlife and the effect on ecosystem services (services that the environment provides to people) for example recreational fishing, bird watching and wildl ife viewing. The EPA is not, however, currently able to quantify or monetize the co
- benefits of reducing mercury exposures affecting provision of ecosystem services.

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. 6 .4 Vegetation Co

- benefits from Reductions in Ambient Ozone
Illustrative emission reduction s trategies that include NO x emission reductions would affect ambient ozone concentrations. Ozone causes discernible injury to a wide array of vegetation (U.S. EPA, 2006a; Fox and Mickler, 1996). Air pollution can affect the environment and affect eco logical systems, leading to changes in the ecological com munity and influencing the diversity, health, and vigor of individual species (U.S. EPA, 2006a). In terms of forest productivity and ecosystem diversity, ozone may be the pollutant with the gre atest potential for regional

- scale forest impacts (U.S. EPA, 2006a). Studies have demonstrated repeatedly that** * E.O. 12866 Review

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- 72 ozone concentrations commonly observed in polluted areas can have substantial impacts on plant function (De Steiguer et al., 1990; Pye, 1988).

Wh en ozone is present in the air, it can enter the leaves of plants, where it can cause significant cellular damage. Like carbon dioxide (CO 2 ) and other gaseous substances, ozone enters plant tissues primarily through the stomata in leaves in a process calle d “ uptake ” (Winner and Atkinson, 1986). Once sufficient levels of ozone (a highly reactive substance), or its reaction products, reaches the interior of plant cells, it can inhibit or damage essential cellular components and functions, including enzyme act ivities, lipids, and cellular membranes, disrupting the plant ’ s osmotic (i.e., water) balance and energy utilization patterns (U.S. EPA, 2006a; Tingey and Taylor, 1982). With fewer resources available, the plant reallocates existing resources away from roo t growth and storage, above ground growth or yield, and reproductive processes, toward leaf repair and maintenance, leading to reduced growth and/or reproduction. Studies have shown that plants stressed in these ways may exhibit a general loss of vigor, wh ich can lead to secondary impacts that modify plants ’ responses to other environmental factors. Specifically, plants may become more sensitive to other air pollutants, or more susceptible to disease, pest infestation, harsh weather (e.g., drought, frost) a nd other environmental stresses, which can all produce a loss in plant vigor in ozone

- sensitive species that over time may lead to premature plant death. Furthermore, there is evidence that ozone can interfere with the formation of mycorrhiza, essential sy mbiotic fungi associated with the roots of most terrestrial plants, by reducing the amount of carbon available for transfer from the host to the symbiont (U.S. EPA, 2006a).
This ozone damage may or may not be accompanied by visible injury on leaves, and likewise, visible foliar injury may or may not be a symptom of the other types of plant damage described above. Foliar injury is usually the first visible sign of injury to plants from ozone exposure and indicates impaired physiological processes in the leaves (Grulke, 2003). When visible injury is present, it is commonly manifested as chlorotic or necrotic spots, and/or increased leaf senescence (accelerated leaf aging). Visible foliar injury reduces the aesthetic value of ornamental vegetation and trees in urban landscapes and negatively affects scenic vistas in protected natural areas.
Ozone can produce both acute and chronic injury in sensitive species depending on the concentration level and the duration of the exposure

. Ozone effects also tend to accumulate over the growing season of the plant, so that even lower concentrations experienced for a longer duration have the potential to create chronic stress on sensitive vegetation. Not all plants, however, are equally sens itive to ozone. Much of the variation in sensitivity between** * E.O. 12866 Review

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- 73 individual plants or whole species is related to the plant ’ s ability to regulate the extent of gas exchange via leaf stomata (e.g., avoidance of ozone uptake through closure of stomata) and the

r elative ability of species to detoxify ozone

- generated reactive oxygen free radicals (U.S. EPA, 2006a; Winner, 1994). After injuries have occurred, plants may be capable of repairing the damage to a limited extent (U.S. EPA, 2006a). Because of the differin g sensitivities among plants to ozone, ozone pollution can also exert a selective pressure that leads to changes in plant community composition. Given the range of plant sensitivities and the fact that numerous other environmental factors modify plant upta ke and response to ozone, it is not possible to identify threshold values above which ozone is consistently toxic for all plants.
Because plants are at the base of the food web in many ecosystems, changes to the plant community can affect associated organi sms and ecosystems (including the suitability of habitats that support threatened or endangered species and below ground organisms living in the root zone). Ozone impacts at the community and ecosystem level vary widely depending upon numerous factors, inc luding concentration and temporal variation of tropospheric ozone, species composition, soil properties and climatic factors (U.S. EPA, 2006a). In most instances, responses to chronic or recurrent exposure in forested ecosystems are subtle and not observab le for many years. These injuries can cause stand

- level forest decline in sensitive ecosystems (U.S. EPA, 2006a, McBride et al., 1985; Miller et al., 1982). It is not yet possible to predict ecosystem responses to ozone with certainty; however, considerabl e knowledge of potential ecosystem responses is available through long

- term observations in highly damaged forests in the U.S. ( U.S. EPA, 2006a).

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. 6 .4.1 Ozone Effects on Forests Air pollution can affect the environment and affect eco logical systems, lead ing to changes in the ecological com munity and influencing the diversity, health, and vigor of individual species (U.S. EPA, 2006a). Ozone has been shown in numerous studies to have a strong effect on the health of many plants, including a variety of co mm ercial and ecologically impor tant forest tree species throughout the United States (U.S. EPA, 2007b).
In the U.S., this data comes from the U.S. Department of Agriculture (USDA) Forest Service Forest Inventory and Analysis (FIA) program. As part of its Pha se 3 program (formerly known as Forest Health Monitoring), FIA looks for visible foliar injury of ozone

- sensitive forest plant species at each ground monitoring site across the country (excluding woodlots and urban trees) that meets certain minimum criteri a. Because ozone injury is cumulative over the course of the growing season, examinations are conducted in July and August, when ozone concentrations and associated injury are typically highest.** * E.O. 12866 Review

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- 74 Monitoring of ozone injury to plants by the U.S. Forest Servi ce has expanded over the last 15 years from monitoring sites in 10 states in 1994 to nearly 1,000 monitoring sites in 41 states in 2002. Since 2002, the monitoring program has further expanded to 1,130 monitoring sites in 45 states. Figure 6

- 16 shows the r esults of this monitoring program for the year 2002 broken down by U.S. EPA Regions. 45 Figure 6

- 17 identifies the counties that were included in Figure 6

- 1 6 , and provides the county

- level data regarding the presence or absence of ozone

related injury. As shown in Figure 6

- 1 6 , large geographic areas of EPA Regions 6, 8, and 10 were not included in the assessment. Ozone damage to forest plants is classified using a subjective five

- category biosite index based on expert opinion, but designed to be equivale nt from site to site. Ranges of biosite values translate to no injury, low or moderate foliar injury (visible foliar injury to highly sensi tive or moderately sensitive plants, respectively), and high or severe foliar injury, which would be expected to res ult in tree

- level or ecosystem

- level responses, respectively (U.S. EPA, 2006a; Coulston, 2004). The highest percentages of observed high and severe foliar injury, which are most likely to be associated with tree or ecosystem

- level responses, are primarily found in the Mid

- Atlantic and Southeast regions. While the assessment showed considerable regional variation in ozone injury, this assessment targeted different

ozone
- sensitive species in different parts of the country with varying ozone sensitivity, which contributes to the apparent regional differences. It is important to note that ozone can have other, more significant impacts on forest plants ( e.g., reduced biomass growth in trees) prior to showing signs of visible foliar injury (U.S. EPA, 2006a).

Asses sing the impact of ground

- level ozone on forests in the U.S. involves understanding the risks to sensitive tree species from ambient ozone concentrations and accounting for the prevalence of those species within the forest. As a way to quantify the risks t o particular plants from ground

- level ozone, scientists have developed ozone

- exposure/tree

- response functions by exposing tree seedlings to different ozone levels and measuring reductions in growth as

“ biomass loss. ” Typically, seedlings are used because t hey are easy to manipulate and measure their growth loss from ozone pollution. The mechanisms of susceptibility to ozone within the leaves of seedlings and mature trees are identical, and the decreases predicted using the seedlings should be related to the decrease in overall plant fitness for mature trees, but the

45 The data are based on averages of all observations collected in 2002, which is the last year for which data are publicly available. For more information, please consult EPA ’ s 2008 Report on the Environment (U.S. EPA, 2008b).** * E.O. 12866 Review

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- 75 c Degree c Degree of Injury: These categories reflect a subjective index based on expert opinion. Ozone can have other, more significant impacts on forest plants (e.g., reduced biomass growth in trees) prior to showing signs of visible foliar injury. Figure 6

- 16. Visible Foliar Injur y to Forest Plants from Ozone in U.S. by EPA Regions a,b,c** * E.O. 12866 Review

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- 76 Figure 6

- 17. Presence and Absence of Visible Foliar Injury, as Measured by U.S. Forest

Service, 2002 Source: U.S. EPA, 2007b
magnitude of the effect may be higher or lower depending on the tree sp ecies (Chappelka and Samuelson, 1998). In areas where certain ozone

- sensitive species dominate the forest community, the biomass loss from ozone can be sig nificant. Experts have identified 2% annual biomass loss as a level of concern, which would cause lo ng term ecological harm as the short

term negative effects on seedlings compound to affect long

- term forest health (Heck and Cowling, 1997).
Ozone damage to the plants including the trees and understory in a forest can affect the ability of the forest to s ustain suitable habitat for associated species particularly threatened and endangered species that have existence value

— a nonuse ecosystem service

— for the public. Similarly, damage to trees and the loss of biomass can affect the forest ’ s provisioning servi ces in the form of timber for various commercial uses. In addition, ozone can cause discoloration of leaves and more rapid senescence (early shedding of leaves), which could negatively affect fall

color tourism because the fall foliage would be less availa ble or less attractive. Beyond the aesthetic damage to fall color vistas, forests provide the public with many other recreational and educational services that may be affected by reduced forest health including hiking, wildlife** * E.O. 12866 Review

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- 77 viewing (including bird watc hing), camping, picnicking, and hunting. Another potential effect of biomass loss in forests is the subsequent loss of climate regulation service in the form of reduced ability to sequester carbon and alteration of hydrologic cycles.

Some of the common tre e species in the United States that are sensitive to ozone are black cherry ( Prunus serotina ), tulip

- poplar ( Liriodendron tulipifera) , and eastern white pine ( Pinus strobus)

. Ozone

- exposure/tree

- response functions have been developed for each of these tree species, as well as for aspen ( Populus tremuliodes ), and ponderosa pine ( Pinus ponderosa ) (U.S. EPA, 2007b).

6

. 6 .4.2 Ozone Effects on Crops Laboratory and field experiments have shown reductions in yields for agronomic crops exposed to ozone, including ve getables (e.g., lettuce) and field crops (e.g., cotton and wheat). Damage to crops from ozone exposures includes yield losses (i.e., in terms of weight, number, or size of the plant part that is harvested), as well as changes in crop quality (i.e., physica l appearance, chemical composition, or the ability to withstand storage) (U.S. EPA, 2007b). The most extensive field experiments, conducted under the National Crop Loss Assessment Network (NCLAN) examined 15 species and numerous cultivars. The NCLAN result s show that

“ several economically important crop species are sensitive to ozone levels typical of those found in the United States ” (U.S. EPA, 2006a). In addition, economic studies have shown reduced economic co
- benefits as a result of predicted reductions in crop yields, directly affecting the amount and quality of the provisioning service provided by these crops, associated with observed ozone levels (Kopp et al., 1985; Adams et al., 1986; Adams et al., 1989). In addition, visible foliar injury by itself can reduce the market value of certain leafy crops (such as spinach, lettuce). According to the Ozone Staff Paper, there has been no evidence that crops are becoming more tolerant of ozone (U.S. EPA, 2007b). Using the Agriculture Simulation Model (AGSIM) ( Taylor, 1994) to calculate the agricultural benefits of reductions in ozone exposure, the EPA estimated that attaining a W126 standard of 13 ppm

- hr would produce monetized benefits of approximately $400 million to $620 million in 2006 (inflated to 2006 dol lars) (U.S. EPA, 2007b). 46

46 These estimates illustrate the value of vegetation effects from a substantial reduction of ozone concentrations, not the marginal change in ozone concentrations anticipated a result of the emission reductions achieved by this rule.** * E.O. 12866 Review

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- 78 6

. 6 .4.3 Ozone Effects on Ornamental Plants Urban ornamental plants are an additional vegetation category likely to experience some degree of negative effects associated with exposure to ambient ozone levels. Several ornamental species have been listed as sensitive to ozone (Abt Associates, 1995). Because ozone causes visible foliar injury, the aesthetic value of ornamental plants (such as petunia, geranium, and poinsettia) in urban landscapes would be reduced (U.S. EPA, 2007b). Sensitive ornamental species would require more frequent replacement and/or increased maintenance (fertilizer or pesticide application) to maintain the desired appearance because of exposure to ambient ozone (U.S. EPA, 2007b). In additio n, many businesses rely on healthy

- looking vegetation for their livelihoods (e.g., horticulturalists, landscapers, Christmas tree growers, farmers of leafy crops, etc.). The ornamental landscaping industry is a multi

- billion dollar industry that affects bo th private property owners/tenants and governmental units responsible for public areas (Abt Associates, 1995). Preliminary data from the 2007 Economic Census indicate that the landscaping services industry, which is primarily engaged in providing landscape care and maintenance services and installing trees, shrubs, plants, lawns, or gardens, was valued at $53 billion (U.S. Census Bureau, 2010). Therefore, urban ornamentals represent a potentially large unquantified benefit category. This aesthetic damage may affect the enjoyment of urban parks by the public and homeowners ’ enjoyment of their landscaping and gardening activities. In addition, homeowners may experience a reduction in home value or a home may linger on the market longer due to decreased aesth etic appeal. In the absence of adequate exposure

- response functions and economic damage functions for the potential range of effects relevant to ornamental plants, we cannot conduct a quantitative analysis to estimate these effects.
We are unable to provid e an estimate of the ozone crop co
- benefits associated with the revised or alternative annual standard s due to data, time, and resource limitations.

6

. 7 References Abt Associates, Inc. 1995. Urban ornamental plants: sensitivity to ozone and potential econo mic losses. U.S. EPA, Office of Air Quality Planning and Standards, Research Triangle Park. Under contract to RADIAN Corporation, contract no. 68

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