3
Air Pollution: Sources, Impacts, and Effects

Concern about the effects of air pollution has existed for centuries. In 14th-century England, regulations were introduced regarding the burning of sea coal, and violators were tortured for producing foul odors. In the United States, the first air pollution regulations also dealt with coal, and in the 19th century, coal and smoke ordinances were passed in Chicago, St. Louis, and Cincinnati. In the 20th century, concern about air pollution in the United States can be traced to severe pollution episodes, such as the 1948 episode in Donora, Pennsylvania (near Pittsburgh), which resulted in nearly 7,000 illnesses and 20 deaths. Although rare, episodes such as those in Donora dramatized the acute health effects of air pollutants. Following the passage of the 1970 Clean Air Act Amendments that required the setting of National Ambient Air Quality Standards (NAAQS)—the imposition of emission standards for hazardous air pollutants and control on mobile source emissions—the United States embarked on a process of improving air quality, while maintaining a growing but changing economy.

Beginning in the 1970s, China started to pay attention to pollution caused from coal combustion. China has experienced rapid economic growth of up to 7-8 percent of GDP per year since the mid-1980s. Within that period, the Law of the People's Republic of China on the Prevention and Control of Atmospheric Pollution was passed by the Committee of People's Congress Council in September of 1987, with amendments in 1995 and in 2000. This short period of fast economic growth has led to higher living standards, but has also caused severe problems in environmental pollution. In recent years the Chinese government has made major efforts to reduce emissions, which have partly compensated for the rapid growth in energy consumption and urbanization. Improving air quality has become an urgent task.



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3 Air Pollution: Sources, Impacts, and Effects Concern about the effects of air pollution has existed for centuries. In 14th- century England, regulations were introduced regarding the burning of sea coal, and violators were tortured for producing foul odors. In the United States, the first air pollution regulations also dealt with coal, and in the 19th century, coal and smoke ordinances were passed in Chicago, St. Louis, and Cincinnati. In the 20th century, concern about air pollution in the United States can be traced to severe pollution episodes, such as the 1948 episode in Donora, Pennsylvania (near Pittsburgh), which resulted in nearly 7,000 illnesses and 20 deaths. Although rare, episodes such as those in Donora dramatized the acute health effects of air pollut- ants. Following the passage of the 1970 Clean Air Act Amendments that required the setting of National Ambient Air Quality Standards (NAAQS)—the imposition of emission standards for hazardous air pollutants and control on mobile source emissions—the United States embarked on a process of improving air quality, while maintaining a growing but changing economy. Beginning in the 1970s, China started to pay attention to pollution caused from coal combustion. China has experienced rapid economic growth of up to 7-8 percent of GDP per year since the mid-1980s. Within that period, the Law of the People's Republic of China on the Prevention and Control of Atmospheric Pollution was passed by the Committee of People's Congress Council in Sep- tember of 1987, with amendments in 1995 and in 2000. This short period of fast economic growth has led to higher living standards, but has also caused severe problems in environmental pollution. In recent years the Chinese government has made major efforts to reduce emissions, which have partly compensated for the rapid growth in energy consumption and urbanization. Improving air quality has become an urgent task. 

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 ENERGY FUTURES AND URBAN AIR POLLUTION Air quality management efforts in the United States are guided by the NAAQS, which are set based on health and welfare effects of the major air pollutants: par- ticulate matter (PM), ozone (O3), carbon monoxide (CO), sulfur dioxide (SO2), nitrogen dioxide (NO2), and lead. Standards for PM have been established for two size categories: PM of less than 10 microns in aerodynamic diameter (PM10) and PM of less than 2.5 microns in diameter. Likewise, air quality management efforts in China are guided by National Air Quality Standards, which have been established for SO2, NO2, and PM10. The current health-based NAAQS are listed in Table 3-1, along with air quality standards adopted by China (Table 3-2) and guidelines established by the World Health Organization (WHO) (Table 3-3). In addition to efforts focused on these major air pollutants, both countries are also concerned with toxic air pollutants that are less ubiquitous. These toxic air pollut- ants (or hazardous air pollutants, as they are called in the Clean Air Act) include benzene and other aromatic compounds from motor vehicles, fuels, and other combustion sources, and mercury from solid waste and coal combustion. China has developed an Air Pollution Index (API) based on its air quality stan- dards, and cities use this tool to report their air quality (see Chapter 4, Table 4-1). Like their national standards, the API contains classes (I-V) which allow cities to comply at various levels. Cities are encouraged to achieve Class II standards TABLE 3-1 U.S. National Ambient Air Quality Standards (NAAQS) Pollutant Primary Stds. Averaging Times Secondary Stds. 9 ppm (10 mg/m3) Carbon monoxide 8-hour None 35 ppm (40 mg/m3) 1-hour None 1.5 µg/m3 Lead Quarterly average Same as primary Nitrogen dioxide 0.053 ppm Annual (arithmetic mean) Same as primary (100 µg/m3) Particulate matter (PM10) Revoked Annual (arith. mean) 150 µg/m3 24-hour Particulate matter (PM2.5) 15.0 µg/m3 Annual (arith. mean) Same as primary 35 µg/m3 24-hour Ozone 0.08 ppm (171 8-hour Same as primary µg/m3) 0.12 ppm (257 1-hour Same as primary µg/m3) (Applies only in limited areas) Sulfur oxides 0.03 ppm (78 µg/ Annual (arith. mean) — m 3) 0.14 ppm (364 24-hour — µg/m3) — 3-hour 0.5 ppm (1300 µg/m3)

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 AIR POLLUTION: SOURCES, IMPACTS, AND EFFECTS TABLE 3-2 China’s Ambient Air Quality Standards (GB 3095-1996) (µg/m 3 unless otherwise noted) Pollutant Class I Standard Class II Standard Class III Standard Averaging Times SO2 20 60 100 Annual 50 150 250 Daily 150 500 700 1-hour TSP 80 200 300 Annual 120 300 500 Daily PM10 40 100 150 Annual 50 150 250 Daily NOx 50 50 100 Annual 100 100 150 Daily 150 150 300 1-hour NO2 40 40 80 Annual 80 80 120 Daily 120 120 240 1-hour CO (mg/m3) 4.00 4.00 6.00 Daily 10.00 10.00 20.00 1-hour O3 120 160 200 1-hour TABLE 3-3 World Health Organization (WHO) Air Quality Guidelines (in µg/m3) Interim Interim Interim Pollutant target 1 target 2 target 3 Standard Averaging Times aPM 10 35 25 15 Annual 2.5 25 75 50 37.5 24-hour 20 PM10 70 50 30 Annual 50 150 100 75 24-hour 100 O3 160 — — 8-hour 40 NO2 — — — Annual 200 — — — 1-hour 20 SO2 125 50 — 24-hour 500 — — — 10-minute aPreferred guideline, according to WHO. SOURCE: WHO, 2006. or better, although as of 2006, nearly 48 percent of urban residents lived in cities not meeting these standards (SEPA, 2007). China is not the only country with varying classes of air quality standards, and WHO recently adopted interim targets to accompany its own guidelines, in order to facilitate implementation in more polluted areas (WHO, 2006). A benefit of the interim targets, and the rationale for China’s classes of standards, is that they allow governments to consider local

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 ENERGY FUTURES AND URBAN AIR POLLUTION circumstances in developing an approach to balance health risks, technological feasibility, economic considerations, and other factors (WHO, 2006). However, by virtue of its formulation in China, in which air quality is reported by class and dominant pollutant (see Chapter 9, Table 9-5), the API has some drawbacks as well. Since the API reflects a comprehensive index of all pollutants, and air quality rankings reflect the lowest level of compliance, cities have tended to focus on a particular pollutant (e.g., PM10) in order to improve their overall rankings. While this clearly provides benefits associated with the reduction of the particular pollutant, it also works against efforts to adopt a multipollutant reduction strategy. It also may lead governments to overlook increasing trends in concentrations of pollutants which are still satisfying a certain criteria; emissions targets in China are established with a target API rank in mind, providing a disincentive to limit emissions other than for the predominant pollutant. This chapter first discusses the effects of air pollution in the United States and China, and then reviews trends in their air pollutant emissions and concentrations, and ends with a discussion of key source-receptor relationships. AIR POLLUTION EFFECTS United States Health Effects of Air Pollution In the United States, air pollution is regulated because of concerns about its impact on human health, visibility, and the environment (NRC, 2004). Economic analysis of air pollution control efforts in the United States indicate that histori- cally, the benefits have far outweighed the costs. In 1997, the U.S. Environmental Protection Agency (EPA) estimated that the costs of control measures undertaken in the United States from 1970 through 1990 totaled nearly $500 billion (1990 dollars). The benefits accruing from the emissions reductions over that same period totaled more than $20 trillion (1990 dollars), outweighing the costs by a ratio of 40 to 1 (EPA, 1997). Control programs adopted more recently still show highly favorable benefit-to-cost balances. In 2005, the EPA adopted the Clean Air Interstate Rule, requiring an additional 70 percent reduction in SO2 emissions and a 60 percent reduction in NOx emissions from large stationary sources by 2015. EPA estimates the cost of these controls will be about $3 billion per year in 2015, while the annual benefits that year will be about $90 billion (EPA, 2005). Analyz- ing the health benefits of any regulation requires flexible, innovative, and multi- disciplinary participation and guidance from scientific experts (NRC, 2002). Over the past 15 years, as the concentrations of CO, SO2, NO2, and Pb have declined, the focus of health studies and control efforts has increasingly turned to PM and O3 as the most important major air pollutant species of concern. Cor- respondingly, the primary focus of this section is on the current understanding of

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 AIR POLLUTION: SOURCES, IMPACTS, AND EFFECTS the health effects of PM and O3 in the United States. This section also discusses the risks of selected toxic air pollutants, which are generally less ubiquitous than the major pollutants mentioned above, but which may also have profound health implications where significant exposures occur (see also Box 3-1). There have been several recent reviews of the improved understanding of the health effects of exposure to PM2.5 (Lippmann et al., 2003; EPA, 2004; Pope and Dockery, 2006). Epidemiological studies have linked exposure to PM2.5 to a range of adverse respiratory effects. Significantly, both short- and long-term PM2.5 exposures are linked to a heightened risk of premature mortality. There is roughly a 10 percent increase in adult mortality rates for every 10 µg/m3 of annual-average PM2.5, a 0.25-1 percent increase per 10 µg/m3 24-hour average PM10, and 0.2-0.8 percent increase per 10 µg/m3 increase in 1-hour peak ozone (Pope et al., 2002; Cohen et al., 2004; WHO, 2006; Smith et al., 2004; HEI, 2006; Ostro et al., 2006). Some short-term studies have also linked “coarse” particles in the size range from 2.5 to 10 microns (PM10-2.5) to premature mortality, but the results for this size fraction are less consistent than for PM2.5 (Brunekreef and Forsberg, 2005). An important change in focus for fine particles (PM2.5) has been from effects on the respiratory system to their role in cardiovascular effects. It is now hypothesized that much of the mortality that is associated with particle exposure results is related to effects on the cardiovascular system (EPA, 2004). This change in focus has led to substantial improvements in the understanding of pathways by which particulate pollutants might induce significant cardiac effects. In terms of morbidity effects, there may be a greater role for larger particles, including particles in the 2.5 to 10 µm size range (PM10-2.5). For example, some epidemiological studies have found that PM10 and PM10-2.5 have greater effects than PM2.5, particularly on emergency room visits (Brunekreef and Forsberg, 2005). Nonetheless, the health impacts of PM2.5 are both better known and larger than for other pollutants, thus countries are beginning to focus more attention on PM2.5 concentrations. Burden of disease estimates are helpful in assessing priori- ties to manage health-related pollutants. Cohen et al. (2005) estimate that outdoor ambient PM2.5 pollution causes 3 percent of mortality from cardiopulmonary disease, roughly 5 percent of mortality from various cancers, and about 1 per- cent of mortality from acute respiratory infections in children 5 years and under, worldwide. This burden is unevenly distributed, though, occurring predominantly in developing countries (and 65 percent in Asia alone). Owing to uncertainties, including the fact that only mortality was considered in the assessment, Cohen et al. suggest that the impact of urban air pollution on burden of disease is likely underestimated. Two age groups, older adults and the very young, are potentially at greater risk for PM-related effects. Epidemiologic studies have generally not shown strik- ing differences between adult age groups. However, some epidemiologic studies have suggested that serious health effects, such as premature mortality, are greater among older populations (e.g., Dockery et al., 1993; Pope et al., 1995). Epidemio-

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 ENERGY FUTURES AND URBAN AIR POLLUTION logic evidence has reported associations with emergency hospital admissions for respiratory illness and asthma-related symptoms in children. In the United States, approximately 22 million people, or 11 percent of the population, have received a diagnosis of heart disease, about 20 percent of the population have hypertension, and about 9 percent of adults and 11 percent of children in the United States have been diagnosed with asthma. In addition, about 26 percent of the U.S. population are under 18 years of age, and about 12 percent are 65 years of age or older. To put the estimates of premature mortality impacts of PM into context, overall health statistics indicate that there are approximately 2.5 million deaths from all causes per year in the U.S. population, with about 100,000 deaths from chronic lower respiratory diseases (Kochanek et al., 2004). EPA estimates the cost of meeting the revised 24-hour PM2.5 standards at $5.4 billion in 2020. This estimate includes the costs of purchasing and installing controls for reducing pollution to meet the standard. It also estimates that meeting the revised 24-hour standard will result in health and visibility benefits ranging from $9 billion to $76 billion a year by 2020. These costs and benefits are in addition to the estimated costs ($7 billion) and benefits ($20-160 billion per year by 2015) associated with meeting the 1997 standards for fine PM (EPA, 2006b). Epidemiological and clinical evidence links short-term exposure to elevated ozone levels to respiratory symptoms and illness, with epidemiological studies also showing a positive association with emergency room visits and hospital admissions (EPA, 2007b). There is also evidence for an association between elevated O3 concentrations and premature mortality (Bell et al., 2004; Gryparis et al., 2004). There are limited health effects that can be ascribed to the other criteria pollutants, given the apparent large influence of PM and ozone. Although EPA does make estimates in the relevant criteria documents, it appears that there are relatively few deaths or serious illnesses arising from exposure to these other pollutants at the levels currently present in the United States. The last assessment of the effects of hazardous air pollutants was performed based on emissions and concentration estimates for 1999. From a national per- spective, benzene is the most significant air toxic for which cancer risk could be estimated, contributing 25 percent of the average individual cancer risk identified in this assessment (EPA, 2006a). Based on EPA’s national emissions inventory, the key sources for benzene are on-road (49 percent) and non-road mobile sources (19 percent). EPA projects that on-road and non-road mobile source benzene emissions will decrease by about 60 percent between 1999 and 2020, as a result of motor vehicle standards, fuel controls, standards for non-road engines and equip- ment, and motor vehicle inspection and maintenance programs. Most of these programs reduce benzene simultaneously with other volatile organic compounds. Diesel PM, which often contains benzene and other toxic pollutants, is currently regulated in California, where estimates attribute 70 percent of cancers resulting from ambient air pollution to diesel PM (SCAQMD, 2000). EPA lists diesel PM as a possible carcinogen, but has not adopted a unit risk factor.

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 AIR POLLUTION: SOURCES, IMPACTS, AND EFFECTS Of the 40 air toxics showing the potential for acute respiratory effects, acro- lein is the most significant, contributing 91 percent of the nationwide average non-cancer hazard identified in EPA’s assessment (SCAQMD, 2000). Note that the health information and exposure data for acrolein include much more uncer- tainty than those for benzene. Based on the national emissions inventory, the key sources for acrolein are open burning, prescribed fires and wildfires (61 percent), and on-road (14 percent) and non-road (11 percent) mobile sources. The apparent dominance of acrolein as a non-cancer “risk driver” in both the 1996 and 1999 national-scale air toxics assessments has led to efforts to develop an effective monitoring test method for this pollutant. EPA projects acrolein emissions from on-road sources will be reduced by 53 percent between 1996 and 2020, as a result of existing motor vehicle standards and fuel controls. EPA’s national air toxics assessment estimates that in most of the country, people have a lifetime cancer risk from breathing air toxics between 1 and 25 in one million (SCAQMD, 2000). This means that out of one million people, between 1 and 25 have an increased likelihood of contracting cancer as a result of breathing air toxics from outdoor sources, if they were exposed to 1999 levels over the course of their lifetime. The assessment estimates that most urban locations in the United States have air toxics lifetime cancer risk greater than 25 in a million. Risks in transportation corridors and in some other locations are greater than 50 in a million. In contrast, one out of every three Americans (330,000 in a million) will contract cancer during a lifetime, when all causes (including exposure to air toxics) are taken into account. Based on these results, the risk of contracting cancer is increased less than 1 percent due to the inhalation of air toxics from outdoor sources. As another comparison, the national risk of contracting cancer from radon exposure is on the order of 1 in 500 (2,000 in a million). Note that risks from human-caused air toxics are commonly viewed differently, because they are involuntary and subject to control, than are risks that are naturally occur- ring or voluntarily assumed. Mercury is a toxic metal that is widely distributed around the globe due to emissions from both anthropogenic and natural sources. Exposure to high levels of mercury is associated with neurologic and kidney damage. For most people, the greatest health risk from mercury is posed by the consumption of fish con- taminated with methyl mercury, which can bioaccumulate up the food chain. Impacts on fetal development due to maternal exposure are of special concern. In the United States, the Center for Disease Control and Prevention’s National Health and Nutrition Examination Survey found that approximately 6 percent of childbearing-aged women had blood mercury levels at or above the reference dose of 5.8 µg/L, an estimated level assumed to be without appreciable harm (CDC, 2004). Most states in the United States have extensive advisories recommending that people limit consumption of fish caught in local waters, in order to avoid excessive mercury intake. In response to concerns about mercury contamination, international efforts are under way to reduce mercury use in manufacturing. In

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 ENERGY FUTURES AND URBAN AIR POLLUTION BOX 3-1 Emissions, Exposure, and Intake Fraction In addition to information on ambient concentrations, it is also useful to have an understanding of human exposure and intake fraction (iF) when considering human health impacts (Bennet et al., 2002). Intake fraction refers to the mass intake of pollutants by people during a given amount of time, relative to the mass of emissions released into the environment. This concept is useful in helping to set priorities in cities—emission sources such as cement mixers in urban areas exhibit high intake fractions, and therefore are more likely to impact human health than are sources which are located further from population centers, or which disperse their pollutants more effectively (e.g., through taller smokestacks). Mobile sources, specifically automobiles, also contribute disproportionately to human exposure, due to their prevalence in population centers, emitting at street level (Laden et al., 2000; Marshall et al., 2005). Of course, simply relocating or redistributing these emissions, while ostensibly reducing the risk in a given urban population, can potentially transfer this risk to other regions. Therefore, decisions to relocate or redistribute sources of emissions must be considered in light of downwind popula- tions and the regional airshed. Studies in the United States and in European cities have underscored the large contribution of mobile sources (Laden et al., 2000; Schwartz et al., 2002; Hoek et al., 2002; Maynard et al. 2007) to air pollution-related mortality. Recent s tudies in China have indicated that, although the electrical power generation sec- tor is responsible for the bulk of SO2 and TSP emissions, its iF value is generally much lower than are three other key industries: mineral production, chemical, and m etallurgy (Wang et al., 2006; Ho and Nielsen, 2007). Improved understanding of the relationship between pollution sources and human exposure will further aid cities in assessing their risks and in developing appropriate strategies to reduce air pollution. the United States, control technology has been required to significantly reduce mercury emissions from medical waste incinerators and from municipal solid waste combustors, and control requirements have recently been adopted for coal- fired power plants. Enironmental Effects The welfare effects of greatest interest include visibility degradation, impacts on crop production and ecosystem health, and materials damage. During the 1970s and 1980s, recognition that lakes and streams in the eastern United States were becoming acidified due to atmospheric deposition led to requirements for power plants to significantly reduce emissions of sulfur and nitrogen oxides (NRC, 2004). Recovery of some lakes and streams in the northeastern United States

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 AIR POLLUTION: SOURCES, IMPACTS, AND EFFECTS is now being observed as a consequence of these emissions reductions, but full recovery will require additional controls and, in some places, will take decades (NAPAP, 2005). Deposition of reactive nitrogen in the form of ammonia and ammonium and nitrate ions is receiving increasing attention in the United States, due to concerns about eutrophication of coastal zones and over-fertilization of sensitive terrestrial ecosystems (NRC, 2004). Visibility is most affected by airborne particulate matter, particularly PM with particle diameters between 0.1 and 1.0 µm. The U.S. goal for visibility is to restore visibility in protected national parks and wilderness areas to natural levels by 2064. In conjunction with the National Park Service, other federal land managers, and state organizations, the EPA has supported visibility monitoring in national parks and wilderness areas since 1988. The monitoring network was originally established at 20 sites, but it has now been expanded to 110 sites that represent all but one (Bering Sea) of the 156 mandatory Federal Class I areas across the country. Annual average visibility conditions (reflecting light extinction due to both anthropogenic and non-anthropogenic sources) vary regionally across the United States. The rural east generally has higher levels of impairment than do remote sites in the west, with the exception of urban-influenced sites such as San Gorgonio Wilderness (California) and Point Reyes National Seashore (California), which have annual average levels comparable to certain sites in the northeast (EPA, 2004). Higher visibility impairment levels in the east are due to generally higher concentrations of anthropogenic fine particles, particularly sulfates, and higher average relative humidity levels. In fact, sulfates account for 60-86 percent of the haziness in eastern sites. Regional trends in Class I area visibility are updated and presented in the EPA’s National Air Quality and Emissions Trends Report (EPA, 2001). Eastern trends for the 20 percent haziest days from 1992-1999 showed about a 16 percent improvement. However, visibility in the east remains significantly impaired, with an average visual range of approximately 20 km on the 20 percent haziest days. In western Class I areas, aggregate trends showed little change during 1990-1999 for the 20 percent haziest days. Average visual range on the 20 percent haziest days in western Class I areas is approximately 100 km. PM does produce some effects on crops and forested ecosystems by the deposition of reactive nitrogen species. At this time, there is limited information on the extent of these effects and their trends. The larger influence on crop pro- duction, forest growth, and other indicators of ecosystem health and function is from ozone. Detrimental effects on vegetation include reduction in agricultural and commercial forest yields, reduced growth and increased plant susceptibility to disease, and potential long-term effects on forests and natural ecosystems. Plants are significantly more sensitive to ozone than are people. Murphy et al. (1999) evaluated benefits to eight major crops associated with several scenarios concern- ing the reduction or elimination of O3 precursor emissions from motor vehicles in the United States. Their analysis reported a $2.8 billion to $5.8 billion (1990 dol-

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0 ENERGY FUTURES AND URBAN AIR POLLUTION lars) benefit from the complete elimination of O3 exposures from all sources (i.e., ambient O3 reduced to a background level assumed to be 0.025 to 0.027 ppm). The EPA is currently considering setting “secondary” welfare-based NAAQS for ozone to address the effects of this pollutant on crops and other vegetation. There is damage to materials from acidic species (gaseous and particulate) and ozone as well as from black carbon particles. With the decreases in the emis- sions that have occurred over the past 35 years, materials effects in the United States have become less important than in the past, and do not currently figure strongly in decisions regarding the setting of secondary air quality standards. China Health Effects Air pollution and its impact on people’s health and the environment is a matter of great concern in China. Heavy reliance on coal in power production and a rapidly growing car fleet, usually in combination with outdated technologies and poor maintenance, have led to concentrations of air pollutants far exceeding the limits of both national air quality standards and the air quality guidelines recommended by the WHO (2006). Nearly all of China’s rural residents and a shrinking fraction of urban residents use solid fuels (biomass and coal) for household cooking and heating. As a result, by the use of global meta-analyses of epidemiological studies, it is estimated that indoor air pollution from solid fuel use in China alone is responsible for ~420,000 premature deaths annually, more than the estimated 300,000 attributed to urban outdoor air pollution in the country (Zhang and Smith, 2005). The major air pollutants monitored in China are suspended particulates, SO 2, and NOx. Health end-points studied in China in association with air pollution include changes in mortality of all causes, of respiratory disease, cardiovascular disease, and cerebrovascular disease, and morbidity, as well as the number of outpatient and emergency room visits. Increases in respiratory and other clinical symptoms and decrease in lung functions and immune functions are also studied. However, in comparison with air monitoring data, data on the effects of human health are limited. Aunan and Pan (2004) specifically summarized the relation- ships between PM10 and SO2 and mortality, hospital admissions, and chronic respiratory symptoms and diseases. They expressed the exposure-response func- tions in terms of percentage change (per unit of exposure), rather than as absolute numbers. They derived the following coefficients for acute effects: a 0.03 per- cent (standard error [S.E.] 0.01) and a 0.04 percent (S.E. 0.01) increase in all- cause mortality per µg/m3 of PM10 and SO2, respectively; a 0.04 percent (S.E. 0.01) increase in cardiovascular deaths per µg/m3 for both PM10 and SO2; and a 0.06 percent (S.E. 0.02) and a 0.10 percent (S.E. 0.02) increase in respiratory deaths per µg/m3 of PM10 and SO2, respectively. For hospital admissions due to

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 AIR POLLUTION: SOURCES, IMPACTS, AND EFFECTS cardiovascular diseases, the obtained coefficients are 0.07 percent (S.E. 0.02) and 0.19 percent (S.E. 0.03) for PM10 and SO2, respectively, whereas the coefficients for hospital admissions due to respiratory diseases are 0.12 percent (S.E. 0.02) and 0.15 percent (S.E. 0.03) for PM10 and SO2, respectively. Exposure-response functions for the impact of long-term PM10 levels on the prevalence of chronic respiratory symptoms and diseases are derived from the results of cross-sectional questionnaire surveys, and indicate a 0.31 percent (S.E. 0.01) increase per µg/m 3 in adults and 0.44 percent (S.E. 0.02) per µg/m3 in children. With some excep- tions, Chinese studies report somewhat lower exposure-response coefficients, as compared to studies in Europe and the United States. Increasing China’s already severe air pollution will substantially increase the incidence of respiratory diseases throughout the country, as air pollution is estimated to be the primary cause of nearly 50 percent of all respiratory ailments (Brunekreef and Holgate, 2002). According to the United Nations Environmental Programme statistics (UNEP, 1999), soot and particle pollution from the burning of coal causes approximately 50,000 deaths per year in China, while some 400,000 people suffer from chronic bronchitis annually in the country’s 11 largest urban areas. The United Nations Development Programme estimated that the death rate from lung cancer in severely polluted areas of China was 4.7-8.8 times higher than in areas with good air quality (UNDP, 2002). Extrapolating from 1995 emission levels and trends, the World Bank estimated that by 2020, China will need to spend approximately $390 billion—or about 13 percent of projected GDP—to pay for the healthcare costs that will accrue solely from the burning of coal (World Bank, 1997). Zaozhuang, a coal-dependent eastern city, was estimated to be spending 10 percent of its GDP on air pollution-related health damages in 2000; and with- out additional controls, this share could rise to 16 percent by 2020 (Wang and Mauzerall, 2004). In Shanghai, which has been taking measures to diversify its fuel mix, it was estimated that health impacts due to particulate air pollution in urban areas in 2001, totaled $625 million, accounting for over 1 percent of the city’s GDP (Kan and Chen, 2004). The Health Effects Institute recently made available a compendium of epidemiologic studies of air pollution in Asia from 1980-2006, including 69 studies from mainland China (HEI, 2006). Visibility There is limited research about the impact of air pollution on visibility in China. Qiu and Yang (2000) analyzed the visibility trends in northern China from 1980 to 1994 based on 0.74 µm aerosol optical depths at five meteorological obser- vations. In Zhengzhou (Figure 3-1) and Geermu, visibility showed an improving trend during this period and a possible reason for this is vertical distribution shifts of aerosol particles up in the troposphere. Visibility at Urumqi, Harbin, Beijing, and Zhengzhou in winter is impaired. At Harbin, the visibility range in summer is about twice that in winter. Cheung et al. (2005) utilized a formula developed

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0 ENERGY FUTURES AND URBAN AIR POLLUTION use ambient concentrations collected at a receptor(s) and possibly other variables and work backwards to the source to estimate source contributions to ambient PM loadings at the receptor location. Receptor methods primarily describe the current situation, since they are observationally based, and therefore, are not used in predicting changes in PM concentrations due to changes in emissions. While receptor models can separate primary from secondary components, they are used most effectively to link primary species observed at a receptor to source types or categories (source apportionment), or individual sources (e.g., a specific emitter) (source attribution), and to quantify (value with an uncertainty estimate) the source contribution at the receptor. Secondary components are usually grouped by compound (e.g., ammonium sulfate, ammonium nitrate), but quantitative separa- tion into source categories is usually not obtained. Source markers or tracers (i.e., usually multiple markers used to identify a given source type) are used to identify primary sources, and these may include inorganic and/or organic species. The three most widely used receptor models are CMB (Friedlander, 1973; Watson et al., 1984), PMF (Paatero, 1997), and UNMIX (Lewis et al., 2003). All three are based on the general mass balance equation and require ambient concen- tration measurements. PMF and UNMIX require only ambient concentrations and the factors developed are interpreted by the investigator as specific source types (e.g., motor vehicle, soil dust, etc.). Specific source information is not required for PMF and UNMIX. CMB requires the assumption that the sources are known and that source profiles (i.e., mass fractions of individual chemical species com- prising the emissions) are available for each source. The EPA has supported the development of these three models for use in the regulatory process, and they are widely used in SIPs being developed across the country.2 Model Integration In most planning efforts, it is useful to employ both source- and receptor- oriented models to be able to cross-compare results. Receptor models can help to identify problems with the source modeling (Core et al., 1982), while source models can provide predictions of the effects of specific control actions. In both cases, efforts have to be made to provide the critical input data. In the case of source models, the biggest issue is obtaining good emissions inventories. It is often a problem to identify all of the sources and to characterize their emissions. In the case of receptor models, there needs to be a program of ambient monitoring for a sufficient number of chemical species over a sufficiently long time frame—so that it provides an appropriate basis for the receptor model applications. 2Current information about these models is available at .

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0 AIR POLLUTION: SOURCES, IMPACTS, AND EFFECTS Applications in China It is important to understand the contribution of each emission source of air pollutants to ambient concentrations, to establish effective measures for risk reduction. Source-oriented and receptor-oriented models have been used to inves- tigate source apportionment and source/receptor relationships for different air pollutants in the urban atmosphere in China since the 1980s (Wang, 1985; Zhao et al., 1991). Most of the methods described above were adapted to differentiate the primary pollutants, such as mineral dust and fugitive dust. Application of source-oriented models in China has been limited until rel- atively recently, due to incomplete emission inventory data. However, some researchers have used international emission data, local meteorological fields, and advanced models to simulate the spatial and temporal distribution of pollutants in the context of international field studies, and more recently to examine local or regional control options. More recently, Chinese researchers have begun to incorporate local emissions inventory data into source-oriented modeling studies (e.g., Chen et al., 2007). For example, M.G. Zhang et al. (Zhang, 2004; Zhang et al., 2005, 2006) applied CMAQ (Models-3 Community Multi-scale Air Quality modeling system) coupled with the Regional Atmospheric Modeling System to East Asia to analyze the production and transport processes of organic carbon (OC), black carbon (BC), and sulfur compounds in the spring of 2001, when two large field campaigns, TRACE-P (TRAnsport and Chemical Evolution over the Pacific) and ACE-Asia (Aerosol Characterization Experiment—Asia), were being conducted over a broad area covering northeastern Asia and the western Pacific (Figure 3-31). Wang et al. (2005) used the STEM-2K1 atmospheric chemistry and transport model with MM5 meteorological fields, anthropogenic emissions estimates from Streets et al. (2003), and biogenic emissions from the GEIA inventory to compare contributions from transportation, power generation, and industry to concentra- tions of ozone, SO2, NOx, and CO in Guangdong Province in March 2001. They concluded that the transportation sector was the largest contributor to ozone levels, and found that in their simulations, ozone formation in urban areas was limited by VOC emissions, whereas ozone formation in rural areas was NO x-limited. Chen et al. (2007) applied the CMAQ model with MM5 meteorological fields to investigate the contributions of transport from surrounding provinces to PM10 pollution in Beijing. The study used county-level emissions inventories for primary PM from the environmental protection administrations of Beijing and the surrounding provinces, and examined simulated and measured PM 10 concentrations for 4 months in 2002. Observed PM10 concentrations were repro- duced relatively well, except for the month of April, when concentrations were underpredicted, because the simulations did not account for extreme sandstorm events. The modeling analysis indicated that transboundary pollution contributes significantly to PM10 concentrations in Beijing, with especially high contributions when pollution levels in Beijing are elevated.

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0 ENERGY FUTURES AND URBAN AIR POLLUTION FIGURE 3-31 Horizontal distributions of average black carbon aerosol concentrations and wind fields for the lowest model layer in the period of March-April 2001. Compared with the application of source-oriented models, receptor-oriented models have been utilized more extensively in China. Particulate pollution is the most serious pollutant in Chinese cities, so models like principle component analysis, absolutely principle component analysis, positive matrix factorization, and chemical mass balance (CMB) have been important analysis tools; and, since the 1980s, these tools have been used in more than 20 cities that have no emissions inventories (Wang, 1985; Zhao et al., 1991; Chen et al., 1994; X.Y. Zhang et al., 2001; Y.H. Zhang et al., 2004). Earlier studies focused on TSP. After requirements shifted to PM10 control in 1996, more studies investigated the source apportion- ment of PM10. Recent studies in Beijing and Hong Kong were aimed at the source apportionment of PM2.5 (Ho et al., 2006, Song et al., 2006a, 2006b). At the same time, source apportionment studies have been extended to specific pollutants like polycyclic aromatic hydrocarbons (PAHs), OC, and EC (Qi et al., 2002; Cao et al., 2005b; Peng et al., 2005; Wan et al., 2007). New methods including genetic algorithms, neural networks, fuzzy set theory, and nested CMB were developed

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0 AIR POLLUTION: SOURCES, IMPACTS, AND EFFECTS and utilized in source apportionment studies (Feng et al., 2002; Li et al., 2000, 2003; Li and Ding 2005). Eleven typical source apportionment studies for PM in Chinese cities are summarized in Appendix C and show that major sources of TSP include coal combustion dust, fugitive (soil) dust, and construction dust. These sources also contribute significantly to PM10 in some cities, especially in northern China. In Hong Kong, a developed city without extensive construction activity or coal uti- lization, receptor model results suggest that secondary aerosol and motor vehicle exhaust are relatively important sources of PM10 as well as of PM2.5. While receptor models are useful for estimating source contributions when accurate emissions inventories are not available, better information is needed to support their use. For example, local source profiles need to be developed step by step; long-term and systematic sampling should be implemented, chemical analysis methods should be compared nationally and internationally, and the source apportionment results should be reconciled with the results from source modeling and from emissions inventories. REFERENCES Aunan, K. and X.C. Pan. 2004. Exposure-response functions for health effects of ambient air pollution applicable for China—a meta-analysis. Science of the Total Environment 329:3-16. Aunan, K., T.K. Berntsen, and H.M. Seip. 2000. Surface ozone in China and its possible impact on agricultural crop yields. Ambio 29:294-301. Stockholm: The Royal Swedish Academy of Sciences. Beijing EPB (Beijing Municipal Environmental Protection Bureau). 2006. Report on the State of the Environment in Beijing 2001-2005. Bell, M.L., A. McDermott, S.L. Zeger, J.M. Samet, and F. Dominici. 2004. Ozone and short- term mortality in 95 U.S. urban communities. Journal of the American Medical Association 292:2372-2378. Bergin, M.H., R. Greenwald, J. Xu, Y. Berta, and W.L. Chameides. 2001. Influence of aerosol dry deposition on photosynthetically active radiation available to plants: A case study in the Yangtze delta region of China. Geophysical Research Letters 28:3605-3608. Brook, J.R., T..F Dann, and R. Vet. 2004. Time-Integrated Particle Measurements: Status in Canada. European Monitoring and Evaluation Program (EMEP) Workshop on Particulate Matter Mea- surement and Modeling. New Orleans, Louisiana. Brunekreef, B. and B. Forsberg. 2005. Epidemiological evidence of effects of coarse airborne particles on health. European Respiratory Journal 26:309-318. Brunekreef, B. and S.T. Holgate. 2002. Air pollution and health. Lancet 360:1233-1242. Cabada, J., S. Pandis, R. Subramanian, A. Robinson, A. Polidori, and B. Turpin. 2004. Estimating the secondary organic aerosol contribution to PM2.5 using the EC tracer method. Aerosol Science and Technology 38(Suppl)1:140-155. Cao, H. 1989. Air pollution and its effects on plants in China. Journal of Applied Ecology 26:763-773. Cao J.J., B. Rong, S.C. Lee, J.C. Chow, K.F. Ho, J.G. Watson, Z.S. An, K. Fung, S.X. Liu, and C.S. Zhu. 2005a. Composition of indoor aerosols at the Emperor Qin’s Terra-cotta Museum, Xi’an, China during summer, 2004. China Particuology 3(3):170-175.

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