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Environmental Pollution 159 (2011) 2251e2264 Contents lists available at ScienceDirect Environmental Pollution journal homepage: www.elsevier.com/locate/envpol Review Nitrogen deposition and its ecological impact in China: An overview q Xuejun Liu a,e, *, Lei Duan b, Jiangming Mo c, Enzai Du d, Jianlin Shen a, Xiankai Lu c, Ying Zhang a, Xiaobing Zhou e, Chune He f, Fusuo Zhang a a College of Resources and Environmental Sciences, China Agricultural University, Beijing 100193, China b Department of Environmental Science and Engineering, Tsinghua University, Beijing 100084, China c South China Botanical Garden, Chinese Academy of Sciences, Guangzhou 510650, China d College of Urban and Environmental Sciences, Peking University, Beijing 100871, China e Xinjiang Institute of Ecology and Geography, Chinese Academy of Sciences, Urumqi 830011, China f Institute of Geographic Sciences and Natural Resources Research,Chinese Academy of Sciences, Beijing 100101, China article info abstract Article history: Received 20 April 2010 Received in revised form 31 July 2010 Accepted 5 August 2010 Keywords: Atmospheric pollution N emission and deposition Critical loads Ecological impact Nitrogen (N) deposition is an important component in the global N cycle that has induced large impacts on the health and services of terrestrial and aquatic ecosystems worldwide. Anthropogenic reactive N (N r ) emissions to the atmosphere have increased dramatically in China due to rapid agricultural, industrial and urban development. Therefore increasing N deposition in China and its ecological impacts are of great concern since the 1980s. This paper synthesizes the data from various published papers to assess the status of the anthropogenic N r emissions and N deposition as well as their impacts on different ecosystems, including empirical critical loads for different ecosystems. Research challenges and policy implications on atmospheric N pollution and deposition are also discussed. China urgently needs to establish national networks for N deposition monitoring and cross-site N addition experiments in grasslands, forests and aquatic ecosystems. Critical loads and modeling tools will be further used in N r regulation. Ó 2010 Published by Elsevier Ltd. 1. Introduction Nitrogen (N) deposition has been an important component in the global N cycle with increasing anthropogenic reactive N (N r ) emissions since the industrial revolution (Vitousek et al., 1997; Galloway et al., 2008). Excess N deposition has aroused concerns about its negative impacts on ecosystem health and services such as loss of biodiversity (Sala et al., 2000; Stevens et al., 2004), eutrophication and N saturation (Aber et al., 1998), soil acidification (Richter and Markewitz, 2001), and increased susceptibility to secondary stresses (Aerts and Bobbink, 1999; Witzell and Shevtsova, 2004). Rates of N deposition have leveled off or stabilized in the US and Europe since the late 1980s or early 1990s with the implementation of stricter legislation to limit atmospheric pollution (e.g. Goulding et al., 1998; NADP, 2000). In contrast, emissions of N r species in China have been increasing continuously since the 1980s mainly due to growing agricultural and industrial activities (Klimont et al., 2001; Zhang et al., 2007, 2009a). These increased N r emissions to the atmosphere have aroused widespread concern on air pollution in q All the co-authors contributed equally to this work. * Corresponding author. E-mail address: xuejun.13500@gmail.com (X. Liu). China (Richter et al., 2005). Although there have been several N deposition monitoring programs and N deposition simulation experiments since the late 1990s (e.g. Wang et al., 2004; Liu et al., 2006; Mo et al., 2006), there are still large gaps in knowledge of the magnitude and potential impacts of atmospheric N deposition on different ecosystems across China. In this review paper we summarize all the published data from N deposition monitoring and modeling, critical loads, and the effects on ecosystems in order to: 1) identify the magnitude and spatio-temporal variability of N deposition; 2) summarize the major impacts of N deposition on terrestrial and aquatic ecosystems; 3) analyze the potential critical loads of N deposition to major Chinese ecosystems; 4) come up with recommendations for research on regulatory strategies for mitigation of atmospheric N r pollution and deposition in China. 2. Emissions and atmospheric concentrations of N r pollutants in China 2.1. Trends in N r emissions to the atmosphere There have been a number of studies on both reduced and oxidized N emissions in China (e.g. Kato and Akimoto, 1992; Sun and Wang, 1997; Tian et al., 2001; Streets et al., 2003; Yamaji 0269-7491/$ e see front matter Ó 2010 Published by Elsevier Ltd. doi:10.1016/j.envpol.2010.08.002

2252 X. Liu et al. / Environmental Pollution 159 (2011) 2251e2264 et al., 2004; Ohara et al., 2007; Wang et al., 2009a; Zhang et al., 2009a). Despite the variation among studies, NH 3 and NO x emissions have shown substantial increases since the early 1980s (Fig. 1 a, b). Compared with 1980, NH 3 emissions (13.7 Tg N yr 1 ) had doubled and NO x emissions (6.0 Tg N yr 1 ) had increased by a factor of 4 by the year 2005 (Zhao et al., 2009a). The rapid increases in both N r species emissions are closely related to intensive agricultural and industrial activities. The majority of anthropogenic NH 3 in China is emitted from N fertilizers (e.g. ammonium bicarbonate and urea) and animal/human excreta (Zhao and Wang, 1994; Zhang et al., 2010a). NO x emissions are derived mainly from fossil fuel combustion processes including power plants, transportation and industry (Streets and Waldhoff, 2000; Streets et al., 2003) with minor contributions from lightning, biomass burning and arable soils (Yan et al., 2003). If this trend continues we can expect the N r emission induced deposition to make a larger contribution to acid rain than that of sulfur (S) deposition in China in the near future. 2.2. Concentrations of major gaseous and particulate N pollutants Results from monitoring of concentrations of NO x and NH 4 þ en and NO 3 eninpm 10 (particulate matter smaller than 10 microns) or N O x emission s N H 3 emission s 7 6 5 4 3 2 1 0 15 12 Tian et al.,2001. China Environ. Yearbook,2008. Streets et al.,2001&2003 Ohara et al.,2007 Kato and Akimoto,1992. Zhang et al.,2007. Zhang et al.,2009a. Van Aardenne, et al.,1999. Klimont et al.,2001&2009 1975 1980 1985 1990 1995 2000 2005 2010 Year 9 6 3 Wang et al.,2009a. Wang et al.,1997. Streets et al.,2003. Klimont et al.,2001. Sun and Wang,1997. Olivier et al.,1998. FRCGC,2007. 0 1975 1980 1985 1990 1995 2000 2005 2010 Year Fig. 1. Total anthropogenic NO x (a) and NH 3 (b) emissions (Tg N yr 1 ) in China during 1980 and 2007. Data sources are from various published references as shown in the Figure. a b total suspended particulates (TSP) in urban and rural areas show heavy N r pollution in major Chinese mega cities (Chan and Yao, 2008) and in some rural or suburban regions such as the North China Plain and Taihu Lake Plain (Ju et al., 2009). Concentrations of major gaseous and particulate N pollutants during 1999 and 2009 are summarized in Table 1. 2.2.1. NH 3 concentrations NH 3 is the most abundant basic gas in the atmosphere. It can react with acidic gases (e.g. H 2 SO 4, HNO 4 and HCl) to form secondary particles and can also return to the land surface by dry deposition not far from the emission sources (Asman et al., 1998; Erisman and Schaap, 2004). There have been few atmospheric NH 3 monitoring results available in China so far. NH 3 concentrations were higher in agricultural and urban regions with large population density in the order North China > South China > Northwest China and the Qinghai-Tibetan Plateau (Table 1). For example, annual mean NH 3 concentrations at two agricultural sites on the North China Plain were 13.5 and 9.5 mgnm 3 (Shen et al., 2009). Cao et al. (2009) and Meng et al. (2010) reported similar NH 3 concentrations in Xi an, Shaanxi province and Houma, Shanxi province, respectively. The high NH 3 concentrations in North and South China are consistent with the high emission densities in these regions (Wang et al., 1997; Zhang et al., 2010a). The regional background NH 3 concentrations (1.5e3.4 mg Nm 3 ) in China (Meng et al., 2010) are also relatively high compared with those at other remote sites worldwide (e.g. 0.1e1 mg Nm 3 ). 2.2.2. NO 2 concentrations NO 2 is one of the major precursors that contribute to acid deposition and plays an important role in the formation of tropospheric ozone by photochemical reactions with non-methane hydrocarbons (NMHC) (Fowler et al., 1998). NO 2 concentrations have been routinely monitored in almost every mega city in China since the late 1990s. According to the 2008 Report on the State of the Environment, NO 2 concentrations were higher in mega cities (e.g. Beijing, Tianjin, Shanghai and Chongqing) than in cities with lower populations in Western China (e.g. Guizhou, Yunnan and Inner Mongolia). These monitoring results are consistent with space observations on NO 2 by satellite (Richter et al., 2005). In recent years vehicle numbers have increased rapidly in the mega cities and emissions from vehicles are considered to be the largest source of NO x in some mega cities such as Tianjin (Zhao and Ma, 2008). Fossil fuel combustion for power generation and industry are also important sources of NO x. High NO 2 concentrations were not confined to urban air but were also found in some rural regions with highly-developed economies. For example, annual NO 2 concentrations were 9.3 mg Nm 3 at Quzhou on the North China Plain (Shen et al., 2009) and 12.8 mgnm 3 at Jurong in the Yangtze River Delta (Su et al., 2009). This is also reflected in the relative high NO 2 concentrations at regional background sites on the North China Plain and in the Yangtze River Delta (Meng et al., 2010). Table 1 Reported concentrations of major gaseous and particulate N r pollutants in China during 1999 and 2009 (mg Nm 3 ). Region NH 3 NO 2 HNO 3 pnh 4 þ pno 3 North China 1.5e44.7 2.3e23.8 0.1e0.6 4.0e13.5 1.0e7.8 South China 0.5e22.8 0.6e35.8 0.1e1.4 0.8e6.5 0.1e3.7 Northwest China 0.1e11.6 0.1e27.7 n.d. n.d. n.d. Qinghai-Tibetan Plateau 0.5e2.7 0.03e0.7 n.d. n.d. n.d. Notes: 1) n.d., not determined; 2) pnh 4 þ and pno 3 dparticulate ammonium and nitrate; 3) Data sources: published journal papers, Ph.D. and M.Sc. theses and governmental reports (see Supporting online material).

X. Liu et al. / Environmental Pollution 159 (2011) 2251e2264 2253 Increased NO x emissions from vehicles and small factories and relatively high NO x emissions from agricultural fields can account for the high NO 2 concentrations in rural areas (Fang et al., 2007, 2009a). In contrast, NO 2 concentrations (0.4e2.3 mg Nm 3 ) were very low at regional background sites in Northeastern China (Longfengshan) and on the Qinghai-Tibetan Plateau (Waliguan) (Meng et al., 2010). 2.2.3. HNO 3 concentrations Atmospheric HNO 3 is very important in N or acid deposition and atmospheric chemistry, but information on HNO 3 concentrations in China is very limited. According to a few measurements, HNO 3 concentrations at province scale seem to coincide with NO 2 concentrations. Hu et al. (2008) reported high HNO 3 concentration (1.4 mgnm 3 ) in summer at a coast site in the Pearl River Delta. Aas et al. (2007) also reported high HNO 3 concentrations at rural sites in Chongqing and Hunan province in South China. Shen et al. (2009) found a mean HNO 3 concentration of 0.6 mgnm 3 at a rural site on the North China Plain. These limited data suggest that atmospheric HNO 3 concentrations in China are relatively low compared with either NH 3 or NO 2 (Table 1). 2.2.4. Particulate NH 4 þ and NO 3 concentrations Though PM 10 has been routinely measured in most of the cities in China, measurements of particulates NH 4 þ and NO 3 are limited. Most of the measurements of particulates NH 4 þ and NO 3 have been conducted at urban sites, especially mega cities, to study the chemical components of PM 10 and/or PM 2.5. Both particulates NH 4 þ and NO 3 concentrations were high in mega cities and in North China compared with those in South China (Table 1). Relatively high concentrations of particulates NH 4 þ and NO 3 at rural sites in China (Aas et al., 2007; Cao et al., 2009; Shen et al., 2009) implied high rates of N deposition in natural and agricultural ecosystems of these regions. Measurements of particulates NH 4 þ and NO 3 in Northwest China and the Qinghai-Tibetan Plateau are still scarce (Table 1). 3. Nitrogen deposition monitoring in China 3.1. Monitoring networks for N deposition in China Although no long-term national deposition monitoring networks for N deposition currently exist in China, there have been a number of studies on acid rain or wet deposition since the early 1980s. From 1981 to 1983 the Chinese National Environment Bureau organized a nationwide campaign for acid rain measurement (Zhao and Sun, 1986). Similar campaigns were also carried out during the early and late 1990s. Since 1999 four Chinese cities (Xiamen, Xi-An, Chongqing and Zhuhai) have joined the Acid Deposition Monitoring Network in East Asia (EANET) (http://www.eanet.cc/product.html). Currently both the Chinese Ministry of Environmental Protection and the National Meteorological Bureau have been running two independent atmospheric deposition monitoring networks since the late-1990s, both of which include more than 300 sites across China (Ding et al., 2004; Anon., 2010). The sites of the former network are mainly distributed around urban areas and the latter are sparsely located in rural or background regions. More recently, China Agricultural University has organized a Nationwide Nitrogen Deposition Monitoring Network (NNDMN) since 2004 (Liu and Zhang, 2009). This network contains about 40 monitoring sites across China, covering cropland, grassland, forest, and urban ecosystems. There are several additional sub-national networks on N deposition monitoring, including the Integrated Monitoring Program on Acidification of Chinese Terrestrial Ecosystems (IMPACT) (Tang et al., 2001) and the World Meteorological Organization Global Atmosphere Watch Precipitation Chemistry Program (WMO/GAW) of China (http://cdc.cma.gov.cn/index.jsp). These networks tend to monitor wet deposition (N input from rainfall and snowfall) but the NNDMN also measures dry deposition of various N r species (Shen et al., 2009). All these networks have different methodologies and quality control systems. For example, the NNDMN cooperates closely with Centre for Ecology & Hydrology (CEH) Edinburgh, UK and runs the same quality control system as CEH. While the WMO/GAW of China adopts the methods used by WMO protocols. In general, the monitoring networks in China lack long-term continuous data and uniform measuring methods which should be improved in the future. 3.2. Patterns of nitrogen deposition in China Using current N deposition monitoring networks and published data, Lü and Tian (2007) reported the spatial pattern of N wet and dry deposition. They found that average N wet plus dry deposition was 12.35 Tg N yr 1 during 1993 and 2003. This estimate is comparable to the total estimated emissions of NO x and NH y (14.67 Tg N yr 1 )in2000(streets et al., 2003), but much greater than modeled N deposition over China (8.82 Tg N yr 1 ) in 1993 (Dentener et al., 2006a). However, the results of Lü and Tian (2007) were most likely underestimated for both N wet and dry deposition. The main reasons for underestimation are that dissolved organic N (DON) in wet deposition and NH 3, HNO 3 and particulates NH þ 4 and NO 3 (pnh þ 4 and pno 3 ) in dry deposition were not taken into account in their study. According to Zhang et al. (2008a), DON in wet deposition averaged 8.6 kg N ha 1 yr 1, approximately 30% of total N wet deposition across 15 sites in China. Song et al. (2005) reported 25.5% of N wet deposition from DON in the Taihu Lake region. Shen et al. (2009) found that NH 3, HNO 3 and pnh þ 4 and pno 3 were important components of dry deposition on the North China Plain, contributing to 57% of total dry deposition. Liu and Zhang (2009) re-estimated the total N deposition in China to be 15 Tg N yr 1 in the 2000s compared with 7.4 and 12.0 Tg N yr 1 in the 1980s and 1990s, respectively. Therefore, the current total N wet and dry deposition rates are likely to be higher than 15 Tg N yr 1 þ if all the N r species (e.g. DON, NH 3, HNO 3 and pnh 4 and pno 3 ) are included in the overall N deposition budgets. Because of the factors outlined above (especially the incomplete measurement of N dry deposition), most published studies have greater wet than dry deposition. For example, Wang et al. (2008a) observed that wet deposition in the loess area accounted for over 90% of the total deposition. In some forest areas, however, dry deposition was greater than wet deposition. In broadleaf forests, dry deposition can explain 67e75% of the total N deposition (Fan et al., 2007a; Hu et al., 2007). Dry deposition fluxes throughout China from 1990 to 2003 increased slightly on average, with an apparent increase in the Southeast and Southwest, and declines were marked in the Northeast and in parts of the Midlands (Lü and Tian, 2007). 3.3. Simulations of N deposition in China Zhao et al. (2009a) simulated N deposition using the Models-3/ Community Multiscale Air Quality (CMAQ) system (V4.4) (Byun and Ching, 1999). This model has previously been modified and proven to be suitable for Chinese regional and urban-scale air quality simulations (Streets et al., 2007; Wang et al., 2008b). The driving meteorological inputs are provided by the fifth-generation NCAR/ Penn State Mesoscale Model (MM5). To validate the reliability of CMAQ modeling, the simulated results were compared with observation values for annual wet S and N deposition from two programs: IMPACTS and EANET (http://www.eanet.cc/product/ datarep/).

2254 X. Liu et al. / Environmental Pollution 159 (2011) 2251e2264 The model calculated N concentration and deposition in each 36 36 km 2 grid in a domain covering most of East Asia, based on the latest Chinese emission inventory of NO x,nh 3,SO 2, and VOC s (Wei et al., 2008; Zhao et al., 2008a, 2009a) and other related emissions inventories outside China (e.g., Streets et al., 2003). Total N flux was calculated as the sum of nitrate, NO x and ammonium, including wet and dry deposition. Northern and eastern China were recognized as the regions receiving the highest N deposition, with a substantial contribution from the high density of energy consumption and emissions. However, some other modeling results showed that total N wet and dry deposition rates peaked over the central south China (Lü and Tian, 2007). This is still an open question for the spatial pattern of N deposition in China because of the lack of systematic monitoring data in the country. Holland et al. (1999) compared pre-industrial and contemporary global distribution of N deposition by a global three-dimension chemical transport model, MOGUNTIA, at a 10 10 resolution. The Eastern U.S.A., Western Europe and Southern Asia (including southeastern China) were three hotspots of high N deposition. Dentener et al. (2006b) assessed global N deposition using different state-of-the-art global atmospheric chemistry models at 2e3 resolutions, which further consolidated the high N deposition in southeast China. Continental scale models of East Asia have horizontal resolutions of 0.5e1 or 40e45 km. Most studies have focused on S deposition or acid deposition, including oxidized S and N(Hao et al., 2001; Halloway et al., 2002; Park et al., 2005). Wang et al. (2008b) and Hayami et al. (2008) carried out modelemodel comparisons between eight atmospheric transport models (ATMs), and some of these models covered both NH x and NO y deposition. High N deposition often occurred in the eastern part of China (Lü and Tian, 2007). Zhang (2009) constructed N budgets and spatial variation of N deposition on the North China Plain (NCP) by an ATM, the FRAME model. On the NCP about 3.4 Tg N or 100 kg N ha 1 yr 1 was deposited as wet and dry deposition in the year 2004, agreeing well with directly measured results in this region (e.g. Zhang et al., 2008b; Shen et al., 2009; He et al., 2007, 2010). Deposition of the reduced N species dominated the NCP budget with 2.7 Tg N, 2.5 times more N than oxidized N deposition, suggesting a large contribution from agricultural sources (Zhang, 2009). 4. Impacts of simulated N deposition on different ecosystems in China The problem of N deposition originates from acid deposition (rain) and this was recognized in the early 1980s (Zhao and Sun, 1986). Studies on the effects of acid or N deposition on ecosystems have become more prominent over the last decade. Numerous simulated N deposition experiments have been conducted in different grassland and forest ecosystems in China (e.g., Mo et al., 2006; Fan et al., 2007b; Xia et al., 2009). Fig. 2 summarizes the locations of N addition experiments conducted in forest, grassland and desert ecosystems across China since the late 1990s. 4.1. Forests China has a large territory with great climatic complexity and spatial variation which sustains a variety of forest ecosystems ranging from boreal forests in the north to tropical rain forests in the south. These forests play an important role in maintaining biodiversity and ecological equilibrium and in providing services for social development. High atmospheric N deposition has been reported in many forest ecosystems (Zhou and Yan, 2001; Zhang et al., 2006; Chen and Mulder, 2007; Hu et al., 2009) with N deposition levels commonly exceeding 20 kg N ha 1 yr 1 in central and east China above which forest health will be seriously Fig. 2. Distribution of N addition experiments in forest, grassland and desert ecosystems across China based on published references (e.g., Fan et al., 2008; Hu et al., 2009; Lin et al., 2007; Mo et al., 2006; Shan, 2008; Shen et al., 2002; Song et al., 2009; Tu et al., 2009; Wu et al., 2009; Yao et al., 2009; Yu et al., 2007; Zhang et al., 2004; Zhang et al., 2009b; Zhang et al., 2010b; Zhou et al., 2004; Zhou et al., in press) and personal communications (Du E.Z., Lu X.K., Li K.H., Jin G.Z., per. com.). threatened (MacDonald et al., 2002; Bobbink et al., 2010). However, studies on the responses of forest ecosystems to elevated N deposition only started in China in the 2000s and the response records are far from complete. A Sino-Norwegian project IMPACTS, launched in 1999, has established monitoring sites at five forest ecosystems in the south of China to monitor atmospheric N deposition and its effects on leaching of N (Chen et al., 2004). To better understand and predict the effects of N deposition on forest ecosystems, DingHuShan Forest Ecosystem Long-term Nitrogen Research Project (DHSLTNR) was established. The experiment covered three tropical/subtropical forest sites (an old-growth monsoon evergreen broadleaf forest, a mixed pine and broadleaf forest and a pine forest) in 2002 (Mo et al., 2006). This experiment aimed to study the responses of forest ecosystems to elevated N deposition. Since then, many N addition experiments have been established in different forest types from south to north China (Fig. 2). Here we mainly focus on the effects of N addition on N dynamics in forest soils, soil acidification, plant growth, biodiversity, litterfall decomposition and flux of greenhouse gases. 4.1.1. Effects on N dynamics and soil acidification It is well established that elevated N deposition will alter N cycling greatly in forest ecosystems. Observations in tropical and temperate forests showed that N inputs to forest floors can affect the status of N in underlying forest soils and increase soil available N content (Fan et al., 2007b,c; Hu et al., 2009; Lu et al., 2009; Xu et al., 2009; Fang et al., 2009a, in press). Nitrogen leaching will occur if N input exceeds soil retention capacity. In five subtropical forested catchments of South China, Chen et al. (2004) found that N leaching from soils was high and almost all N leached as NO 3 enin high N deposition sites (35 kg N ha 1 yr 1 ). Fang et al. (2009a) also found that experimental N additions (50e150 kg N ha 1 yr 1 ) increased dissolved inorganic N (DIN) leaching in all three subtropical forests, with 25e66% of added N leached over a 3-year period. In addition, leaching of dissolved organic N (DON) increased under high N input conditions (Fang et al., 2009b). Elevated N deposition can increase the acidification potential of forest ecosystems and decrease soil buffering capacity (Vogt

X. Liu et al. / Environmental Pollution 159 (2011) 2251e2264 2255 et al., 2006; Lu et al., 2009). In a mature subtropical forest of southern China, Lu et al. (2009) found that the ecosystem was sensitive to high N addition, and continuous two-year N additions (50e150 kg N ha 1 yr 1 )significantly enhanced soil acidification, Al 3þ mobilization, and leaching of base cations from soils, the typical characteristics of N saturation (Aber et al., 1998). Fan et al. (2007b) also found that exchangeable base cations (e.g. Ca 2þ and Mg 2þ ) decreased with increasing N addition in a subtropical Chinese fir plantation after three years of N manipulation (60e240 kg N ha 1 yr 1 ). In addition, the nitrate leached out was accompanied by positively charged counter-ions, the base cations K þ,ca 2þ and Mg 2þ, resulting in the further acidification of the leached soil, or hydrogen and aluminum ions, which may cause the acidification of receiving systems. 4.1.2. Effects on plant growth Changes in N cycling and soil quality induced by N deposition will affect plant growth. There have been only a few studies on the responses of plant growth to N addition. These responses have been found to be related to plant demand for N. In N-limited forests N deposition can meet plant demand for N and improve plant nutrient status, with an increased photosynthetic capacity and stimulation of plant growth. However, excess N will result in nutrient imbalance in trees, disturb N metabolism, reduce net photosynthesis, and restrict plant growth (Li et al., 2005; Lu et al., 2006, 2007; Mo et al., 2008a). By studying the seedling growth response of two tropical tree species (Schima superba and Cryptocarya concinna) to N addition, Mo et al. (2008a) found that net photosynthetic rate and biomass production were increased by lower N addition (50e100 kg N ha 1 yr 1 ) but both species were negatively affected by higher N addition. In an N-saturated tropical forest, Lu et al. (2006, 2007) found that N addition significantly increased foliar N in three understory plants and additional N accumulated as organic N (e.g. free amino acids) and N addition inhibited net photosynthetic capacity in these plants. The negative effect of N addition was also found on the belowground biomass production in the same forest, which showed that fine root biomass decreased significantly with increasing levels of N addition (Mo et al., 2008b). In a 3-year study, Duan et al. (2009) observed that elevated CO 2 concentration and N deposition increased biomass accumulation but the responses varied among tree species. Nitrogen deposition stimulated aboveground biomass accumulation with decreasing root: shoot ratio and elevated CO 2 concentration significantly increased biomass allocation to belowground or aboveground biomass of different tree species. 4.1.3. Effects on biodiversity Biodiversity can be significantly affected by N addition (Clark and Tilman, 2008). Although the impacts of N deposition on biodiversity have generated wide international interest, there are only a few long-term research projects in China (e.g., the DingHuShan Long-Term Nitrogen Research or DHSLTNR) (Lu et al., 2008). These studies show that N deposition can alter species diversity, and excessive N can reduce species diversity. Studies in an old-growth tropical forest of southern China showed that high N addition levels (e.g. >100 kg N ha 1 yr 1 ) significantly reduced understory species diversity and the mechanism for change appeared to be N deposition-mediated soil acidification rather than fertilization (Lu et al., in press). In the same forest, Xu et al. (2006) found that N addition also significantly decreased soil fauna diversity. However, some studies have shown that species richness was increased by low N addition levels. For example, Xu et al. (2005) found that 16- months of N addition (50e100 kg N ha 1 yr 1 )increasedsoil fauna diversity in the pine forest of Dinghushan. These changes may be related to soil N status, vegetation composition and time of N addition. A few studies have shown that soil microorganisms are very sensitive to N addition. Xue et al. (2007) found that short-term N addition significantly increased soil bacterial numbers but decreased soil fungal counts. A lower soil N status and different species requirements for N may be attributable to shifts in functional microbial communities. 4.1.4. Effects on litter decomposition Nitrogen has long been recognized as an important factor regulating litter decomposition. Litter decomposition is a key step in nutrient cycling. Studies on the effects of N addition on litter decomposition have shown different results, ranging from positive to negative responses of the decomposition rate and elemental release, depending on ecosystem N status, litter quality, species, and time of N deposition (Mo et al., 2006, 2007a, 2008c; Song et al., 2007a; Deng et al., 2007; Fan et al., 2008). In studies of subtropical forests, Mo et al. (2006) found that litter decomposition rates exhibited no significant positive and even some negative responses to N addition in N-saturated mature forests, but showed significant positive responses in N-limited disturbed and rehabilitated forests. Nitrogen deposition has significant cumulative effects on litter decomposition and the initial effects of N addition change over time (Fang et al., 2007). Effects of N addition on litter decomposition also varied depending on the nutrient status of the litter (Mo et al., 2008c). 4.1.5. Effects on the fluxes of greenhouse gases Few studies have been conducted on the effects of elevated N deposition on the fluxes of greenhouse gases (GHGs, CO 2,CH 4, and N 2 O) from forest soils, and the response pattern depends on forest type, N status of the soil, and the level of N deposition. In N-saturated mature forests, high N addition can reduce soil respiration (CO 2 emissions) (Mo et al., 2008b), increase soil N 2 O emissions (Zhang et al., 2008c), and decrease CH 4 uptake rates (Zhang et al., 2008d). However, studies in N-limited pine forests show that N addition has no significant effects on soil respiration or CH 4 uptake in spite of increased N 2 O emissions as result of high N addition (Mo et al., 2007b; Zhang et al., 2008c,d). 4.2. Grasslands Grasslands in China cover about 40% of the total territory and diverse grassland types range from vast, continuous areas of temperate grasslands in arid and semi-arid regions to alpine grasslands on the Tibetan Plateau, and to small areas of the warm temperate and tropical regions (Chen and Chen, 2007). The vast area and wide distribution of Chinese grasslands, however, have been largely ignored in terms of global N deposition. Most studies on the responses of grassland ecosystems to N deposition have been restricted to the semi-arid temperate steppes (e.g. the Inner Mongolia grasslands) in northern China. Current studies include responses of plant growth, biodiversity, ecosystem carbon exchange, and N transformation processes. 4.2.1. Effects on plant growth Several studies have found that N is a major limiting factor in the growth of grasslands (Xia et al., 2009). Nitrogen addition may lessen the N limitation by increasing soil N availability and thus stimulate plant growth (Pan et al., 2004; Bai et al., 2010). In Leymus chinensis communities in typical steppe of Inner Mongolia, Pan et al. (2005) found that N addition increased significantly the density, height, aboveground biomass, belowground biomass and total biomass of Leymus chinensis. With the greater increase of aboveground biomass than belowground biomass, N addition

2256 X. Liu et al. / Environmental Pollution 159 (2011) 2251e2264 significantly reduced the root: shoot ratio of the affected plants (Pan et al., 2005; Zhang et al., 2010b). Therefore, N addition not only increased the biomass of Leymus chinensis population but also changed the resource partitioning among plant parts. Plant functional traits were greatly changed by N addition at the same time. Observations in a mature typical steppe ecosystem showed that N addition led to increased foliar N concentration, total chlorophyll content, and specific leaf area (SLA) in these plants to some extent, but plant species differed significantly in their responses to increased N addition rates (Wan et al., 2008; Huang et al., 2009). In addition, added N can change productivity and competitive balance between C 3 (Leymus chinensis) and C 4 (Chloris virgata) plants. Leymus chinensis accumulated more biomass under N treatments and the final biomass of Chloris virgata was less impacted than Leymus chinensis (Niu et al., 2008). In a 4-year N addition study, Xia et al. (2009) found that the gross ecosystem productivity increased in spite of the reduction in the biomass of forbs. 4.2.2. Effects on biodiversity Plant diversity commonly responds to N addition as a loss of richness accompanied by a shift in plant functional group composition (Huang et al., 2009; Bai et al., 2010). N enrichment allows species with acquisitive resource-use strategies to exclude those with conservative resource-use strategies without other stresses (e.g. drought). As found by Bai et al. (2010) in a mature community of Inner Mongolia grasslands, N addition led to a large reduction in species richness accompanied by increased dominance of early successional annuals and loss of perennial grasses and forbs at all N input rates. In addition, with increasing N uptake, long-term N addition would decrease N re-sorption proficiency and make some dominant species less dependent on N re-sorption, while other nutrients (e.g., P) or resources (e.g., water and light) could become limiting factors for plant growth (Huang et al., 2009). As a result, interspecific differences in N re-sorption may influence the positive feedback between species dominance and N availability. Reduced microbial diversity (e.g. functional diversity) has been observed in a semi-arid temperate steppe in response to high levels of N addition but a low level of N addition stimulated microbial diversity (Zhang et al., 2008e). An N-addition-mediated decline in ph could be largely responsible for the decline in diversity. Zhang et al. (2008e) suggested a critical N loading level between the rates of 160 and 320 kg N ha 1 yr 1 in the temperate steppe, beyond which the microbial community would shift its response to N addition. However, this high critical load may only occur in severely degraded steppe because of soil N deficiency. Further studies are required in other grasslands. 4.2.3. Effects on net ecosystem carbon exchange Net ecosystem C exchange (NEE) represents the balance between gross ecosystem productivity (GEP) and ecosystem respiration (ER). It is addressed well in the changes in NEE response to N addition in a temperate steppe in northern China. N addition has been observed to increase NEE in the two hydrologically contrasting growing seasons because of the larger stimulation in GEP than in ER (Xia et al., 2009). However, the magnitude of N stimulation on NEE declined over time. Niu et al. (2010) considered that the temporal decline in the N stimulation of NEE resulted from an N-induced shift in species composition. In any case, the positive effects of N on NEE suggest that N deposition under global climate change may increase C sequestration in the temperate steppe (Niu et al., 2009, 2010). In addition, the interactive effects of water/ warming and N addition on NEE were also studied. The N responses of NEE under warming were suppressed (Xia et al., 2009), but there are no additive effects of water and N addition on NEE (Niu et al., 2009). 4.2.4. Effects on N transformation processes As a part of N cycling, N addition significantly impacts N transformation processes (N mineralization and nitrification). Nitrogen addition increased soil inorganic N (ammonium and nitrate N) and rates of net N mineralization and nitrification, but reduced soil microbial biomass C (Yu et al., 2007; Zheng et al., 2008). Zhang et al. (2008e) have reported that 3-year N addition (0e640 kg N ha 1 yr 1 ) caused gradual or step increases in soil NH 4 þ en, NO 3 en, net N mineralization and nitrification in the early growing season. In a typical steppe dominated by a Leymus chinensis community, however, N addition also showed dose effects. Zhang et al. (2009b) found that low N addition (e.g. 17.5 kg N ha 1 yr 1 )stimulatedn mineralization but high N addition (280 kg N ha 1 yr 1 ) inhibited N mineralization. The mechanism behind these should be clarified in the future. 4.3. Croplands Almost all croplands can be regarded as N-saturated ecosystems due to continual application of chemical N fertilizers. Nitrogen deposition onto croplands is relatively low compared with N fertilizer inputs. However, N deposition may make a substantial contribution to nutrient budgets in some intensively managed croplands (He et al., 2010) or in some low input agroecosystems. Furthermore, heavy N deposition is normally associated with acid rain and atmospheric N r pollution, both of which could lead to yield decline in croplands. 4.3.1. Nutrient contribution of N deposition to croplands Nitrogen is an essential nutrient for plant growth and NH 4 þ en, NO 3 en are two major forms of N taken up by plants. Nitrogen deposition (mainly as NH 4 þ en and NO 3 en) is important N source in agroecosystems (He et al., 2007). Zhang (2006) reported up to 58 kg N ha 1 yr 1 wet deposition of NH 4 þ en and NO 3 en in the highly-developed Shanghai region. He et al. (2010) found that total airborne N inputs to a maizeewheat rotation system on the NCP ranged from 99 to 117 kg N ha 1 yr 1, while plant available N from deposition for maize and wheat was about 52 kg N ha 1 yr 1, accounting for 50% of the total N deposition or 31% of total N uptake by the two crop species. Based on some earlier measurements, Zhu (1997) estimated approximately 1.5 Tg annual N input from wet deposition plus irrigation to croplands of China in the 1980s. Liu and Zhang (2009) estimated total N input from wet and dry deposition and found 15 Tg N yr 1 for the whole of China in the 2000s, equal to about 50% of N fertilizer use in the country. Assuming that one third of N deposition occurred in the agricultural areas, 5.0 Tg annual N deposition may have returned to croplands of China since the 2000s. This is an important nutrient input to croplands from the atmosphere. 4.3.2. Negative effects of N deposition on croplands High N addition level of N deposition especially acid deposition might produce some negative impacts in agroecosystems. Earlier studies on the effect of acid deposition on croplands were carried out in the late 1970s especially in south China (Zhang et al., 1997; Hu and Su, 1999). High N deposition together with acid rain might decrease yield and quality of crops by influencing soil properties and soil microorganisms (Lin, 2005). The impact of N deposition on soils depends on the form of the deposition, the N demands/requirements of crops, and soil chemistry, mineralogy and moisture status (Hornung, 1995). Increased N deposited above background levels will initially be utilized, resulting in stimulation in plant growth. However, increased N availability can eventually result in induced deficiencies of other nutrients (Zhou and Ogura, 1996; Qiu et al., 1997). Deposited NH 4 þ en can be nitrified to NO 3 en, resulting in

X. Liu et al. / Environmental Pollution 159 (2011) 2251e2264 2257 enhanced leaching of base cations and soil acidification (Xu et al., 2002; Guo et al., 2010). Increased N availability can also lead to frost and drought sensitivity and increased risk of pest and disease infestation (Xu, 1995). Moreover, high NH 3 concentrations and deposition induced by excess N fertilization may cause NH 3 toxicity to crops, with symptoms of injury showing in the leaves (Fangmeier et al., 1994). Anthropogenic N r pollution and deposition may result in yield loss and quality decline of crops. But these negative effects on croplands have been observed mainly in hotspots of N deposition (e.g. areas close to intensive livestock farms, N fertilizer factories or power plants). 4.4. Aquatic and coastal ecosystems China has more than 1500 rivers with watersheds larger than 1000 km 2 and more than 2800 lakes with surface areas greater than 1km 2. The Bohai Sea, Yellow Sea, East China Sea and South China Sea cover a total area of about 4.7 million km 2. Nitrogen emissions and deposition are expected to be high in coastal provinces with rapid economic growth. However, only a few studies on N deposition have been reported in the Yellow Sea, East China Sea and Bohai Sea and in Taihu Lake (e.g., Zhang and Liu, 1994; Wan et al., 2002; Song et al., 2005). Current studies are mainly focused on the effects of N deposition on N cycling, phytoplankton productivity, biodiversity and toxic alga blooms. 4.4.1. Effects on N cycling in aquatic and coastal ecosystems Nitrogen deposition has been confirmed to play a significant role in the N cycles of some marine/coastal and aquatic ecosystems in China. Total DIN fluxes of atmospheric input through wet and dry deposition were estimated to be 0.93 Tg N yr 1 in the South Yellow Sea (surface area 2.8 10 5 km 2 )and1.75tgnyr 1 in the East China Sea (surface area 7.7 10 5 km 2 ), values of which were as high as 0.94 and 1.21 times the riverine N inputs, respectively (Wan et al., 2002). In the surface waters of the West Yellow Sea atmospheric deposition was the main source of new nutrients and 65% of the total DIN input was from N deposition (Zhang and Liu, 1994). Bashkin et al. (2002) estimated that 1.06 Tg N yr 1 was deposited onto the Yellow and BohaiSeas (surface area 4.2 10 5 km 2 ), with nearly equal contributions from N wet and dry deposition found. In some riverine and aquatic ecosystems in China, N deposition has been found to be a large component of the total N input, sometimes accounting for most of the N input. For instance, total inorganic N flux from precipitation into the Yangtze, Yellow and Pearl River valleys were 3.75, 0.64 and 1.29 Tg N yr 1 with NH 4 en: NO 3 enratiosof3:1,3:1 and 4:1, respectively (Xing and Zhu, 2002). In Taihu Lake (surface area 2338 km 2 ), N flux of total wet N deposition was 2.75 10 3 Tg N yr 1 on average during 1987 and 1988, accounting for 6.7% of the total N input to the lake area (Sun and Huang, 1993); and these numbers increased to 6.56 10 3 Tg N yr 1 and 13.6% during 2002 and 2003 (Song et al., 2005). In addition, wet deposition of DON in Taihu Lake was 1.67 10 3 Tg N yr 1, contributing to about one quarter of the total N input from deposition (Song et al., 2005). High atmospheric N inputs have both positive and negative impacts on biological processes (e.g. new production, phytoplankton competitive interactions, and toxic algal blooms) depending on the nutrient situation in these ecosystems. 4.4.2. Effects on primary production Nitrogen input through atmospheric deposition has positive effects on the new production in China, especially in oligotrophic marine and aquatic ecosystems (e.g. central Yellow Sea). In-situ incubation experiments showed that phytoplankton species flourished in response to nutrient additions and chlorophyll-a increased significantly when rainwater was added (Zou et al., 2000). In the Yellow Sea, NO 3 en in wet deposition was estimated to account for about 4.3e9.2% of the NO 3 en requirement for the annual new production and three times higher production would be expected if dry NO 3 en deposition, and wet and dry NH 4 þ en deposition are included (Chung et al., 1998). In the east coastal Yellow Sea, about 115 mg C m 2 of new production would be expected annually from the input of nutrients in rainwater (Zou et al., 2000). NH 4 þ en appears to be the more important limiting nutrient compared to NO 3 en in rainwater and chlorophyll-a in NH 4 þ en incubations were twice as high as that in NO 3 en incubations (Zou et al., 2000). Insitu experiments showed that the potential impacts of NO 3 en and urea on chlorophyll-a concentration, primary productivity and their size structures were different and had seasonal variations (Zhu et al., 2008). However, primary production was most likely phosphorus (P)-limited in most of the Chinese estuaries and coastal waters (Huang et al., 1989; Zhang, 1994; Zou et al., 2001). For instance, in the coastal Yellow Sea and Jiaozhou Bay, phytoplankton was most likely P limited which appears to be a result of both low absolute PO 4 3e - concentration and a high N:P ratio. In these coastal eutrophic regions, the atmospheric wet deposition had limited impacts on production but will further increase the N/P ratio (Zou et al., 2000). Simulation experiments of NO 3 en addition in seawater showed an increase in the partial pressure of CO 2 in seawater, which would weaken the intensity of the carbon sink (Zhang et al., 2006). 4.4.3. Effects on phytoplankton diversity Phytoplankton competitive interactions and composition can be affected by N inputs including that from N deposition. In-situ experiments showed that NO 3 en addition had different effects on netphytoplankton, nanophytoplankton and picophytoplankton at different times of year (Wang and Jiao, 2002), indicating that additional N inputs may regulate the composition of the phytoplankton species. The species of planktonic diatoms and Shannon s index decreased (Qu et al., 2000) when N and P concentrations and the N:P ratios increased artificially. 4.4.4. Relationship to harmful alga bloom High atmospheric N and P deposition loads may supply nutrients for harmful algae during blooms in aquatic and marine/coast areas in China. Cyanobacterial blooms in Taihu Lake occurred at the end of April 2007 and had crucial impacts on the livelihoods of millions of local people, especially through effects on drinking water quality. Annual bulk deposition rates of total N and total P during 2007 in Taihu Lake were estimated to be 30 kg N ha 1 and 0.84 kg P ha 1, providing substantial nutrient inputs necessary for cyanobacterial blooms in northern Taihu Lake during summer and autumn (Zhai et al., 2009). Nitrogen sources (e.g. NO 3 en and NH 4 þ en) and N:P ratios were reported to have large impacts on the growth of toxic bloom-associated algae (e.g. Prorocentrum donghaiens and Phaeocystis globosa) (Ou and Lü, 2006; Cai et al., 2009). In oligotrophic regions of continental shelves (e.g. the Yellow Sea), episodic atmospheric wet deposition may supply the major nutrients (N and P) and trace elements necessary to stimulate harmful blooms (Zhang, 1994). Although nutrient input from atmospheric deposition (e.g. N, P and S) was much smaller than the riverine input in Changjiang River estuary, clear seasonal variations in wet deposition could have potential impacts on red tides in spring, summer and autumn (Zhang et al., 2003). 5. Empirical critical loads in different Chinese ecosystems Critical load is defined as a quantitative estimate of an exposure to one or more pollutants below which significant harmful effects on specified sensitive elements of the environment do not occur

2258 X. Liu et al. / Environmental Pollution 159 (2011) 2251e2264 according to present knowledge (UBA, 2004) and this has been used to negotiate emissions reductions (Hettelingh et al., 1995). Empirical critical loads have been well summarized for Europe and the United States (Bobbink et al., 2003) but there have been few studies in China. The empirical critical loads of some forests and grasslands are shown in Table 2. We realize that large amount of uncertainty exists as results of the very limited studies and short-term observations under high N rates in China. Some studies (e.g., the first three in Table 2) were carried out where historical N deposition has already been very high, and perhaps even higher than the actual critical load. The researchers may have therefore missed the opportunity to observe the ecosystem changes that had already occurred. In addition, the effects of N deposition might not be distinguished from other anthropogenic impacts such as soil acidification by higher sulfur deposition and soil degradation by over-grazing. However, the critical loads listed are a first attempt to summarize the available information in China. Empirical critical loads were determined by compiling reported field observations of detrimental ecological effects and by noting the deposition levels at which the effects occurred (Bobbink et al., 2002). In recent years N addition experiments have been carried out in China for some forests and grasslands (Table 2). Similar to experiments in Europe and North America, several plots were established in forests and/or grasslands and were treated with NO 3 en, NH 4 þ enorboth(nh 4 NO 3 ) at different doses (including no N addition as control). It was assumed that high N input would be necessary for significant impacts in the N addition experiments (Mo et al., 2006; Fan et al., 2007b,c; Lin et al., 2007). Therefore the minimum dose was usually equal to or several times higher than the estimated atmospheric N deposition. The experiments have been of at least one year s duration and some that are located in long-term research stations have already existed for more than five years and will continue in the future (Lin et al., 2007; Lu et al., 2007, in press; Wan et al., 2008). There are very limited studies on N effects in China and they all have been included in the present review. A range of critical loads was assigned to each ecosystem studied, with the higher limit equal to the lowest input level at which a response occurs, and the lower equal to the highest input level at which no significant effects occurred. It should be noted that the total N inputs include both the N added in fertilizer and also the N deposited from the atmosphere (Table 2). The total deposition was simulated through the USEPA CMAQ (Models-3/Community Multiscale Air Quality) model (V4.4) (Byun and Ching, 1999) in each 36 36 km 2 grid in a domain covering most of East Asia (Zhao et al., 2009a). As shown in Table 2, the critical loads of forests varied considerably from 10 to 30 kg N ha 1 yr 1 for temperate deciduous forests to 170e300 kg N ha 1 yr 1 for subtropical coniferous plantations; while they were also low (<100 kg N ha 1 yr 1 ) for temperate coniferous forests and subtropical forests (both coniferous and broad-leaved). The critical loads of grasslands varied from <50 kg N ha 1 yr 1 for typical temperate steppes and alpine steppes to 150e250 kg N ha 1 yr 1 for subtropical grasslands. Other grassland types with lower critical loads (<100 kg N ha 1 yr 1 ) include desert steppes and temperate dry grasslands (Table 2). The values of the critical loads above were based on the biological or chemical response of an ecosystem such as physiological variation, reduced biodiversity, elevated nitrate leaching, and changes in soil microorganisms, to varying level of N inputs. The critical loads of forests and grasslands obtained in China are much higher than the values of natural forests and grasslands in Europe, i.e., 10e15 and 10e30 kg N ha 1 yr 1, respectively (Bobbink et al., 2003). One explanation for the difference is that loads would be larger in subtropical ecosystems than in the temperate ecosystems of Europe. Other factors may include specific ecosystems (e.g., warm-humid subtropical forests having more capacity for cycling N inputs) and the shortcomings of experiment design that mentioned above. In a previous study critical loads of nutrient N were calculated and mapped using the steady state mass balance (SSMB) method (Duan et al., 2001). The SSMB method calculates the critical load of an ecosystem over the long-term based on defining acceptable values for fluxes out of the ecosystem (acceptable/critical N leaching or NO 3 en concentration in leachate) (UBA, 2004). The recommended values for European use were applied (UBA, 2004) because there has been no research on the critical limits of N in Chinese ecosystems. The range of values extracted from the previous map for each vegetation type is also shown in Table 2. The two sets of values showed positive correlation and were comparable for temperate deciduous forest, subtropical evergreen broadleaved forest, typical temperate steppe, and alpine steppe. However, the empirical critical loads were much higher than those calculated for other vegetation types. Although it is too early to doubt the applicability of critical limits on the basis of limited observations, the critical limits of N leaching need updating according to the empirical critical loads for future application of SSMB in China. One important reason for the large difference between the calculated critical loads and the empirical values may be underestimation of denitrification during the SSMB calculation for the subtropical forest ecosystems. As shown in Table 2, emissions of nitrous oxide (N 2 O) increased with enhanced N input in desert steppe (Shan, 2008). Preliminary measurements also indicated that the forest is a significant source of N 2 O(Chen and Mulder, 2007), an important greenhouse gas. In a subtropical forest in southwest China the observed net N uptake by the vegetation was found to be relatively small and most N was leached to the groundwater after nitrification of NH 4 þ en in the surface soil. However, in the stream only small fluxes of N were recorded in the annual runoff (Larseen et al., submitted for publication). This suggests that denitrification is important in removing N from the ecosystem in gaseous form. New criteria instead of critical leaching (e.g., critical N 2 O emission) may be appropriate. Based on the median of critical load range of each vegetation type in Table 2, a critical load map for N deposition in China was drawn (Fig. 3). The distribution of croplands which were insensitive to atmospheric N deposition and assigned a very high critical load of >200 kg N ha 1 yr 1, is also shown in the map. The empirical critical loads were lower in northwest China and higher in the southeast (Fig. 3). The lowest critical load of N occurred mainly in the northeast and in some part of the north, followed by northwest China, especially on the Qinghai-Tibetan Plateau and in the east of Inner Mongolia (Fig. 3). Nitrogen deposition in northeast and northwest China was commonly very low (e.g., <15 kg N ha 1 yr 1 ). In contrast, the critical loads of N in the southeast, where high N deposition existed, were relatively high (subtropical ecosystems in south China) to very high (agricultural ecosystems in the north). 6. Research needs and recommendations China s current economy (GDP) is eight times of that in the 1980s and is expected to continue to increase in the next few decades (Liu and Diamond, 2008). As a result, national anthropogenic N r emissions mainly from agricultural production and fossil fuel combustion are likely to increase substantially in the near future. On the other hand, the government is making more effort to control environmental pollution by improvement of air quality in mega cities as a result of hosting the 2008 Beijing Summer Olympic Games (Wang et al., 2009b; Chan and Yao, 2008). There are strong research needs to forecast future N r emission trends in China

X. Liu et al. / Environmental Pollution 159 (2011) 2251e2264 2259 Table 2 Summary of N critical loads for some forests and grasslands in China. Vegetation type Site Dominant species N deposition (kg N ha 1 yr 1 ) N input (kg N ha 1 yr 1 ) CL Main responses References (kg N ha 1 yr 1 ) a Forests Subtropical coniferous plantation Subtropical monsoon evergreen broad-leaved forest Shaxian, Sanming, Fujian (117 43 0 E, 26 31 0 N) Dinghushan Biosphere Reserve, Zhaoqing, Guangdong (1l2 33 0 E, 23 l0 0 N) Subtropical coniferous forest TieshanpingForest Park, Chongqing (106 41 0 E, 29 37 0 N) Subtropical evergreen broad-leaved forest LiangfengaoForest Park, Muchuan, Sichuan (103 47 0 E, 28 29 0 N) Temperate coniferous forest ChangbaishanForest Research Station, Jilin (127 42 0 E, 41 4l 0 N) Cunninghamia l anceolata Schima superba Temperate deciduous forest Fusong, Jilin (127 29 0 E, 42 20 0 N) Populus alba, Betula platyphyl Grasslands Typical temperate steppe Inner Mongolia Grassland Ecosystem Research Station (IMGERS), Xilin River Basin, Inner Mengonia (ll6 40 0 E, 43 32 0 N) Temperate dry grassland Yunwushan Grassland Natural Reserve, Ningxia Subtropical grassland Dongchuan Mudflow Monitoring Station, Xiaojiang River Basin, Yunnan Alpine meadow Maqu Grassland Resarch Station, Gansu (102 10 0 E, 34 01 0 N) Haibei Research Station for Alpine Meadow, Qinghai (37 37 0 N, 101 19 0 E) 53 120e240 170e300 (70e140) Decrease in litter decomposition and needle K, Ca, and Mg content 38 50e100 90e140 (30e70) Change in photosynthetic and physiologic characteristics of dominant understory species Pinus massoniana 42 <40 40e80 (15e30) N leaching, biomass decrease in ground vegetation Neolitsea aurata 18 <50 20e70 (30e70) Decrease in nutrient release from the litter and the decomposition of lignin and cellulose Fan et al., 2007a, b; Liu et al., 2008 Fang et al., 2005; Lu et al., 2006; 2007; Xu et al., 2005 Lin et al., 2007 Song et al., 2007a,b; Song et al., 2009 Pinus koraiensis 12 25e50 40e60 (15e30) Decrease in soil microorganism Zhao et al., 2008b, 2009b 7 0e25 10e30 (15e30) Decrease in soil microorganism Zhao et al., 2008b, 2009b Leymus chinensis 4 105e175 110e180 for degraded; < 50 for natural (15e30) Peak value reached for specific leaf area, leaf N content, and total chlorophyll content Thymus mongolicus 3 50e100 50e100 (15e30) Thymus mongolicus community replaced by Stipa hungeana Wan et al., 2008; Bai et al., 2010; Pan et al., 2004, 2005 Cheng et al., 1996 Heteropogon Contortus 4 150e250 150e250 (30e70) Gramineae dominant Zhang et al., 2004 Kobresia humilis 2 <150 <150 (15e30) Decrease in biodiversity Yao et al., 2009; Shen et al., 2002 Desert steppe Siziwangqi, Wulanchabu, Inner Stipa breviflora 4 <100 <100 (<15) Increase in N2O emission Shan, 2008 Mengonia (111 54 0 E, 41 47 0 N) Alpine steppe North to Qinghai Lake, Qinghai Stipapur purea 2 <50 <50 (15e30) Decrease in biodiversity Zhou et al., 2004 a Values in bracket are critical loads of nutrient N calculated by the steady state mass balance (SSMB) method (Duan et al., 2001).

2260 X. Liu et al. / Environmental Pollution 159 (2011) 2251e2264 Fig. 3. Estimated critical loads for N deposition in various ecosystems in China. The results are based on the data from Table 2 and the vegetation map of China. Blank means no data. considering all forms of trade-offs in N emission increase or decrease due to expanded intensive agriculture (e.g. improved nutrient management and change of land use) and wider application of innovative techniques in traffic and industrial emission reduction. It is difficult to evaluate the effect of pollution control measures on atmospheric N r (and other pollutants) concentrations and their deposition over the whole country without a national atmospheric deposition monitoring network. There is an urgent requirement to organize a long-term national deposition network (like the NADP) to monitor N wet and dry deposition across the country using uniform monitoring methods. Cross-site N addition experiments along with typical forests, grasslands and aquatic ecosystems are urgently required. Such long-term experiments could provide information on the impact of elevated N deposition on both terrestrial and marine ecosystems against the background of global change. Modeling tools are very useful for quantifying atmospheric N deposition (including both spatial and temporal variations) and its impact on natural and semi-natural ecosystems at different scales. For modeling approaches, the most important issue is qualitative assessment of results which is normally realized by comparing the simulated data with the measured data in the same region. It is crucial to reduce gaps between modeled and measured results by improving the understanding of atmospheric Nr emission, transport and deposition processes (e.g., by optimizing model parameters) in the future. Critical loads are useful to help propose N regulation strategies and decrease N r emissions. However, there are only limited short-term results on critical loads to soil acidification and biodiversity. We appeal for systematic and long-term field studies on critical levels and critical loads for grasslands, forests and aquatic ecosystems. International collaboration on N deposition measurements, ecosystem responses and modeling is also required. In addition, education is required to improve public awareness of environmental protection, especially atmospheric N r pollution and deposition. Nitrogen regulation tools and strategies should be recommended and taken into account when policy-makers consider the mitigation of anthropogenic N r emissions and their negative effects. In summary, N deposition has to some extent become an indicator of anthropogenic N r emissions induced by an expanding Chinese economy. Nitrogen deposition can lead to increases in environmental nutrient inputs in intensive agricultural ecosystems and also produce detrimental effects in many natural and seminatural ecosystems in less populated regions. We believe that heavy N deposition can be used as an important nutrient resource in croplands. On the other hand, the potential risk of N deposition on grasslands, forests and aquatic ecosystems should be controlled within acceptable levels (below critical loads) by substantially reducing the N r emissions to the environment. Acknowledgements We thank Dr. Peter Christie (Belfast, UK) for his linguistic corrections. This work was supported by One Hundred Person Project of the Chinese Academy of Sciences (304), the Innovative Group Grants from NSFC (30821003), the Program for New Century Excellent Talents in University (NCET-06-0111) and the National Natural Science Foundation of China (40771188, 41071151, 30970521). Appendix. Supplementary data Supplementary data associated with this article can be found, in the online version, at doi:10.1016/j.envpol.2010.08.002. References Aas, W., Shao, M., Jin, L., Larssen, T., Zhao, D.W., Xiang, R.J., Zhang, J.H., Xiao, J.S., Duan, L., 2007. Air concentrations and wet deposition of major inorganic ions at five nonurban sites in China, 2001e2003. Atmospheric Environment 41, 1706e1716. Aber, J.D., McDowell, W., Nadelhoffer, K., Magill, A., Berntson, G., Kamakea, M., McNulty, S., Currie, W., Rustad, L., Fernandez, I., 1998. Nitrogen saturation in temperate forest ecosystems: hypotheses revisited. Bioscience 48, 921e934. Aerts, R., Bobbink, R., 1999. Chapter 4: the impact of atmospheric nitrogen deposition on vegetation processes in terrestrial, non-forest ecosystems. In: Langan, Simon J. (Ed.), The Impact of Nitrogen Deposition on Natural and Seminatural Ecosystems. Kluwer Academic Publishers, The Netherlands, p. 88.