*Primary author to whom correspondence should be sent. Email: dstreets@anl.gov

Black carbon (also known as elemental carbon or soot), together with organic carbon, comprises the aerosol component of incompletely combusted carbonaceous fuels. Despite its small size (< 1 Fm), black carbon may contribute to climate change, soiling of building materials, and negative health effects. Black carbon is estimated to have an atmospheric residency time of 3-7 days (Liousse et al., 1996) to several weeks (Parungo et al., 1994; Wolff, 1981). In order to properly deal with these environmental and economic problems, the sources and quantities of black carbon emissions in China must be correctly estimated.

Black carbon is the portion of these aerosols that primarily absorbs solar radiation (Liousse et al., 1996; Penner et al., 1993). This property is in contrast to most other aerosols, which are thought to have light-scattering properties that may to some degree counteract the global warming effects of carbon dioxide and other products of incomplete combustion (carbon monoxide, methane, non-methane hydrocarbons). The indirect climate effects of black carbon may also tend to a net warming of the atmosphere. It is believed that black carbon affects cloud albedo, its strong hydrophilic qualities increasing the number of cloud condensation nuclei. It has been estimated that the solar-absorption effect of black carbon may counteract the cooling effect of other aerosols significantly, even bringing the total effect to a net warming of 0.5E C from a net cooling of 1.2EC (Liousse et al., 1996).

Black carbon also causes other, more immediately evident, problems. Its light-absorbing properties reduce atmospheric visibility (Wolff, 1981) and these fine particles may coat building materials, damaging the appearance of homes, public buildings, and historic landmarks (Hamilton and Mansfield, 1991).

Despite all the possible climate implications, these may not be the most severe effects from black carbon emissions in China. The small diameter (< 1 Fm) of black carbon particles allows for deep penetration into the lungs. There the particles may "slow the clearance mechanisms and provide absorption sites for toxic pollutants." (Hamilton and Mansfield, 1991). With nearly three-quarters of black carbon emissions in China originating in the domestic sector— where exposure levels are greatest—and the unvented or poorly vented conditions in most Chinese household stoves, there is obvious cause for concern.

Black carbon emissions for this inventory were calculated using fuel consumption data and sector-specific black carbon (BC) emission factors. By examining emission factors from several sources (Tables 1 and 2) to determine which sector/fuel combinations have the highest emission rates and combining this information with data on Chinese energy consumption, based on data in the RAINS-Asia model, version 7.02 (Foell et al., 1995), the most significant sector/fuel combinations were determined.

It was found that five sectors (domestic, industry, agricultural combustion, power, and transportation) contribute most heavily to BC emissions. Within the domestic sector, emissions from all primary fuels (except fuel oil and gas, due to their very low emission factors) are calculated. There are, then, five fuel categories within the domestic sector. These are: hard coal, derived coal or coal briquettes, fuelwood, crop residues, and animal waste. Within the transportation sector, both medium distillate oils and heavy fuel oils are found to be significant. For each of the other sectors only one fuel type contributes significantly to BC emissions. Industrial and power emissions are greatest from hard coal combustion; and agricultural burning is, of course, of crop residues.

For each regional and sector/fossil fuel combination, energy values are estimated from the RAINS-Asia energy model, assuming a fixed rate of exponential growth as described in Streets et al. (1998). This method is used in order to preserve the regional specificity in the model. Several sources (Tables 1 and 2) are used to derive emission factors for black carbon. Those that were determined to best represent actual emissions in China are used in this work. These emission factors are then combined with 1995 energy consumption data for Chinese regions to yield black carbon emissions. Figure 1 identifies the 27 regions into which China is divided in the RAINS-Asia model. Specific carbon contents, black carbon emission factors, and control measures are applied as follows.

Black carbon emissions from the domestic sector are calculated for five fuel types: hard coal, derived coal (coal briquettes), fuelwood, crop residues, and animal waste.

Regional hard coal energy values for 1995 are derived as described below. Domestic consumption of hard coal in 1995 was 4,699 PJ, 18.7% of total coal consumption. A carbon content of 25.8 kt-C/PJ (IPCC, 1997) is applied to the energy value to yield total carbon content. The black carbon emission factor for hard coal of 1.75% is used to calculate total emissions; because there is no control technology in the domestic sector, no control factor is applied to emissions in this sector. Total hard coal emissions in 1995 for the domestic sector, as calculated here, are 2.12 mt-C (million metric tonnes (10¹² g) of carbon).

Derived coal emissions are calculated in a similar way. In 1995, derived coal was beginning to be introduced in selected areas of China. Total domestic consumption was 75 PJ, only 0.3% of total coal consumption, or 1.6% of residential coal consumption. A carbon content value of 27.0 kt-C/PJ is used to convert energy to total carbon. A black carbon emission factor of 0.7% is applied to account for the cleaner burning of coal that has been washed and processed—60% fewer particulate emissions compared to unwashed hard coal (Xie, 1991). Total black carbon emissions in this category are small, less than 1% of domestic, hard coal emissions (14.2 kt-C).

Because biofuel energy data are not currently available in the RAINS-Asia model, 1995 energy consumption rates were taken from other sources (Streets and Waldhoff, in press; Streets et al., 1998). Total carbon content for each of the three fuels (fuelwood, crop residues, and animal waste) is calculated by multiplying energy data by the conversion rate of 29.9 kt-C/PJ (IPCC, 1997). Fuel-specific emission factors are then applied. These are 0.25% for fuelwood, 0.22% for both crop residues and animal waste (Table 2). This type of domestic burning is uncontrolled, so no control factors are used in calculating total biofuel emissions (0.69 mt-C).

Large industrial boilers combust fuels more completely than do smaller domestic devices, so black carbon emissions for most fuels tend to be quite low. In addition, many industrial plants in China use some form of particulate emission control, mainly electrostatic precipitators (ESP). In this study we assume that the ESP collection efficiency for black carbon is approximately equal to the total particulate collection rate (IEA, 1997). Black carbon emissions in the industrial sector are found to be significant only from hard coal combustion.

BC emissions for coal combustion in the industrial and power generation sectors are calculated as follows:

BC_{Power + Industry} = (E_Power + E_Industry) * 25.8 kt-C/PJ * 1.75 * (1 - Collection Rate)

Where BC is black carbon emissions, E is energy in PJ (10¹⁵ J), and the collection rate is an average percent collection of all particles in industry and power generation, by province (China Environment Yearbook, 1996). Then black carbon emissions from the power generation sector are calculated as follows (see later discussion for the derivation of the power-sector collection efficiency of 96.4%):

BC_Power = (E_Power) * 25.8 kt-C/PJ * 1.75 * (1 - 0.964)

Thus:

BC_Industry = BC_{Power + Industry} - BC_Power

In other words, the standard carbon content of 25.8 kt-C/PJ (IPCC, 1997) is applied to the energy data from both the industrial and power generation sectors. The base emission factor of 1.75% is multiplied by total carbon content to determine total uncontrolled black carbon emissions. Controlled industrial emissions are calculated using a two-step process for each region. First, combined black carbon emissions from power generation and industrial combustion were calculated based on the 1995 average provincial particulate collection rates found in the China Environment Yearbook (1996). Next, emissions from the power sector were calculated as described below. These emissions were subtracted from the total (power and industrial) emissions to yield industrial emissions. In this way, the higher rates of particulate collection found in the power sector are used, while preserving the provincial differences in emission collection available from the China Environment Yearbook (1996). Industrial BC emissions are thus estimated to be 0.88 mt-C.

Power generation in China is a relatively low emitter of black carbon. This is due in part to the high combustion efficiency, and in part to particulate controls that are more common in this sector than the industrial or domestic sectors. All new power plants and most old plants use electrostatic precipitators to collect particulate emissions. China has made great strides in collecting particulate matter from coal-fired power plants in recent years. In 1983, only 25 generating units were equipped with high-performance electrostatic precipitators. By 1991, this had increased to 236 units and by 1993 to 316 units (SSTC, 1995; Zhai, 1992). Today, almost all new and renovated power plants employ ESPs. The goal of the eighth Five-Year Plan (1990-1995) was to raise the average collection efficiency of particulate matter from coal-fired power plants to 95%. In fact, this goal was reached in 1993 (SSTC, 1995). Extrapolating from the historical data (SSTC, 1995; Zhai, 1992; Xie, 1991), we estimate that in 1995 the average particulate-matter removal efficiency from coal-fired power plants was 96.4%.

As in other sectors, coal consumption for 1995 is estimated from data for 1990 and 2000 from the RAINS-Asia model—6,487 PJ, or 25.8% of total coal combustion in China. These energy data were converted to total carbon content using the rate of 25.8 kt-C/PJ (IPCC, 1997). Total carbon values were multiplied by the base emission rate of 1.75% to yield black carbon emissions. The control rate of 96.4% was then applied to account for the particulate controls. The total black carbon emissions from the power generating sector in China in 1995 are estimated to be 105 kt-C.

Two fuel categories were examined for the transportation sector. Medium distillate fuels are considered to be primarily diesel-burning highway vehicles and contributed 497 PJ, or 68.8% of total energy from diesel sources for this sector. Heavy fuel-oil is primarily burned in diesel locomotives, and the energy consumption was estimated to be 225 PJ. Energy data were converted to carbon content using the value 20.2 kt-C/PJ (IPCC, 1997). Diesel fuel combustion in the transportation sector has a relatively high emission factor of 0.21%. The black carbon emission factor is applied to total fuel carbon to estimate emissions. Total black carbon emissions for the transportation sector are very small in 1995, less than 1% of total black carbon emissions, at 31 kt-C.

Burning of crop residues in the field as a means of disposal is fairly common in China. Black carbon emissions from this source, though small, are significant. Several assumptions have to be made in order to estimate this source of black carbon. Crop production statistics, available by province and crop (FAO,1998), are combined with crop-to-residue ratios (Lu, 1993) to calculate available regional crop residues. Although other crop residues may be burned in the fields, emissions are calculated based on data for the three largest crops: rice, wheat, and corn. Approximately 80% of crop residues are combusted in Asia, including those used for domestic fuel (Crutzen and Andreae, 1990). Residues burned in fields make up approximately 23% of total residues. Although it is believed that rice husks comprise the majority of residues burned in fields, no data were found on precise quantities, so a flat rate of 23% of available residues is applied to rice, wheat, and corn residues. Total mass is converted from mass of "dry matter" to carbon content using the conversion factor of 45% (IPCC, 1997). This amount is converted to black carbon by multiplying by the appropriate emission factor: rice - 0.13%, wheat - 0.20%, and corn - 0.16% (Liousse, 1996). Total black carbon emissions from combustion of crop residues in fields are estimated to be 75 kt-C.

Our inventory places total black carbon emissions in China in 1995 in the range of 3.9 mt-C. Preliminary results suggest that the majority of black carbon emissions originate in the domestic sector (Table 3). This is due largely to the high rate of incomplete combustion and uncontrolled nature of domestic burning. Nearly three-quarters (72.2%) of total BC emissions come from the domestic sector, with coal combustion generating the highest share (54.6% of total emissions or 2.12 mt-C) and biofuels contributing about half that amount (almost 17.6% of total emissions or 0.69 mt-C). Industrial coal combustion generates slightly more black carbon than domestic biofuel use, totaling 0.88 mt-C, or 22.4% of total emissions. Combustion of crop residues as a means of disposal contributes about 75 kt-C (1.9%). Power generation contributes a similarly small amount, about 105 kt-C (2.7%). In 1995 the transportation sector’s contribution to black carbon emissions is negligible, less than 1% of total emissions, only 31 kt-C. However, since emissions in this sector are expected to rise considerably by 2020, values for 1995 are included for comparison purposes. In this way, Chinese distribution differs significantly from emissions in the United States, where the transportation sector contributes the majority of black carbon emissions, due to the high particulate emission factor from diesel combustion and the fact that there are effective particulate controls on most other sources.

With a combined total of 40% of BC emissions, the three RAINS-Asia regions (Figure 1) with the highest emissions of black carbon are HEHE (Hebei, Anhui, and Henan), NEPL (Heilongjiang, Liaoning, and Jilin), and SICH (Sichuan) at 0.68, 0.50, and 0.38 mt-C, respectively. These are the three most populous regions, and all three regions are also centers of industry in China. Since domestic and industrial combustion are the first and second largest contributors to BC emissions, it is not surprising that these three regions have exceptionally high emissions.

Table 4 compares the work of Penner et al. (1993) to this work. Penner et al. estimated total black carbon emissions for 1980 using two methods. The first method employed regional SO₂-to-BC ratios to calculate black carbon emissions, based on SO₂ inventories, with a result of 3.7 mt-C for China. The second method used energy data and emission factors, similar to the method used in this work, and the results were 2.7 mt-C. The second method results in emissions more consistent with this work (3.9 mt-C in 1995), allowing for energy growth over the 15-year difference between the two estimates. Despite the large absolute differences in emissions, the order of magnitude is compatible (10¹² g) and the percent emissions by fuel type are fairly consistent. Parungo et al. (1994) also estimate the annual black carbon emissions in China to be on the order of 1 _ 10⁶ to 1 _ 10¹² g. This work seems to agree with the larger estimate, finding approximately 3.9 _ 10¹² g in 1995.

Emissions projections for 2020 are calculated on the basis of 2020 energy projections from the RAINS-Asia model and emission factors of technology/fuel combinations likely to be representative of China in 2020. For some sectors, the energy projections are combined with emission factors for typical performance of similar facilities in the United States in 1990. This method assumes that Chinese technology will trail the United States by approximately 30 years, and is useful when it is not possible to develop an average emission rate because of fundamental and unpredictable changes in technology structure. Typically, emissions are estimated at the national level and then shared out to the regions using appropriate sector/fuel regional energy consumption data for 2020 from the RAINS-Asia model. Table 5 contains the projected regional and national BC emissions data for all major emitting sectors in 2020. The sector-specific methodology for 2020 is as follows.

Coal is presently combusted directly in a variety of inefficient stoves and hearths for the purposes of residential cooking and heating. It is expected that such use will continue in the future, though at a lower growth rate than energy and population projections would indicate, because of the penetration of cleaner fuels, including coal briquettes (derived coal, or DC), natural gas, and electricity. The RAINS-Asia model projects an increase in domestic coal combustion from 4,699 PJ in 1995 to 7,009 PJ in 2020.

Due to the inefficiency of combustion in most of the rural applications in China, present-day emissions of BC from this source are the highest and contributed 53% of total BC emissions in 1995. Expectations for the future are that more-efficient cook stoves will increasingly penetrate the Chinese market, leading to more complete combustion and lower BC emissions. To develop an appropriate emission factor it is necessary to look back in time in the U.S., because domestic coal combustion is virtually nonexistent today. We have adopted the value of 378 kg/10⁹ Btu of total particulates (Gray, 1982; EPA, 1995 ); this is a value typical for the best combustors in the U.S. in the late 1970s and is appropriate for China in 2020. This translates to a BC emission factor of 49.5 kt/10¹⁵ Btu (Gray, 1982) and an emissions estimate of 329 kt-C for China in 2020. Thus, through the application of more-efficient combustors, an increase in domestic coal consumption of 49.2% between 1995 and 2020 translates to a reduction in BC emissions of 85%. By 2020 the direct residential combustion of coal contributes only 17% of total BC emissions.

A portion of present-day direct combustion of coal will be replaced by the combustion of coal briquettes, consumption of which is expected to grow rapidly from 75 PJ in 1995 to 600 PJ in 2020. Because briquettes are generally burned efficiently today whenever they are used, the emission factor is expected to stay relatively constant in the future, leading to a 2020 national emissions estimate of 128 kt-C.

The regional distribution of domestic, derived-coal combustion in 2020 in the RAINS-Asia model is distorted, because of extrapolation from an unusual base-case situation. We expect that coal briquettes will be available in all urban areas by 2020 and will be especially valuable in the smaller, more remote cities where penetration of natural gas and electricity for domestic use may still be lacking. Rather than use the 2020 RAINS-Asia regional distributions, therefore, we have chosen to use the present-day (1991) distribution of urban populations within regions to allocate emissions from the domestic use of coal briquettes (China Statistical Yearbook, 1992).

It is difficult to predict how much biofuel will be consumed in China in 2020. As cleaner burning, more-efficient fossil fuels and electricity become more available, especially in rural areas, biofuel use is expected to be phased out. However, recent trends and examination of causative factors lead us to believe that this phase-out might not be as rapid as is commonly believed. Several assumptions are used to estimate consumption of biofuels and the resulting black carbon emissions. The rate of per capita consumption is calculated from data for previous years (Sinton et al., 1996; Lu, 1993). All rates show biofuel use to be declining on a per capita basis. Because it is believed that biofuels will play less of a role in the future energy needs of China, the greatest level of present-day decline (-0.01 kgce cap^-1 year^-1) is applied to each of the three main types of biofuel: fuelwood, crop residues, and animal waste (1995 total values are converted to per capita values). Per capita totals are then multiplied by 1,450 million, the Asian Development Bank’s projection for Chinese population in 2020 (Siddiqi et al., 1994). Total energy values are multiplied by appropriate emission factors (assumed to remain at 1995 levels) to yield 642 kt-C from domestic biofuel combustion, which is allocated to the regional level based on the 1995 distribution.

The industrial use of coal represents one of the most polluting forms of energy use in China today. Some 13,888 PJ of coal were combusted in 1995 in facilities ranging in size from large manufacturing facilities in the heavy industries to small, "cottage" industries that are barely distinguishable from residences. Whereas many of the former facilities do utilize some form of particulate controls (wet scrubbers, cyclones, or possibly electrostatic precipitators) even today, the latter almost certainly use no form of emissions cleanup and vent waste gases directly to the atmosphere. In many of the coal-burning regions in China, even quite large industrial facilities use no particulate cleanup devices and contribute much to the particulate burden of the surrounding air.

In 1995, BC emissions from industrial coal combustion were about 876 kt-C, somewhat lower than domestic coal combustion per unit of energy released, partly because of greater average combustion efficiency and partly because of emission controls. Also, whereas residential coal use is widely dispersed across rural and suburban areas, industrial coal use tends to be concentrated in or close to the major urban centers. Industrial coal combustion is estimated to have contributed 22% of total BC emissions in 1995.

To predict what industrial coal use will look like in China in 2020 is virtually impossible. The projections of RAINS-Asia show more than a doubling of industrial coal use over 1995 levels, reaching 30,855 PJ by 2020, driven by economic factors. Though impossible to quantify, the expected future trends are: (a) a reduction or elimination of very small industrial coal-burning enterprises on either profitability or environmental grounds; (b) a consolidation of industrial production in large, western-style, manufacturing facilities; (c) replacement of much of the equipment in existing medium-sized facilities with more modern equipment on energy-efficiency grounds; and (d) the rapid penetration of efficient particulate cleanup devices (ESPs or baghouses) on all large boilers and burners.

Because the pace and extent of these transformations cannot be easily predicted, we again assume that China in 2020 will have reached the stage of the U.S. industrial coal industry as it existed in 1990. In that year, the U.S. consumed about 1.71 quads or 1,804 PJ of industrial steam coal (US DOE, 1990). National emissions of PM₁₀ in 1990 were about 76.2 kt (EPA, 1997). Converting this value to BC emissions (Gray, 1982; EPA, 1992; EPA, 1997) and adjusting it to the China 2020 consumption value of 30,855 PJ, yields a national BC emissions estimate of 304 kt-C. In similar fashion to the situation for domestic coal combustion, therefore, we find an increase in industrial coal combustion of 122% between 1995 and 2020 translating to a reduction in BC emissions of 65% between 1995 and 2020–due to the combination of modernization, restructuring, and environmental control. The share of industrial BC emissions in 2020 thus drops to 16.3%.

The relatively small projected combustion of oil products for industrial energy production in 2020, coupled with the high efficiency of the combustion devices and the low proportion of BC in the total particulate fraction, yields negligibly small contributions from these sources.

A large increase in coal combustion for electric power generation in China is expected. Between 1995 and 2020, the RAINS-Asia model projects an increase from 6,487 PJ to 17,655 PJ. Today, almost all major power plants employ particulate control devices; only a few older, smaller ones are uncontrolled. New plants being built today use highly efficient electrostatic precipitators, while some older plants use the less-efficient wet-particle scrubbers or cyclones. Because of this level of control, 1995 BC emissions from power plants are quite low–105 kt-C, or 2.6% of total.

By 2020, we can anticipate that all the uncontrolled or poorly controlled power plants will have been retired or upgraded, such that the average level of particulate emissions approaches the performance of the best electrostatic precipitators. On this premise, we can again approximate Chinese emission rates in 2020 to U.S. emission rates from this sector in 1990. In that year, U.S. power plants consumed 16,943 PJ of coal (US DOE, 1990) and emitted 240 kt of PM₁₀ (EPA, 1997). In similar fashion to the calculation for industrial coal use, we therefore estimate emissions of BC from power generation in China in 2020 to be 59 kt-C.

Though coal use is growing rapidly, the advent of stricter particulate control measures reduces BC emissions by 44% between 1995 and 2020. The sectoral share of the total, however, remains almost the same at 3%.

By far the most important development in China between now and 2020, as regards particulate emissions, is the expected switch to diesel engines for a large proportion of the auto, truck, and rail fleet. Diesel engines are energy efficient, generally low polluting, and relatively cheap. With liquid fuels expected to be at a premium in China in the future, the energy efficiency advantage is highly valued. Because of this, diesel engines and fuels figure prominently in all Chinese projections of the energy future. A big drawback, however, is the high rate of fine particle emissions. In 1995, the extent of diesel engine usage for transportation was quite small, and BC emissions were only about 31 kt, or less than 1% of total.

The RAINS-Asia model projects a huge increase in the use of middle distillate (MD) oils (diesel oil) for transportation purposes from 497 PJ in 1995 to 6,707 PJ in 2020. This is consistent with the burgeoning of the Chinese transportation system and the trend toward diesel fuel (Tsinghua University, 1997) We estimate that diesel automobiles will capture 25% of the light-duty vehicle market in China by 2020 and diesel trucks will capture 70% of the heavy-duty vehicle market (Streets et al., 1998).

Gray (1982) estimates the BC emission factors for diesel automobiles to be 51 kg/10⁹ Btu and diesel trucks 33 kg/10⁹ Btu. If we develop splits for automobiles and trucks based on the expected number of vehicles on the road in 2020 and the average annual mileage driven by the two classes of vehicles, we obtain BC emissions of 177 kt-C for diesel automobiles and 104 kt-C for diesel trucks in 2020. This is such a large increase that BC emissions from highway vehicles contribute 15.1% of the total in 2020.

In addition, diesel-powered locomotives are expected to continue to replace coal-fired locomotives for rail haulage. We project an increase in diesel fuel (HF/TRA) usage for rail transport from 225 PJ in 1995 to 1,537 PJ in 2020. Using Gray’s BC emission factor of 48.8 kg/10⁹ Btu for diesel-fueled locomotives, an estimate for 2020 of 71 kt-C of BC is obtained. In total, the transportation share of national BC emissions rises from 0.8% in 1995 to 18.9% in 2020. Obviously, solving the problem of fine particulate emissions from diesel vehicles is particularly important for China’s future.

Because combustion of crop residues in the field is used as a means of disposing of this waste, and not as an energy source, it has the potential to vary significantly over a period of years. Efforts are currently being made to encourage farmers to plow under these wastes to increase the organic content and fertility of the soil.

Several factors needed to be examined in order to estimate emissions from this sector. The first factor is the projected Chinese grain production in 2020. The World Bank series, "China 2020" predicts that about 636 mt of grains (rice, wheat, and "coarse" grains) will be harvested in 2020. This is expected to produce a total of 737 mt of grain residues. It is further assumed that efforts to educate farmers on the benefits of plowing-under versus burning crop residues would have a significant impact on farmers’ behavior, reducing the percent of residue burned in the fields in 2020 to one-half the rate in 1995 (from 23% to 11.5%). The black carbon emission factors for this uncontrolled type of burning are assumed to remain equal to the 1995 rates. When this emission factor is applied to the total estimated crop residues subjected to burning, the result is 56 kt-C. Because of a lack of provincial detail for 2020, regional distribution is assumed to be the same as in 1995.

Black carbon emissions in 2020 are projected to be 1.87 mt-C (Table 5), less than one-half the total for 1995. The sectoral distribution is spread much more evenly in 2020 than 1995, though domestic emissions continue to dominate with nearly 60% of total black carbon emissions (Figure 2). Fossil fuel emissions in the domestic sector are only 457 kt-C. The decrease is primarily from hard coal, a fuel that is expected to burn much more cleanly in 2020 than 1995. There is expected to be relatively significant increases in black carbon emissions from derived coal, but not nearly enough to balance the large decrease in hard-coal emissions. Domestic fossil-fuel emissions are less than one-quarter of the 1995 total for this sector. Though black carbon emissions due to biofuel combustion are expected to decrease, this sector’s share of total emissions will rise due to the significant decline in total emissions.

Emissions in the industrial and power generation sectors are expected to decline significantly in this time period (65.3% and 44.5%, respectively). This decrease is due entirely to expected improvements in emission control technology, a point that is made especially clear by the fact that energy use is projected to more than double in this same time frame.

Black carbon emissions from field combustion are also expected to decrease by 2020. This decrease is based on the assumption that farmers will burn fewer crop residues in the fields as a means of disposal. Emission factors will remain fairly constant, and it is unlikely any pollution controls will be instituted in this type of open burning, so any decrease in emissions must be based on a decline in combustion activity for this sector.

The largest increases in Chinese black carbon emissions in 2020 will be in the transportation sector. Emissions are expected to be 353 kt-C in 2020 (18.9% of total emissions), compared to only 31 kt-C in 1995. It is possible to avoid much of the growth in particulate emissions from diesel combustion by using cleaner-burning diesel engines, or by installing particulate controls. If this is done, much of the 11-fold increase in black carbon emissions may be averted.

1995

2020

The authors are grateful to The World Bank for permission to use Figure 1. This work was funded by the National Aeronautics and Space Administration as part of the China-MAP program.

Burnet, P. G., Edmisten, N. G., Tiegs, P. E., Houck, J. E., and Yoder, R. A. (1986) Journal of the Air Pollution Control Association 36 1012-1018.

Butcher, S. S. and Ellenbecker, M. J. (1982) Journal of the Air Pollution Control Association 32 380-384.

China Statistical Yearbook, China Statistical Publishing House, 1992.

China Environment Yearbook, Zhang, L., ed., 1996.

China Science and Technology Newsletter, No. 43 (1995) The State Science and Technology Commission, People’s Republic of China.

Crutzen, P.R. and Andreae. M. O. (1990) Science 250 1669-1678.

Dasch, J. M. (1982) Environmental Science Technology 6 638-645.

Dasch, J. M. and Cadle, S. H. (1989) Aerosol Science and Technology 10 236-248.

Foell, W., Amann, M., Carmichael, G., Chadwick, M., Hettelingh, J. P., Hordijk, L., and Dianwu, Z. (1995) RAINS-Asia: An Assessment Model for Air Pollution in Asia, Final Report, The World Bank.

Food and Agriculture Organization, (1995) FAOSTAT Agricultural Data,

Gray, H.A., "Control of Atmospheric Fine Primary Carbon Particle Concentrations," EQL Report 23, California Institute of Technology, Pasadena, CA, February 1986.

Hamilton, R. S., and Mansfield, T. A. (1991) Atmospheric Environment 25A 715-723.

International Energy Agency (1997) "Particulate control handbook for coal-fired plants," IEA Coal Research, London, UK.

IPCC (1997) "Greenhouse Gas Inventory Workbook, Vol. 2" J. T. Houghton et al. eds., Bracknell, UK.

Liousse, C., Penner, J. E., Chuang, C., Walton, J. J., and Eddleman, H. (1996) Journal of Geophysical Research 101 19,411-19,432.

Lu, Y. (1993) Fueling One Billion: An Insider’s Story of Chinese Energy Policy Development, Washington Institute Press, Washington, D.C.

Muhlbaier, J. L. and Williams, R. L. in "Particulate Carbon: Atmospheric Life Cycle," Wolff and Klimisch, eds. (1982).

Mumford, J. L., Harris, D., Williams, B. K., Chuang, J. C., and Cooke, M. (1987) Environmental Science Technology 21 308-311.

Ogren, J. A., and Charlson, R. S. (1984) Tellus 36 262-271.

Parungo, F., Nagamoto, C., Zhou, M. Y., Hansen, A. D. A., and Harris, J. (1994) Atmospheric Environment 28 3251-3260.

Penner, J. E., Eddleman, H., and Novakov, T. (1993) Atmospheric Environment 27A 1277-1295.

Siddiqi, T. A., Streets, D. G., Wu, Z. and He, J. (1994) National Response Strategy for Global Climate Change: People’s Republic of China, Asian Development Bank, Report No. TA 1690-PRC.

Sinton, J. E., ed. (1996) China Energy Databook, Report No. LBL-32822 Rev. 4, Berkeley, CA.

Streets, D. G. and Waldhoff, S. T. (1998) Energy—The International Journal, in press.

Streets, D. G., Waldhoff, S. T., and He, K. "Emissions of Carbon Monoxide in China in 1995 and 2020" Draft Final Report to NASA China-MAP Project, September 10, 1998.

Tsinghua University and others (1997) "China’s Strategies for Controlling Motor Vehicle Emissions," Report to the World Bank under China Environmental Technical Assistance Project B-9-3.

Turco, R. P., Toon, O. B., Whitten, R. C., Pollack, J. B., and Hamill, P. in "Precipitation Scavenging, Dry Deposition, and Resuspension," Pruppacher, Semonin, and Slinn, eds. (1983).

U.S. Department of Energy (1990) Annual Energy Outlook, DOE Report DOE/EIA-0383(90), Washington, D.C.,.

U.S. Environmental Protection Agency (1982) Compilation of Air Pollutant Emission Factors, 3rd Ed., EPA Report AP-42, Research Triangle Park, NC.

U.S. Environmental Protection Agency (1992), National Air Quality and Emissions Trends Report, 1991, EPA Report 450-R-92-001, Research Triangle Park, NC.

U.S. Environmental Protection Agency (1995) AP-42, http://www.epa.gov.

U.S. Environmental Protection Agency (1997) National Air Pollutant Emission Trends, 1900-1996, EPA Report EPA-454/R-97-011, Research Triangle Park, NC.

U.S. Environmental Protection Agency Working Group (1996) National Air Pollutant Emission Trends, 1900-1995, Office of Air Quality Planning and Standards, Research Triangle Park, NC.

Wolff, G. T. (1981) Journal of the Air Pollution Control Association 31 935-939.

The World Bank (1997) China 2020: At China’s Table, Washington D.C.

Xie, S. (1991) Energy and Environment in China, Presented at Energy and Environment in the Asia-Pacific Region, East-West Center, Honolulu, Hawaii.

Zhai, R. (1992) Electric Power Industry and Environment Protection in China, Presented at US-China Conference on Energy, Environment, and Market Mechanisms, Seattle, Washington and San Francisco, California.