Below is the uncorrected machine-read text of this chapter, intended to provide our own search engines and external engines with highly rich, chapter-representative searchable text of each book. Because it is UNCORRECTED material, please consider the following text as a useful but insufficient proxy for the authoritative book pages.
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)âthe imposition of emission standards for hazardous air pollutants and control on mobile source emissionsâthe 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. 61
62 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 Âquality stan- dards, and cities use this tool to report their air quality (see Chapter 4, TableÂ 4â1). Like their national standards, the API contains classes (I-V) which allow Âcities to comply at various levels. Cities are encouraged to achieve ClassÂ II standards TABLE 3-1â U.S. National Ambient Air Quality Standards (NAAQS) Pollutant Primary Stds. Averaging Times Secondary Stds. Carbon monoxide 9 ppm (10 mg/m3) 8-hour None 35 ppm (40 mg/m3) 1-hour None Lead 1.5 Âµg/m3 Quarterly average Same as primary Nitrogen dioxide 0.053 ppm Annual (arithmetic mean) Same as primary (100 Âµg/m3) Particulate matter (PM10) Revoked Annual (arith. mean) Â 150 Âµg/m3 24-hour Â Particulate matter (PM2.5) 15.0 Âµg/m3 Annual (arith. mean) Same as primary 35 Âµg/m3 24-hour Â Ozone 0.08 ppm (171 8-hour Same as primary Âµg/m3) 0.12 ppm (257 1-hour Same as primary Âµg/m3) (Applies only in limited areas) Sulfur oxides 0.03 ppm (78 Âµg/ Annual (arith. mean) â m 3) 0.14 ppm (364 24-hour â Âµg/m3) â 3-hour 0.5 ppm (1300 Âµg/m3)
AIR POLLUTION: SOURCES, IMPACTS, AND EFFECTS 63 TABLE 3-2â Chinaâs Ambient Air Quality Standards (GB 3095-1996) (Âµg/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 3-3â World Health Organization (WHO) Air Quality Guidelines (in Âµg/m3) Interim Interim Interim Pollutant target 1 target 2 target 3 Standard Averaging Times aPM 35 25 15 10 Annual 2.5 75 50 37.5 25 24-hour PM10 70 50 30 20 Annual 150 100 75 50 24-hour O3 160 â â 100 8-hour NO2 â â â 40 Annual â â â 200 1-hour SO2 125 50 â 20 24-hour â â â 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 Chinaâs classes of standards, is that they allow governments to consider local
64 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 60Â 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
AIR POLLUTION: SOURCES, IMPACTS, AND EFFECTS 65 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 Âµg/m3 of annual-average PM2.5, a 0.25-1 percent increase per 10 Âµg/m3 24-hour average PM10, and 0.2-0.8 percent increase per 10 Âµg/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 âcoarseâ 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 Âµm 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-
66 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 $5.4Â 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 24âhour standard will result in health and visibility benefits ranging from $9Â 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 EPAâs national emissions inventory, the key sources for benzene are on-road (49 percent) and non-road mobile sources (19Â 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.
AIR POLLUTION: SOURCES, IMPACTS, AND EFFECTS 67 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 EPAâs 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 ârisk driverâ 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. EPAâs 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 Preventionâs National Health and Nutrition Examination Survey found that approximately 6 percent of c Â hildbearing-aged women had blood mercury levels at or above the reference dose of 5.8 Âµg/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
68 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 citiesâemission 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 Â 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 Â 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 Âpollution. 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. Environmental 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
AIR POLLUTION: SOURCES, IMPACTS, AND EFFECTS 69 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 Âµm. 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 EPAâs 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. ÂMurphy et al. (1999) evaluated benefits to eight major crops associated with Âseveral 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-
70 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 âsecondaryâ 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 peopleâs health and the environment is a m Â atter 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 Chinaâs 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 0.03Â per- cent (standard error [S.E.] 0.01) and a 0.04 percent (S.E. 0.01) increase in all- cause mortality per Âµg/m3 of PM10 and SO2, respectively; a 0.04 percent (S.E. 0.01) increase in cardiovascular deaths per Âµg/m3 for both PM10 and SO2; and a 0.06Â percent (S.E. 0.02) and a 0.10 percent (S.E. 0.02) increase in respiratory deaths per Âµg/m3 of PM10 and SO2, respectively. For hospital admissions due to
AIR POLLUTION: SOURCES, IMPACTS, AND EFFECTS 71 cardioÂvascular diseases, the obtained coefficients are 0.07 percent (S.E. 0.02) and 0.19Â 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 Âµg/m 3 in adults and 0.44 percent (S.E. 0.02) per Âµg/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 Chinaâs 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 Âstatistics (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 Âsuffer from chronic bronchitis annually in the countryâs 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 billionâor about 13 percent of projected GDPâto pay for the healthcare costs that will accrue solely from the burning of coal (World Bank, 1997). ÂZaozhuang, 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 M Â auzerall, 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 cityâs 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 Î¼m 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
72 ENERGY FUTURES AND URBAN AIR POLLUTION r= FIGURE 3-1â Variation characteristics of year-averaged aerosol optical depth and Âvisibility over Zhengzhou during 1980-1994. SOURCE: Qiu and Yang, 2000. 3-1 new higher res jpeg Sept 14 in the U.S. Interagency Monitoring to Protect Visual Environment (IMPROVE) study to calculate the contribution of organic matter and sulfate in PM2.5 to the light extinction in the Pearl River Delta region, China. The fine sulfate is a larger contributor with humid climates. Sun et al. (2006) characterized the chemical compositions of PM2.5 and PM10 in haze-fog episodes in Beijing. The serious air pollution in haze-fog episodes was strongly correlated with meteorological condi- tions and with the emissions of pollutants from anthropogenic sources. Cultural Relics There are more than 30 World Cultural Heritage sites in China and most of them are suffering damage from air pollution. From 2002 to 2005, the State Administration of Cultural Heritage initiated a project to investigate the corro- sion status of cultural collections inside most of the museums in China. About 10 percent of collections were damaged seriously by indoor air pollution. Out- door cultural relics, like grottos and carved stones, are eroded predominantly by acid rain. However, very little research has focused on indoor air pollution and environmental assessment in Chinaâs museums. For example, Christoforou et al. (1996) developed computer-based models to simulate the air flow into the caves
AIR POLLUTION: SOURCES, IMPACTS, AND EFFECTS 73 and particle deposition within the Yungang caves. It was found that horizontal surfaces within caves 6 and 9 at Yungang would become completely covered by a full monolayer of particles in less than half a year, under the April condi- tions studied there, and will be soiled even more rapidly under annual average conditions. Cao et al. (2005a) investigated the indoor air quality in Emperor Qin's Terra-Cotta Museum in Xiâan, China, during summer 2004. The average levels of indoor PM2.5 and total suspended particles (TSPs) were 108.4 and 172.4 Âµg/m3. Sulfate ((32.4Â±6.2) percent), organics ((27.7Â±8.0) percent), and geological material ((12.5Â±3.4) percent) dominated indoor PM 2.5, followed by ammonium ((8.9Â±2.8) percent), nitrate ((7.0Â±2.9) percent), and elemental car- bon (EC, (3.9Â±1.5) percent). High concentrations of acidic aerosols will erode the terra-cotta warriors and horses, especially in the summer season with high temperature (30Â°C) and relative humidity (70 percent) and with undesirable solar radiation inside the museum. Agriculture Effects of air pollution on crops have been studied in China since the 1970s. Cao (1989) reported the effects of short-term exposure of three sensitive groups of crops to SO2 and hydrofluoric acid (HF). The sensitive species are reported to have suffered 5 percent injury at 880Â±1430 mg/m3 SO2 and 12Â±48 mg/m3 HF during an 8-hour exposure. Black carbon is the aerosol most responsible for reducing atmospheric trans- parency and visibility by so much in China that agricultural productivity is reduced by an estimated 10-20 percent (Chameides et al., 1999), with additional loss from black carbon deposited on plant leaves (Bergin et al., 2001). Chameides et al. (1999) estimated that regional haze in China is currently depressing optimal yields of ~70 percent of the crops grown in China by at least 5-30 percent. Chang and Hu (1996) reported that the average yield reduction for vegetables in Chongqing is 24 percent. Shu et al. (1993) estimated the annual cost of for- est damage by acid rain in Guangxi province to be $80 million. Ou et al. (1996) estimated the economic loss due to acid rain damages to crops and materials in the Xiamen area to be about $6 million, which equals about 1 percent of the GDP for the area. Chang and Hu (1996) reported that the annual damage from air pollution in Chongqing in 1993 was about $220 million, which is 4.4 percent of the GDP. A recent study on the impacts of ozone on agriculture in China found that reductions in crop yields in 1990 were less than 3 percent for most grain crops, but that predictions for 2020 suggested that crop losses for soybeans and spring wheat might reach 20 and 30 percent, respectively (Aunan et al., 2000). Wang and Mauzerall (2004) estimated that due to O3 concentrations in 1990, China lost 1-9Â percent of their yield of wheat, rice, and corn and 23-27 percent of their yield of soybeans, with an associated value of 1990 US$3.5 billion. In 2020, assuming
74 ENERGY FUTURES AND URBAN AIR POLLUTION no change in agricultural production practices, grain loss due to increased levels of O3 pollution is projected to increase to 2-16 percent for wheat, rice, and corn, and 28-35 percent for soybeans; the associated economic costs are expected to increase 82 percent by 2020, compared to 1990 costs. Acid Deposition Acid rain emerged as an important environmental problem in China in the late 1970s. The region with the most serious acid rain problem in China is in the southwest; approximately 90 percent of the monitoring stations with a mean pH of less than 5.6 are located south of the Yangtze River (Zhao et al., 1988). The deposition of sulfur is in some places higher than what was reported from the "black triangle" in central Europe in the early 1980s (Larssen et al., 2006). Since 1989, an acid rain network including 88 stations has been operated by the China Meteorological Administration. Based on the observational data, Chinaâs acid rain problem experienced two stages in the last two decades. The first stage was characterized by rapid economic development and increasing acid deposition through the mid-1990s, followed by a period of relative stability which persists today. Although no further worsening has been observed in recent years, the situ- ation is far from being ameliorated (Ding, 2004). EMISSIONS AMOUNTS AND TRENDS United States In recent years, concern about air pollution in the United States has evolved to a focus on chronic exposures to non-lethal concentrations of air pollutants. Since the late 1960s and early 1970s, emission controls on industrial and vehicular sources of air pollution have resulted in dramatic improvements in some facets of air quality. For example, in the 1960s, most industrialized urban areas in the eastern United States exceeded air quality standards for ambient concentrations of sulfur dioxide. But, after three decades of control efforts, directed largely at industrial sources, there are now no areas in the United States that exceed the NAAQS for SO2. Carbon Monoxide (CO) There are now only a few areas in the United States that continue to exceed NAAQS for carbon monoxide (NRC, 2003), due largely to control strategies that reduced CO emissions (per kilometer traveled) from motor vehicles by more than 95 percent. Figure 3-2 shows the emissions of CO from 1983 to 2002.
AIR POLLUTION: SOURCES, IMPACTS, AND EFFECTS 75 FIGURE 3-2â CO emissions by source, 1983-2002. SOURCE: EPA, 2006b. Nitrogen Oxides (NOx) Oxides of nitrogen arise from combustion sources including mobile and stationary sources. Although reductions of NOx have resulted in meeting the NAAQS, further emission reductions have been mandated to reduce ozone and particulate matter concentrations. Figure 3-3 shows the EPAâs estimates of reduc- tions in NOx emissions over the past 20 years. Some recent studies have suggested that these estimates may overestimate the amount of reductions achieved from mobile sources in recent years (Parrish, 2006). Sulfur Dioxide (SO2) For SO2, the major sources include the combustion of coal and oil, smelting of non-ferrous metals, petroleum refining, and other industrial processes. The overall trends in estimated SO2 emissions are shown in Figure 3-4. Sulfur dioxide emissions are thought to be relatively well known, because they are dominated by large stationary sources with direct emissions monitors. As with NOx, emissions of SO2 are generally higher in the eastern United States than in the less densely populated western half of the country. It should be noted that reductions in NOx and SO2 have occurred while the total generation of electricity has increased. Similarly, since 1970, aggregate emissions traditionally associated with vehicles have significantly decreased (with
76 ENERGY FUTURES AND URBAN AIR POLLUTION FIGURE 3-3â NOx emissions by source, 1983-2002. SOURCE: EPA, 2006b. FIGURE 3-4â SO2 emissions by source, 1983-2002. SOURCE: EPA, 2006b.
AIR POLLUTION: SOURCES, IMPACTS, AND EFFECTS 77 the exception of NOx), even as vehicle miles traveled (VMT) have increased by approximately 149 percent. NOx emissions from vehicles increased between 1970 and 1999 by 16 percent, due mainly to emissions from light-duty trucks and heavy-duty vehicles. However, as recent trends show, vehicle travel is having a smaller and smaller impact on emissions as a result of stricter engine and fuel standards, even with additional growth in VMT. PM10 At this time, relatively few locations in the United States are in violation of the PM10 NAAQS. Emissions trends for PM10 are shown in Figure 3-5. The trends data reflect emissions from fuel combustion, transportation, and industrial production, but do not include emissions of PM10 from âfugitiveâ sources, includ- ing windblown dust and agricultural activities. PM2.5 With the promulgation of a NAAQS for PM2.5 in 1997, concern for fine par- ticles and their emission became a higher priority. Figure 3-6 presents the trends in the primary PM2.5 emissions where there is a truncated time series, since it was FIGURE 3-5â PM10 emissions by source, 1988-2003. SOURCE: EPA, 2006b.
78 ENERGY FUTURES AND URBAN AIR POLLUTION FIGURE 3-6â PM2.5 emissions by source, 1993-2002. SOURCE: EPA, 2006b. not until 1993 that this size fraction was separately inventoried. Most of the PM 2.5 is secondary in nature and thus emissions of SO2, NOx, and VOC (volatile organic compounds) that serve as the precursors to the secondary fine particles are also important to understanding fine particle concentrations (Turpin and Huntzicker, 1995; Cabada et al., 2004). Assessing PM2.5 sources is a complicated issue, as the fractions (primary versus secondary) vary by pollution sources, meteorology, and local geography (Seinfeld and Pankow, 2003; Kanakidou et al., 2005; Pun and Seigneur, 2007). Moreover, reductions in precursors will not necessarily lead to reductions in PM2.5 (Gaydos et al., 2005; Vayenas et al., 2005; Robinson et al., 2007), and therefore accurate emissions inventories for all precursors are criti- cal, as are their inclusion in monitoring and modeling, in order to analyze and characterize PM2.5 formation. Lead (Pb) The final criteria pollutant for which there are direct emissions into the atmo- sphere is lead in total suspended particles. The trend for Pb in TSP is presented in Figure 3-7. It can be seen that there was a very sharp drop in lead emissions during the early 1980s, as lead was phased out of gasoline. Complete removal was achieved by 1988, except for Alaska where leaded fuel ceased to be sold in 1991. Today, the principal remaining lead emissions arise from primary and
AIR POLLUTION: SOURCES, IMPACTS, AND EFFECTS 79 FIGURE 3-7â Lead emissions by source, 1982-2002. SOURCE: EPA, 2006b. secondary lead smelting and from other non-ferrous metal processes, rather than from gasoline combustion (see Box 3-2). Mercury Mercury is regulated in the United States as a hazardous air pollutant. The EPA estimates that from 1990 to 1999, anthropogenic emissions of mercury in the United States declined from 220 short tons per year to 115 short tons per year, due largely to reductions in emissions from medical waste incinerators and municipal waste combustors (EPA, 2007a). As of 1999, utility coal boilers were the largest source of mercury emissions, producing about 45 short tons per year of emissions in the United States. The Clean Air Mercury Rule adopted in 2005 would limit emissions from this sector to 38 short tons in 2010 and to 15 short tons in 2018. More rapid reductions may be achieved in response to state-level regulations. Volatile Organic Compounds (VOCs) VOCs are regulated in the United States as an important precursor of ozone, and because some VOCs are toxic in their own right. Emissions of VOCs have
80 ENERGY FUTURES AND URBAN AIR POLLUTION BOX 3-2 Mobile Source Emissions in the United States Emissions from motor vehicles are grouped into two main categories: (1)Â majorÂ gaseous and particulate air pollutants, which can be found in relatively high amounts in the atmosphere; and (2) air toxics, which usually are found in smaller concen- trations, but are carcinogens, causing cancers and other adverse human health effects at very low levels. The major gaseous and particulate pollutants to which motor vehicles contribute include carbon monoxide (CO); ozone (O3)âthrough its atmospheric precursors volatile organic compounds (VOCs) and nitrogen oxides (NOx); fine particulate matter, PM10 and PM2.5; and nitrogen dioxide (NO2). Trans- portation sources account for a small share of total SOx emissions, but emissions from vehicles increased from 1970 to 1999, largely due to increased emissions from non-road vehicles. The toxics emitted from motor vehicles include aldehydes (acetaldehyde, formaldehyde, etc.), benzene, 1,3-butadiene, and a large family of substances known as polycyclic organic matter (including polycyclic aromatic hydrocarbons, or PAHs). On-road vehicles are the dominant source of emissions from the transporta- tion sector. Light-duty gasoline-powered vehicles (cars and trucks) account for the predominant share of CO, NOx, and VOC emissions from on-road vehicles, but, in the United States, their share of all criteria pollutants has been decreasing over time. declined due to a variety of regulatory actions, including more stringent controls on exhaust emissions and evaporative fuel losses from vehicles, better control of emissions in the refining and distribution of motor vehicle fuels, and the imposition of controls on a variety of industrial and mobile emissions sources under regulatory programs designed to address toxic air pollution. The trends in estimated VOC emissions are shown in Figure 3-8. China China has conducted emission control since the 1980s, beginning later than in the United States. At first, emission control was focused on dust emissions and then on soot emissions, because of economic limitations. The central government has encouraged desulfurization measures for power plants and industrial emis- sions, but little action was taken prior to 2000. More stringent policies, such as an enhanced levy system, have been implemented by the government in recent years. The reduction of SO2 emissions and improvement of ambient air quality for SO2 has been observed in some big cities, such as Beijing. China currently has nationwide statistics of emission amounts only on three major pollutants: SO2, soot, and industrial dust. Industrial soot emission refers
AIR POLLUTION: SOURCES, IMPACTS, AND EFFECTS 81 FIGURE 3-8â VOC emissions by source, 1983-2002. SOURCE: EPA, 2006b. to PM produced as a result of fuel combustion (and includes power plants as an industrial source). Other soot refers to fuel combustion from all other social and economic activities and operation of public facilities. It is calculated on the basis of estimated coal consumption by households and others. Industrial dust refers to the volume of PM emitted by industrial processes and suspended in the air for a given period of time, such as dust from refractory material of iron and steel works, dust from coke-screening systems and sintering machines of coke plants, dust from lime kilns, and dust from cement production in building material enterprises, but excluding soot and dust emitted from power plants. In this section, most historical data on emissions only go back to 1997. While data exist for years prior to 1997, those data do not capture emissions from townships and counties and, therefore, are less useful when compared to more recent data. As discussed below, during 1994-2005, emissions of major air pollutants in China increased year after year. Among them, the emission of air pollutants from industrial sources had a dramatic increase. This is closely related to the sustained growth of the national economy and the mode of economic growth in China. Meanwhile, emission of air pollutants from residential sources was under fairly good control, resulting in a steadily declining trend of the emissions of sulfur dioxide and soot. This is related to the gradual implementation of clean energy strategies and integrated pollution prevention and control programs in cities. But exhaust emissions of automobiles were rising rapidly and are becoming a major
82 ENERGY FUTURES AND URBAN AIR POLLUTION source of urban atmospheric pollution. See Box 3-3 at the end of this section for more information on mobile sources in China. SO2 In recent years, owing to sustained economic growth, particularly in heavy industrial sectors such as steel, cement, and aluminum, Chinaâs industrial energy consumption, and consequently SO2 emissions, have steadily increased (Fig- ure 3â9). Over this same period, residential emissions have remained relatively stable. Streets and Waldhoff (2000) estimated that SO2 emissions are projected to increase from 25.2 mt in 1995 to 30.6 mt in 2020, provided emission controls are implemented on major power plants; if this does not happen, emissions could increase to as much as 60.7 mt by 2020. Emissions are concentrated in the popu- lated and industrialized areas of China: the northeastern plain, the east central and southeastern provinces, and the Sichuan Basin (Streets and Waldhoff, 2000). In 2004, over 52 percent of all Chinaâs SO2 emissions came from just nine provinces: Shandong, Hebei, Shanxi, Guizhou, Sichuan, Henan, Jiangsu, Inner Mongolia, and Guangdong. Shandong province emitted the largest amount of SO2 from industrial sources, approximately 8.2 percent of Chinaâs total industrial SO 2 emissions. This is a result of energy-intensive heavy industrial activity, but also of high-sulfur content in the local coal (SSB, 2006). Guizhou province reported the largest amount of SO2 emissions from residential sources, approximately 19.7 percent of Chinaâs total residential SO2 emissions, due to the prevalence of coal combustion for cooking and heating, as well as high-sulfur content (up to 2.5 percent) in the local coal (Ministry of Commerce, 2007; Xiao and Liu, 2004). An examination of the five most heavily polluting industrial sectors from 2000-2004 reveals that their SO2 emissions comprised, on average, more than 80 percent of Chinaâs total, while their economic contributions averaged less than one-third of the total (SEPA, 2005a). Since 2004, SO2 emissions from coal-fired power plants are estimated at 9.3 Mt in China, accounting for 49 percent of total industrial emissions. The top five provinces for power plant SO2 emission are Shandong, Hebei, Henan, Inner Mongolia, and Shanxi, which emitted 33.8 percent of total SO 2 emissions from power plants in China. Of 1,196 coal-fired power plants included in the statistics on emissions, only 425 have installed desulfurization facilities; they only reduced 1.57 million tons of SO2 emissions, which is much below their desulfurization capacity. Thus, China still has significant room for reducing emissions of SO 2. Production/supply of electric power, gas, and water; non-metal mineral production; smelting/Â pressing of ferrous metals; raw chemicals and chemical products; and smelting/pressing of non- f Â errous metals.
AIR POLLUTION: SOURCES, IMPACTS, AND EFFECTS 83 BOX 3-3 Mobile Sources in China Between 2000 and 2004 the passenger car fleet in China more than doubled and the diesel truck fleet has also grown rapidly: heavy-duty vehicles more than tripled between 1998 and 2002, and light-duty diesel trucks doubled in the same time frame. The motorcycle fleet, which includes many highly polluting two-stroke vehicles, also doubled over the same time period (He et al., 2005). Rapid urbaniza- tion and the improvement of the standard of living are expected to stimulate the purchase and use of motor vehicles, which will have a significant impact on urban air pollution. In general, mobile sources are currently contributing 45-60 percent of NOx emissions, 40-90 percent of VOC emissions, and about 80-90 percent of CO emissions in typical Chinese cities (Wang et al., 2005). More important than the relative contribution of vehicles and other sources to overall emissions is the con- tribution of motor vehicle emissions to personal exposure; because vehicles emit at ground level near the breathing zones of people, their emissions are frequently more important than is reflected by the total tonnage emitted. In particular, in urban centers, along roadsides, and especially in urban street canyons in crowded central business districts, mobile sources can contribute 2 to 10 times as much pollution as in general background situations. 3000 2500 2000 10,000 Tons 1500 1000 500 0 1997 1998 1999 2000 2001 2002 2003 2004 2005 Year Total SO 2 Industrial SO 2 Domestic SO 2 FIGURE 3-9â Trends of emission amounts of SO2 in China, 1997-2005. SOURCE: SEPA, 2001a, 2005b, 2006. 3-9 eps file, not a jpeg as is other version
84 ENERGY FUTURES AND URBAN AIR POLLUTION Soot As shown in Figure 3-10, industrial soot emissions decreased remarkably in the last decade of the 20th century. But, these emissions, like SO2 emissions, have increased steadily in recent years. The provinces where emissions exceeded 600,000 tons in order were Shanxi, Sichuan, Henan, Hebei, and Inner Mongolia. These five provinces emitted 37.5 percent of total soot emissions in China. The major emission sectors are electric power, non-metal mineral products, and smelting and pressing of ferrous metals, which accounted for 67 percent of the total emissions from industry; among them electric power contributed 44.2 percent. Industrial Dust The emission of industrial dust has not changed much since the year 2000. The trend is decreasing (see Figure 3-10), probably due to the effectiveness of dust removal by industries. Figure 3-11 shows the emissions of industrial dust by province in 2004. The provinces where emissions exceeded 600,000 tons per year were Hunan, Hebei, Henan, and Shanxi, in that order. Their emissions contributed 31.4 percent of the total industrial emissions. The non-metal mineral sector emit- ted 70.2 percent of the total dust emission from Chinaâs industries and the smelting and pressing of ferrous metals sector contributed 15.2 percent. 1800 1600 1400 10,000 Tons 1200 1000 800 600 400 200 0 1997 1998 1999 2000 2001 2002 2003 2004 2005 Year Total soot Industrial soot Domestic soot Industrial dust FIGURE 3-10â Trends of emission amounts of soot and industrial dust in China, 1997â2005. 3-10
AIR POLLUTION: SOURCES, IMPACTS, AND EFFECTS 85 Tibet Hainan Shanghai Tianjin Beijing Qinghai Ningxia Jilin Heilongjiang Yunnan Gansu Xinjiang Fujian Chongqing Guizhou Zhejiang City Hubei Liaoning Jiangsu Shaanxi Jiangxi Inner Mongolia Guangdong Shandong Sichuan Anhui Guangxi Shanxi Henan Hebei Hunan 0 10 20 30 40 50 60 70 10,000 tons FIGURE 3-11â The distribution of industrial dust emission amount by province in 2004. SOURCE: SEPA, 2005a. 3-11 Mercury Researchers at Tsinghua University and Argonne National Laboratory have collaborated to develop multiyear inventories of mercury emissions in China, covering the period from 1995 to 2003. They estimate that mercury emissions from anthropogenic sources in China totaled about 700 tons in 2003, up from about 550 tons in 1995. The major sources of anthropogenic mercury emissions were non-ferrous metal smelting and coal combustion, with the latter contribut- ing about 260 tons in 2003, mainly from the industrial and power sectors. While emissions rose nationwide, Wu et al. (2006) concluded that mercury emissions in some provinces had declined, for example in Liaoning Province, due to reduced metal smelting, and in Beijing Province due to an increased use of pollution control technology in the power sector.
86 ENERGY FUTURES AND URBAN AIR POLLUTION AMBIENT CONCENTRATIONS United States In the United States, ambient concentrations are measured for the major criteria pollutants, CO, NO2, SO2, O3, PM, and lead in total suspended particles, as well as for selected VOCs and other air toxics. Approximately 25 years of data are available from 1980 through 2005. Carbon Monoxide (CO) Carbon monoxide pollution is primarily related to emissions from motor vehicles. In the United States, improved catalytic converter technology coupled with improved fuel quality has led to a steady decline in ambient concentrations as shown in Figure 3-12. Nitrogen Dioxide (NO2) The NAAQS for oxides of nitrogen are designated for NO2. The trend in ambient NO2 concentrations in the United States is shown in Figure 3-13. The NAAQS is currently met at all of the locations where NO2 is monitored. The highest concentrations are in southern and central California. 14 12 10 Concentration (ppm) National Standard 8 6 4 2 0 1980 1985 1990 1995 2000 2005 FIGURE 3-12â CO concentrations, 1980-2005, based on annual second maximum 8-hour average. National trend based on 152 sites. SOURCE: EPA, 2007b. 3-12
AIR POLLUTION: SOURCES, IMPACTS, AND EFFECTS 87 0.06 National Standard 0.05 Concentration (ppm) 0.04 0.03 0.02 0.01 0 1980 1985 1990 1995 2000 2005 FIGURE 3-13â NO2 concentrations, 1980-2005, based on annual arithmetic average. National trend based on 88 sites. SOURCE: EPA, 2007b. 3-13 Sulfur Dioxide (SO2) The ambient concentrations of SO2 have also declined over the past two decades (Figure 3-14). The major drop in ground-level SO2 occurred in the 1970s, and all sites in the United States now meet the NAAQS for SO2. PM10 As discussed previously, airborne particulate matter is characterized as PM 10 and PM2.5. PM10 concentrations have declined in the United States since a stan- dard was imposed for this size class in 1987, as shown in Figure 3-15. PM2.5 Data for PM2.5 have been collected over a much shorter time period, with national trend data available since 1999, as shown in Figure 3-16. The most sig- nificant concentrations are in southern and central California and in many of the large urban areas of the midwestern and eastern United States (e.g., Pittsburgh). The United States has also been making measurements of fine particle com- position. Beginning in 1988, the IMPROVE program has been making particle
88 ENERGY FUTURES AND URBAN AIR POLLUTION 0.04 National Standard 0.03 Concentration (ppm) 0.02 0.01 0 1980 1985 1990 1995 2000 2005 FIGURE 3-14â SO2 concentrations, 1980-2005, based on annual arithmetic average. N Â ational trend based on 163 sites. SOURCE: EPA, 2007b. 3-14 160 National Standard 140 120 Concentration (Âµg/m3) 100 80 60 40 20 0 90 91 92 93 94 95 96 97 98 99 00 01 02 03 04 05 19 19 19 19 19 19 19 19 19 19 20 20 20 20 20 20 FIGURE 3-15â PM10 concentrations, 1990-2005, based on seasonally weighted annual average. National trend based on 435 sites. SOURCE: EPA, 2007b. 3-15
AIR POLLUTION: SOURCES, IMPACTS, AND EFFECTS 89 30 25 Concentration (Âµg/m 3 ) National Standard 20 15 10 5 0 1999 2000 2001 2002 2003 2004 2005 FIGURE 3-16â PM2.5 concentrations, 1999-2005, based on seasonally weighted annual average. National trend based on 658 sites. SOURCE: EPA, 2007b. 3-16 composition measurements in national parks, forests, and other areas where visibility degradation is a concern. In 2001, the IMPROVE network, which had focused primarily on the west, added sites in the eastern and central United States. Data from this network indicate that in the eastern United States, the rural PM 2.5 is dominated by sulfate with some nitrate and carbon, while the western particle compositions are more dominated by nitrate in California and carbonaceous aero- sol in much of the rest of the region. A second composition monitoring network, the Speciation Trends Network, was initiated by the EPA in 2000 to 2001 to provide composition data from urban sites across the United States. The compo- sition patterns in these urban locations are presented in Figure 3-17. Sulfate and carbonaceous particles are prominent in urban aerosols measured at the eastern STN sites. Urban sites in California show a relatively large influence from nitrate compared to eastern cities. Lead (Pb) Lead in total suspended particulate decreased sharply in the 1980s, as the use of leaded gasoline declined. Figure 3-18 presents the trends in airborne Pb concentration.
90 ENERGY FUTURES AND URBAN AIR POLLUTION FIGURE 3-17â Annual average PM2.5 concentrations (Âµg/m3) and particle type in urban areas, 2002. 2.5 Concentration (Âµg/m3)3-2 2 National Standard 1.5 1 0.5 0 1980 1985 1990 1995 2000 2005 FIGURE 3-18â Lead concentrations, 1980-2005, based on annual maximum quarterly average. National trend based on 16 sites. SOURCE: EPA, 2007b. 3-18
AIR POLLUTION: SOURCES, IMPACTS, AND EFFECTS 91 Ozone Ozone is a secondary pollutant formed in the atmosphere from reactions of nitrogen oxides and VOCs. The trends in ozone concentrations are shown in Figure 3-19, and the geographical distribution of measured ozone concentrations are given in Figure 3-20. Considering country-wide trends, ozone concentrations have declined modestly since 1980. However, many parts of the United States do not meet the current ozone NAAQS. Until the early 1990s, much of the focus in ozone control was on the reduc- tion of VOC concentrations. The 1990 Clean Air Act Amendments required areas in moderate to severe non-attainment of the ozone standard to make additional measurement of the VOC precursors, as part of the Photochemical Assessment Monitoring Program Stations network. The trends in VOC values from 1995 to 2001 are shown in Figure 3-21. It can be seen that there was a greater decrease in VOC concentrations than in the commensurate levels of ozone. In many locations, it is now recognized that control of NOx is a critical pathway to ozone control. However, it is clear that for many areas it will be difficult to achieve the current 8-hour ozone standard, and the current value is under review in consideration of recommendations from the Clean Air Scientific Advisory Committee to reduce the 8-hour standard concentra- tion to no greater than 70 ppb. 0.15 Concentration (ppm) 0.1 National Standard 0.05 0 1980 1985 1990 1995 2000 2005 FIGURE 3-19â Ozone concentrations, 1980-2005, based on annual fourth maximum 8âhour average. National trend based on 286 sites. SOURCE: EPA, 2007b. 3-19
92 ENERGY FUTURES AND URBAN AIR POLLUTION FIGURE 3-20â Trend in 8-hour O3 levels, 1983-2002, averaged across EPA regions, based on annual fourth maximum 8-hour average. SOURCE: EPA, 2003. FIGURE 3-21â Median percent change (1995-2001) at PAMS monitors for selected VOC species. SOURCE: EPA, 2003.
AIR POLLUTION: SOURCES, IMPACTS, AND EFFECTS 93 China In China, ambient concentrations are measured for the major criteria pol- lutants, CO, NOx, SO2, TSP/ PM10, O3, and Pb. China issued its first ambient air quality standards in 1982 and revised them in 1996 and 2000. The National Air Quality Standard of China divides the standard levels into three grades. For protecting residential health, it is required that the ambient air quality must achieve the Grade II level. Monitoring is the responsibility of local environmental protection bureaus, which typically report air quality in terms of the standard(s) achieved (Ref. to Chapter 4 of this report). Among the criteria pollutants, only four are routinely monitored for the whole country. Monitoring of ozone and lead is not required for every city, so they are not included in the national air quality statistics. Results shown in Figure 3-22 from the routine monitoring network of 360 cities for the period from 1999-2005 reveal that the air quality has improved, but nearly 40 percent of urban areas do not meet the Grade II air quality standards. China has high levels of SO2 and TSP due in large part to coal use. Mean- while, the number of motor vehicles has increased substantially since the mid- 1980s, primarily in urban areas and in city clusters, leading to increased emis- sions of NOx, VOCs, and particulates, and causing higher levels of ozone in the summer in urban areasâand much higher levels of inhalable particulates (PM10 70.0 60.0 50.0 40.0 Percent 30.0 20.0 10.0 0.0 1999 2000 2001 2002 2003 2004 2005 Meeting Class I & Class II CNAAQS Failing to meet Class III CNAAQS FIGURE 3-22â The percentage of 360 cities achieving different levels of air quality stan- dards from 1999-2005. SOURCES: SEPA, 2000, 2001a, 2001b, 2002, 2003, 2004, 2005a, 2005b, 2006.
94 ENERGY FUTURES AND URBAN AIR POLLUTION and PM2.5) throughout the country. In general, urban air quality across China basically remains stable against a background of rapid economic growth and accelerated urbanization. In addition, the air quality of some cities enjoyed some improvement. CO Carbon monoxide was observed in high concentration in Chinaâs large cities. Recently, strict emission standards and improved catalytic converter technology have led to a steady decline in ambient concentrations as shown in Figure 3-23. NOx /NO2 In China, the concentration limit for NOx was replaced by a limit for NO2 in 2000. As shown in Figure 3-24, after the year 2000, the ambient air concentra- tion of NO2 usually did not exceed the Grade II standard. Ambient concentrations were relatively high in some major cities, such as Guangzhou, Beijing, Shanghai, Hangzhou, Harbin, Urumqi, Nanjing, Chengdu, and Wuhan. It is important to note that, following the revision in 2000 to measure NO2 instead of NOx, most monitoring sites were not adjusted, and thus their locations, often at street level, are not suitable to accurately measure the true NOx concentra- tions in the urban atmosphere. NOx concentrations are generally higher than the NO2 measurements reveal, but are no longer being reported, and the result has 5.0 4.5 4.0 3.5 3.0 mg/m3 2.5 2.0 1.5 1.0 0.5 0.0 1982 1984 1986 1988 1990 1992 1994 1996 1998 2000 2002 2004 2006 Beijing Guangzhou Dalian FIGURE 3-23â Carbon monoxide ambient concentration trend in three large cities in China. SOURCE: Beijing, Guangzhou, and Dalian EPB annual reports. 3-23
AIR POLLUTION: SOURCES, IMPACTS, AND EFFECTS 95 FIGURE 3-24â Average annual urban NOx concentrations in China from 1990-2002. NOTE: NOx and NO2 concentrations in 2000 are not exact because of the lack of the data of some cities. SOURCE: Hao and Wang, 2005. 3-24 been that many cities, while ostensibly satisfying Grade II standards for NO 2, are in reality suffering from NOx pollution. Furthermore, these cities have focused on other criteria pollutants which exceed the Grade II standard and have paid much less attention to NOx emissions reduction strategies. Recent satellite observations have also provided a new insight into Chinaâs air pollution. In reviewing NO2 column data from two satellites, GOME and SCIAMACHY, from the period 1996-2002, a significant increase on the order of 50 percent was observed over China, which suggested a larger and more rapid increase in NO2 emissions than local and national inventories might suggest (Richter et al., 2005; Irie et al., 2005). Wang et al. (2007) used data from the Dutch-Finnish Ozone Monitoring Instrument to show that aggressive measures taken by Beijing to restrict motor vehicle traffic during the 2006 Sino-African Summit resulted in NOx emission reductions of more than 40 percent. SO2 SO2 is a major air pollutant in China because nearly 70 percent of the fuel used is coal. Routine monitoring of SO2 started in the 1970s. Figure 3-25 shows
96 ENERGY FUTURES AND URBAN AIR POLLUTION FIGURE 3-25â Average annual urban SO2 concentrations in China from 1990-2002. SOURCE: Hao and Wang, 2005. that in the northern cities and southern cities, the average concentration of SO 2 decreased dramatically from 1990 to 2000, at which point average concentrations began to rise again (see also Figure 3-26). The higher SO2 levels in the northern cities are the result of a combination of factors. Most regions of China experience higher SO2 concentrations in the winter as a result of increased heating needs, as coal continues to be the dominant source. Northern cities have a greater heating load, but they also have a large concentration of energy-intensive industries which contribute to higher year-round SO2 emissions. TSP/PM10 Figure 3-27 shows that concentrations of TSP have declined across China since 1990. This has been accomplished as a result of improved regulation, the installation of pollution controls (e.g., fabric filters), and the shuttering or relocat- ing of certain industries. Still, many cities continue to face challenges in reducing PM concentrations. Since 2000, high concentrations of PM10 have been the most frequent cause of Class II violations in China. In Beijing, the annual average level of PM 10 fluctu- ated around 160 Âµg/m3 from 2000 to 2004 (Beijing EPB, 2006). Megacities such as Beijing, Shanghai, and Guangzhou are frequently among the cities of the world with the highest levels of airborne PM (UNEP, 2002). Large areas of China are exposed to high levels of particulate pollution.
AIR POLLUTION: SOURCES, IMPACTS, AND EFFECTS 97 0.45 0.40 0.35 0.30 0.25 0.20 0.15 0.10 0.05 0.00 1980 1985 1990 1995 2000 2005 Beijing Dalian Guangzhou Shanghai Shenyang Tianjin Wuhan Chongqing FIGURE 3-26â Average annual SO2 concentrations in several large cities, China, 1980- 2005 (mg/m3). 3-26 FIGURE 3-27â Average annual urban TSP concentration in China from 1990-2002. SOURCE: Hao and Wang, 2005.
98 ENERGY FUTURES AND URBAN AIR POLLUTION PM2.5 China has not yet set standards for PM2.5. Unlike other air pollutants men- tioned before, all the information reported here for PM2.5 comes from individual research studies. In China, much work has been done to characterize ambi- ent concentration, chemical composition, size distribution, optical properties, s Â easonal variation, horizontal and vertical profiles, transport, and source-receptor relationships. Ambient air quality measurements of PM2.5 have been made in megacities, such as Beijing, Shanghai, and Guangzhou, and at the regional scale, for example, in the Pearl River Delta region and theYangtze River Delta region. Results show that PM2.5 has very high concentration levels, sometimes close to100 Î¼g/m3. The ratio of PM2.5 to PM10 is about 50-70 percent. Thus, PM2.5 is an important air pollutant in urban areas and is especially important in regional pollution (Figure 3-28). Ozone There are only a few cities in China routinely monitoring for ozone, includ- ing Beijing and Lanzhou in Gansu province. In fact, high concentrations of 200 1 PM2.5 PM10 PM2.5/PM10 PM2.5 STD 0.8 150 Concentration, mg/m3 0.6 PM2.5/PM10 100 0.4 50 0.2 0 0 1997-1999 1994-1995 2000 2000-07 2000-11 2002 1999-11 Qingdao Beijing Pearl River Delta Yangtze FIGURE 3-28â PM2.5 and PM10 mass concentration level in several regions of China (annual average). SOURCE: Peking University. 3-28
AIR POLLUTION: SOURCES, IMPACTS, AND EFFECTS 99 ground-level ozone have been observed for many years in several of Chinaâs urban areas. In the mid-1970s, photochemical smog first appeared in the Xigu petroleum industry district in Lanzhou. In 1986, Beijing also experienced photochemical smog in the summer. There the O3 level gradually increased later in the day exceeding Class II of the CNAAQS. Researchers at Peking University measuring the diurnal variations of episodic ground-level ozone found that concentrations have increased sharply since the 1990s, and often exceed 200 ppb (Figure 3-29). Systematic monitoring data collected by the Beijing Municipal Environment Monitoring Center showed that the hours of O3 concentration exceeding Grade II and the days of exceeding Grade II are still very high (see Figure 3-30). More recently, some southern cities, especially coastal cities, have faced the threat of photochemical smog in the summer and fall, even on the regional scale. Apr-June,1982 (ZGC) June,1987(city average) June,1993(ZGC) June,1997(ZGC) June,2000(ZGC) Aug 10-24,2003(Olympic site) 300.0 250.0 200.0 O3 (Âµg/m3 ) 150.0 100.0 50.0 0.0 0:00 3:00 6:00 9:00 12:00 15:00 18:00 21:00 Beijing Time FIGURE 3-29â The diurnal variation and trends in the episodic concentrations of ambient O3 measured in Zhongguancun (ZGC), Beijing (1982-2003), a northwest suburb of the city, about 20 km from Tianâanmen square. The 2008 Olympic Games site is about 4Â km north of ZGC. The yellow line indicates the 1-hour average O3 concentration at grade II, according to the national ambient air quality standards of China (2000 amendment to GB3095-1996). 3-29 SOURCE: Shao et al., 2006.
100 ENERGY FUTURES AND URBAN AIR POLLUTION 90 500 80 70 400 60 300 Hours 50 Days 40 200 30 20 100 10 0 0 1997 2001 2002 2003 2004 2005 Year Days exceeding Hours exceeding Grade II Grade II FIGURE 3-30â The hours and days of ozone concentration exceeding NAAQS Grade II in different years. SOURCE: Beijing EPB, 2006. 3-30 The Pearl River Delta region, Guangzhou, and surrounding areas have frequently experienced high O3 concentrations (Zhang et al., 1998). A similar study in the Yangtze River Delta region showed that high ozone concentrations are also often found at sites some distance removed from urbanized or industrial regions (Shao et al., 2006). SOURCE RECEPTOR RELATIONSHIPS In the United States, emissions reductions planning efforts are often made at the state level in State Implementation Plans (SIPs). SIPs are designed to reduce concentrations of ambient pollutants to below current NAAQS, and must make a showing that the plan will achieve the desired goal. To effectively and efficiently meet this objective, a quantitative understanding of source impacts at receptor locations is required. This information then guides policy makers at various levels within government in developing the most efficient and cost-effective approaches for attaining the health- and welfare-based air quality standards. Historically, emissions reduction strategies are guided by various types of statistical and mathematical modeling to identify and to âquantitativelyâ estimate a value with an associated uncertainty, and to apportion the source impacts at a given receptor location(s). Two general approaches are used to estimate source contributions quantita- tively at a receptor location(s). These are source-oriented and receptor-oriented methods (NRC, 2004). Source-oriented approaches start from the source of the emissions and use models or modeling systems to describe the transport, trans-
AIR POLLUTION: SOURCES, IMPACTS, AND EFFECTS 101 formation, and fate of those emissions from the source to the receptor. Receptor- oriented approaches begin with ambient concentrations measured at a receptor site(s) and use statistical or similar approaches to estimate source contributions at the site(s). Both source-oriented and receptor-oriented approaches have limita- tions, but they complement each other, so the use of both is advised to obtain the best estimates of the impact of sources or source types at receptor locations. Source-Oriented Approaches Source-oriented approaches start from the emissions source of the Âpollutant and work forward in time (rather then starting from the receptor and work- ing backward in time) to estimate the contribution of a source(s) at a receptor location(s). Source-oriented approaches use advanced mathematical models or modeling systems to describe the fate of those emissions between the source and the receptor (Russell, 2007; Seigneur and Moran, 2004, and references within). The most advanced source-oriented modeling systems, called chemical trans- port models (CTMs), consist of three major components: an emission model, a meteorological model, and an atmospheric process model (chemistry model). Each model may be composed of several modules. Ambient concentration data are typically not incorporated directly into CTMs (except to provide initial and boundary conditions) but are used for model performance evaluation (Seigneur et al., 2000; Seigneur and Moran, 2004; Russell, 2007). ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ Morris et al. (2004) reported ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ on a comparison of the Community Multiscale Air Quality (CMAQ) model and the CAMx model, which are state-of-the-art models that are widely used in the United States, using high-time-resolution data. Sulfate is well reproduced; organic carbon (OC), using a factor to convert OC to organic material of 1.8, is reproduced reasonably well; while nitrate and EC are overestimated by both models. Most importantly, CTM can be used to predict changes in PM mass and in components observed at a receptor, due to future or predicted changes in emis- sions (e.g., emissions reductions recommended through a SIP). Source-oriented models also effectively link source and receptor for secondary PM air pollutants (e.g., sulfate, nitrate, SOA) (Russell, 2007) but are not as effective for tracking primary species. Thus, there is synergy between source- and receptor-oriented models, with the application of both providing the most accurate picture of the impact of source emissions at receptor locations. Receptor-Oriented Approaches Receptor approaches are observationally based and may involve simple analyses such as time-series analysis, or correlation between or among pollutants; or they may involve more complex multivariate approaches usually referred to as receptor modeling (Hopke, 1991, 2003; Seinfeld and Pandis, 1998; Brook et al., 2004; and references within these publications). In all cases, receptor approaches
102 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 Âemitter) (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. 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 frameâso that it provides an appropriate basis for the receptor model applications. Current information about these models is available at <http://www.epa.gov/scram001/Âreceptorindex. htm>.
AIR POLLUTION: SOURCES, IMPACTS, AND EFFECTS 103 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 p Â ollutants 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 ExperimentâAsia), 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.
104 ENERGY FUTURES AND URBAN AIR POLLUTION FIGURE 3-31â 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 Âcomponent 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). ÂEarlier 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
AIR POLLUTION: SOURCES, IMPACTS, AND EFFECTS 105 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 Chinaâa 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 S Â ciences. 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 26:763â773. 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 Qinâs Terra-cotta Museum, Xiâan, China during summer, 2004. China Particuology 3(3):170-175.
106 ENERGY FUTURES AND URBAN AIR POLLUTION Cao, J.J., F. Wu, J.C. Chow, S.C.Â Lee, Y.Â Li, S.W.Â Chen, Z.S.Â An, K.K.Â Fung, J.G.Â Watson, C.S.Â Zhu, and S.X.Â Liu. 2005b. Characterization and source apportionment of atmospheric organic and elemental carbon during fall and winter of 2003 in Xiâan, China. Atmospheric Chemistry and Physics 3127-3137. CDC (Centers for Disease Control and Prevention). 2004. Blood mercury levels in young children and childbearing-aged womenâUnited States, 1999-2002. Morbidity and Mortality Weekly Report 53(43):1018-1020. Chameides, W. L., H. Yu, S.C. Liu, M. Bergin, X. Zhou, L. Mearns, G. Wang, C.S. Kiang, R.D. Saylor, C. Luo, Y. Huang, A. Steiner, and F. Giorgi. 1999. Case study of the effects of atmospheric aero- sols and regional haze on agriculture: An opportunity to enhance crop yields in China through emission controls? Proceedings of the National Academy of Sciences 96:13626-13633. Chang, Y. and T. Hu. 1996. The Economic Cost of Damage from Atmospheric Pollution in Chongqing. Report of Chongqing Environmental Protection Bureau. Chen, D.S., S.Y. Cheng, L. Liu, T. Chen, and X.R. Guo. 2007. An integrated MM5-CMAQ model- ing approach for assessing trans-boundary PM10 contribution to the host city of 2008 Olympic summer gamesâBeijing, China. Atmos. Environ. 41:1273-1250. Chen, S.L., Y.B. Zheng, Q. Zhao, et al. 1994. Source apportionment of TSP in Chongqing. Chongqing Environmental Sciences 16(4):30-32 (In Chinese). Cheung, H.C., T. Wang, K. Baumann, and H. Guo. 2005. Influence of regional pollution outflow on the concentrations of fine particulate matter and visibility in the coastal area of southern China. Atmospheric Environment 39:6463-6474. Christoforou, C.S., L.G. Salmon, and G. Cass. 1996. Fate of atmospheric particles within the Buddhist cave temples at Yungang China. Environmental Science and Technology 30:3425-3434. Cohen, A.J., H.R. Anderson, B. Ostro, K.D. Pandey, M. Krzyzanowski, N. KÃ¼nzli, K. Gutschmidt, C.A. Pope III, I. Romieu, J.M. Samet, and K.R. Smith. 2004. Mortality impacts of urban air pollution, in M. Ezzati, A.D. Rodgers, A.D. Lopez, C.J.L. Murray (eds.), Comparative Quan- tification of Health Risks: Global and Regional Burden of Disease due to Selected Major Risk Factors. Geneva: World Health Organization, Vol. 2, pp. 1353-1433. Cohen, A.J., H.R. Anderson, B. Ostra, K.D. Pandev, M. Krzyzanowski, N. KÃ¼nzli, K. Gutschmidt, A. Pope, I. Romieu, J.M. Samet, and K. Smith. 2005. The global burden of disease due to outdoor air pollution. Journal of Toxicology and Environmental Health Part A 68:1-7. Core, J.E., P.L. Hanrahan, and W.M. Cox. 1982. Particulate dispersion model evaluation: A new a Â pproach using receptor models. Journal of the Air Pollution Control Association 1142-1147. Ding, G. 2004. Database from the acid rain network of China Meteorological Adminstration and its preliminary analyses. Journal of Applied Meteorological Science 15(Z1):85-94 (in Chinese). Dockery, D.W., C.A. Pope III, X. Xu, J.D. Spengler, J.H. Ware, M.E. Fay, B.G. Ferris, Jr., and F.E. Speizer. 1993. An association between air pollution and mortality in six U.S. cities. New E Â ngland Journal of Medicine 329:1753-1759. EPA (U.S. Environmental Protection Agency). 1997. The Benefits and Costs of the Clean Air Act, 1970 to 1990. Washington, D.C.: U.S. Environmental Protection Agency. EPA. 2001. National Air Quality and Emissions Trends Report, 1999. Washington, D.C.: U.S. Envi- ronmental Protection Agency. EPA. 2003. Latest Findings on National Air Quality: 2002 Status and Trends. Washington, D.C.: U.S. Environmental Protection Agency. EPA. 2004. Air Quality Criteria for Particulate Matter, October. Washington, D.C.: U.S. Environmen- tal Protection Agency. EPA. 2005. Regulatory Impact Analysis for the Final Clean Air Interstate Rule, EPA-452/R-05-002. Washington, D.C.: Office of Air and Radiation. EPA. 2006a. ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ National-Scale Air Toxics Assessment For 1999: Estimated Emissions, Concentrations and Risk. Washington, D.C.: Office of Air and Radiation.
AIR POLLUTION: SOURCES, IMPACTS, AND EFFECTS 107 EPA. 2006b. Regulatory Impact Analysis Of EPAâs Final Revisions to the National Ambient Air Quality Standards for Particle Pollution (Particulate Matter)ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ . Washington, D.C.: Office of Air and Radiation. EPA. 2007a. Controlling Power Plant Emissions: Emissions Progress, http://www.epa.gov/mercury/ control_emissions/emissions.htm. EPA. 2007b. Review of National Ambient Air Quality Standards for Ozone. Final Staff Paper, J Â anuary. Feng, Y.C., Z.P. Bai, and T. Zhu. 2002. Principle and application of nested chemical mass balance. Environmental Science 23(Suppl):103-108 (in Chinese). Friedlander, S.K. 1973. Chemical element balances and identification of air pollution sources. Envi- ronmental Science and Technology 7:235-240. Gaydos, T.M., C.O. Stanier, and S.N. Pandis. 2005. Modeling of in situ ultrafine atmospheric particle formation in the eastern United States. Journal of Geophysical Research 110:D07S12. Gryparis, A., B. Forsberg, K. Katsouyanni, A. Analitis, G. Touloumi, J. Schwartz, E. Samoli, S. ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ Medina, H.R. Anderson, E.M. Niciu, H.E. Wichmann, B. Kriz, M. Kosnik, J. Skorkovsky, J.M. Vonk, and Z. DÃ¶rtbudak. 2004. Acute effects of ozone on mortality from the âAir Pollution and Health: A European Approachâ project. American Journal of Respiratory and Critical Care Medicine 170:1080-1087. Hao, J.M. and L.T. Wang. 2005. Improving urban air quality in China: Beijing case study. Journal of the Air and Waste Management Association 55(9):1298-1305. He, K. et al. 2005. Oil consumption and CO2 emissions in Chinaâs road transport: Current status, future trends, and policy implications. Energy Policy 33:1499-1507. HEI (Health Effects Institute). 2006. PAPA-SAN: Public Health and Air Pollution in AsiaâScience ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ Access on the Net, http://www.healtheffects.org/Asia/papasan-overview.htm. Ho, K.F., J.J. Cao, S.C. Lee, and C.K. Chan. 2006. Source apportionment of PM 2.5 in urban area of Hong Kong. Journal of Hazardous Materials B138:73-85. Ho, M.S., and C.P. Nielsen, eds. 2007. Clearing the Air: The Health and Economic Damages of Air Pollution in China. Cambridge, Mass.: MIT Press. Hoek, G., B. Brunekreef, S. Goldbohm, P. Fischer, and P.A. van den Brandt. 2002. Association between mortality and indicators of traffic-related air pollution in the Netherlands: A cohort study. Lancet 360:1203-1209. Hopke, P.K., ed. 1991. Receptor Modeling for Air Quality Management. Amsterdam: Elsevier Sci- ence Publishers. Hopke, P.K. 2003. Recent Developments in Receptor Modeling. Journal of Chemometrics 17:255â265. Irie, H., K. Sudo, H. Akimoto, A. Richter, J.P. Burrows, T. Wagner, M. Wenig, S. Beirle, Y. Kondo, V.P. Sinyakov, and F. Goutail. 2005. Evaluation of long-term tropospheric NO2 data obtained by GOME over East Asia in 1996-2002. Geophysical Research Letters 32(11):L11810. Kan, H.D. and B.H. Chen. 2004. Particulate air pollution in urban areas of Shanghai China: Health- based economic assessment. Science of the Total Environment 322:71-79. Kanakidou, M., J.H. Seinfeld, S.N. Pandis, I. Barnes, F.J. Dentener, M.C. Facchini, R. Van Dingenen, B. Ervens, A. Nenes, C.J. Nielsen, E. Swietlicki, J.P. Putaud, Y. Balkanski, S. Fuzzi, J. Horth, G.K. Moortgat, R. Winterhalter, C.E.L. Myhre, K. Tsigaridis, E. Vignati, E.G. Stephanou, and J. Wilson. 2005. Organic aerosol and global climate modeling: A review. Atmospheric Chemistry and Physics, 5:1053-1123. Kochanek, K.D., S.L. Murphy, R.N. Anderson, and C. Scott. 2004. Deaths: Final data for 2002. National Vital Statisitics Reports; vol. 53, no. 5. Hyattsville, MD: National Center for Health Statistics. Laden, F., L.M. Neas, D.W. Dockery, and J. Schwartz. 2000. Association of fine particulate matter from different sources with daily mortality in six U.S. cities. Environmental Health Perspec- tives 108(10).
108 ENERGY FUTURES AND URBAN AIR POLLUTION Larssen, T., E. Lydersen, D.G. Tang, Y. He, J.X. Gao, H.Y. Liu, L. Duan, H.M. Seip, R.D. Vogt, J. Mulder, M. Shao, Y.H. Wang, H. Shang, X.S. Zhang, S. Solberg, W. Aas, T. Okland, O. Eilertsen, V. Angell, Q.R. Liu, D.W. Zhao, R.J. Xiang, J.S. Xiao, and J.H. Luo. 2006. Acid rain in China. Environmental Science & Technology 40(2):418-425. Lewis, C.W., G.A. Norris, T.L. Conner, and R.C. Henry. 2003. Source apportionment of Phoenix PM2.5 aerosol with the unmix receptor model. Journal of the Air Waste Management Associa- tion 53:325-338. Li, Z.Y. and H.K. Ding. 2005. Application of BP neural network to sources apportionment of atmo- spheric particulates. Environmental Monitoring in China 21(2):74-76 (in Chinese). Li, Z.Y. and L.H. Peng. 2000. Source apportionment of atmospheric particulates based on genetic algorithm. Research of Environmental Sciences 13(6):19-21 (in Chinese). Li, Z.Y., C.J. Ni, and J. Ding. 2003. Application of rough sets theory to sources apportionment of atmospheric particulates. Journal of Sichuan University 35(4):112-114 (in Chinese). Lippmann, M., M. Frampton, J. Schwartz, D. Dockery, R. Schlesinger, P. Koutrakis, J. Froines, A. Nel, J. Finkelstein, J. Godleski, J. Kaufman, J. Koenig, T. Larson, D. Luchtel, L.-J.S. Liu, G. OberdÃ¶rster, A. Peters, J. Sarnat, C. Sioutas, H. Suh, J. Sullivan, M. Utell, E. Wichmann, and J. Zelikoff. 2003. The U.S. Environmental Protection Agency Particulate Matter Health Effects Research Centers Program: A midcourse report of status, progress, and plans. Environmental Health Perspectives 111(8). Marshall, J.D., S.K. Teoh, and W.W. Nazaroff. 2005. Intake fraction of nonreactive vehicle emissions in US urban areas. Atmospheric Environment 39(7):1363-1371. Maynard, D., B.A. Coull, A. Gryparis, and J. Schwartz. 2007. Mortality risk associated with short-term exposure to traffic particles and sulfates. Environmental Health Perspectives 115(5):751â755. Ministry of Commerce. 2007. Guizhou: Planned investments of over 200 million Yuan to address issues of residential coal use, May 30. Morris, R.E., B. Lau, T.W. Tesche, D. McNally, C. Loomis, G. Stella, G. Tonnesen, and Z. Wang. 2004. VISTAS Emissions and Air Quality ModelingâPhase I Task 4cd Report: Model Perfor- mance Evaluation and Model Sensitivity Tests for Three Phase I Episodes, August. Murphy, J.J., M.A. Deluki, D.R. McCubbin, and H.J. Kim. 1999. The cost of crop damage caused by ozone air pollution from motor vehicles. Journal of Environmental Management 55:273-289. NAPAP (National Acid Precipitation Assessment Program). 2005. National Acid Precipitation Assess- ment Program Report to Congress: An Integrated Assessment. NRC (National Research Council). 2002. Estimating Public Health Benefits of Air Pollution Regula- tions. Washington, D.C.: National Academy Press. NRC. 2003. Managing Carbon Monoxide in Meteorological and Topographical Problem Areas. Washington, D.C.: The National Academies Press. NRC. 2004. Air Quality Management in the United States. Washington, D.C.: The National Acad- emies Press. Ostro, B., H. Tran, and J. Levy. 2006. The Health Benefits of Reduced Tropospheric Ozone in Cali- fornia. Journal of the Air and Waste Management Association 56:1007-1021. Ou, S., L. Pan, and M. Zhuang. 1996. Estimation of the effects and economic loss caused by acid deposition in Xiamen area. Res. Environ. Sci. 9(5):57-59 (in Chinese). Paatero, P. 1997. Least squares formulation of robust, non-negative factor analysis. Chemometrics and Intelligent Laboratory Systems 37:23-35. Parrish, D.D. 2006. Critical evaluation of US on-road vehicle emission inventories. Atmospheric Environment 40:2288-2300. Peng, L., J. Li, and T. Zhu. 2005. Carbon isotope characteristics of polycyclic aromatic hydrocarbons in atmospheric particles and their source analysis in urban area of Zhengzhou City. China En- vironmental Science 25(1):106-109 (in Chinese). Pope, C.A. and D.W. Dockery. 2006. Health effects of fine particulate air pollution: Lines that con- nect. Journal of Air and Waste Management Association 56(6).
AIR POLLUTION: SOURCES, IMPACTS, AND EFFECTS 109 Pope, C.A. III, M.J. Thun, and M.M. Namboordiri. 1995. Particulate air pollution as a predictor of mortality in a prospective study of U.S. adults. American Journal of Respiratory and Critical Care Medicine 151:669-674. Pope, C.A. III, R.T. Burnett, M.J. Thun, E.E. Calle, D. Krewski, K. Ito, and G.D. Thurston. 2002. Lung cancer, cardiopulmonary mortality, and long-term exposure to fine particulate air pollu- tion. Journal of the American Medical Association 287:1132-1141. Pun, B.K. and C. Seigneur. 2007. Investigative modeling of new pathways for secondary organic aerosol formation. Atmospheric Chemistry and Physics 7:2199-2216. Qi, S.H., G.Y. Sheng, J.M. Fu, Z.S. Wang, and Y.S. Min. 2002. Source apportionment experiment of polycyclic aromatic hydrocarbons (PAHs) in aerosols from Macao. China Environmental S Â cience 22(2):118-122 (in Chinese). Qian, Y. and F. Giorgi. 2000. Regional climatic effects of anthropogenic aerosols? The case of South- western China. Geophysical Research Letters 27:3521-3524. Qian, Y., C.B. Fu, R.M. Hu, and Z.F. Wang. 1996. Effects of industrial SO2 emission on temperature ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ variation in China and East Asia. Clim. Environ. Res. 2:143-149 (in Chinese). Qian, Y., L.R. Leung, S.J. Chan, and F. Giorgi. 2003. Regional climate effects of aerosols over China: Modeling and observation. Tellus 55B:914-934. Qiu, J.H. and L.Q. Yang. 2000. Variation characteristics of atmospheric aerosol optical depths and visibility in North China during 1980-1994. Atmospheric Environment 34:603-609. Richter, A., P. Burrows, H. Nues, C. Granier, and U. Niemeijer. 2005. Increase in tropospheric nitro- gen dioxide over China observed from space. Nature 437:129-130. Robinson, A.L., N.M. Donahue, M.K. Shrivastava, E.A. Weitkamp, A.M. Sage, A.P. Grieshop, T.E. Lane, J.R. Pierce, and S.N. Pandis. 2007. Rethinking organic aerosols: Semivolatile emissions and photochemical aging. Science 315(5816):1259-1262. Russell, A.G.Â 2007. EPA supersites program-related emissions-based particulate matter modeling: Applications and advances. Journal of the Air and Waste Management Association.Â Accepted for publication. SCAQMD (South Coast Air Quality Management District). 2000. Multiple Air Toxics Exposure Study (MATES-II) Final Report, March. Schwartz, J., F. Laden, and A. Zanobetti. 2002. The Concentrationâresponse relation between air pollution and daily deaths. Environmental Health Perspectives 110(10). Seigneur, C. and M.D. Moran. 2004. Using models to estimate particle concentration. Chapter 8 in Particulate Matter Science for Policy Makers: A NARSTO Assessment, P. McMurry, M. Shep- herd, and J. Vickery, eds. Cambridge: Cambridge University Press. Seigneur, C., B. Pun, P. Pai, J. Louis, P.A. Solomon, C. Emery, R. Morris, M.S. Zahniser, D.R. W Â orsnop, P. Koutrakis, W. White, and I. Tombach. 2000. Guidance for the performance evalu- ation of three-dimensional air quality modeling systems for particulate matter and visibility. Journal of the Air and Waste Management Association 50:588-599. Seinfeld, J.H. and S.N. Pandis. 1998. Atmospheric Chemistry and Physics: From Air Pollution to Global Change. New York: John Wiley and Sons. Seinfeld, J.H. and J.F. Pankow. 2003. Organic atmospheric particulate material. Annual Review of Physical Chemistry 54:121-140. SEPA (State Environmental Protection Administration). 2000. Report on the State of the Environment in China 1999. SEPA. 2001a. Annual Report of Environmental Statistics 2000. SEPA. 2001b. Report on the State of the Environment in China 2000. SEPA. 2002. Report on the State of the Environment in China 2001. SEPA. 2003. Report on the State of the Environment in China 2002. SEPA. 2004. Report on the State of the Environment in China 2003. SEPA. 2005a. Annual Report of Environmental Statistics 2004. SEPA. 2005b. Report on the State of the Environment in China 2004. SEPA. 2006. Report on the State of the Environment in China 2005.
110 ENERGY FUTURES AND URBAN AIR POLLUTION SEPA. 2007. Report on State of the Environment in China 2006. Shao, M., X. Tang, Y. Zhang, and W. Li. 2006. City clusters in China: Air and surface water pollution. Frontiers in Ecology and the Environment 4(7):353-361. Shu, J., H. Cao, Y. Gao, et al. 1993. Ecological criteria for Masson pine effected by acidic precipitation and evaluation of economic loss. China Environ. Sci. 13(4):279-283 (in Chinese). Smith, K.R., S. Mehta, and M. Maeusezahl-Feuz. 2004. Indoor smoke from household solid fuels, in Comparative Quantification of Health Risks: Global and Regional Burden of Disease due to Selected Major Risk Factors. M. Ezzati, A.D. Rodgers, A.D. Lopez, and C.J.L Murray, eds. Geneva: World Health Organization, vol. 2, 1435-1493. Song, Y., Y.H. Zhang, S.D. Xie, L.M. Zeng, M. Zheng, L.G. Salmon, M. Shao, and S. Slanina. 2006a. Source apportionment of PM2.5 in Beijing by positive matrix factorization. Atmospheric Envi- ronment 40:1526-1537. Song. Y., S.D. Xie, Y.H Zhang, et al. 2006b. Source apportionment of PM2.5 in Beijing using principal component analysis/absolute principal component scores and UNMIX. Science of the Total Environment 372:278-286. SSB (Shandong Province Statistical Bureau). 2006. Shandong Statistical Yearbook 2005. Streets, D.G. and S.T. Waldhoff. 2000. Present and future emissions of air pollutants in China: SO 2, NOx and CO. Atmospheric Environment 34:363-374. Streets, D.G. et al. 2001. ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ Black carbon emissions in China. Atmospheric Environment 35:4281-4296. Streets, D.G., T.C. Bond, G.R. Carmichael, S.D. Fernandes, Q. Fu, D. He, Z. Klimont, S.M. Nelson, N.Y. Tsai, M.Q. Wang, J.H Woo, and K.F. Yarber. 2003. A year-2000 inventory of gaseous and primary aerosol emissions in Asia to support TRACE-P modeling and analysis. Journal of Geophysics Research 108(D21):8809. Sun, Y.L., G.S. Zhuang, A.H. Tang, et al. 2006. Chemical characteristics of PM2.5 and PM10 in haze- fog episodes in Beijing. Environmental Science and Technology 40:3148-3155. Turpin, B.J. and J.J. Huntzicker. 1995. Identification of secondary organic aerosol episodes and quan- titation of primary and secondary organic aerosol concentrations during SCAQS. Atmospheric Environment 29:3527-3544. UNDP (United Nations Development Programme). 2002. China Report on Human Development 2002. Beijing, China: Financial Economy Press. UNEP (United Nations Environmental Programme). 1999. Global Environment Outlook 2000. L Â ondon, UK: Earthscan Publications Ltd. UNEP. 2002. Global Environmental Outlook 2003. London, UK: Earthscan Publications Ltd. Vayenas, D.V., S. Takahama, C.I. Davidson, and S.N. Pandis. 2005. Simulation of the thermodynamics and removal processes in the sulfate-ammonia-nitric acid system during winter: Implications for PM2.5 control strategies. Journal of Geophysical Research 110:D07S14. Wan, X.L., J.W. Chen, F.L. Tian, et al. 2007. Source apportionment of PAHs in atmospheric particu- lates of Dalian: Factor analysis with nonnegative constraints and emission inventory analysis. Atmospheric Environment 40(34):6666-6675. Wang, M.X. 1985. Using factor analysis to study aerosol sources. Chinese Journal of Atmospheric Sciences 9(1):73-81 (in Chinese). Wang, S.X., J.M. Hao, M.S. Ho, J. Li, and Y.Q. Lu. 2006. Intake fractions of industrial air pollutants in China: Estimation and application. Science of the Total Environment 354:127-141. Wang, X., G. Carmichael, D. Chen, Y. Tang, and T. Wang, 2005. Impacts of different emissions sources on air quality during March 2001 in the Pearl River Delta (PRD) region. Atmospheric Environment 39:5227-5241. Wang, X.P. and D.L.Mauzerall. 2004. Characterizing distributions of surface ozone and its impact on grain production in China, Japan and South Korea: 1990 and 2020. Atmospheric Environment 38:4383-4402. Wang, Y., M.B. McElroy, K.F. Boersma, H.J. Eskes, and J.P. Veefkind. 2007. Traffic restrictions associated with Sino-African summit: Reductions of NOx detected from space, Geophysical Research Letters 34:L08814.
AIR POLLUTION: SOURCES, IMPACTS, AND EFFECTS 111 Watson, J.G., J.A. Cooper, and J.J. Huntzicker. 1984. The effective variance weighting for least squares calculations applied to the mass balance receptor model. Atmospheric Environment 18:1347â1356. WHO (World Health Organization). 2006. WHO Air Quality Guidelines for Particulate Matter, Ozone, Nitrogen Dioxide and Sulfur Dioxide: Global Update 2005. World Health Organiza- tion, Geneva. World Bank. 1997. Chinaâs Environment in the New Century: Clear Water, Blue Skies. The World Bank, Washington, D.C. Wu, D., X.X. Tie, C.C. Li, et al. 2005. An extremely low visibility event over the Guangzhou region: ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ A case study. Atmospheric Environment 39:6568-6577. Wu, J., W.M. Jiang, C.B. Fu, et al. 2004. Simulation of the radiative effect of black carbon aerosols and the regional climate responses over China. Advances in Atmospheric Science 21(4):637-649. Wu, Y., S.X. Wang, D.G. Streets, J.M. Hao, M. Chan, and J.K. Jiang. 2006. Trends in anthropo- genic Â mercury emissions in China from 1995 to 2003. Environmental Science Technology 40:5312-5318. Xiao, H.Y. and C.Q. Liu. 2004. Chemical characteristics of water-soluble components in TSP over Guiyang, SW China, 2003. Atmospheric Environment 38(37):6297-6306. Xu, Q. 2001. Abrupt change of the mid-summer climate in central east China by the influence of atmospheric pollution. Atmospheric Environment 35:5029-5040. Zhang, D. 2006. Source apportionment of TSP in ambient air in Shuozhou center urban. Shanxi Chemical Industry 26(3):61-63 (in Chinese). Zhang, J. and K.R. Smith. 2005. Indoor air pollution from household fuel combustion in China: A review. The 10th International Conference on Indoor Air Quality and Climate, 1-26. Zhang, M.G. 2004. Modeling of organic carbon aerosol distributions over east Asia in the springtime. China Particuology, 2(5):192-195. Zhang, M.G., Y.F. Xu, R.J. Zhang, et al. 2005. Emission and concentration distributions of black Âcarbon aerosol in East Asia during the spring time. Chinese Journal of Geophysics, 48(1):55-61. Zhang, M.G., Y.F. Pu, R.J. Zhang, et al. 2006. Simulation of sulfur transport and transformation in East Asia with a comprehensive chemical transport model. Environmental Modelling & Soft- ware, 21(6):812-820. Zhang, X.Y., J.J. Cao, et al. 2001. Particulate pollution control in Xiâan. ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ In: Chinese Urban Air Pol- lution Control. Chinese Science & Technology Press. 222-293 (in Chinese). Zhang, Y., K. Shao, X. Tang, and J. Li. 1998. The Study of Urban Photochemical Smog Pollution in China. Acta Scicentiarum Naturalum Universitis Pekinesis 34(2-3):392-400. Zhang, Y.H., X.L. Zhu, L.M. Zeng, et al., 2004. Source apportionment of fine-particle pollution in Beijing, In: Urbanization, Energy, and Air Pollution in China. National Academy of Engineer- ing and National Research Council. Washington, D.C.: The National Academies Press. Pp. 139-153. Zhao, D., J. Xiong, Y. Xu, and W.H. Chan. 1988. Acid rain in southwestern China. Atmospheric Environment (1967) 22(2):349-358. Zhao, D.S., M.X. Wangï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ , et al. 1991. Coal-combustion air pollution ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ andï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ aerosolï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ . Chinese Environ- mental Press 1-404 ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ï¿½ (in Chinese). Zhou, X.J., W.L. Li, and Y.F. Luo. 1998. Numerical simulation of the aerosol radiative forcing and regional climate effect over China. Scientia Atmospherica Sinica 22(4):418-427 (in Chinese).