A high-resolution ammonia emission inventory in China


  • Xin Huang,

    1. State Key Joint Laboratory of Environmental Simulation and Pollution Control, Department of Environmental Science, Peking University, Beijing, China
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  • Yu Song,

    1. State Key Joint Laboratory of Environmental Simulation and Pollution Control, Department of Environmental Science, Peking University, Beijing, China
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  • Mengmeng Li,

    1. State Key Joint Laboratory of Environmental Simulation and Pollution Control, Department of Environmental Science, Peking University, Beijing, China
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  • Jianfeng Li,

    1. State Key Joint Laboratory of Environmental Simulation and Pollution Control, Department of Environmental Science, Peking University, Beijing, China
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  • Qing Huo,

    1. State Key Joint Laboratory of Environmental Simulation and Pollution Control, Department of Environmental Science, Peking University, Beijing, China
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  • Xuhui Cai,

    1. State Key Joint Laboratory of Environmental Simulation and Pollution Control, Department of Environmental Science, Peking University, Beijing, China
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  • Tong Zhu,

    1. State Key Joint Laboratory of Environmental Simulation and Pollution Control, Department of Environmental Science, Peking University, Beijing, China
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  • Min Hu,

    1. State Key Joint Laboratory of Environmental Simulation and Pollution Control, Department of Environmental Science, Peking University, Beijing, China
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  • Hongsheng Zhang

    1. Laboratory for Atmosphere-Ocean Studies, Department of Atmospheric and Oceanic Science, School of Physics, Peking University, Beijing, China
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[1] The existence of gas-phase ammonia (NH3) in the atmosphere and its interaction with other trace chemical species could have a substantial impact on tropospheric chemistry and global climate change. China is a large agricultural country with an enormous animal population, tremendous nitrogen fertilizer consumption and, consequently, a large emission of NH3. Despite the importance of NH3 in the global nitrogen (N) cycle, considerable inaccuracies and uncertainty exist regarding its emission in China. In this study, a comprehensive NH3 emission inventory was compiled for China on a 1 km × 1 km grid, which is suitable for input to atmospheric models. We attempted to estimate NH3 emissions accurately by taking into consideration as many native experiment results as possible and parameterizing the emission factors (EFs) by the ambient temperature, soil acidity and other factors. The total NH3emission in China was approximately 9.8 Tg in 2006. The emission sources considered included livestock excreta (5.3 Tg), fertilizer application (3.2 Tg), agricultural soil (0.2 Tg), nitrogen-fixing plants (0.05 Tg), crop residue compost (0.3 Tg), biomass burning (0.1 Tg), urine from rural populations (0.2 Tg), chemical industry (0.2 Tg), waste disposal (0.1 Tg) and traffic (0.1 Tg). The regions with the highest emission rates are located in Central and Southwest China. Seasonally, the peak ammonia emissions occur in spring and summer.

1. Introduction

[2] Ammonia (NH3) is a very important alkaline constituent in the atmosphere, and it has a wide variety of impacts. A large part of atmospheric aerosols consist of sulfate neutralized to various extents by NH3. NH3 is a key precursor to the neutralization of HNO3 and H2SO4 in the atmosphere to form the secondary inorganic aerosols (NH4)2SO4, NH4HSO4 and NH4NO3, which contribute to ambient particulate matter [Seinfeld and Pandis, 2006], as shown in following chemical reaction equations:

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Measurements have indicated that the mass of secondary sulfate, nitrate and ammonium could account for 25–60% of the total PM2.5 (fine particles in the ambient air, 2.5 μm or less in aerodynamic diameter) mass in China [He et al., 2001; Niu et al., 2006; Fang et al., 2009]. Results from recent studies show that these fine particles can be deposited deep in the lungs, which may lead to increased morbidity [Dockery et al., 1993]. Fine particles also can result in regional visibility degradation by light scattering [Ye et al., 2011]. Global climate change due to ambient particulate matter has gained increasing attention in recent years. These aerosol particles have a significant effect on radiative forcing by both direct scattering and the absorption of solar radiation and modification of the shortwave reflective properties of clouds [Charlson et al., 1992; Martin et al., 2004], thereby exerting a cooling effect on the planet. In addition, NH3, as an important alkaline gas in the atmosphere, can react readily with acidic substances and neutralize the acidity of precipitation. Gaseous ammonia and ammonium compounds in particles are eventually deposited to the soil or water bodies. NH4+ deposited to the soil may either be taken up by plant roots or be subjected to nitrification yielding NO3, as shown in following equations. Both processes result in proton release into the soil [Krupa, 2003]. Consequently, ammonia emission is also known as one of the causes of soil acidification, eutrophication and even the perturbation of ecosystems [Van Breemen et al., 1983; Pearson and Stewart, 1993].

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The most important ammonia emitters are livestock waste and fertilizer, which jointly contribute over 57% globally and over 80% in Asia [Zhao and Wang, 1994; Bouwman et al., 1997; Streets et al., 2003]. China is a country with dense rural populations, and it has the world's top-ranking agricultural production. Both of these factors contribute significantly to the global ammonia emissions budget. Previous researchers have demonstrated that approximately 14 Tg NH3is released in China annually, making it one of the most important emitters, with a contribution of 20% to the global budget and 55% of Asian emissions. What is worth mentioning, global ammonia column was mapped based on the observation of IASI sensor (Infrared Atmospheric Sounding Interferometer) onboard EUMETSAT's Meteorological Operational Satellite (MetOp-A), which also demonstrated that China is an important emitter [Clarisse et al., 2009].

[3] There are several ammonia inventories for China [Zhao and Wang, 1994; Olivier et al., 1998; Xing and Zhu, 2000; Streets et al., 2003; Ohara et al., 2007]. However, these inventories have weaknesses. First, they often use uniform and European-based EFs for the entire country. Ammonia volatilization, especially from nitrogen fertilizer application and animal waste management systems, depend strongly on soil acidity and ambient temperature [Kirk and Nye, 1991; European Environment Agency (EEA), 2009]. The EFs should be parameterized by such factors. Second, not all ammonia emission sources are included during estimation, such as vehicle exhaust and wastewater treatment. This omission could introduce inaccuracies in the estimation of total emissions. Third, the prevailing inventories often have coarse temporal and spatial resolutions. Agricultural activities such as fertilization and crop burning exhibit seasonal variations due to planting and harvesting practices. Ammonia volatilization also strongly depends on the temperature, which fluctuates seasonally. Hence, a temporal resolution of one year tends toward overestimation in winter and underestimation in summer. Spatially, the existing inventories are mostly compiled at coarse spatial resolutions, such as at the provincial level [Wang et al., 1997], at 1° × 1° [Bouwman et al., 1997] and at a 30 min × 30 min grid [Streets et al., 2003]. For a modeling study of the NH3 cycle and air quality simulation, a more detailed inventory is preferred.

[4] A comparison of measured and predicted NH4NO3 concentrations also indicates that the existing inventories might be overestimated by approximately 20–75% in East Asia [Kim et al., 2006]. Therefore, an accurate ammonia emission inventory is needed to reflect various sources with a fine spatial and temporal resolution for China, especially for air environmental quality simulation. In this study, a comprehensive ammonia inventory for the year 2006 is developed. The sources of ammonia considered include the following: (1) farmland ecosystem (soil and nitrogen-fixing plant emission, fertilizer application and crop residue compost), (2) livestock wastes (free-range, intensive and grazing rearing systems), (3) biomass burning (forest and grass fires, crop residue burning and fuel wood combustion), (4) excrement of rural populations; 5) chemical industry (ammonia synthesis and nitrogen fertilizer production), (6) waste disposal (wastewater and solid waste treatment), and (7) traffic sources. The activity data are from province-specific statistical data sets, with the exception of a MODIS burned area product (MCD45A1), which was used in the assessment of biomass burning emissions. The EFs are characterized by seasons and locations, in which the ammonia volatilization from nitrogen fertilizer and animal agriculture is parameterized by the ambient temperature, soil acidity and other crucial influential factors.

2. Data and Methods

[5] Ammonia emission is calculated as a product of activity data and corresponding EFs. The important sources and its activity data are listed in Table 1, more detailed information could be found in Table S1 in the auxiliary material.

Table 1. Activity Data for Various Sources
Activity Data Set and Its DescriptionAmount
Synthetic fertilizer application (synthetic fertilizer consumption)22.4 Tg
Agriculture soil (cultivated areas)1.2 × 108 hectare
Nitrogen-fixing crop (cultivated areas)2.1 × 107 hectare
Compost of crop residues (mass composted)235 Tg
Livestock wastes (animal population)1.2 × 1010
Biomass burning (burned biomass or burned area)411 Tg burned biomass; 2118 km2 burned area
Excrement of rural population (rural population using tatty latrine)2.6 × 108
Chemical industry (production)49 Tg synthetic ammonia; 39 Tg N nitrogen fertilizers
Waste disposal (waste amount)2.0 × 1010 m3 wastewater; 78 Tg solid waste
Traffic sources (vehicle population)1.6 × 108

2.1. Farmland Ecosystem

[6] Farmland ecosystems are an important source of ammonia emission. These emitters include synthetic fertilizer application, agricultural soil, nitrogen-fixing plants and the compost of crop residues.

2.1.1. Synthetic Fertilizer Application

[7] In China, a large agricultural country with intensive agricultural areas and conventional cultivation practices, N fertilizers are applied extensively, with more than 20 Tg of N fertilizers applied every year [National Bureau of Statistics of China (NBSC), 2008a]. Recent studies have indicated that 10–30% of applied N is lost though NH3 volatilization [Ju et al., 2009]. NH3 volatilization depends substantially upon fertilizer types. In China, urea and ammonium bicarbonate (ABC) are the two dominant N fertilizers, followed by ammonium nitrate (AN) and ammonium sulfate (AS). Urea and ABC accounted for approximately 69% and 26%, respectively, of the total chemical N fertilizer consumption in 2006 [NBSC, 2007c, 2007e]. Hence, the N fertilizers considered in our study are classified into five categories: urea, ABC, AN, AS and other (including calcium ammonium nitrate and ammonium phosphates). The loss rate strongly depends on numerous factors, including soil properties (pH, calcium content, water content, buffer capacity and porosity), meteorological conditions (temperature, wind speed and precipitation), time of application in relation to a crop canopy and the method of application [Bouwman et al., 1997].

[8] In this study, ammonia emissions were estimated by multiplying consumption and gridded (1 km × 1 km) condition-specific EFs for five kinds of fertilizer as well as 16 major plants. The rates of fertilizer consumption were obtained as the product of cultivated area and the application rate of plants. The plants to which considerable amounts of N fertilizers are applied include early rice, semi-late rice, late rice, non-glutinous rice, wheat, maize, bean, potato, peanut, oil crop, cotton, beet, sugarcane, tobacco, vegetables and fruits [NBSC, 2007c, 2007e].

[9] The type of fertilizer, the soil properties, the fertilization method, the application rate and temperature were introduced as parameters to develop EFs for specific conditions. For each kind of fertilizer, a reference emission factor (EF0), referring to existing native measurements, was adjusted by soil pH, application rate, temperature and fertilization method to develop a specific EF for a given scenario, as shown in following equation:

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where EFi is the emission factor for a specific condition; EF0i is the reference emission factor for a type i N fertilizer; CFpH is the correction factor for different soil acidity; CFrate is the correction factor for different application rates; CFT is the correction factor for diverse temperatures; and CFmethodis the correction factor for the fertilization method, including basal dressing and top dressing. The monthly and gridded EFs were developed considering all the factors. Ammonia emissions could then be calculated as the product of monthly fertilizer consumption and corresponding condition-specific EFs. Fertilizer Type

[10] The potential ammonia emission from each fertilizer type often varies widely. ABC is a highly volatile compound that is extensively used in China. The ammonia loss from the direct use of ABC could be up to 30–40% of the applied N [Zhu et al., 1989]. Urea, which is less volatile than ABC, is first converted to ammonium bicarbonate by the enzyme urease, which takes approximately 2–3 days in agricultural soil. Other N fertilizers, such as AN and AS, have a much lower ammonia emission potential (<10% of applied N) [Sutton et al., 1995; Van der Hoek, 1998]. The EF0 for urea and ABC were based on experiments carried out in Henan and Jiangsu Province through the micrometeorological method [Cai et al., 1986; Zhu et al., 1989]. The EF0for other less prevalent fertilizers refers to the up-to-date and reliable EFs provided by the European Environment Agency [EEA, 2009], as shown in Table 2.

Table 2. EFs, Expressed as Percentage of Volatilized NH3-N From Applied Fertilizer-N, and Correction Coefficients for Different N Fertilizer Categories
Fertilizer CategoriesMeasured EFsCorrection Coefficients
Acid SoilAlkaline SoilCFrateaCFTbCFmethodc
Others0.8f0.8f1.180.010.32 Soil Property

[11] The volatilization of ammonia greatly depends on the soil properties, including bulk density, volumetric water content, CEC, soil acidity and urease activity. The results from the analysis of covariance demonstrate that the volatilization is most strongly correlated with pH [Li and Ma, 1993; Corstanje et al., 2008]. Several publications have shown the disparity of ammonia volatilization from urea and ABC in different soils [Cai et al., 1986; Zhu et al., 1989]. For fertilizers such as AS, which form sparingly soluble salts with calcium and thus increase the dissolution of calcium carbonate, the emission rate will be greater in the alkaline soil, whereas for AN, the emission rate is nearly unaffected by the soil pH. Several studies support that ammonia loss progressively increases with an increase in soil pH, and linear regression was used to yield specific EFs in some studies [Whitehead and Raistrick, 1991; Bouwman et al., 2002; Fan et al., 2005; Jones et al., 2007]. In our study, linear interpolation method was used to specify EFs by assuming that a linear relationship between ammonia volatilization and soil pH. A soil pH with a spatial resolution of 1 km in China was provided by the Harmonized World Soil Database (available from http://www.iiasa.ac.at/Research/LUC/External-World-soil-database/HTML/). Application Rate

[12] The application rate was used as a modifier for EFs for various types of fertilizers. Ju et al. [2009] reported an increase in ammonia emission with increasing application rate. Experiments on the relationship between application rate and ammonia volatilization are limited. Correction was done only when application rate is relatively high. A value of 200 kg N ha−1 was considered the high rate, and a value of 1.18 for CFrate was used for a province with a high application rate [Li et al., 2002; Song et al., 2004; Fan et al., 2006]. For regions where the rate was lower than 200 kg N ha−1, no correction was made. Temperature

[13] As the partial pressure of NH3 in solution increases exponentially with temperature, we would expect a strong relationship between temperature and ammonia emission. Several studies have noted an increase in emission from N fertilizers with increasing air temperature. EEA [2009] have summarized empirical volatilization rates as functions of spring air temperature for various fertilizers except ABC. We derived a relationship of emission rate and temperature for ABC based on the available native experimental data [Lu et al., 1980], as listed in Table 2. Monthly average temperatures for a 1 km grid were developed on the basis of NCEP (National Centers for Environmental Prediction) final analysis data set. Fertilization Method

[14] The method of fertilization is another important factor affecting the volatilization of ammonia. We considered two common fertilization methods: basal dressing and top dressing. For the same fertilizer, the EF for basal dressing is approximately 32% of that for top dressing [Lu et al., 1980; Qu, 1980; Fillery et al., 1986; Zhang and Zhu, 1992; Li and Ma, 1993; Cai et al., 2002]. The application time and amount of fertilizer used as basal dressing and top dressing are referenced from the Chinese planting information network (http://www.zzys.gov.cn/) and census data and investigation results [NBSC, 2007a, 2007e; Wang et al., 2008].

2.1.2. Agriculture Soil

[15] Many organisms in soils are involved in the decomposition of organic matter excrete NH3 directly or N compounds which could be readily hydrolyzed to NH3. The NH3 fluxes from soil are related closely to the biological activity in the soil and other meteorological factors [Bouwman et al., 1997]. There is limited information on ammonia emissions from cultures without fertilizers in China. Background emission levels of ammonia from agricultural soil were reviewed by Yan et al. [2003], and an EF value of 1–2 kg NH3-N ha−1 yr−1was recommended as the loss of ammonia in temperate regions. The province-level area of arable land refers to the census data [NBSC, 2007e]. The provincial ammonia emission from agricultural soil was calculated as a product of arable land and EF.

2.1.3. Nitrogen-Fixing Crop

[16] Nitrogen-fixing crops, even if not given fertilizer-N, have been estimated to fix as much N as fertilized crops. Thus, emissions of NH3may be expected to be similar to those from fertilized agricultural crops. The most widely cultivated nitrogen-fixing plants in China are beans (mostly soybean, accounting 77% cultivated area), peanuts and green manure. The ammonia emissions may be approximately estimated as follows:

display math

where ENH3 is the emitted NH3from N-fixing crops;Aiis the area covered by N-fixing cropsi; NFi is the nitrogen fixation rate of type iN-fixing crops; andEFNH3 is the ammonia EF for crops i.

[17] The province-level cultivated areas of these three nitrogen-fixing crop groups were provided by the National Bureau of Statistics of China [NBSC, 2006, 2007e]. The N fixation rates of these crops are 105, 120 and 130 kg N/ha/yr, respectively [Kuenzler and Craig, 1986; Yan et al., 2003]. Data on NH3 fluxes over legume crops are sparse, but the 0.01 kg NH3/kg N suggested by EEA was used in our estimation [EEA, 2006].

2.1.4. Compost of Crop Residues

[18] Composting refers to the use of both aerobic and anaerobic microbial processes to degrade waste materials for beneficial reuse. In the rural China, crop straws and other agricultural residues are often composted to add soil nutrients. A large amount of ammonia is released during this biological process. The ammonia emissions from this source are assessed by multiplying the amount and the corresponding EFs. Provincial crop residue production can be estimated based on crop production and production-to-residue ratios [Lal, 2005; NBSC, 2007e]. The portion of crop residues for composting refers to an investigation covering most of the provinces in China [Gao and Ma, 2002]. The ammonia EF for the composting operation is 0.32 kg/ton [Stephen et al., 2004].

2.2. Livestock Wastes

[19] Nitrogen in animal excrement in the form of urea can rapidly hydrolyze to form ammonium carbonate and then volatilize as gaseous ammonia. Early inventories of ammonia emission from livestock wastes were calculated as the product of livestock numbers and annual EFs per animal [Buijsman et al., 1987]. More recent inventories have improved the single EFs per animal with specific EFs for the various phases of manure management, including animal housing, manure storage, manure spreading and the grazing stage [Pain et al., 1998; Zhang et al., 2010]. Recently, the mass-flow approach has been widely used at the national and global scale [Webb et al., 2006; Beusen et al., 2008]. Ammonia is emitted from a pool of ammoniacal nitrogen (TAN), which is not increased during manure management, but it is depleted by the losses though ammonia volatilization, leaching and other pathways during different stages. This approach enables the rapid estimation of the consequences of abatement at the upstream stage on the ammonia losses at the downstream stage, which is more accurate than the previous methods [Webb and Misselbrook, 2004].

[20] Based on the mass-flow methodology in the EEA's inventory guidebook [EEA, 2009], we analyzed the transformation and migration of nitrogen in the animal husbandry (Figure 1). The ammonia emissions from each stage of livestock manure management are affected by many factors, such as species, gender, age, bodyweight, nitrogen content of the feed, housing structure, manure storage system, spreading technique, time spent outside or inside and meteorological conditions [Zhang et al., 2010]. Although numerous experiments have been performed on ammonia emission during livestock production, few measurements have been recorded in China. Consequently, China-specific EFs must be recalculated on the basis of the native practice of livestock production and environmental conditions.

Figure 1.

N flows in the manure management system and ammonia volatilization (TAN: ammoniacal nitrogen).

[21] In China, animals are raised in several completely different systems, including free-range, intensive and grazing system. Ammonia emissions were calculated separately for these three systems. The typical animal categories in China are listed inTable 3. Gender and age are also taken into account in the definition of livestock subcategories. The available statistical data, official surveys (http://www.caaa.cn/) and industry reports for husbandry were collected to obtain the number of each livestock class [Editorial Office of China Animal Industry Yearbook (EOCAIY), 2007]. The annual excretions of TAN by each livestock class were derived based on the rearing period [China Agricultural University (CAU), 1997; Zhou et al., 2010], the daily amount of excrement [CAU, 1997; Ministry of Environmental Protection of People's Republic of China (MEPPRC), 2001; K. Wang et al., 2009], the nitrogen content and the percent of TAN [CAU, 1978; Webb and Misselbrook, 2004; EEA, 2009] (Table 3). In the following stepwise procedure of manure management, slurry and solid manures were calculated separately. Monthly emissions estimation was conducted based on temperature-depended EFs (Table S2). We assumed that the animal population is same during every month of the whole year since the monthly fluctuation of meat productions is very small (http://www.caaa.cn/). Monthly specific EFs were developed using provincial monthly average temperature from NCEP final analysis data set.

Table 3. Parameters Used in Estimates of Annual TAN Excretion per Animal for Each Livestock Class
Livestock ClassPerioda (day)Excrementb (kg/day/cap)Nitrogen Contentsc (%)TAN Contentd (%)
Beef <1 year3655.007.00.900.3860
Beef >1 year36510.0020.00.900.3860
Dairy cows <1 year3655.007.00.900.3860
Dairy cows >1 year36519.0040.00.900.3860
Goat <1 year3650.661.51.350.7560
Goat >1 year3650.752.61.350.7550
Sheep <1 year3650.661.51.350.7560
Sheep >1 year3650.752.61.350.7550
Fattening pig753.201.50.400.3470
Laying hen365-0.12-1.6370
Laying duck365-0.13-1.1070
Laying goose365-0.13-0.5570
Meat duck55-0.10-1.1070
Meat goose70-0.10-0.5570

2.2.1. Housing

[22] Housing emission is related to the portion of excreta deposited inside buildings. Various livestock classes have different fractions of time spent in buildings (Table S2). Emissions from different locations were calculated by multiplying the amount of TAN and EFs. Ammonia volatilization from the housing stage depends on factors such as housing conditions, humidity, temperature and animal waste handling practices [Anderson et al., 2003]. In our study, we characterized the EFs by flooring type, waste handling practice and temperature variations. Deep litter flooring is commonly used in free-range and grazing systems, but solid flooring is widespread in intensive large-scale farming. The differences between these methods are summarized byHutchings et al. [2001]. The handling practice of waste in the housing stage has been surveyed in some provinces in China [Liu et al., 2008], and the EFs for slurry and solid were derived from previous studies [Webb and Misselbrook, 2004; Department for Environment, Food and Rural Affairs, 2005; EEA, 2009]. The temperature adjustment of the volatilization rate was accomplished though the diverse EFs at different temperature intervals [Mannebeck and Oldenburg, 1991; Koerkamp et al., 1998].

2.2.2. Storage

[23] According to the actual treatment of housing waste in China, usually no treatment is given for the excrement deposited outside buildings, and those excreted inside are usually settled in two ways: biogas generation and composting. The handling practices of manure storage are different in different parts of China. According to Liu et al.'s investigation, approximately 10% more manure is used to produce biogas in the Southern provinces in the free-range rearing system [Liu et al., 2008]. Unfortunately, there is limited data on the fraction of different storage methods in the intensive rearing system, for which half of the manure was assumed to be handled as slurry. According to the EEA's guidelines [EEA, 2009], the immobilization of TAN, the emission of other nitrogenous gases (N2O, NO and N2) and the leaching loss of nitrogen are other notable pathways of nitrogen transformation. The parameters involved in these processes are all listed in Table S2.

2.2.3. Spreading

[24] Manure is extensively applied to cropland as basal fertilizer in China [Liu et al., 2008; Li et al., 2009]. The estimate refers to the approach recommended by EEA but is improved by adding the parameter Xfeed, which is defined as the portion of manure used for feed production. The reason for this addition is that large-scale henneries and hoggeries often use the manure as feed after treatments. The EFs for different livestock classes were derived from existing experiments and expert judgments (Table S2) [Webb and Misselbrook, 2004; EEA, 2009].

2.3. Miscellaneous Sources

2.3.1. Biomass Burning

[25] Biomass burning could release significant amounts of ammonia [Andreae and Merlet, 2001]. The types of biomass burning in this study include forest and grass fires, crop residue in-field burning, domestic combustion and fuelwood combustion. The emissions from crop residues and fuelwood were estimated based on the statistical data [NBSC, 2007e, 2008b]. The emissions from forest and grassland fires were estimated by combining a newly available MODIS (Moderate Emission Resolution Imaging Spectroradiometer) burned area product (MCD45A1) with the combustion completeness considering the impact of vegetation moisture content [Song et al., 2010]. The EFs for each type of biomass are summarized in Table 4 [Li et al., 2007].

Table 4. Ammonia EFs of Miscellaneous Anthropogenic Sources
Biomass burning   
   Forest fires2.9g NH3/kgSong et al. [2010]
   Grassland fires0.7g NH3/kgSong et al. [2010]
   Crop residues burning0.37(wheat) 0.68(maize) 0.52(others)g NH3/kgLi et al. [2007]
   Fuelwood combustion1.3g NH3/kgAndreae and Merlet [2001]
Human excrement0.787kg NH3/yr/capBuijsman et al. [1987], Möller and Schieferdecker [1989], Yin et al. [2010]
Chemical industry   
   Synthetic ammonia0.01kg NH3/tonEEA [2009]
   Nitrogen fertilizers production5.0kg NH3/tonAsman [1992]
Waste disposal   
   Wastewater0.003g NH3/m3Yin et al. [2010]
   Landfill0.560kg NH3/tonYin et al. [2010]
   Compost1.275kg NH3/tonStephen et al. [2004]
   Incineration0.210kg NH3/tonYin et al. [2010]
   Light-duty gasoline vehicles0.026g NH3/kmStephen et al. [2004]
   Heavy-duty gasoline vehicles0.028g NH3/kmStephen et al. [2004]
   Light-duty diesel vehicles0.004g NH3/kmStephen et al. [2004]
   Heavy-duty diesel vehicles0.017g NH3/kmStephen et al. [2004]
   Motorcycles0.007g NH3/kmStephen et al. [2004]

2.3.2. Excrement of Rural Populations

[26] China's rural areas are widely distributed with tatty latrines (roughly constructed latrines without any flushing treatment). Feces in these tatty latrines accumulate in the open air and are subsequently used as fertilizer, which is another important ammonia emitter. The amount of human waste used for spreading was obtained based on the daily excrement of children and adults, the rural population and the fraction of tatty latrines in each province [NBSC, 2007a, 2007d]. The EFs were the average of several publications [Buijsman et al., 1987; Möller and Schieferdecker, 1989; Yin et al., 2010], which are listed in Table 4.

2.3.3. Chemical Industry

[27] The chemical industry is also a gaseous ammonia source, particularly the production of synthetic ammonia and nitrogen fertilizers. These emissions were estimated by multiplying the production and corresponding EFs. The available census data for the chemical industry were provided by the national statistical bureau [NBSC, 2007b]. The EFs were provided by EEA and EPA reports [Stephen et al., 2004; EEA, 2009].

2.3.4. Waste Disposal

[28] Ammonia emissions from wastewater treatment plants and refuse processing plants were also included; these emissions should not be neglected, especially in urban areas [Anderson et al., 2003]. The waste disposal types considered were wastewater treatment, landfill, compost and incineration of solid waste. These data were obtained from the available yearbook [Statistics and Census Service of Macao SAR Government, 2007; NBSC, 2007d] (http://www.dgbas.gov.tw), and the corresponding EFs were derived from existing publications [Stephen et al., 2004; Yin et al., 2010].

2.3.5. Traffic Sources

[29] There is evidence that traffic may also be an important source of ammonia in urban areas [Ianniello et al., 2010; Vega et al., 2010; Meng et al., 2011]. The vehicles taken into consideration include light-duty and heavy-duty gasoline vehicles, light-duty and heavy-duty diesel vehicles and motorcycles. The emissions were calculated as the product of the number of operating vehicles [China's Association of Automobile Manufacturers, 2007], mileage [Cai and Xie, 2007] and ammonia emissions per kilometer for each vehicle category [Stephen et al., 2004].

2.4. Gridded and Monthly Emissions

[30] The emissions were compiled at a 1 km × 1 km spatial resolution by the various types. The fertilization and wild fire emissions were already calculated based on gridded activity data and EFs, so no allocation was needed. The other emissions were presented at the provincial level first and then allocated using land cover, rural population, and other proxies. More specifically, emissions from farmland ecosystems, with the exception of fertilization, were allocated by land cover class. Emissions from free-range and intensive livestock production were redistributed based on rural population density, and the grassland land cover class was used to allocate grazing emission. Emission from the field burning of crop residue was gridded by using the MODIS thermal anomalies/fire products (MOD14A2 and MYD14A2). Emissions from biofuel combustion and human waste were also regridded on the basis of rural population distribution. The remaining sources were equally distributed in each grid. After the spatial allocation of all emissions was completed, an aggregation was conducted to sum up all emissions.

[31] The monthly emission estimation for fertilizer application and livestock waste were elaborated in detail in sections 2.1 and 2.2. Efforts were also devoted to temporally allocate ammonia emission from biomass burning. MCD45A1 (monthly burned area product), MOD14A2 and MYD14A2 products (8-day thermal anomalies/fire products) were utilized to ascertain the timing of different kind of biomes, more details could be found in two existing studies [Song et al., 2010; Huang et al., 2012]. Emissions from other minor sources were equally divided into 12 months.

3. Result and Discussion

3.1. Ammonia Source Apportionment

[32] The ammonia emissions at the individual province level in 2006 from different sources are presented in Table 5. As shown, the total ammonia emission was 9.8 Tg, contributing approximately 15% and 35% to global and Asian emissions, respectively, based on the previous assessment at the global and Asian scale [Bouwman et al., 1997; Streets et al., 2003]. The ammonia emission from China is much higher than elsewhere, exceeding Europe (5 Tg) and the USA (4 Tg) by approximately 96% and 145%, respectively [Bouwman et al., 1997; Pain et al., 1998; Anderson et al., 2003].

Table 5. Regional-Specific Ammonia Emissions (Gg) From Different Sources in 2006
 FertilizerAgricultural SoilN-Fixing CropCompostLivestockBiomass BurningHuman ExcrementChemical IndustryWaste DisposalTraffic
Inner Mongolia79.
Hong Kong0.

[33] The most important contributor is livestock manure management (5.3 Tg), accounting for approximately 54% of the total budget. Next is the fertilizer application (3.2 Tg), which is responsible for 33% of emissions. The less important sources at the national scale are the following: soil 0.2 Tg (2%), N-fixing crops 0.05 Tg (1%), compost of crop residues 0.3 Tg (3%), biomass burning 0.1 Tg (1%), human excrement 0.2 Tg (2%), the chemical industry 0.2 Tg (2%), waste disposal 0.1 Tg (1%) and traffic 0.1 Tg (1%).

[34] The emission characteristics in China reflect its unique agricultural structure and farming practice. For Europe and the USA, livestock overwhelmingly dominate the inventory. In China, however, both fertilizer and livestock play significant roles. The types of fertilizer used in China are quite different from those used in developed countries. Urea and ABC, which are highly volatile compounds, are the two most widely used N fertilizers in China. In contrast, less volatile fertilizers such as anhydrous ammonia and ammonium nitrate are much more popular in North America and Europe (http://www.fertilizer.org/ifa/ifadata). The incredibly large consumption and high volatility of urea and ABC make the ammonia emission from fertilizer application as important as that from the livestock in China. At the same time, the feature of husbandry in China is also distinct, being characterized by multiple rearing systems. The most important is the free-range system, accounting for 65% of the total. In most rural areas, livestock are raised in a traditional way; that is, a small number of animals are fed by individual families and are allowed to roam freely instead of being contained in any manner. Next is intensive rearing (approximately 30%), which refers to the process of raising livestock in confinement at high stocking density; in this process, the farm operates as a factory. This type of farming is extensive in broiler (73%), dairy cow (57%) and pig (41%) production [EOCAIY, 2007]. Grazing, a relatively less important system, only dominates in the northern and western parts of China. Consequently, the free-range system in rural areas is the greatest contributor, accounting for 67% of the emissions of livestock. The largest NH3 emission livestock category is cattle, which emitted 1.9 Tg NH3 per year, followed by laying hens and pigs, which both emitted 0.7 Tg NH3 per year.

[35] Other sources were not as notable for the whole country, as shown in Figure 2. However, some of these sources could become significant at the provincial scale. For instance, the contribution of biomass burning is a bit higher in Heilongjiang and Jilin provinces (3%), where forest fires frequently occur. Human waste is responsible for greater portions in Guizhou and Shaanxi provinces because of the larger rural populations and the ubiquitous tatty latrines (sanitary latrines only account for only 30% and 35%, respectively). In Shanxi and Hainan provinces, 8% and 7% of ammonia emissions are caused by chemical industry production, respectively, which is higher than the average of 2% for the whole country. Both waste disposal and traffic source are far more prevalent in urban areas such as Beijing and Shanghai, with, on average, 12 and 7 times greater contributions than elsewhere, respectively.

Figure 2.

Source contributions (%) to ammonia emissions in China.

3.2. Spatial and Temporal Distribution

[36] The spatial pattern of ammonia emission is illustrated in Figure 3. It is easy to see the high emission rates in Henan, Shandong, Hebei, and Jiangsu provinces and in Eastern Sichuan. The emission contributions from Henan, Shandong, Hebei and Sichuan provinces, all of which have high agriculture production, accounted for 12.7%, 9.2%, 8.2% and 6.5%, respectively. The highest emission rates of NH3 were mainly caused by fertilizer application and livestock, which jointly contributed 93%, 88%, 91% and 87% for these four major emitters, respectively.

Figure 3.

Spatial distribution of ammonia emissions in 1 km grid cell (kg yr−1).

[37] Table 5 lists the emission amounts from fertilizer application in each province (the pattern is shown in Figure S1 in 1 km grid cells). The intensity of NH3 volatilization from synthetic N fertilizers was concentrated in Henan, Shandong, Hebei and Jiangsu province, where the crop cultivation is most developed over the whole country. The crop production in these four provinces is the greatest in China, with a total contribution of 33% of China's crop production, and they consumed approximately 30% of the N fertilizer for the whole country in 2006 [NBSC, 2007e]. For example, the ammonia emission rates in Henan province reached a maximum value of 7.2 kg per square kilometer of arable land. The farmers here usually adhere to fertilization during plant cultivation. The application rate in this region was incredibly high, and consequently, the nitrogen loss was tremendous [Ju et al., 2009]. According to an investigation carried out by the government [NBSC, 2007c], the application rate of N fertilizer in Henan province is 1.1–3.4 times higher than the national average for staple crops. For ABC, the gap was even greater (1.1–9.3 times). Among the various kinds of crops, sugarcane and vegetable have the highest fertilization rates (335 and 264 kg N per hectare, respectively), which are double the rates of rice and wheat and are approximately tenfold higher than those of beans and peanuts. In addition, vegetables are cultivated extensively in China, with a sowing area that is just behind cereal crops. Over 20% of ammonia loss originates from vegetable fertilization. The North China Plain, with a 25,624,000 and 5,754,000 ha area under cultivation of cereal crops and vegetables, respectively, is responsible for 43% of the NH3 emissions from fertilization in China. The cultivated land in Guangdong, Fujian, Hunan and Jiangxi provinces also shows large NH3 volatilization. Paddy fields are common in this region, with rice being the dominant crop and contributing the most emissions (approximately 40% for Jiangxi province). The smaller emitters are located mostly in western China, with a minimal amount of arable land or low use of synthetic nitrogenous fertilizers.

[38] The ammonia emissions from livestock during 2006 are also presented in Table 5 and Fig. S2. The high NH3 emission areas, with rates over 2000 kg NH3 per square kilometer, were concentrated in Henan, Shandong, Hebei and East Sichuan province. Approximately 5.3 Tg NH3was released in China due to livestock rearing, 0.6 Tg of which was emitted from Henan province, followed by Hebei (0.5 Tg), Shandong (0.5Tg) and Sichuan (0.4 Tg). These four provinces are well-known for their large animal population, providing approximately 36% of China's husbandry production [NBSC, 2007e]. Many kinds of animals are extensively bred in Henan, Hebei and Shandong provinces, including those raised for beef, dairy, pork and poultry. Among them, the greatest contributors are laying hens and beef, which contribute 26% and 24%, respectively, to the total emissions of these four provinces. Inner Mongolia, Tibet and Xinjiang province, where sheep are widely raised, also emit remarkable ammonia emissions related to sheep manure management, accounting for over 40% of emissions of these provinces. Cattle are widely raised in Northeast China and are responsible for 53% and 51% of the ammonia emissions in Liaoning and Heilongjiang provinces, respectively. Sichuan province is another large emitter, with an emission rate of 0.4 Tg NH3 per year. Cattle and pig production released 54% and 20% NH3 among all livestock emission in Sichuan province, respectively.

[39] The seasonal distribution of ammonia emissions was in general agreement with that of temperature and agricultural timing (Figure 4). Generally, the maximum emission was in summer, and nearly 42% and 28% of the fertilization and livestock emissions, respectively, occurred during this season because of the high temperature and dense fertilization. Streets et al. [2003]also showed a roughly similar seasonal pattern of ammonia emission, except for fluctuations from June to August and September to November. The disparity was attributed to the timing of different plants. Various crops were planted on the same farmland in rotation over large areas in China, particularly south part, such as the winter wheat-summer maize rotation system practiced in the North China Plain. Summer plants such as maize are usually seeded in June with the application of base fertilizer, and the topdressing fertilizer is applied two months later, and the peak emissions correspondingly occur during June and August (http://www.zzys.gov.cn/). Moreover, fertilizer is densely applied for semi-late rice in June with emissions of 0.16 Tg NH3 (42% of total emissions in June), and cotton is intensively fertilized in August (0.14 Tg, 35% of total emissions in August), which also leads to peak volatilization in June and August. The high emission rates in September and November could be attributed to the basal dressing and top dressing of wheat, with 0.11 Tg and 0.15 Tg NH3 emissions respectively. The high ammonia volatilization rate from livestock also occurs in summer. There is little variation of animal population among the different seasons. The increasing emissions during the time period from June to August were caused by larger EFs associated with the substantial increase of temperature. This tendency could be mostly attributed to emissions from laying hens, which were responsible mostly for the monthly variation because their EFs change more rapidly with temperature than other species [Mannebeck and Oldenburg, 1991]. Although other emission sources such as biomass burning also have seasonal variations, temporal disparities were not considered in our estimate due to their relatively small contributions.

Figure 4.

Monthly ammonia emissions in 2006.

3.3. Comparison With Existing Result

[40] Several inventories of ammonia emissions have been established in China for different base years in support of various activity data set and EFs. Among them, the ammonia emission inventory developed by Streets et al. [2003] (available at http://www.cgrer.uiowa.edu/EMISSION_DATA_new/) has been widely used in air quality modeling [Zhang et al., 2007; Lin et al., 2008].

[41] Our estimation was compared with the previous emission inventories listed in Table 6. As presented, the results are generally 40–50% higher than ours. These disparities are caused mainly by the emission estimations from the fertilizer volatilization. Our estimate has smaller emissions and larger spatial disparities. The fertilizer consumptions are comparable, and the discrepancy is mostly caused by the distinct EFs employed. Zhao and Wang [1994], Streets et al. [2003], Yan et al. [2003] and S. Wang et al. [2009] used uniform EFs for the whole country, and these values were based on European experiments or foreign expert judgments rather than native measurements in China. Zhang et al. [2011] spatially and temporally characterized the EFs at the county level using the National Ammonia Reduction Strategy Evaluation System (NARSES) model developed for agriculture emissions in the UK. Similarly, many parameters have been introduced based on measurements in Europe. As mentioned above, the ammonia volatilization from chemical fertilizer strongly depends on numerous environmental conditions. The soil pH values, temperatures and agricultural practices are completely different between Europe and China. Our estimation made full use of local experiments, and it adjusted the EFs for different crops under local conditions. For instance, the EFs for Urea and ABC are identical, at 15% and 30%, respectively, in Streets et al.'s [2003]paper, while in our calculation, the EFs can reach up to 35–40% for non-glutinous rice, which is top-dressed during June to July, and then decrease to lower than 5% for early rice that is basal-dressed during February to March. Consequently, the disparity between their results and the estimates made in this study are to be expected. The comparisons of province-level emissions are demonstrated inFigure 5a. The largest emitters are concentrated in the Hebei, Henan, Shandong, Jiangsu and Sichuan provinces, in accordance with previous results [Streets et al., 2003; Zhang et al., 2011]. Henan and Shandong were given more weight in our budget, with 18% and 10% of the total fertilizer emissions, while the corresponding contributions in the inventory of Streets et al. are 10% and 7%, respectively. Simultaneously, the contribution of Jiangsu Province declines from 10% in Streets et al.'s result to 8% in our estimate. This difference could be attributed to the involvement of the soil acidity and the application rate. More emissions occurred in Henan and Shandong provinces because of alkaline soil and higher application rates, which were not considered in previous studies and might cause underestimation in this area. Jiangsu has fewer emissions in our results due to the lower soil pH and the smaller application rate.

Table 6. Comparison of the Ammonia Emissions (Tg yr−1) in Our Study With Other Published Results
 Base YearTotalFertilizerHusbandryBiomass BurningOthers
This study20069.
Zhao and Wang [1994]199013.66.44.2-3.0
Yan et al. [2003]1995-4.6---
Streets et al. [2003]200013.
Ohara et al. [2007]2000--5.50.5-
S. Wang et al. [2009]200513.45.17.0-1.3
Zhang et al. [2011]2005-4.3---
Figure 5.

Comparison of provincial-level results on (a) fertilizer and (b) livestock emissions.

[42] Comparisons of emissions related to animal manure management at the national scale and the province level are indicated in Table 6 and Figure 5b. Our estimation is comparable to the results of Zhao and Wang [1994], Streets et al. [2003] and Ohara et al. [2007], at approximately 5.0 Tg/yr. The value given by S. Wang et al. [2009] was 7.0 Tg/yr. The differences among these inventories mainly resulted from the selection of EFs. For previous livestock ammonia inventories, the EFs were either presented in loss rates per capita and then multiplied by population [Zhao and Wang, 1994; Streets et al., 2003; Ohara et al., 2007] or were recalculated for distinct stages as the portion of ammonia [S. Wang et al., 2009; Zhang et al., 2010]. However, most of the previous inventories used Europe-based EFs, introducing significant inaccuracies [Streets et al., 2003]. Our research considered three different rearing systems with Chinese characteristics. The EFs were also disaggregated both spatially and temporally according to the climate conditions and local manure treatment practices. Region-specific EFs made it possible to accurately represent the emission distribution. Spatially, Henan and Hebei provinces, two biggest emitters, released 0.6 and 0.5 Tg NH3 in our estimate, respectively, comparable with the results given by Zhang et al. [2010]; the corresponding value in Streets et al.'s paper is 0.3 and 0.3 Tg, as shown in Figure 5b. Higher emissions in our study were caused by larger animal population in these two provinces in 2006 than 2000 (base year of Streets et al.'s inventory) [NBSC, 2007a]. In contrast, Sichuan province plays a much more important role in Streets et al.'s inventory due to the high assessment of N excrement from pig production. N excrement is age-dependent, especially for pigs [Liu et al., 2008]. Sichuan, with a large amount of pig production, might be overemphasized in Streets et al.'s budget due to the neglect of disparities among excrements in different growth phases.

[43] Compared with the previous ammonia inventories, many more sources were considered, including biomass burning and waste disposal. Although the total emissions of these miscellaneous sources were relatively small, some play an important role at the regional scale. For instance, waste disposal contributed almost 32% of the ammonia emission in Shanghai. Therefore, it is not acceptable to neglect these sources during calculation. Most existing inventories ignored these small sources, thus resulting in inaccurate estimates of some areas.

[44] Our results are also compared with the satellite data revealing global NH3 column distribution [Clarisse et al., 2009]. Elevated columns are observed in Northeast China and the North China Plain through satellite monitoring, similar to our inventory. However, little ammonia is found in South China and Sichuan province according to satellite observation. The difference between our estimation and satellite observed distributions is first attributed to the fact that satellite monitoring suffers signal impairments caused by meteorological effects such as water vapor, clouds [Garcia et al., 2008]. In South China and Sichuan province, higher amount of rain and cloud cover might introduce some uncertainties. Second, concentration distribution could not be always consistent with emission pattern. As mentioned in section 1, ammonia is very reactive and consequently short-lived in the atmosphere. It could react with many acid materials like SO2. Sichuan and Guizhou province are highly polluted areas with large amounts of SO2 emission [Zhang et al., 2009]. Simultaneously, both temperature and humidity are high there, which further accelerate NH3 depletion. The disparities between our estimation and satellite results are to be expected.

3.4. Uncertainty

[45] Emission uncertainty is associated with activity data and EFs. Activity data were obtained through various first-hand investigations and from statistical information. Local conditions and situations were also considered in the process of EF correction. Nevertheless, there are still large uncertainties, particularly for fertilization and livestock emissions, due to the extremely large values and the numerous parameters involved in the EF adjustment. Generally, for economic activities such as fossil fuel use and industrial and agricultural production, which are often reasonably accurate, the coefficient variation (CV) is usually 5–10%. Activities such as biofuel use, waste burning and landfilling are less well-documented [Zhao et al., 2011]. The distribution and uncertainties for activity levels for various sources are listed in Table S3 based on existing publications or assumptions made for this study. The EFs show a substantial variation according to the type of process [Olivier et al., 1998]. The EFs for NH3 emissions from biofuel combustion and chemical industry production have larger fluctuations (CV ≥ 100%), and others have smaller uncertainties, such as livestock (CV: 50%) and waste burning (CV: 50%). Further details can be found in Table S3. We ran 20 000 Monte Carlo simulations to estimate the range of fire emissions with a 95% confidence interval. The estimated emission range was 6.5–12.5 Tg/yr.

4. Conclusion

[46] We developed a comprehensive and fine emission inventory of ammonia in China by combining the statistical data set with MODIS products. EFs specified temporally and spatially were used to enhance the accuracy of this estimation. The total emissions for ammonia in China was 9.8 Tg, of which nitrogen fertilizer application and livestock manure management were the largest contributors, at 3.2 Tg and 5.3 Tg, respectively. Other sources accounted for 13% of the total, but some may play an important role in certain locations. This inventory covered the whole year of 2006, and it showed obvious seasonal variability. The lowest emissions occurred in winter, while the highest emissions were in summer. These emissions were significantly associated with temperature variations and the timing of farming practices. At the province level, Henan was the largest contributor, mostly because of the robust agricultural activity, with both crop production and animal husbandry. This work provides a detailed ammonia emission inventory for China on a 1 km × 1 km grid, and it can be used for global and regional air quality simulation. This inventory can be accessed through our website (http://phy-en.pku.edu.cn/pdf/inventory/NH3-2006.rar) or by contacting the corresponding author.


[47] The MCD45A1 burned area product was provided by the University of Maryland. The MODIS NDVI, VCF, and 8-day thermal anomalies/fire products were provided by the Land Process Distributed Active Archive Center (LPDAAC), USA. The China Land Cover product was provided by the Environmental and Ecological Science Data Center for West China, National Natural Science Foundation of China. The 1 km population distribution data set was developed by the Data Center for Resources and Environmental Sciences Chinese Academy of Sciences (RESDC). This study was funded by National Natural Science Foundation of China (40975088 and 70821140353), the Public Welfare Projects for Environmental Protection (200809018), and National Program on Key Basic Research Project of China (973) under grant 2010CB428501.