High-resolution ammonia emissions inventories in China from 1980–2012

Ammonia (NH 3 ) can interact in the atmosphere with other trace chemical which can lead to detrimental environmental consequences, such as the formation of ﬁne particulates and ultimately global climate change. China is a major agricultural country, and livestock numbers and nitrogen fertilizer use have increased drastically 5 since 1978, following the rapid economic and industrial development experienced by the country. In this study, comprehensive NH 3 emissions inventories were compiled for China for 1980–2012. In a previous study, we parameterized emissions factors (EFs) considering ambient temperature, soil acidity, and the method and rate of fertilizer application. In this study, we reﬁned these EFs by adding the e ﬀ ects of wind speed and 10 new data from ﬁeld experiments of NH 3 ﬂux in cropland in northern China. We found that total NH 3 emissions in China increased from 5.9 to 11.2 Tg from 1980 to 1996, and then decreased to 9.5 Tg in 2012. The two major contributors were livestock manure and synthetic fertilizer application, which contributed 80–90 % of the total emissions. Emissions from livestock manure rose from 2.87 Tg (1980) to 6.17 Tg (2005), and then 15 decreased to 5.0 Tg (2012); beef cattle were the largest source followed by laying hens and pigs. The remarkable downward trend in livestock emissions that occurred in 2007 was attributed to a decrease in the numbers of various livestock animals, including beef cattle, goats, and sheep. Meanwhile, emissions from The We 20 combined this condition-speciﬁc EF with that from Cai et al. (1986) and Zhu et al. (1989)


Introduction
Ammonia (NH 3 ) is an important reactive nitrogen (N) compound, and has wide impacts on both atmospheric chemistry and ecosystems. As an alkaline gas in the atmosphere, it can readily neutralize both sulfate and nitric acid to form ammonium sulfate and ammonium nitrate, which are the major constituents of secondary inorganic aerosols 5 (Behera and Sharma, 2012). Kirkby et al. (2011) found that atmospheric NH 3 could substantially accelerate the nucleation of sulfuric acid particles, thereby contributing to the formation of cloud condensation nuclei. The total mass of secondary ammonium salts accounts for 25-60 % of particulate matter less than or equal to 2.5 µm in aerodynamic diameter (PM 2.5 ) (Ianniello et al., 2011;He et al., 2001;Fang et al., 2009), and 10 large amounts of this fine PM not only cause air pollution but also have a significant effect on radiative forcing (Charlson et al., 1992;Martin et al., 2004). In addition, the sinking of NH 3 into terrestrial and aquatic ecosystems can directly or indirectly cause severe environmental issues, such as soil acidification, eutrophication of water bodies, and even a decrease in biological diversity (Matson et al., 2002;Pearson and Stew-15 art, 1993). When deposited into soils, NH 3 compounds can be converted into nitrate (NO − 3 ) through nitrification, simultaneously releasing protons into the soil, resulting in soil acidification (Krupa, 2003).
Livestock manure and synthetic fertilizer represent the most important sources of NH 3 emissions, jointly accounting for more than 57 % of global emissions and more 20 than 80 % of total emissions in Asia (Bouwman et al., 1997;Streets et al., 2003;Zhao and Wang, 1994). Previous studies have verified that China emits a considerable proportion of the total global NH 3 emissions budget due to its intensive agricultural activities (Streets et al., 2003). A major agricultural country, China has undergone rapid industrialization and urbanization since the Chinese government implemented its eco-Introduction the growth of crops, has also undergone a considerable, sustained increase. According to figures from the International Fertilizer Industry Association (Zhang et al., 2012), synthetic fertilizer production has increased 3-fold during the past three decades, from 10 million t in 1980 to 43 million t in 2012. Several factors have contributed to the dramatic rise in the use of synthetic fertilizers. First, their use grew dramatically in the 5 latter half of the 20th century in most parts of the world, as farmers increasingly expected to achieve higher crop yields. Second, N over-fertilization has been common, resulting in higher NH 3 volatilization loss, especially in the North China Plain and Taihu region (Xiong et al., 2008;Ju et al., 2009). In addition, due to farmers' increasing labor costs and income from off-farm activities, traditional farmyard manure has been gradually eliminated in much of China and replaced by synthetic fertilizers (Ma et al., 2009;Zhang et al., 2011). As a consequence, the surge in NH 3 emissions from synthetic fertilizer application during this period has been inevitable. Meanwhile, since 1980, when China began developing a series of policies to support livestock production, the industry has undergone rapid growth driven by the increasing demand for beef, pork, mutton, 15 milk, and wool (Zhou et al., 2007). For example, in 2006, there were 56 million head of slaughtered cattle in China, representing a 16-fold increase from the number in 1980; the number of poultry in production increased 12-fold during this same period (EO-CAIY, 2007). The flourishing livestock industry has produced large volumes of manure that releases gaseous NH 3 through N hydrolyzation and volatilization. In conclusion, 20 a marked increase in NH 3 emissions from livestock manure and synthetic fertilizer are expected from 1980 to the present, but specific data on annual emissions and variation in emissions are lacking. Changes induced by anthropogenic activities can significantly influence the global N cycle (Vitousek et al., 1997). Therefore, to better understand the evolution of the global 25 N budget and the impacts on living systems, it is essential to quantify NH 3 emissions during recent decades in China. Moreover, the compilation of multi-year regional and national NH 3 emissions inventories would also help elucidate the causes of severe air pollution in China. In a previous study, we developed a comprehensive NH 3 inventory for 2006 to show the monthly variation and spatial distribution of NH 3 emissions in China based on a bottom-up method (Huang et al., 2012). Our method had several advantages over previous inventories. First, emissions factors (EFs) characterized by ambient temperature, soil acidity, and other crucial influences based on typical local agricultural prac-5 tices were used to parameterize NH 3 volatilization from synthetic fertilizer and animal manure. In addition, we included as many different types of emission sources as possible, such as vehicle exhaust and waste disposal. Our NH 3 emissions inventory was compared with some recent studies to show its reliability. Paulot et al. (2014) used the adjoint of a global chemical transport model (GEOS-Chem) to optimize NH 3 emissions estimation in China; the results were similar to our previous study (Huang et al., 2012). In addition, the distribution of the total NH 3 column in eastern Asia retrieved from measurements of the Infrared Atmospheric Sounding Interferometer (IASI) aboard the European METeorological OPerational (MetOp) polar orbiting satellites (Van Damme et al., 2014) was also in agreement with the spatial pattern of NH 3 emissions calcu-15 lated in our previous study. Hence, our bottom-up emissions inventory appears to be reliable, and the method can be used to estimate NH 3 emissions in China.
In this study, we made further improvements based on this bottom-up method and considered the following sources of NH 3 : (1) farmland ecosystems (synthetic fertilizer application, soil and N fixing, and crop residue compost); (2)

Methods and data
NH 3 emissions were calculated as a product of the activity data and corresponding condition-specific EFs, according to the following equation: where E (NH 3 ) is the total NH 3 emissions; i , p, and m represent the source type, the 5 province in China, and the month, respectively; A i ,p,m is the activity data of a specific condition; and EF i ,p,m is the corresponding EF. The emissions were allocated to a 1 km × 1 km spatial resolution on the basis of land cover, rural population, and other proxies.

Synthetic fertilizer application
NH 3 volatilization from synthetic fertilizers represents an important pathway of N release from the soil, resulting in large losses of soil and plant N (Harrison and Webb, 2001). We classified the synthetic fertilizers used in Chinese agriculture as urea, ammonium bicarbonate (ABC), ammonium nitrate (AN), ammonium sulfate (AS), and oth-15 ers (including calcium ammonium nitrate, ammonium chloride, and ammonium phosphates). NH 3 emissions were estimated by multiplying gridded (1 km×1 km) EFs for five types of fertilizer by synthetic fertilizer use, which was calculated as the product of cultivated area and the application rate to crops (EOCAY, 1981(EOCAY, -2013Zhang et al., 2012;NBSC, 2003NBSC, -2013b

Wind speed
In addition to temperature, wind speed is a meteorological parameter that affects the partial pressure of NH 3 by regulating the exchange of NH 3 between the soil/floodwater and the air, thereby influencing NH 3 volatilization (Bouwman et al., 2002). Several previous studies have shown that high winds significantly influence NH 3 volatilization (Den-5 mead et al., 1982;Fillery et al., 1984;Freney et al., 1985). In this study, we followed the approach of Gyldenkaerne et al. (2005) to introduce the effects of wind speed on NH 3 volatilization from synthetic fertilizer application. The original EFs were multiplied by a factor that was an exponential function of wind speed. The monthly average wind speeds for a 1 km × 1 km grid were based on the final analysis dataset of the National 10 Centers for Environmental Prediction (NCEP).

Improvement of the EF for soil pH
Huang et al. (2012) used the measurements of Cai et al. (1986) and Zhu et al. (1989) to develop a linear relationship between NH 3 volatilization and soil pH that was applied to correct the EFs for soil acidity in different regions of China. To refine the correction so 15 that it was nationally representative, we referred to recent results from field experiments to further adjust this relationship. Huo et al. (2015) conducted fertilization experiments in a representative area of farmland in the North China Plain in Hebei Province from 25 March to 6 May 2012, and measured a soil pH and average air temperature of 8.2 and 15 • C, respectively. The results yielded an NH 3 EF for urea of 12 % ± 3 %. We 20 combined this condition-specific EF with that from Cai et al. (1986) and Zhu et al. (1989) to refine the correction for soil pH.

Livestock waste
A mass-flow approach has been widely used to estimate NH 3 emissions from livestock waste (Beusen et al., 2008;Velthof et al., 2012). Livestock waste can produce Introduction

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Interactive Discussion
Discussion Paper | Discussion Paper | Discussion Paper | Discussion Paper | ammoniacal N (TAN), which can be converted into gaseous NH 3 or lost through other pathways during manure management (Webb and Misselbrook, 2004;Webb et al., 2006). We estimated livestock emissions by multiplying TAN at four different stages of manure management: outdoor, housing, manure storage, and manure spreading onto farmland (Pain et al., 1998) with the corresponding EFs. TAN is the product of the daily 5 amount of manure produced (kg day −1 capita −1 ), N content (%), and TAN content (%) ( Table S1), all of the parameters for which were assumed to have not changed during the 30 year period. In addition, we also considered three main animal-rearing systems in China: free-range, intensive, and grazing. The first two systems are extensively implemented in most rural areas of the country. The free-range system is characterized 10 by small-scale rearing belonging to individual families and has been rapidly developed over recent decades (http://www.caaa.cn/). Based on animal husbandry yearbooks, we defined an intensive rearing system as that where the number of a single livestock class on a single farm was larger than a certain value (Table S2). The number of livestock in each class from 1980 to 2012 was provided by official 15 statistical data and husbandry industry reports (EOCAIY, 1999(EOCAIY, -2013EOCAY, 1981EOCAY, -2013. We applied specific EFs in different situations and took the ambient temperature from the NCEP final analysis dataset into account in the stage in which the manure was stored indoors. 20 The other minor NH 3 emission sources included agricultural soil, N-fixing plants, the compost of crop residues, biomass burning, excrement waste from rural populations, the chemical industry, waste disposal, traffic sources, and NH 3 escape from thermal power plants. The data sources and EFs for each source type are summarized in Ta (EEA, 2013). We roughly estimated the amount of NH 3 escape by multiplying the total flue gas released in power plants nationwide by the maximum allowable concentration of NH 3 carried in flue gas (NEA, 2011;CAEPI, 2013).

Results and discussion
3.1 Annual NH 3 emissions 5 Over the past 30 years, China has undergone dramatic changes and significant economic development, and the annual variation in NH 3 emissions has changed correspondingly. Figure 1 illustrates the trends in total NH 3 emissions, which are divided into fertilizer application, livestock waste, and other minor sources. Total emissions increased from 5.9 to 11.2 Tg between 1980 and 1996, then decreased to 9.5 Tg in 2012.
The most important contributor was livestock manure management, accounting for approximately 50 % of the total budget. Due to the extremely high consumption and high volatility of ABC and urea, synthetic fertilizer application was responsible for 30-43 % of the total emissions, second only to livestock manure. However, in Europe and the United States, where less-volatile synthetic fertilizers such as AN and AS are more 15 popular (Bouwman and VanderHoek, 1997), livestock manure overwhelmingly dominate the NH 3 emissions inventory (Ferm, 1998). These two primary sources combined accounted for 80-90 % of the total emissions budget, with other minor sources accordingly accounting for 10-20 %. million, and 0.9 to 10 billion respectively, from the 1980s to mid-2000s (Fig. S1 in the Supplement). Correspondingly, NH 3 emissions increased from 2.9 Tg in 1980 to 6.2 Tg in 2005, more than doubling during this period, and then decreased to 5.0 Tg in 2012 (Fig. 2). We divided livestock NH 3 emissions from 1980 to 2012 into four phases. In the first phase (1980)(1981)(1982)(1983)(1984)(1985)(1986)(1987)(1988)(1989)(1990), emissions steadily increased with a mean growth rate of approximately 3 %. Free-range production contributed most of the emissions (the population of free-range animals represented more than 90 % of the major livestock animals; EOCAY, 1991). The second phase (1991)(1992)(1993)(1994)(1995)(1996) saw the most rapid increase in emissions, and the growth rate rose to 10 % between the years 1994 and 1995. In 1992, China began to implement a reform of the socialist market economic system, which 10 had previously driven livestock production (CAAA, 2009), and accordingly, more NH 3 emissions from livestock waste were emitted. However, in 1997, there was a 0.5 Tg decrease in livestock emissions, compared to the those in 1996. This observed decline could be attributed to the Asian financial crisis, which started in 1997 and may have a detrimental effect on the development of the Chinese livestock industry. From 15 1998 to 2005, as the third phase, NH 3 emissions continually rose in conjunction with an increase in livestock production due to improvements in cultivation technique and increases in market demand (Zhang et al., 2003). Compared to the first two phases, the contribution of intensive rearing systems to total emissions also increased, with the population of intensively reared animals representing nearly 20 % of the major livestock   (2012), without significant changes. Table S3 presents the interannual emissions of the typical livestock categories. Among them, beef cattle were consistently the largest NH 3 emitter, contributing an annual mean 1.9 Tg NH 3 ; pigs, laying hens, goats, and sheep were also major contrib-10 utors to the total emissions in this period. Poultry had the most rapid growth rate in NH 3 emissions. Nevertheless, a marked downward trend has appeared since 2007 for several major NH 3 sources, including beef cattle, goats, and sheep, which led to the decrease in total emissions discussed above.

Emissions from synthetic fertilizer application
15 Figure 3 shows the estimations of NH 3 emissions from synthetic fertilizer application for the period 1980-2012. Annual levels consistently increased from 1980 (2.1 Tg) to 1996 (4.7 Tg), and then declined from 1996 to 2012 (2.8 Tg). ABC and urea were the major sources; NH 3 release from other synthetic fertilizers, such as AS and AN, made a negligible contribution to emissions (< 0.1 %). In general, the interannual variation in 20 emissions reflects the changes in farming practices in China. First, the relative contribution of urea and ABC has gradually changed over recent decades. During the 1980s, ABC represented a substantial fraction of synthetic fertilizers used in China, and because of its high volatilization (Zhu et al., 1989), emissions from this kind of chemical fertilizer dominated in this period. However, ABC was inefficient for crop production 25 because of the low N content (17 % N) and high N loss. In the mid-1990s, China introduced the technology of urea production, which promoted its widely application (Zhang et al., 2012) placed ABC and become the dominant chemical fertilizer used in cropland over the last 20 years. In 1980, 3.0 million and 5.1 million t of urea and ABC, respectively, were produced; by 2012, these values had changed to approximately 28.8 million and 3.4 million t, accounting for approximately 66.7 and 7.9 % of total synthetic fertilizer production in China, respectively (Fig. S2). Because NH 3 volatilization from ABC is more than 5 2-fold that from urea (Roelcke et al., 2002;Cai et al., 1986), the increasing proportion of urea application relative to that of ABC has caused the decrease in total synthetic fertilizer emissions observed from the mid-1990s onwards. As shown in Fig. 3, although emissions from urea application increased by 1.0 Tg from 1996 to 2012, those from ABC fell by nearly 3.0 Tg. In addition, there have been seasonal disparities in NH 3 10 emissions from synthetic fertilizers, caused by variation in both the temperature and timing of fertilizer application for different crops. Generally, NH 3 volatilization began to rise in April with increasing temperatures, and the highest emissions occurred in summer (June-August), which can be attributed to the high temperatures and intensive application of fertilizer. The seasonal distribution of synthetic fertilizer emissions was 15 almost constant from 1980 to 2012, corresponding to stability in the seasonal distribution of agricultural activities.  1980, 1996, 2006, and 2012 are illustrated in Fig. 4. Emissions from livestock and synthetic fertilizer dominated the total inventories. Specifically, the proportion of emissions from synthetic fertilizers of total emissions peaked at 42.3 % in 1996, and then started to decrease in the following years, which can be attributed to changes in the types of fertilizer used.

5
Furthermore, as mentioned above, the Asian financial crisis and the resulting depression in the livestock market led to a decline in the proportion of emissions from livestock after 1997 and 2006, respectively. The contributions of the chemical industry, waste disposal, traffic, and NH 3 escape from thermal power plants reached the peak values of 3.3, 2.8, 1.8, and 1 %, respectively, in 2012. tial distribution of the total emissions. As mentioned above, NH 3 volatilization from synthetic fertilizer application initially increased rapidly, reaching its peak value in the mid-1990s, and then consistently decreased. This decadal pattern can be observed in the temporal-spatial distribution of NH 3 emissions from fertilizers (middle column in Fig. 5 (Richter and Roelcke, 2000), as well as in Sichuan, which ranked first in ABC application among all provinces in the 1980s. In 2000, NH 3 emission rates from cultivated land in Shandong, Henan, Anhui, and Jiangsu provinces (the North China Plain) generally exceeded 3000 kg km −2 , but decreased to less than 2000 kg km −2 in 2012 due to a reduction in the use of ABC. The total usage of ABC fertilizer in these 5 provinces in 2000 was 5-fold that in 2012. Hubei, Hunan, Jiangxi, and Guangdong provinces, covering China's major rice-production areas, displayed significant growth in NH 3 volatilization from 1980 to 1990, with emissions approximately doubling; however, emissions reduced after the mid-1990s possibly because of the transition of fertilizer usage from ABC to urea in these provinces. In contrast to the variation observed in the 10 areas mentioned above, NH 3 emissions in the Northeast Plain encompassing Jilin, Heilongjiang, and Inner Mongolia, and in Xinjiang Province, have consistently increased in recent decades. From the 1990s onwards, grain production in the Northeast Plain entered a rapid growth period, accompanied by an increasing demand for synthetic fertilizers. 15 The spatial distribution of NH 3 emissions from livestock waste was similar to that from synthetic fertilizers, with high emission rates in eastern China, East Sichuan, and parts of Xinjiang. In the 1980s, Sichuan was the largest emitter among all of the provinces, accounting for more than 10 % of emissions from livestock manure management, followed by Inner Mongolia and Henan. More than half of NH 3 emissions from livestock Introduction sions in these three provinces were beef cattle and laying hens, which contributed 38 and 19 % in 2000, respectively. Beef cattle and goats were extensively bred in Henan, Shandong, Sichuan, and Hebei provinces, and in 2012 the decrease in their population caused a corresponding decrease in NH 3 emissions from livestock manure in these provinces, by approximately 0.14, 0.14, 0.02, and 0.09 Tg, respectively. In addi-5 tion, although emissions from grazing rearing system were less significant than those from other systems (free-range and intensive) in general, they did become significant in northern Inner Mongolia, central and southern Xinjiang, west-central Qinghai, western Sichuan, and large areas of Tibet.

Comparison with previous studies
Our NH 3 emissions inventories provide a detailed description of interannual variation from 1980 to 2012 in China. A comparison between this study and the Regional Emission Inventory in Asia (REAS) is presented in Fig. 6 sions in our study was generally consistent with that in REAS before 1996. However, after that year the annual trend of emissions in our study differed from those in REAS. In addition, the NH 3 emissions in REAS were generally higher than those in our study. These differences are likely attributable to differences in the estimations of synthetic fertilizer emissions, discussed below. 20 In REAS, NH 3 emissions from animal manure applied as fertilizer were included as a category of fertilizer emissions . NH 3 from the application of animal waste onto croplands was 2.8 Tg in 2000 in REAS, accounting for approximately 60 % of the total fertilizer emissions in that year. To render these two inventories comparable, we excluded the application of animal waste from the fertilizer emissions in REAS 25 using the value for 2000. A comparison of the emissions from synthetic fertilizer application is presented in Fig. 7. We found that the REAS values were 20-50 % higher than ours in 2000-2005, and  responsible for the differences of total emissions between REAS and our study in the 2000s. It should be noted that in REAS 2.1, the agricultural emissions were extrapolated from REAS 1.1 for 2000 using the corresponding activity data of the target year (Kurokawa et al., 2013), which could have resulted in considerable inaccuracies due to various missing parameters. On the other hand, the discrepancy possibly originated 5 from the treatment of the types of fertilizer and the corresponding EFs considered in the estimation methods. As mentioned above, the types of fertilizer applied have changed substantially since 1997. Although the total amount of synthetic fertilizers increased significantly, the proportion of highly volatile ABC consistently decreased, which could be responsible for the marked decline in NH 3 emissions. However, REAS considered only the total fertilization activities rather than the change in fertilizer types so the emissions in REAS continued to increase in recent years. Moreover, we took into account the local environmental conditions (soil pH, wind speed etc.) and agricultural practices, and used fields results from Chinese studies to correct the EFs, whereas REAS employed only uniform EFs based on European studies, and applied these across the 15 whole of China. Fu et al. (2015) recently estimated synthetic fertilizer NH 3 emissions at approximately 3.0 Tg in 2011 using the bi-directional CMAQ model coupled to an agro-ecosystem model, which is similar to our value of 2.8 Tg for the same year and supports the reliability of our inventories. Table 3 shows the comparison of the emissions from livestock waste in our study 20 with previous ones. Our results are generally in close agreement with those of Wang (1994), Streets et al. (2003), and Yamaji et al. (2004). The majority of the previous inventories used European-based EFs, which could introduce significant inaccuracies. Our study employed a mass-flow approach, and considered three different livestock rearing systems, as well as four phases of manure management based on 25 local agricultural practices. The EFs used in our study were also refined according to environmental conditions. Hence, our estimations employed more realistic parameters, and the differences between the present study and previous ones are expected. For the total NH 3 emissions, our result (10.1 Tg) was close to that reported (10.4 Tg) using ACPD 15,2015 High-resolution ammonia emissions inventories in China from 1980-2012 a global 3-D chemical transport model for 2005(Paulot et al., 2014. The global NH 3 column mapped with the observations of IASI sensor also demonstrated excellent qualitative agreement with the spatial distribution of our estimated emissions (Van Damme et al., 2014). More specifically, several emissions hotspots, including the North China Plain, Sichuan and Xinjiang provinces (near Ürümqi and in Dzungaria), and the 5 region around the Tarim Basin were also detected by the IASI sensor.

Uncertainty
Uncertainties in NH 3 emissions originated from the values used for both the activity and EFs. Huang et al. (2012) summarized the possible sources of uncertainty in the emissions inventory, including extremely high activity data for fertilizer use and livestock, the numerous parameters involved in the EF adjustment, and large variation (≥ 100 %) in the coefficients of biofuel combustion and chemical industry production. Furthermore, our method used constant EFs for estimating 30 year inventories, which may introduce additional uncertainties. First, the application rate and synthetic fertilization method may have changed during recent decades because Chinese farmers have 15 come to expect higher grain production within limited areas of cropland, which may lead to uncertainties in NH 3 loss per unit area. Second, although we considered interannual changes in the percentage of intensive rearing systems to livestock emissions, manure management, which was divided into four phases in our method, could have also changed over time because it was affected by many factors including the N con-20 tent of the feed, housing structure, manure storage system, spreading technique, and time spent outside or indoors (Zhang et al., 2010). In addition, over recent decades, excessive synthetic fertilizer use has caused significant soil acidification in China (Guo et al., 2010), but our inventories did not consider the influence on NH 3 volatilization, which may lead to more deviation in the emissions estimation. We ran 20 000 Monte 25 Carlo simulations to estimate the range of NH 3 emissions with a 95 % confidence interval for 1980, 1990, 2000, and 2012. The estimated emission ranges were 4.5-7.4, 6.3-11.1, 8.0-13.4, and 7.5-12.1 Tg yr −1 , respectively. 15,2015 High-resolution ammonia emissions inventories in China from 1980-2012 to changes in the types of synthetic fertilizer applied and livestock manure management. These factors were the two major NH 3 sources, accounting for more than 80 % of total NH 3 emissions, while demonstrating different temporal trends. Emissions from synthetic fertilizers initially rose, from 2.1-4.72 Tg, in the period 1980-1996, and then decreased to 2.81 Tg by 2012, which was caused by a change in the relative contri-10 butions of urea and ABC consumption to total emissions. In contrast to synthetic fertilizer emissions, emissions from livestock, ranging from 2.9-6.1 Tg from 1980 to 2012, rose until 2005, but significantly decreased after 2006. Other sources were insignificant in the total budget, and exhibited distinct variation during this period. NH 3 emissions generally peaked in the spring and summer, corresponding to planting schedules and 15 ambient temperatures. At the regional level, the spatial patterns of the total emissions have generally been consistent over recent decades, with high emissions rates of more than 2000 kg km −2 concentrated in Hebei, Shandong, Henan, Jiangsu, Anhui, and East Sichuan provinces, which represent the major areas of intensive agriculture in China. Compared to NH 3 emissions in REAS, our results are more reliable because we con-20 sidered more parameters when calculating specific EFs according to local ambient conditions and agricultural practices. It should be noted that gaps still exist in these inventories due to uncertainties in the activity data, EFs, and related parameters, especially for earlier years. As many samples as possible should be used in statistical censuses, and more local field studies 25 should be implemented for better estimates of the EFs to reduce uncertainties. Such high-resolution inventories can be used in global and regional modeling to simulate atmospheric aerosol formation, explore the impacts of NH 3 emissions on air quality, ACPD ACPD ACPD 15,2015 Phys., 11, 10803-10822, doi:10.5194/acp-11-10803-2011Phys., 11, 10803-10822, doi:10.5194/acp-11-10803- , 2011 Ju, X. T., Xing, G. X., Chen, X. P., Zhang, S. L., Zhang, L. J., Liu, X. J., Cui, Z. L., Yin, B., Christie, P., Zhu, Z. L., and Zhang, F. S.: Reducing environmental risk by improving n management in intensive Chinese agricultural systems, P. Natl. Acad. Sci. USA, 106, 3041-3046, doi:10.1073/pnas.0902655106, 2009. Kirkby, J., Curtius, J., Almeida, J., Dunne, E., Duplissy, J., Ehrhart, S., Franchin, A., Gagne, S., 20 Ickes, L., Kurten, A., Kupc, A., Metzger, A., Riccobono, F., Rondo, L., Schobesberger, S., Tsagkogeorgas, G., Wimmer, D., Amorim, A., Bianchi, F., Breitenlechner, M., David, A., Dommen, J., Downard, A., Ehn, M., Flagan, R. C., Haider, S., Hansel, A., Hauser, D., Jud, W., Junninen, H., Kreissl, F., Kvashin, A., Laaksonen, A., Lehtipalo, K., Lima, J., Lovejoy, E. R., Makhmutov, V., Mathot, S., Mikkila, J., Minginette, P., Mogo, S., Nieminen, T., Onnela, A.,