4.1.6 Human Welfare Impact Assessment PM and ozone have numerous documented effects on environmental quality that affect human welfare. These welfare effects include direct damages to property, either through impacts on material structures or by soiling 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. EPA's Criteria Documents for PM and ozone list numerous physical and ecological effects known to be linked to ambient concentrations of these pollutants ( US EPA, 1996a; 1996b). This section describes individual effects and how we quantify and monetize them. These effects include changes in commercial crop and forest yields, visibility, and nitrogen deposition to estuaries.

<IP_ADDRESS> Visibility Benefits Changes in the level of ambient particulate matter caused by the reduction in emissions from the IAQR will change the level of visibility in much of the Eastern U. S. Visibility directly affects people's enjoyment of a variety of daily activities. Individuals value visibility both in the places they live and work, in the places they travel to for recreational purposes,

and at sites of unique public value, such as the Great Smokey Mountains National Park. This section discusses the measurement of the economic benefits of visibility. 26A change of less than 10 percent in the light extinction budget represents a measurable improvement in visibility, but may not be perceptible to the eye in many cases. Some of the average regional changes in visibility are less than one deciview ( i. e. less than 10 percent of the light extinction budget), and thus less than perceptible. However, this does not mean that these changes are not real or significant. Our assumption is then that individuals can place values on changes in visibility that may not be perceptible. This is quite plausible if individuals are aware that many regulations lead to small improvements in visibility which when considered together amount to perceptible changes in visibility.

27The Clean Air Act designates 156 national parks and wilderness areas as Class I areas for visibility protection.

4 76 It is difficult to quantitatively define a visibility endpoint that can be used for valuation.

Increases in PM concentrations cause increases in light extinction. Light extinction is a measure of how much the components of the atmosphere absorb light. More light absorption means that the clarity of visual images and visual range is reduced, ceteris paribus. Light absorption is a variable that can be accurately measured. Sisler ( 1996) created a unitless measure of visibility based directly on the degree of measured light absorption called the deciview. Deciviews are standardized for a reference distance in such a way that one deciview corresponds to a change of about 10 percent in available light. Sisler characterized a change in light extinction of one deciview as " a small but perceptible scenic change under many circumstances." Air quality models were used to predict the change in visibility, measured in deciviews, of the areas affected by the preliminary control options. 26 EPA considers benefits from two categories of visibility changes: residential visibility and recreational visibility. In both cases economic benefits are believed to consist of both use values and non use values. Use values include the aesthetic benefits of better visibility,

improved road and air safety, and enhanced recreation in activities like hunting and birdwatching. Non use values are based on people's beliefs that the environment ought to exist free of human induced haze. Non use values may be a more important component of value for recreational areas, particularly national parks and monuments.

Residential visibility benefits are those that occur from visibility changes in urban,

suburban, and rural areas, and also in recreational areas not listed as federal Class I areas. 27 For the purposes of this analysis, recreational visibility improvements are defined as those that occur specifically in federal Class I areas. A key distinction between recreational and residential benefits is that only those people living in residential areas are assumed to receive benefits from residential visibility, while all households in the U. S. are assumed to derive some 28For details of the visibility estimates discussed in this chapter, please refer to the benefits technical support document for the Nonroad Diesel rulemaking ( Abt Associates 2003).

29 An SAB advisory letter indicates that" many members of the Council believe that the Chestnut and Rowe study is the best available." ( EPA SAB COUNCIL ADV 00 002, 1999) However, the committee did not formally approve use of these estimates because of concerns about the peer reviewed status of the study. EPA believes the study has received adequate review and has been cited in numerous peer reviewed publications ( Chestnut and Dennis, 1997).

4 77 benefit from improvements in Class I areas. Values are assumed to be higher if the Class I area is located close to their home. 28 Only two existing studies provide defensible monetary estimates of the value of visibility changes. One is a study on residential visibility conducted in 1990 ( McClelland, et.

al., 1993) and the other is a 1988 survey on recreational visibility value ( Chestnut and Rowe,

1990a; 1990b). While there are a number of other studies in the literature, they were conducted in the early 1980s and did not use methods that are considered defensible by current standards. Both the Chestnut and Rowe and McClelland et al studies utilize the contingent valuation method. There has been a great deal of controversy and significant development of both theoretical and empirical knowledge about how to conduct CV surveys in the past decade. In EPA's judgment, 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 benefits of visibility changes in recreational areas. 29 This study serves as an essential input to our estimates of the benefits of recreational visibility improvements in the primary benefits estimates. Consistent with SAB advice, EPA has designated the McClelland,

et al. study as significantly less reliable for regulatory benefit cost analysis, although it does provide useful estimates on the order of magnitude of residential visibility benefits (

EPASAB COUNCIL ADV 00 002, 1999). Residential visibility benefits are therefore only included as a sensitivity estimate in Appendix 9 B ( to be completed for the Supplemental Analysis).

The Chestnut and Rowe study measured the demand for visibility in Class I areas managed by the National Park Service ( NPS) in three broad regions of the country:

California, the Southwest, and the Southeast. Respondents in five states were asked about their willingness to pay 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 visibility levels in these photographs were later converted to deciviews for the current analysis. The survey data collected were used to estimate a WTP 4 78 equation for improved visibility. In addition to the visibility change variable, the estimating equation also included household income as an explanatory variable.

The Chestnut and Rowe study did not measure values for visibility improvement in Class I areas outside the three regions. Their study covered 86 of the 156 Class I areas in the U. S. We can infer the value of visibility changes in the other Class I areas by transferring values of visibility changes at Class I areas in the study regions. However, these values are not as defensible and are thus presented only as a sensitivity analysis ( to be completed for the Supplemental Analysis). A complete description of the benefits transfer method used to infer values for visibility changes in Class I areas outside the study regions is provided in the benefits TSD for the Nonroad Diesel rulemaking ( Abt Associates, 2003).

The estimated relationship from the Chestnut and Rowe study is only directly applicable to the populations represented by survey respondents. EPA used benefits transfer methodology to extrapolate these results to the population affected by the proposed IAQR. A general willingness to pay equation for improved visibility ( measured in deciviews) was developed as a function of the baseline level of visibility, the magnitude of the visibility improvement, and household income. The behavioral parameters of this equation were taken from analysis of the Chestnut and Rowe data. These parameters were used to calibrate WTP for the visibility changes resulting from the IAQR. The method for developing calibrated WTP functions is based on the approach developed by Smith, et al. ( 2002). Available evidence indicates that households are willing to pay more for a given visibility improvement as their income increases ( Chestnut, 1997). The benefits estimates here incorporate Chestnut's estimate that a 1 percent increase in income is associated with a 0.9 percent increase in WTP for a given change in visibility.

Using the methodology outlined above, EPA estimates that the total WTP for the visibility improvements in Southeastern Class I areas brought about by the IAQR is $ 880 million in 2010 and $ 1,400 million in 2015. This value includes the value to households living in the same state as the Class I area as well as values for all households in the U. S. living outside the state containing the Class I area, and the value accounts for growth in real income.

We examine the impact of expanding the visibility benefits analysis to other areas of the country in a sensitivity analysis to be completed for the Supplemental Analysis. 4 79 The benefits resulting from visibility improvements in Southeastern Class I areas under the Proposed IAQR are presented in Figure 4 2. This figure presents these benefits both in terms of the total benefits modeled for each of the Class I areas ( i. e., the " Park Benefits" map)

and the benefits realized by the populations in each of the 48 contiguous states ( i. e., the " State Benefits" map). The latter results reflect the willingness to pay of state residents for visibility improvements occuring in Class I areas in the Southeastern United States.

One major source of uncertainty for the visibility benefit estimate is the benefits transfer process used. Judgments used to choose the functional form and key parameters of the estimating equation for willingness to pay for the affected population could have significant effects on the size 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 results.

<IP_ADDRESS> Agricultural, Forestry and other Vegetation Related Benefits The Ozone Criteria Document notes that " ozone affects vegetation throughout the United States, impairing crops, native vegetation, and ecosystems more than any other air pollutant" ( US EPA, 1996). Changes in ground level ozone resulting from the preliminary control options are expected to impact crop and forest yields throughout the affected area.

Well developed techniques exist to provide monetary estimates of these benefits to agricultural producers and to consumers. These techniques use models of planting decisions,

yield response functions, and agricultural products supply and demand. The resulting welfare measures are based on predicted changes in market prices and production costs. Models also exist to measure benefits to silvicultural producers and consumers. However, these models have not been adapted for use in analyzing ozone related forest impacts. As such, our analysis ( to be completed for the Supplemental Analysis) provides monetized estimates of agricultural benefits, and a discussion of the impact of ozone changes on forest productivity, but does not monetize commercial forest related benefits.

<IP_ADDRESS>.1 Agricultural Benefits. Laboratory and field experiments have shown reductions in yields for agronomic crops exposed to ozone, including vegetables ( e. g., lettuce)

and field crops ( e. g., cotton and wheat). The most extensive field experiments, conducted under the National Crop Loss Assessment Network ( NCLAN) examined 15 species and numerous cultivars. The 4 80 2015 V is ibility B e n e f its In M illio n s o f D o lla rs 0  0 .05 0 .05 0.1 0 .1  0 .25 0 .25 0.4 0 .4  1 .5 ____ ___ ____

0 1 5 1 5  40 4 0  55 5 5  130 1 3 0  800 S tate P ark Shenandoah Jam e s R iv e r F ace L in v ille G o rg e G reat Sm o ky M o u n tain s Sw an Q uarter S h in in g R o c k Jo y c e K ilm e r S lick rock C o h o tta S ipsey W o lf Is la n d O k e fenokee S t. M arks C hassahow itz ka Ev e rglad es B reton M am m oth C av e 0  0 .20 0 .2  0 .5 0 .5  0 .8 0 .8  1 .0 1 .0  1 .82 C h a n g e in D e c iv iew S ta te le v e l dollar v a lues reflect the w illingne s s to p ay o f s ta te re s id e n ts fo r v is ibility im p ro v em ents o ccurrin g in C lass I are a s in the southea s te rn United S ta te s . T he re a re 16 C lass I a reas in the southe a s t w hich acco u n t for ap p ro xim ately 17% o f total p ark v is its in the United S ta tes. S ta te dollar b enefits a re d riven b y p o p u latio n and p ro xim ity to p a rks.

Pa rk d o lla r b e n e f its a re d riv e n b y deg re e o f v is ibility im provem en t and p a rk visitatio n .

S ta te B en efits P ark B en e f its Figure 4.2. Visib ility Improvements in Southeastern Class I Areas 4 81 NCLAN results show that " several economically important crop species are sensitive to ozone levels typical of those found in the U. S." ( US EPA, 1996). In addition, economic studies have shown a relationship between observed ozone levels and crop yields ( Garcia, et al., 1986).

Due to data limitations, we were unable to assess ozone related agricultural benefits associated with the proposed IAQR. However, we will be assessing these benefits for the Supplemental Analysis and for the analysis of the final IAQR.

<IP_ADDRESS>.2 Forestry Benefits. Ozone also has been shown conclusively to cause discernible injury to forest trees ( US EPA, 1996; Fox and Mickler, 1996). In our previous analysis of the HD Engine/ Diesel Fuel rule, we were able to quantify the effects of changes in ozone concentrations on tree growth for a limited set of species. Due to data limitations, we were not able to quantify such impacts for this analysis. We plan to assess both physical impacts on tree growth and the economic value of those physical impacts in our analysis of the final rule. We will use econometric models of forest product supply and demand to estimate changes in prices, producer profits and consumer surplus. These benefits will be estimated for the final IAQR.

<IP_ADDRESS>.3 Other Vegetation Effects. An additional welfare benefit expected to accrue as a result of reductions in ambient ozone concentrations in the U. S. is the economic value the public receives from reduced aesthetic injury to forests. There is sufficient scientific information available to reliably establish that ambient ozone levels cause visible injury to foliage and impair the growth of some sensitive plant species ( US EPA, 1996c, p. 5 521).

However, present analytic tools and resources preclude EPA from quantifying the benefits of improved forest aesthetics.

Urban ornamentals represent an additional vegetation category likely to experience some degree of negative effects associated with exposure to ambient ozone levels and likely to impact large economic sectors. In the absence of adequate exposure response functions and economic damage functions for the potential range of effects relevant to these types of vegetation, no direct quantitative economic benefits analysis has been conducted. It is estimated that more than $ 20 billion ( 1990 dollars) are spent annually on landscaping using ornamentals ( Abt Associates, 1995), both by private property owners/ tenants and by governmental units responsible for public areas. This is therefore a potentially important welfare effects category. However, information and valuation methods are not available to allow for plausible estimates of the percentage of these expenditures that may be related to impacts associated with ozone exposure. 4 82 The EGU standards, by reducing NO X emissions, will also reduce nitrogen deposition on agricultural land and forests. There is some evidence that nitrogen deposition may have positive effects on agricultural output through passive fertilization. Holding all other factors constant, farmers' use of purchased fertilizers or manure may increase as deposited nitrogen is reduced. Estimates of the potential value of this possible increase in the use of purchased fertilizers are not available, but it is likely that the overall value is very small relative to other health and welfare effects. The share of nitrogen requirements provided by this deposition is small, and the marginal cost of providing this nitrogen from alternative sources is quite low.

In some areas, agricultural lands suffer from nitrogen over saturation due to an abundance of on farm nitrogen production, primarily from animal manure. In these areas, reductions in atmospheric deposition of nitrogen represent additional agricultural benefits.

Information on the effects of changes in passive nitrogen deposition on forests and other terrestrial ecosystems is very limited. The multiplicity of factors affecting forests,

including other potential stressors such as ozone, and limiting factors such as moisture and other nutrients, confound assessments of marginal changes in any one stressor or nutrient in forest ecosystems. However, reductions in deposition of nitrogen could have negative effects on forest and vegetation growth in ecosystems where nitrogen is a limiting factor ( US EPA,

1993).

On the other hand, there is evidence that forest ecosystems in some areas of the United States are nitrogen saturated ( US EPA, 1993). Once saturation is reached, adverse effects of additional nitrogen begin to occur such as soil acidification which can lead to leaching of nutrients needed for plant growth and mobilization of harmful elements such as aluminum.

Increased soil acidification is also linked to higher amounts of acidic runoff to streams and lakes and leaching of harmful elements into aquatic ecosystems.

<IP_ADDRESS> Benefits from Reductions in Materials Damage The preliminary control options that we modeled are expected to produce economic benefits in the form of reduced materials damage. There are two important categories of these benefits. Household soiling refers to the accumulation of dirt, dust, and ash on exposed surfaces. Criteria pollutants also have corrosive effects on commercial/ industrial buildings and structures of cultural and historical significance. The effects on historic buildings and outdoor works of art are of particular concern because of the uniqueness and irreplaceability of many of these objects. 4 83 Previous EPA benefit analyses have been able to provide quantitative estimates of household soiling damage. Consistent with SAB advice, we determined that the existing data ( based on consumer expenditures from the early 1970' s) are too out of date to provide a reliable enough estimate of current household soiling damages ( EPA SAB Council ADV 003,

1998) to include in our base estimate. We calculate household soiling damages in a sensitivity estimate that will be completed as part of the Supplemental Analysis.

EPA is unable to estimate any benefits to commercial and industrial entities from reduced materials damage. Nor is EPA able to estimate the benefits of reductions in PMrelated damage to historic buildings and outdoor works of art. Existing studies of damage to this latter category in Sweden ( Grosclaude and Soguel, 1994) indicate that these benefits could be an order of magnitude larger than household soiling benefits.

<IP_ADDRESS> Benefits from Reduced Ecosystem Damage The effects of air pollution on the health and stability of ecosystems are potentially very important, but are at present poorly understood and difficult to measure. The reductions in NO X caused by the final rule could produce significant benefits. Excess nutrient loads,

especially of nitrogen, cause a variety of adverse consequences to the health of estuarine and coastal waters. These effects include toxic and/ or noxious algal blooms such as brown and red tides, low ( hypoxic) or zero ( anoxic) concentrations of dissolved oxygen in bottom waters, the loss of submerged aquatic vegetation due to the light filtering effect of thick algal mats, and fundamental shifts in phytoplankton community structure ( Bricker et al., 1999).

Direct functions relating changes in nitrogen loadings to changes in estuarine benefits are not available. The preferred WTP based measure of benefits depends on the availability of these functions and on estimates of the value of environmental responses. Because neither appropriate functions nor sufficient information to estimate the marginal value of changes in water quality exist at present, calculation of a WTP measure is not possible.

If better models of ecological effects can be defined, EPA believes that progress can be made in estimating WTP measures for ecosystem functions. These estimates would be superior to avoided cost estimates in placing economic values on the welfare changes associated with air pollution damage to ecosystem health. For example, if nitrogen or sulfate loadings can be linked to measurable and definable changes in fish populations or definable indexes of biodiversity, then CV studies can be designed to elicit individuals' WTP for changes in these effects. This is an important area for further research and analysis, and will require close collaboration among air quality modelers, natural scientists, and economists. 4 84 4.2 Benefits Analysis   Results Applying the impact and valuation functions described in Section C to the estimated changes in ozone and PM described in Section B yields estimates of the changes in physical damages ( i. e. premature mortalities, cases, admissions, change in light extinction, etc.) and the associated monetary values for those changes. Estimates of physical health impacts are presented in Table 4 16. Monetized values for both health and welfare endpoints are presented in Table 4 17, along with total aggregate monetized benefits. All of the monetary benefits are in constant year 1999 dollars.

Not all known PM and ozone related health and welfare effects could be quantified or monetized. The monetized value of these unquantified effects is represented by adding an unknown " B" to the aggregate total. The estimate of total monetized health benefits is thus equal to the subset of monetized PM and ozone related health and welfare benefits plus B,

the sum of the nonmonetized health and welfare benefits.

Total monetized benefits are dominated by benefits of mortality risk reductions. The primary analysis estimate projects that the proposed rule will result in 9,600 avoided premature deaths in 2010 and 13,000 avoided premature deaths in 2015. The increase in benefits from 2010 to 2015 reflects additional emission reductions from the standards, as well as increases in total population and the average age ( and thus baseline mortality risk) of the population. Note that unaccounted for changes in baseline mortality rates over time may lead to reductions in the estimated number of avoided premature mortalities. 4 85 Table 4 16. Reductions in Incidence of Adverse Health Effects Associated with Reductions in Particulate Matter and Ozone Associated with the Proposed IAQRa Endpoint 2010 2015 PM related Endpoints Premature mortalityb Long term exposure ( adults, 30 and over)

Long term exposure ( infant, < 1 yr) 9,600 13,000 22 29 Chronic bronchitis ( adults, 26 and over) 5,200 6,900 Non fatal myocardial infarctions ( adults, 18 and older) 13,000 18,000 Hospital admissions   Respiratory ( all ages) c 4,200 5,800 Hospital admissions   Cardiovascular ( adults, > 18) d 3,700 5,000 Emergency Room Visits for Asthma ( 18 and younger) 7,000 9,200 Acute bronchitis ( children, 8 12) 12,000 16,000 Lower respiratory symptoms ( children, 7 14) 140,000 190,000 Upper respiratory symptoms ( asthmatic children, 9 18) 490,000 620,000 Asthma Exacerbations ( asthmatic children, 6 18) 190,000 240,000 Work loss days ( adults, 18 65) 1,000,000 1,300,000 Minor restricted activity days ( adults, age 18 65) 6,100,000 7,900,000 Ozone related Endpoints Hospital Admissions   Respiratory Causes ( adults, 65 and older) e 630 1,500 Hospital Admissions   Respiratory Causes ( children, under 2 years) 380 840 Emergency Room Visits for Asthma ( all ages) 120 250 Minor restricted activity days ( adults, age 18 65) 280,000 610,000 School absence days ( children, age 6 18) 180,000 390,000 a Incidences are rounded to two significant digits.

b Premature mortality associated with ozone is not separately included in this analysis. It is assumed that the Impact function for premature mortality captures both PM mortality benefits and any mortality benefits associated with other air pollutants.

c Respiratory hospital admissions for PM includes admissions for COPD, pneumonia, and asthma.

d Cardiovascular hospital admissions for PM includes total cardiovascular and subcategories for ischemic heart disease, dysrhythmias, and heart failure.

e Respiratory hospital admissions for ozone includes admissions for all respiratory causes and subcategories for COPD and pneumonia. 4 86 Table 4 17. Results of Human Health and Welfare Benefits Valuation for the Proposed IAQR ( millions of 1999 dollars) a, b Endpoint Pollutant 2010 2015 Premature mortalityc Long term exposure, ( adults, > 30yrs)

3% discount rate PM $ 53,000 $ 77,000 7% discount rate $ 50,000 $ 72,000 Long term exposure ( child < 1yr) $ 130 $ 180 Chronic bronchitis ( adults, 26 and over) PM $ 1,900 $ 2,700 Non fatal myocardial infarctions 3% discount rate PM $ 1,100 $ 1,500 7% discount rate $ 1,000 $ 1,400 Hospital Admissions from Respiratory Causes O3 and PM $ 85 $ 130 Hospital Admissions from Cardiovascular Causes PM $ 78 $ 110 Emergency Room Visits for Asthma O3 and PM $ 2.0 $ 2.6 Acute bronchitis ( children, 8 12) PM $ 4.3 $ 5.7 Lower respiratory symptoms ( children, 7 14) PM $ 2.3 $ 3.0 Upper respiratory symptoms ( asthmatic children, 9 11) PM $ 13 $ 17 Asthma exacerbations PM $ 8.0 $ 10 Work loss days ( adults, 18 65) PM $ 140 $ 170 Minor restricted activity days ( adults, age 18 65) O3 and PM $ 320 $ 440 School absence days ( children, age 6 11) O3 $ 13 $ 28 Worker productivity ( outdoor workers, age 18 65) O3 $ 8.1 $ 17 Recreational visibility ( Southeastern Class I areas) PM $ 880 $ 1,400 Monetized Totald Base estimate 3% discount rate O3 and PM $ 58,000+ B $ 84,000+ B 7% discount rate $ 54,000+ B $ 79,000+ B a Monetary benefits are rounded to two significant digits.

b Monetary benefits are adjusted to account for growth in real GDP per capita between 1990 and the analysis year ( 2010 or 2015).

c Valuation assumes the 5 year distributed lag structure described earlier. Results reflect the use of two different discount rates; a 3 percent rate which is recommended by EPA's Guidelines for Preparing Economic Analyses ( US EPA, 2000c), and 7 percent which is recommended by OMB Circular A 94 ( OMB, 1992).

d B represents the monetary value of the nonmonetized health and welfare benefits. A detailed listing of unquantified PM, ozone, and mercury related health effects is provided in Table XI B. 1. 4 87 Our estimate of total monetized benefits in 2010 for the proposed rule is $ 58 billion using a 3 percent discount rate and $ 54 billion using a 7 percent discount rate. In 2015, the monetized benefits are estimated at $ 84 billion using a 3 percent discount rate and $ 79 billion using a 7 percent discount rate. Health benefits account for 98 percent of total benefits,

mainly because we are unable to quantify most of the non health benefits. The monetized benefit associated with reductions in the risk of premature mortality, which accounts for $ 53 billion in 2010 and $ 77 billion in 2015, is over 90 percent of total monetized health benefits.

The next largest benefit is for reductions in chronic illness ( CB and non fatal heart attacks),

although this value is more than an order of magnitude lower than for premature mortality.

Hospital admissions for respiratory and cardiovascular causes, visibility, minor restricted activity days, work loss days, school absence days, and worker productivity account for the majority of the remaining benefits. The remaining categories account for less than $ 10 million each, however, they represent a large number of avoided incidences affecting many individuals.

A comparison of the incidence table to the monetary benefits table reveals that there is not always a close correspondence between the number of incidences avoided for a given endpoint and the monetary value associated with that endpoint. For example, there are 100 times more work loss days than premature mortalities, yet work loss days account for only a very small fraction of total monetized benefits. This reflects the fact that many of the less severe health effects, while more common, are valued at a lower level than the more severe health effects. Also, some effects, such as hospital admissions, are valued using a proxy measure of WTP. As such the true value of these effects may be higher than that reported in Table 4 16.

Ozone benefits are in aggregate positive for the nation. However, due to ozone increases occurring during certain hours of the day in some urban areas, there is a dampening of overall ozone benefits in both 2010 and 2015, although the net incidence and benefits estimates for all health effects categories are net positive. Overall, ozone benefits are low relative to PM benefits for similar endpoint categories because of the increases in ozone concentrations during some hours of some days in certain urban areas.

4.3 Discussion This analysis has estimated the health and welfare benefits of reductions in ambient concentrations of particulate matter and ozone resulting from reduced emissions of NOx and SO 2 from affected EGUs. The result suggests there will be significant health and welfare 4 88 benefits arising from the regulation of emissions from EGUs in the U. S. Our estimate that 13,000 premature mortalities would be avoided in 2015, when emission reductions from the regulation are fully realized, provides additional evidence of the important role that pollution from the EGU sector plays in the public health impacts of air pollution.

To examine the importance of specific assumptions and analytical choices we made for this analysis, we will be providing a number of sensitivity analyses in an appendix to be completed for the upcoming Supplemental Analysis of the proposed rule. In addition, there are other uncertainties that we could not quantify, such as the importance of unquantified effects and uncertainties in the modeling of ambient air quality. Inherent in any analysis of future regulatory programs are uncertainties in projecting atmospheric conditions, and sourcelevel emissions, as well as population, health baselines, incomes, technology, and other factors. The assumptions used to capture these elements are reasonable based on the available evidence. However, data limitations prevent an overall quantitative estimate of the uncertainty associated with estimates of total economic benefits. If one is mindful of these limitations, the magnitude of the benefit estimates presented here can be useful information in expanding the understanding of the public health impacts of reducing air pollution from EGUs.

The U. S. EPA will continue to evaluate new methods and models and select those most appropriate for the estimation the health benefits of reductions in air pollution. It is important to continue improving benefits transfer methods in terms of transferring economic values and transferring estimated Impact functions. The development of both better models of current health outcomes and new models for additional health effects such as asthma and high blood pressure will be essential to future improvements in the accuracy and reliability of benefits analyses ( Guo et al., 1999; Ibald Mulli et al., 2001). Enhanced collaboration between air quality modelers, epidemiologists, and economists should result in a more tightly integrated analytical framework for measuring health benefits of air pollution policies. The Agency welcomes comments on how we can improve the quantification and monetization of health and welfare effects and on methods for characterizing uncertainty in our estimates. 5 1 SECTION 5 QUALITATIVE ASSESSMENT OF NONMONETIZED BENEFITS 5.1 Introduction This proposal will result in benefits in addition to the enumerated human health and welfare benefits resulting from reductions in ambient levels of PM and ozone. This rule will also result in benefits that we were unable to monetize. This chapter discusses welfare benefits associated with reduced acid deposition, reduced eutrophication in water bodies, and the reduced health and welfare effects due to the deposition of mercury. Welfare benefits including visibility benefits, agricultural, forestry and other benefits due to reductions in ozone levels, and benefits from reductions in materials damage are discussed in chapter 4 of this report. In contrast to the benefits discussed, it is also possible that this proposal will lessen the benefits of passive fertilization for forest and terrestrial ecosystems where nutrients are a limiting factor and for some croplands.

5.2 Atmospheric Deposition of Sulfur and Nitrogen   Impacts on Aquatic, Forest, and Coastal Ecosystems Atmospheric deposition of sulfur and nitrogen, more commonly known as acid rain,

occurs when emissions of SO 2 and NO x react in the atmosphere ( with water, oxygen, and oxidants) to form various acidic compounds. These acidic compounds fall to earth in either a wet form ( rain, snow, and fog) or a dry form ( gases and particles). Prevailing winds transport the acidic compounds hundreds of miles, often across state and national borders. Acidic compounds ( including small particles such as sulfates and nitrates) cause many negative environmental effects. These pollutants

acidify lakes and streams,

harm sensitive forests, and

harm sensitive coastal ecosystems.

The effect of atmospheric deposition of acids on freshwater and forest ecosystems depends largely on the ecosystem's ability to neutralize the acid ( Driscoll et al., 2001). This is referred 5 2 to as an ecosystem's acid neutralizing capacity ( ANC). Acid neutralization occurs when positively charged ions such as calcium, potassium, sodium, and magnesium, collectively known as base cations, are released. As water moves through a watershed, two important chemical processes act to neutralize acids. The first involves cation exchange in soils, a process by which hydrogen ions from the acid deposition displace other cations from the surface of soil particles, releasing these cations to soil and surface water. The second process is mineral weathering, where base cations bound in the mineral structure of rocks are released as the minerals gradually break down over long time periods. As the base cations are released by weathering, they neutralize acidity and increase the pH level in soil water and surface waters. Acid deposition, because it consists of acid anions ( e. g., sulfate, nitrate), leaches some of the accumulated base cation reserves from the soils into drainage waters. The leaching rate of these base cations may accelerate to the point where it significantly exceeds the resupply via weathering ( Driscoll et al., 2001).

Soils, forests, surface waters and aquatic biota ( fish, algae, and the rest), and coastal ecosystems share water, nutrients, and other essential ecosystem components and are inextricably linked by the chemical processes described above. For example, the same base cations that help to neutralize acidity in lakes and streams are also essential nutrients in forest soils, meaning that cation depletion both increases freshwater acidification and decreases forest productivity. Similarly, the same nitrogen atom that contributes to stream acidification can ultimately contribute to coastal eutrophication as it travels downstream to an estuarine environment. Therefore, to understand the full effects of atmospheric deposition, it is necessary to recognize the interactions between all of these systems.

5.2.1 Freshwater Acidification Acid deposition causes acidification of surface waters. In the 1980s, acid rain was found to be the dominant cause of acidification in 75 percent of acidic lakes and 50 percent of acidic streams. Areas especially sensitive to acidification include portions of the Northeast ( particularly the Adirondack and Catskill Mountains, portions of New England, and streams in the mid Appalachian highlands) and Southeastern streams. Some high elevation Western lakes, particularly in the Rocky Mountains, have become acidic, especially during snowmelt.

However, although many Western lakes and streams are sensitive to acidification, they are not subject to continuously high levels of acid deposition and so have not become chronically acidified ( NAPAP, 1990). 5 3 ANC, a key indicator of the ability of the water and watershed soil to neutralize the acid deposition it receives, depends largely on the watershed's physical characteristics:

geology, soils, and size. Waters that are sensitive to acidification tend to be located in small watersheds that have few alkaline minerals and shallow soils. Conversely, watersheds that contain alkaline minerals, such as limestone, tend to have waters with a high ANC.

As acidity increases, aluminum leached from the soil flows into lakes and streams and can be toxic to aquatic species. The lower pH levels and higher aluminum levels that result from acidification make it difficult for some fish and other aquatic species to survive, grow,

and reproduce. In some waters, the number of species of fish able to survive has been directly correlated to water acidity. Acidification can also decrease fish population density and individual fish size ( U. S. Department of the Interior 2003).

Recent watershed mass balance studies in the Northeast reveal that loss of sulfate from the watershed exceeds atmospheric sulfur deposition ( Driscoll et al., 2001). This suggests that these soils have become saturated with sulfur, meaning that the supply of sulfur from deposition exceeds the sulfur demands of the ecosystem. As a result, sulfur is gradually being released or leached from the watershed into the surface waters as sulfate. Scientists now expect that the release of sulfate that previously accumulated in watersheds will delay the recovery of surface waters in the Northeast that is anticipated in response to the recent SO 2 emission controls ( Driscoll et al., 2001).

A recent study at a stream in the Catskill Mountains found that stream nitrate concentrations were positively correlated to mean annual air temperature but not to annual nitrogen deposition ( Murdoch et al., 1998). This research suggests that, in nitrogen saturated soils, microbial processes ( nitrogen mineralization and nitrification), which are sensitive to changes in temperature and moisture, are the primary factors controlling nitrate leaching,

rather than atmospheric deposition or vegetation uptake of nitrogen. Therefore, declines in nitrogen deposition in nitrogen saturated soils may not immediately lead to improvements in stream water chemistry ( Murdoch et al., 1998).

A major study of the ecological response to acidification is taking place in the Bear Brook Watershed in Maine. Established in 1986 as part of the EPA's Watershed Manipulation Project, the project has found that experimental additions of sulfur and nitrogen to the watershed increased the concentrations of both sulfate and nitrate in the West Bear Brook stream. Stream water concentrations of several other ions, including base cations,

aluminum, and ANC, changed substantially as well ( Norton et al., 1999). During the first year 5 4 of treatment, 94 percent of the nitrogen added experimentally to the Bear Brook watershed was retained, while the remainder leached into streams as nitrate. Nitrogen retention decreased to about 82 percent in subsequent years ( Kahl et al., 1993, 1999). Although the forest ecosystem continued to accumulate nitrogen, nitrate leaching into the stream continued at elevated levels throughout the length of the experiment. This nitrate contributed to both episodic and chronic acidification of the stream. This and other similar studies have allowed scientists to quantify acidification and recovery relationships in eastern watersheds in much more detail than was possible in 1990.

The Appalachian Mountain region receives some of the highest rates of acid deposition in the United States ( Herlihy et al., 1993). The acid base status of stream waters in forested upland watersheds in the Appalachian Mountains was extensively investigated in the early 1990s ( e. g., Church et al. [ 1992], Herlihy et al. [ 1993], Webb et al. [ 1994], van Sickle and Church [ 1995]). A more recent assessment of the southern Appalachian region from West Virginia to Alabama identified watersheds that are sensitive to acid deposition using geologic bedrock and the associated buffering capacity of soils to neutralize acid. The assessment found that approximately 59 percent of all trout stream length in the region is in areas that are highly vulnerable to acidification, and that 27 percent is in areas that are moderately vulnerable ( SAMAB, 1996). Another study estimated that 18 percent of potential brook trout streams in the mid Appalachian Mountains are too acidic for brook trout survival ( Herlihy et al., 1996). Perhaps the most important study of acid base chemistry of streams in the Appalachian region in recent years has been the Virginia Trout Stream Sensitivity Study ( Webb et al., 1994). Trend analyses of these streams indicate that few long term sampling sites are recovering from acidification, most are continuing to acidify, and the continuing acidification is at levels that are biologically significant for brook trout populations ( Webb et al., 2000).

<IP_ADDRESS> Water/ Watershed Modeling Researchers have used models to help them understand and predict atmospheric,

environmental, and human health responses to acid deposition for well over 20 years. Since 1990, watershed modeling capabilities have also improved as researchers are continuing to refine and expand models that project acidification of waterbodies. Unlike the response of air quality and deposition to changes in emissions, lakes and streams take years to decades to fully reflect reductions in acid deposition. In some cases, soil chemistry has been significantly altered and ions must either build up or be leached out before the chemistry can return to its pre acidification status. Therefore, lake and stream conditions are presented for 2030. 5 5 <IP_ADDRESS> Description of the MAGIC Model and Methods A number of mathematical models of soil and surface water acidification in response to atmospheric deposition were developed in the early 1980s ( e. g., Christopherson and Wright [ 1981]; Christopherson et al. [ 1982]; Schnoor et al. [ 1984]; Booty and Kramer [ 1984];

Goldstein et al. [ 1984]; Cosby et al. [ 1985a, b, c]). These models were based on process level information about the acidification process and were built for a variety of purposes ranging from estimating transient water quality responses for individual storm events to estimating chronic acidification of soils and base flow surface water. One of these models ( MAGIC   the Model of Acidification of Groundwater In Catchments; Cosby et al. [ 1985a, b, c]) has been in use now for more than 15 years. MAGIC has been applied extensively in North America and Europe to both individual sites and regional networks of sites and has also been used in Asia,

Africa, and South America. The utility of MAGIC for simulating a variety of water and soil acidification responses at the laboratory, plot, hillslope, and catchment scales has been tested using long term monitoring data and experimental manipulation data. MAGIC has been widely used in policy and assessment activities in the United States and in several countries in Europe.

<IP_ADDRESS> Model Structure MAGIC is a lumped parameter model of intermediate complexity, developed to predict the long term effects of acidic deposition on surface water chemistry. The model simulates soil solution chemistry and surface water chemistry to predict the monthly and annual average concentrations of the major ions in these waters. MAGIC consists of the following: 1) a section in which the concentrations of major ions are assumed to be governed by simultaneous reactions involving sulfate adsorption, cation exchange,

dissolutionprecipitation speciation of aluminum, and dissolution speciation of inorganic carbon; and 2) a mass balance section in which the flux of major ions to and from the soil is assumed to be controlled by atmospheric inputs, chemical weathering, net uptake, and loss in biomass and losses to runoff. At the heart of MAGIC is the size of the pool of exchangeable base cations in the soil. As the fluxes to and from this pool change over time owing to changes in atmospheric deposition, the chemical equilibria between soil and soil solution shift to give changes in surface water chemistry. The degree and rate of change of surface water acidity thus depend both on flux factors and the inherent characteristics of the affected soils.

Cation exchange is modeled using equilibrium ( Gaines Thomas) equations with selectivity coefficients for each base cation and aluminum. Sulfate adsorption is represented 5 6 by a Langmuir isotherm. Aluminum dissolution and precipitation are assumed to be controlled by equilibrium with a solid phase of aluminum trihydroxide. Aluminum speciation is calculated by considering hydrolysis reactions as well as complexation with sulfate, fluoride,

and dissolved organic compounds. Effects of carbon dioxide on pH and on the speciation of inorganic carbon are computed from equilibrium equations. Organic acids are represented in the model as tri protic analogues. Weathering rates are assumed to be constant. Two alternate mechanisms are offered for simulation of nitrate and ammonium in soils: either 1) first order equations representing net uptake and retention or 2) a set of equations and compartments describing process based nitrogen dynamics in soils controlled by soil nitrogen pools. Input output mass balance equations are provided for base cations and strong acid anions, and charge balance is required for all ions in each compartment. Given a description of the historical, current, and expected future deposition at a site, the model equations are solved numerically to give long term reconstructions of surface water chemistry ( for complete details of the model see Cosby et al. [ 1985 a, b, c], [ 2001]).

MAGIC has been used to reconstruct the history of acidification, to examine current patterns of recovery, and to simulate the future trends in stream water acidity in both individual catchment and regional applications at a large number of sites across North America and Europe ( e. g., Beier et al. [ 1995]; Cosby et al. [ 1985b, 1990, 1995, 1996, 1998];

Ferrier, et al. [ 2001]; Hornberger et al. [ 1989]; Jenkins et al. [ 1990]; Moldan et al. [ 1998];

Norton et al. [ 1992]; Whitehead et al. [ 1988, 1997]; Wright et al. [ 1990, 1994, 1998]).

<IP_ADDRESS> Model Implementation Atmospheric deposition and net uptake release fluxes for the base cations and strong acid anions are required as inputs to the model. These inputs are generally assumed to be uniform over the catchment. Atmospheric fluxes are calculated from concentrations of the ions in precipitation and the rainfall volume into the catchment. The atmospheric fluxes of the ions must be corrected for dry deposition of gas, particulates, and aerosols and for inputs in cloud/ fog water. The volume discharge for the catchment must also be provided to the model. In general, the model is implemented using average hydrologic conditions and meteorological conditions in annual or seasonal simulations ( i. e., mean annual or mean monthly deposition); precipitation and lake discharge are used to drive the model. Values for soil and surface water temperature, partial pressure of carbon dioxide, and organic acid concentrations must also be provided at the appropriate temporal resolution. 5 7 As implemented in this project, the model is a two compartment representation of a catchment. Atmospheric deposition enters the soil compartment, and the equilibrium equations are used to calculate soil water chemistry. The water is then routed to the stream compartment, and the appropriate equilibrium equations are reapplied to calculate runoff chemistry.

Once initial conditions ( initial values of variables in the equilibrium equations) have been established, the equilibrium equations are solved for soil water and surface water concentrations of the remaining variables. These concentrations are used to calculate the lake discharge output fluxes of the model for the first time step. The mass balance equations are ( numerically) integrated over the time step, providing new values for the total amounts of base cations and strong acid anions in the system. These in turn are used to calculate new values of the remaining variables, new lake discharge fluxes, and so forth. The output from MAGIC is thus a time trace for all major chemical constituents for the period of time chosen for the integration.

<IP_ADDRESS> Calibration Procedure The aggregated nature of the model requires that it be calibrated to observed data from a system before it can be used to examine potential system response. Calibration is achieved by setting the values of certain parameters within the model that can be directly measured or observed in the system of interest ( called " fixed" parameters). The model is then run ( using observed atmospheric and hydrologic inputs) and the simulated values of surface water and soil chemical variables ( called " criterion" variables) are compared to observed values of these variables. If the observed and simulated values differ, the values of another set of parameters in the model ( called " optimized" parameters) are adjusted to improve the fit.

After a number of iterations, the simulated minus observed values of the criterion variables usually converge to zero ( within some specified tolerance). The model is then considered calibrated. If new assumptions ( or values) for any of the fixed variables or inputs to the model are subsequently adopted, the model must be recalibrated by readjusting the optimized parameters until the simulated minus observed values of the criterion variables again fall within the specified tolerance.

Calibrations are based on volume weighted mean annual or seasonal fluxes for a given period of observation. The length of the period of observation used for calibration is not arbitrary. Model output will be more reliable if the annual flux estimates used in calibration are based on a number of years rather than just 1 year. There is a lot of year to year 5 8 variability in atmospheric deposition and catchment runoff. Averaging over a number of years reduces the likelihood that an " outlier" year ( very dry, etc.) is the primary data on which model forecasts are based. On the other hand, averaging over too long a period may remove important trends in the data that the model needs to simulate.

The calibration procedure requires that stream water quality, soil chemical and physical characteristics, and atmospheric deposition data be available for each catchment. The water quality data needed for calibration are the concentrations of the individual base cations ( Ca, Mg, Na, and K) and acid anions ( Cl, SO 4, and NO 3) and the pH. The soil data used in the model include soil depth and bulk density, soil pH, soil cation exchange capacity, and exchangeable bases on the soil ( Ca, Mg, Na, and K). The atmospheric deposition inputs to the model must be estimates of total deposition, not just wet deposition. In some instances,

direct measurements of either atmospheric deposition or soil properties may not be available for a given site with stream water data. In these cases, the required data can often be estimated by assigning soil properties based on some landscape classification of the catchment and assigning deposition using model extrapolations from some national or regional atmospheric deposition monitoring network.

Soil Physical and Chemical Properties. Soil data for model calibration are usually derived as a really averaged values of soil parameters within a catchment. If soils data for a given location are vertically stratified, the soils data for the individual soil horizons at that sampling site can be aggregated based on horizon, depth, and bulk density to obtain single vertically aggregated values for the site, or the stratified data can be used directly in the model.

Total Atmospheric Deposition. Total atmospheric deposition consists of three components: wet deposition, the flux of ions occurring in precipitation; dry deposition,

resulting from gaseous and particulate fluxes; and cloud/ fog deposition ( which can be particularly important in mountainous inland areas or moderate highlands in areas adjacent to oceans or seas). Estimates of precipitation volume and ionic concentrations in precipitation can be used to calculate wet deposition for a site. Observations of dry deposition or cloud/ fog deposition are very infrequent. The approach usually used to quantify these components relies on some estimate of the ratio of estimated total deposition to the observed wet deposition for important ions ( e. g., sulphate, nitrate, and ammonium ions). These ratios ( called dry deposition factors) are then used to calculate total deposition from the observed wet deposition data. 5 9 Historical Loading. Calibration of the model ( and estimation of the historical changes at the sites) requires a temporal sequence of historical anthropogenic deposition. Our current understanding of ecosystem responses to acidic deposition suggests that future ecosystem responses can be strongly conditioned by historical acidic loadings. Thus, as part of the model calibration process, the model should be constrained by some measure of historical deposition to the site. However, such long term, continuous historical deposition data do not exist. The usual approach is to use historical emissions data as a surrogate for deposition. The emissions for each year in the historical period can be normalized to emissions in a reference year ( a year for which observed deposition data are available). Using this scaled sequence of emissions,

historical deposition can be estimated by multiplying the total deposition estimated for each site in reference year by the emissions scale factor for any year in the past to obtain deposition for that year.

<IP_ADDRESS> MAGIC Modeling Results Watershed modeling undertaken for IAQR projects that, under IAQR, 1 percent of northeastern lakes would be chronically acidic in 2030. In contrast, the same model used to analyze existing control programs projects 6 percent of northeastern lakes would be chronically acidic in 2030. The modeling projects that, under IAQR, 28 percent of northeastern lakes would be episodically acidic in 2030, compared to 25 percent in 2030 under existing control programs. For Adirondack lakes, a subset of northeastern lakes, the signals of surface water chemical recovery are much stronger. Under IAR, no Adirondack lakes would be chronically acidic, and 64 percent would be episodically acidic in 2030, as opposed to 12 percent chronically acidic and 52 percent episodically acidic in 2030 under current control programs.

Because of the age and types of soils in many high elevation areas of the southeast,

streams in that region are more frequently characterized by a delayed response to changes in deposition. For the ecosystems modeled in this region, 17 percent of streams are currently chronically acidic, and this level stays the same under IAQR 2030; the proportion of episodically acidic streams increases from 19 percent under current conditions to 23 percent under IQAR, which reflects a decrease in the proportion of nonacidic streams from 64 percent under current conditions to 60 percent under IQAR in 2030. It is important to note that,

under the Base Case, the proportion of nonacidic streams decreases even further, dropping from 64 percent under current conditions to 58 percent in 2030. Thus, in the southeast,

IQAR would slow the deterioration of stream health ( episodically acidic) expected under the 5 10 Base Case and would prevent additional streams from becoming chronically acidic. Results of the MAGIC modeling are summarized in Table 5 1.

5.2.2 Forest Ecosystems Our current understanding of the effects of acid deposition on forest ecosystems has come to focus increasingly on the effects of biogeochemical processes that affect plant uptake,

retention, and cycling of nutrients within forested ecosystems. Research results from the 1990s indicate that documented decreases in base cations ( calcium, magnesium, potassium,

and others) from soils in the northeastern and southeastern United States are at least partially attributable to acid deposition ( Lawrence et al., 1997; Huntington et al., 2000). Base cation depletion is a cause for concern because of the role these ions play in acid neutralization and,

in the case of calcium, magnesium, and potassium, their importance as essential nutrients for tree growth. It has been known for some time that depletion of base cations from the soil interferes with the uptake of calcium by roots in forest soils ( Shortle and Smith 1988). Recent research indicates it also leads to aluminum mobilization ( Lawrence et al., 1995), which can have harmful effects on fish ( US Dept. of Interior 2003).

The plant physiological processes affected by reduced calcium availability include cell wall structure and growth, carbohydrate metabolism, stomatal regulation, resistance to plant pathogens, and tolerance of low temperatures ( DeHayes et al., 1999). Soil structure, macro and micro fauna, decomposition rates, and nitrogen metabolism are also important processes that are significantly influenced by calcium levels in soils. The importance of calcium as an indicator of forest ecosystem function is due to its diverse physiological roles, coupled with the fact that calcium mobility in plants is very limited and can be further reduced by tree age,

competition, and reduced soil water supply ( McLaughlin and Wimmer 1999).

A clear link has now been established in red spruce stands between acid deposition,

calcium supply, and sensitivity to abiotic stress. Red spruce uptake and retention of calcium is affected by acid deposition in two main ways: leaching of important stores of calcium from needles ( DeHayes et al., 1999) and decreased root uptake of calcium due to calcium depletion from the soil and aluminum mobilization ( Smith and Shortle, 2001; Shortle et al., 1997;

Lawrence et al., 1997). Acid deposition leaches calcium from mesophyll cells of 1 year old red spruce needles ( Schaberg et al., 2000), which in turn reduces freezing tolerance ( DeHayes et al., 1999). These changes increase the sensitivity of red spruce to winter injuries 5 11 under normal winter conditions in the Northeast, result in the loss of needles, and impair the overall health of forest ecosystems ( DeHayes et al., 1999). Red spruce must also expend more metabolic energy to acquire calcium from soils in areas with low calcium/ aluminum ratios, resulting in slower tree growth ( Smith and Shortle, 2001).

Losses of calcium from forest soils and forested watersheds have now been documented as a sensitive early indicator of the soil response to acid deposition for a wide range of forest soils in the United States ( Lawrence et al., 1999; Huntington et al., 2000).

There is a strong relationship between acid deposition and leaching of base cations from hardwood forest ( e. g., maple, oak) soils, as indicated by long term data on watershed mass balances ( Likens et al., 1996; Mitchell et al., 1996), plot and watershed scale acidification experiments in the Adirondacks ( Mitchell et al., 1994) and in Maine ( Norton et al., 1994;

Rustad et al., 1996), and studies of soil solution chemistry along an acid deposition gradient from Minnesota to Ohio ( MacDonald et al., 1992).

Although sulfate is the primary cause of base cation leaching, nitrate is a significant contributor in watersheds that are nearly nitrogen saturated ( Adams et al., 1997). Recent studies of the decline of sugar maples in the Northeast demonstrate a link between low base cation availability, high levels of aluminum and manganese in the soil, and increased levels of tree mortality due to native defoliating insects ( Horsley et al., 2000). The chemical composition of leaves and needles may also be altered by acid deposition, resulting in changes in organic matter turnover and nutrient cycling.

5.2.3 Coastal Ecosystems Since 1990, a large amount of research has been conducted on the impact of nitrogen deposition to coastal waters. It is now known that nitrogen deposition is a significant source of nitrogen to many estuaries ( Valigura et al., 2001; Howarth 1998). The amount of nitrogen entering estuaries due to atmospheric deposition varies widely, depending on the size and location of the estuarine watershed and other sources of nitrogen in the watershed. For a handful of estuaries, atmospheric deposition of nitrogen contributes well over 40 percent of the total nitrogen load; however, in most estuaries for which estimates exist, the contribution from atmospheric deposition ranges from 15 to 30 percent. The area with the highest deposition rates stretches from Massachusetts to the Chesapeake Bay and along the central Gulf of Mexico coast.

Nitrogen is often the limiting nutrient in coastal ecosystems. Increasing the levels of nitrogen in coastal waters can cause significant changes to those ecosystems. Approximately 5 12 60 percent of estuaries in the United States ( 65 percent of the estuarine surface area) suffer from overenrichment of nitrogen, a condition known as eutrophication ( Bricker et al., 1999).

Symptoms of eutrophication include changes in the dominant species of plankton ( the primary food source for many kinds of marine life) that can cause algal blooms, low levels of oxygen in the water column, fish and shellfish kills, and cascading population changes up the food chain. Many of the most highly eutrophic estuaries are along the Gulf and mid Atlantic coasts, overlapping many of the areas with the highest nitrogen deposition, but there are eutrophic estuaries in every region of the coterminous U. S. coastline.

5.3 Benefits of Reducing Mercury Emissions According to baseline emission estimates, the sources affected by this proposal would emit approximately 45.1 tons of mercury per year in 2010. This estimate is specific to fossilfired electric generating units in excess of 25 megawatt capacity. The proposed regulation would reduce approximately 10.6 tons of mercury ( or 23.5 percent) from the 2010 baseline,

11.8 tons of mercury ( or 26.3 percent) from the 2015 baseline, and 14.3 tons ( or 32 percent)

from the 2020 baseline at affected electric generating units.

Mercury emitted from utilities and other natural and man made sources is carried by winds through the air and eventually is deposited to water and land. Recent estimates ( which are highly uncertain) of annual total global mercury emissions from all sources ( natural and anthropogenic) are about 5,000 to 5,500 tons per year ( tpy). Of this total, about 1,000 tpy are estimated to be natural emissions and about 2,000 tpy are estimated to be contributions through the natural global cycle of re emissions of mercury associated with past anthropogenic activity. Current anthropogenic emissions account for the remaining 2,000 tpy.

Point sources such as fuel combustion; waste incineration; industrial processes; and metal ore roasting, refining, and processing are the largest point source categories on a world wide basis. Given the global estimates noted above, U. S. anthropogenic mercury emissions are estimated to account for roughly 3 percent of the global total, and U. S. utilities are estimated to account for about 1 percent of total global emissions. Mercury exists in three forms:

elemental mercury, inorganic mercury compounds ( primarily mercuric chloride), and organic mercury compounds ( primarily methylmercury). Mercury is usually released in an elemental form and later converted into methylmercury by bacteria. Methylmercury is more toxic to humans than other forms of mercury, in part because it is more easily absorbed in the body ( EPA, 1996). 5 13 If the deposition is directly to a water body, then the processes of aqueous fate,

transport, and transformation begin. If deposition is to land, then terrestrial fate and transport processes occur first and then aqueous fate and transport processes occur once the mercury has cycled into a water body. In both cases, mercury may be returned to the atmosphere through resuspension. In water, mercury is transformed to methylmercury through biological processes and for exposures affected by this rulemaking, methylmercury is considered to be the form of greatest concern. Once mercury has been transformed into methylmercury, it can be ingested by the lower trophic level organisms where it can bioaccumulate in fish tissue ( i. e.,

concentrations of mercury remain in the fish's system for a long period of time and accumulates in the fish tissue as predatory fish consume other species in the food chain). Fish and wildlife at the top of the food chain can, therefore, have mercury concentrations that are higher than the lower species, and they can have concentrations of mercury that are higher than the concentration found in the water body itself. In addition, when humans consume fish contaminated with methylmercury, the ingested methymercury is almost completely absorbed into the blood and distributed to all tissues ( including the brain); it also readily passes through the placenta to the fetus and fetal brain ( EPA, 2001a).

Based on the findings of the National Research Council, EPA has concluded that benefits of Hg reductions would be most apparent at the human consumption stage, as consumption of fish is the major source of exposure to methylmercury. At lower levels,

documented Hg exposure effects may include more subtle, yet potentially important,

neurodevelopmental effects. Figure 5 1 shows how emissions of mercury can transport from the air to water and impact human health and ecosystems. 30 Cardiovascular, immune, and reproductive system problems in adults are potential effects as the literature is either contradictory or incomplete.

5 14 Emissions Reductions Reduce Atmospheric Transport and Deposition Reduce Ecosystem Transport and Methylation Reduce Human and Wildlife Exposure Reduce Health Impacts Fishing ° commercial ° recreational ° subsistence Mercury transforms into methylmercury in soils and water, then can bioaccumulate in fish Atmospheric deposition Lake methylation Ocean methylation volatilization volatilization Largest impacts on young children Impacts include:

° Impaired motor and cognitive skills ° Potential cardiovascular,

immune, and reproductive system problems in adults Power Plant Emissions Humans and wildlife affected primarily by eating contaminated fish Wet and Dry Deposition Figure 5 1. How Emissions of Mercury Can Impact Human Health and Ecosystems30 5 15 Some subpopulations in the U. S., such as: Native Americans, Southeast Asian Americans, and lower income subsistence fishers, may rely on fish as a primary source of nutrition and/ or for cultural practices. Therefore, they consume larger amounts of fish than the general population and may be at a greater risk to the adverse health effects from Hg due to increased exposure. In pregnant women, methylmercury can be passed on to the developing fetus, and at sufficient exposure may lead to a number of neurological disorders in children. Thus, children who are exposed to low concentrations of methylmercury prenatally may be at increased risk of poor performance on neurobehavioral tests, such as those measuring attention, fine motor function, language skills, visual spatial abilities ( like drawing),

and verbal memory. The effects from prenatal exposure can occur even at doses that do not result in effects in the mother. Mercury may also affect young children who consume fish contaminated with Hg. Consumption by children may lead to neurological disorders and developmental problems, which may lead to later economic consequences.

Monitoring the concentrations of mercury in the blood of women of child bearing age can help identify the proportion of children who may be at risk. EPA's reference dose ( RfD)

for methylmercury is 0.1 micrograms per kilogram body weight per day, which is approximately equivalent to a concentration of 5.8 parts per billion mercury in blood.

Although the prenatal period is the most sensitive period of exposure, exposure to mercury during childhood also could pose a potential health risk ( NAS, 2000).

Figure 5 2 shows reported concentrations of mercury in blood of women of childbearing age from the National Health and Nutrition Examination Survey ( NHANES)

( EPA, 2003b). The data presented are for total mercury, which includes methylmercury and other forms of mercury. Total blood mercury is a reasonable indicator of methylmercury exposure in people who consume fish and have no significant exposure to inorganic or elemental mercury ( JAMA, April 2003). Thus the measured concentrations are a good indication of methylmercury concentrations. From this survey, about 8 percent of women of child bearing age had at least 5.8 parts per billion of mercury in their blood in 1999 2000. 5 16 Figure 5 2. Concentrations of Mercury in Blood of Women of Childbearing Age 5 17 50 58 BMDL 85 BMD 5.8 RfD Figure 5 3 shows relative values of the BMD, BMDL and the RfD. The data show a Benchmark Dose ( BMD) BMD at 85 ppb. The BMD is the dose or concentration that produced a doubling of the number of children with a response at the 5th percentile of the population. In this case, the changes evaluated were changes on neuropsychological testing batteries ( i. e. the Boston Naming Test). In determining the RfD, EPA started with the BMD ( 85 ppb) and then used the 95% lower confidence limit to arrive at the 58 ppb BMDL. EPA then applied a composite uncertainty factor of 10 to calculate a final RfD of 5.8 ppb. The uncertainty factor adjustment was used to account for pharmacokinetic and pharmacodynamic uncertainty and variability.

Figure 5 3. Relative Values of BMD, BMDL, and the RfD ( Values in ppb)

In response to potential risks of mercury contaminated fish consumption, EPA and FDA have issued fish consumption advisories which provide recommended limits on consumption of certain fish species for different populations. EPA and FDA are currently developing a joint advisory that has been released in draft form. This newest draft FDA EPA fish advisory recommends that women and young children reduce the risks of Hg consumption in their diet by moderating their fish consumption, diversifying the types of fish they consume,

and by checking any local advisories that may exist for local rivers and streams. This collaborative FDA EPA effort will greatly assist in educating the most susceptible populations. Additionally, the reductions of Hg from this regulation may potentially lead to fewer fish consumption advisories ( both from federal or state agencies), which will benefit the fishing community. As Figure 5 4 shows, currently 44 states have issued fish consumption advisories for non commercial fish for some or all of their waters due to contamination of mercury. The scope of FCA issued by states varies considerably, with some warnings applying to all water bodies in a state and others applying only to individual lakes and streams. Note 5 18 that the absence of a state advisory does not necessarily indicate that there is no risk of exposure to unsafe levels of mercury in recreationally caught fish. Likewise, the presence of a state advisory does not indicate that there is a risk of exposure to unsafe levels of mercury in recreationally caught fish, unless people consume these fish at levels greater than those recommended by the fish advisory.

Reductions in methylmercury concentrations in fish should reduce exposure,

subsequently reducing the risks of mercury related health effects in the general population, to children, and to certain subpopulations. Fish consumption advisories ( FCA) issued by the States may also help to reduce exposures to potential harmful levels of methylmercury in fish ( although some studies have shown limited knowledge of and compliance with advisories by at risk populations ( May and Burger, 1996; Burger, 2000)). To the extent that reductions in mercury emissions reduces the probability that a water body will have a FCA issued, there are a number of benefits that will result from fewer advisories, including increased fish consumption, increased fishing choices for recreational fishers, increased producer and consumer surplus for the commercial fish market, and increased welfare for subsistence fishing populations. 5 19 Figure 5 4. Mercury Reductions By State In 2015 There is a great deal of variability among individuals in fish consumption rates;

however, critical elements in estimating methylmercury exposure and risk from fish consumption include the species of fish consumed, the concentrations of methylmercury in the fish, the quantity of fish consumed, and how frequently the fish is consumed. The typical U. S.

consumer eating a wide variety of fish from restaurants and grocery stores is not in danger of consuming harmful levels of methylmercury from fish and is not advised to limit fish consumption. Those who regularly and frequently consume large amounts of fish, either marine or freshwater, are more exposed. Because the developing fetus may be the most sensitive to the effects from methylmercury, women of child bearing age are regarded as the population of greatest interest. The EPA, Food and Drug Administration, and many States have issued fish consumption advisories to inform this population of protective consumption levels.

The EPA's 1997 Mercury Study RTC supports a plausible link between anthropogenic releases of Hg from industrial and combustion sources in the U. S. and methylmercury in fish.

However, these fish methylmercury concentrations also result from existing background concentrations of Hg ( which may consist of Hg from natural sources, as well as Hg which has been re emitted from the oceans or soils) and deposition from the global reservoir ( which 5 20 includes Hg emitted by other countries). Given the current scientific understanding of the environmental fate and transport of this element, it is not possible to quantify how much of the methylmercury in locally caught fish consumed by the U. S. population is contributed by U. S.

emissions relative to other sources of Hg ( such as natural sources and re emissions from the global pool). As a result, the relationship between Hg emission reductions from Utility Units and methylmercury concentrations in fish cannot be calculated in a quantitative manner with confidence. In addition, there is uncertainty regarding over what time period these changes would occur. This is an area of ongoing study.

Given the present understanding of the Hg cycle, the flux of Hg from the atmosphere to land or water at one location is comprised of contributions from: the natural global cycle;

the cycle perturbed by human activities; regional sources; and local sources. Recent advances allow for a general understanding of the global Hg cycle and the impact of the anthropogenic sources. It is more difficult to make accurate generalizations of the fluxes on a regional or local scale due to the site specific nature of emission and deposition processes. Similarly, it is difficult to quantify how the water deposition of Hg leads to an increase in fish tissue levels.

This will vary based on the specific characteristics of the individual lake, stream, or ocean. 6 1 SECTION 6 COMPARISON OF BENEFITS AND COSTS The estimated social costs to implement the proposed IAQR, as described in the cost analysis document, are approximately $ 2.9 billion annually and $ 3.7 billion annually for 2010 and 2015, respectively ( 1999$). Thus, the net benefits ( social benefits minus social costs) of the program in 2010 are approximately $ 55 + B billion annually in 2010 and $ 80 + B billion annually in 2015 ( 1999$). ( B represents the sum of all unquantified benefits and disbenefits.)

Therefore, implementation of the proposed rule is expected, based purely on economic efficiency criteria, to provide society with a significant net gain in social welfare, even given the limited set of health and environmental effects we were able to quantify. Addition of ozone, directly emitted PM 2.5, mercury, acidification, and eutrophication related impacts would increase the net benefits of the proposed rule. As discussed in section IX of the notice for this rulemaking, we did not complete air quality modeling that precisely matches the IAQR region. We anticipate that any differences in estimates presented due to the modeling region analyzed will be small. Table 6 1 presents a summary of the benefits, costs, and net benefits of the proposed rule. 6 2 Table 6 1. Summary of Annual Benefits, Costs, and Net Benefits of the Inter State Air Quality Rule Description 2010 ( billions of 1999 dollars) 2015 ( billions of 1999 dollars)

Social costsa $ 2.9 $ 3.7 Social benefits b, c Ozone related benefits $ 0.1 $ 0.1 PM related health benefits $ 56.8 + B $ 82.3 + B Visibility benefits $ 0.9 $ 1.4 Net benefits ( benefits costs) b, c, d $ 55 + B $ 80 + B a Note that costs are the annual total costs of reducing pollutants including NOx and SO2.

b As the table indicates, total benefits are driven primarily by PM related health benefits. The reduction in premature fatalities each year accounts for over 90 percent of total benefits. Benefits in this table are associated with NOx and SO2 reductions.

c Not all possible benefits or disbenefits are quantified and monetized in this analysis. B is the sum of all unquantified benefits and disbenefits. Potential benefit categories that have not been quantified and monetized are listed in Table 1 4.

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