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Review

Nitrogen deposition and its ecological impact in China: An overviewq

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

aCollege of Resources and Environmental Sciences, China Agricultural University, Beijing 100193, ChinabDepartment of Environmental Science and Engineering, Tsinghua University, Beijing 100084, Chinac South China Botanical Garden, Chinese Academy of Sciences, Guangzhou 510650, ChinadCollege of Urban and Environmental Sciences, Peking University, Beijing 100871, ChinaeXinjiang Institute of Ecology and Geography, Chinese Academy of Sciences, Urumqi 830011, Chinaf Institute of Geographic Sciences and Natural Resources Research,Chinese Academy of Sciences, Beijing 100101, China

a r t i c l e i n f o

Article history:Received 20 April 2010Received in revised form31 July 2010Accepted 5 August 2010

Keywords:Atmospheric pollutionN emission and depositionCritical loadsEcological impact

a b s t r a c t

Nitrogen (N) deposition is an important component in the global N cycle that has induced large impactson the health and services of terrestrial and aquatic ecosystems worldwide. Anthropogenic reactive N(Nr) 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 impactsare of great concern since the 1980s. This paper synthesizes the data from various published papers toassess the status of the anthropogenic Nr emissions and N deposition as well as their impacts on differentecosystems, including empirical critical loads for different ecosystems. Research challenges and policyimplications on atmospheric N pollution and deposition are also discussed. China urgently needs toestablish national networks for N deposition monitoring and cross-site N addition experiments ingrasslands, forests and aquatic ecosystems. Critical loads and modeling tools will be further used in Nr

regulation.� 2010 Published by Elsevier Ltd.

1. Introduction

Nitrogen (N) deposition has been an important component in theglobal N cycle with increasing anthropogenic reactive N (Nr) emis-sions since the industrial revolution (Vitousek et al., 1997; Gallowayet al., 2008). Excess N deposition has aroused concerns about itsnegative impacts on ecosystem health and services such as loss ofbiodiversity (Sala et al., 2000; Stevens et al., 2004), eutrophicationand N saturation (Aber et al., 1998), soil acidification (Richterand Markewitz, 2001), and increased susceptibility to secondarystresses (Aerts and Bobbink, 1999; Witzell and Shevtsova, 2004).Rates of N deposition have leveled off or stabilized in the US andEurope since the late 1980s or early 1990s with the implementationof stricter legislation to limit atmospheric pollution (e.g. Gouldinget al., 1998; NADP, 2000). In contrast, emissions of Nr species inChina have been increasing continuously since the 1980smainly dueto growing agricultural and industrial activities (Klimont et al., 2001;Zhang et al., 2007, 2009a). These increased Nr emissions to theatmosphere have aroused widespread concern on air pollution in

China (Richter et al., 2005). Although there have been several Ndeposition monitoring programs and N deposition simulationexperiments 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 ofthemagnitude andpotential impacts of atmosphericNdeposition ondifferent ecosystems across China. In this review paper wesummarize all the published data fromN depositionmonitoring andmodeling, critical loads, and the effects on ecosystems in order to: 1)identify the magnitude and spatio-temporal variability of N depo-sition; 2) summarize themajor impacts of N deposition on terrestrialand aquatic ecosystems; 3) analyze the potential critical loads of Ndeposition to major Chinese ecosystems; 4) come up with recom-mendations for research on regulatory strategies for mitigation ofatmospheric Nr pollution and deposition in China.

2. Emissions and atmospheric concentrationsof Nr pollutants in China

2.1. Trends in Nr emissions to the atmosphere

There have been a number of studies on both reduced andoxidized N emissions in China (e.g. Kato and Akimoto, 1992; Sunand Wang, 1997; Tian et al., 2001; Streets et al., 2003; Yamaji

q All the co-authors contributed equally to this work.* Corresponding author.

E-mail address: [email protected] (X. Liu).

Contents lists available at ScienceDirect

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journal homepage: www.elsevier .com/locate/envpol

0269-7491/$ e see front matter � 2010 Published by Elsevier Ltd.doi:10.1016/j.envpol.2010.08.002

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, NH3 and NOx emis-sions have shown substantial increases since the early 1980s (Fig. 1a, b). Compared with 1980, NH3 emissions (13.7 Tg N yr�1) haddoubled and NOx emissions (6.0 Tg N yr�1) had increased by a factorof 4 by the year 2005 (Zhao et al., 2009a). The rapid increases inboth Nr species emissions are closely related to intensive agricul-tural and industrial activities. The majority of anthropogenic NH3 inChina is emitted fromN fertilizers (e.g. ammonium bicarbonate andurea) and animal/human excreta (Zhao and Wang, 1994; Zhanget al., 2010a). NOx emissions are derived mainly from fossil fuelcombustion processes including power plants, transportation andindustry (Streets and Waldhoff, 2000; Streets et al., 2003) withminor contributions from lightning, biomass burning and arablesoils (Yan et al., 2003). If this trend continues we can expect the Nremission induced deposition to make a larger contribution to acidrain 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 NOx and NH4þeN

and NO3�eN in PM10 (particulatematter smaller than 10microns) or

total suspended particulates (TSP) in urban and rural areas showheavy Nr pollution in major Chinese mega cities (Chan and Yao,2008) and in some rural or suburban regions such as the NorthChina Plain and Taihu Lake Plain (Ju et al., 2009). Concentrations ofmajor gaseous and particulate N pollutants during 1999 and 2009are summarized in Table 1.

2.2.1. NH3 concentrationsNH3 is the most abundant basic gas in the atmosphere. It can

react with acidic gases (e.g. H2SO4, HNO4 and HCl) to formsecondary particles and can also return to the land surface by drydeposition not far from the emission sources (Asman et al., 1998;Erisman and Schaap, 2004). There have been few atmosphericNH3monitoring results available in China so far. NH3 concentrationswere higher in agricultural and urban regions with large populationdensity in the order North China > South China > Northwest Chinaand the Qinghai-Tibetan Plateau (Table 1). For example, annualmean NH3 concentrations at two agricultural sites on the NorthChina Plainwere 13.5 and 9.5 mg Nm�3 (Shen et al., 2009). Cao et al.(2009) and Meng et al. (2010) reported similar NH3 concentrationsin Xi’an, Shaanxi province and Houma, Shanxi province, respec-tively. The high NH3 concentrations in North and South China areconsistent with the high emission densities in these regions (Wanget al., 1997; Zhang et al., 2010a). The regional background NH3concentrations (1.5e3.4 mg N m�3) in China (Meng et al., 2010) arealso relatively high compared with those at other remote sitesworldwide (e.g. 0.1e1 mg N m�3).

2.2.2. NO2 concentrationsNO2 is one of the major precursors that contribute to acid

deposition and plays an important role in the formation of tropo-spheric ozone by photochemical reactions with non-methanehydrocarbons (NMHC) (Fowler et al., 1998). NO2 concentrationshave been routinely monitored in almost every mega city in Chinasince the late 1990s. According to the 2008 Report on the State ofthe Environment, NO2 concentrations were higher in mega cities(e.g. Beijing, Tianjin, Shanghai and Chongqing) than in cities withlower populations in Western China (e.g. Guizhou, Yunnan andInner Mongolia). These monitoring results are consistent withspace observations on NO2 by satellite (Richter et al., 2005). Inrecent years vehicle numbers have increased rapidly in the megacities and emissions from vehicles are considered to be the largestsource of NOx in some mega cities such as Tianjin (Zhao and Ma,2008). Fossil fuel combustion for power generation and industryare also important sources of NOx. High NO2 concentrations werenot confined to urban air but were also found in some rural regionswith highly-developed economies. For example, annual NO2concentrations were 9.3 mg N m�3 at Quzhou on the North ChinaPlain (Shen et al., 2009) and 12.8 mg N m�3 at Jurong in the YangtzeRiver Delta (Su et al., 2009). This is also reflected in the relative highNO2 concentrations at regional background sites on the NorthChina Plain and in the Yangtze River Delta (Meng et al., 2010).

0

1

2

3

4

5

6

7

1975 1980 1985 1990 1995 2000 2005 2010

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

ON

xs

noi

ssi

me

a

Year

0

3

6

9

12

15

1975 1980 1985 1990 1995 2000 2005 2010

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.

b

HN

3s

noi

ssi

me

Year

Fig. 1. Total anthropogenic NOx (a) and NH3 (b) emissions (Tg N yr�1) in China during1980 and 2007. Data sources are from various published references as shown in theFigure.

Table 1Reported concentrations of major gaseous and particulate Nr pollutants in Chinaduring 1999 and 2009 (mg N m�3).

Region NH3 NO2 HNO3 pNH4þ pNO3

�

North China 1.5e44.7 2.3e23.8 0.1e0.6 4.0e13.5 1.0e7.8South China 0.5e22.8 0.6e35.8 0.1e1.4 0.8e6.5 0.1e3.7Northwest 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) pNH4þ and pNO3

�dparticulate ammonium andnitrate; 3) Data sources: published journal papers, Ph.D. and M.Sc. theses andgovernmental reports (see Supporting online material).

X. Liu et al. / Environmental Pollution 159 (2011) 2251e22642252


Increased NOx emissions from vehicles and small factories andrelatively high NOx emissions from agricultural fields can accountfor the high NO2 concentrations in rural areas (Fang et al., 2007,2009a). In contrast, NO2 concentrations (0.4e2.3 mg N m�3) werevery low at regional background sites in Northeastern China(Longfengshan) and on the Qinghai-Tibetan Plateau (Waliguan)(Meng et al., 2010).

2.2.3. HNO3 concentrationsAtmospheric HNO3 is very important in N or acid deposition and

atmospheric chemistry, but information on HNO3 concentrations inChina is very limited. According to a few measurements, HNO3concentrations at province scale seem to coincide with NO2concentrations. Hu et al. (2008) reported high HNO3 concentration(1.4 mg Nm�3) in summer at a coast site in the Pearl River Delta. Aaset al. (2007) also reported high HNO3 concentrations at rural sitesin Chongqing andHunan province in South China. Shen et al. (2009)found a mean HNO3 concentration of 0.6 mg N m�3 at a rural site onthe North China Plain. These limited data suggest that atmosphericHNO3 concentrations in China are relatively low compared witheither NH3 or NO2 (Table 1).

2.2.4. Particulate NH4þ and NO3

� concentrationsThough PM10 has been routinely measured in most of the cities

in China, measurements of particulates NH4þ and NO3

� are limited.Most of the measurements of particulates NH4

þ and NO3� have been

conducted at urban sites, especially mega cities, to study thechemical components of PM10 and/or PM2.5. Both particulates NH4

þ

and NO3� concentrations were high in mega cities and in North

China comparedwith those in South China (Table 1). Relatively highconcentrations of particulates NH4

þ and NO3� at rural sites in China

(Aas et al., 2007; Cao et al., 2009; Shen et al., 2009) implied highrates of N deposition in natural and agricultural ecosystems of theseregions. Measurements of particulates NH4

þ and NO3� 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 monitoringnetworks for N deposition currently exist in China, there have beena number of studies on acid rain or wet deposition since the early1980s. From1981 to 1983 the ChineseNational Environment Bureauorganized a nationwide campaign for acid rain measurement (Zhaoand Sun, 1986). Similar campaigns were also carried out during theearly and late 1990s. Since 1999 four Chinese cities (Xiamen, Xi-An,Chongqing and Zhuhai) have joined the Acid DepositionMonitoringNetwork in East Asia (EANET) (http://www.eanet.cc/product.html).Currently both the Chinese Ministry of Environmental Protectionand the National Meteorological Bureau have been running twoindependent atmospheric deposition monitoring networks sincethe late-1990s, both of which include more than 300 sites acrossChina (Ding et al., 2004; Anon., 2010). The sites of the formernetwork are mainly distributed around urban areas and the latterare sparsely located in rural or background regions. More recently,China Agricultural University has organized a Nationwide NitrogenDeposition Monitoring Network (NNDMN) since 2004 (Liu andZhang, 2009). This network contains about 40 monitoring sitesacross China, covering cropland, grassland, forest, and urbanecosystems. There are several additional sub-national networks onN deposition monitoring, including the Integrated MonitoringProgram on Acidification of Chinese Terrestrial Ecosystems(IMPACT) (Tang et al., 2001) and the World Meteorological Organi-zation Global Atmosphere Watch Precipitation Chemistry Program

(WMO/GAW) of China (http://cdc.cma.gov.cn/index.jsp). Thesenetworks tend tomonitor wet deposition (N input from rainfall andsnowfall) but the NNDMN also measures dry deposition of variousNr species (Shen et al., 2009). All these networks have differentmethodologies and quality control systems. For example, theNNDMN cooperates closely with Centre for Ecology & Hydrology(CEH) Edinburgh, UK and runs the same quality control system asCEH. While the WMO/GAW of China adopts the methods used byWMO protocols. In general, the monitoring networks in China lacklong-term continuous data and uniformmeasuring methods whichshould be improved in the future.

3.2. Patterns of nitrogen deposition in China

Using current N deposition monitoring networks and publisheddata, Lü and Tian (2007) reported the spatial pattern of N wet anddry deposition. They found that average N wet plus dry depositionwas 12.35 Tg N yr�1 during 1993 and 2003. This estimate iscomparable to the total estimated emissions of NOx and NHy(14.67 Tg N yr�1) in 2000 (Streets et al., 2003), but much greaterthan 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 deposi-tion. The main reasons for underestimation are that dissolvedorganic N (DON) in wet deposition and NH3, HNO3 and particulatesNH4

þ and NO3� (pNH4

þ and pNO3�) in dry deposition were not taken

into account in their study. According to Zhang et al. (2008a), DONin 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 Lakeregion. Shen et al. (2009) found that NH3, HNO3 and pNH4

þ andpNO3

� were important components of dry deposition on the NorthChina Plain, contributing to 57% of total dry deposition. Liu andZhang (2009) re-estimated the total N deposition in China to be15 Tg N yr�1 in the 2000s compared with 7.4 and 12.0 Tg N yr�1 inthe 1980s and 1990s, respectively. Therefore, the current total Nwet and dry deposition rates are likely to be higher than15 Tg N yr�1 if all the Nr species (e.g. DON, NH3, HNO3 and pNH4

þ

and pNO3�) are included in the overall N deposition budgets.

Because of the factors outlined above (especially the incompletemeasurement of N dry deposition), most published studies havegreater wet than dry deposition. For example, Wang et al. (2008a)observed that wet deposition in the loess area accounted for over90% of the total deposition. In some forest areas, however, drydeposition was greater than wet deposition. In broadleaf forests,dry deposition can explain 67e75% of the total N deposition (Fanet al., 2007a; Hu et al., 2007). Dry deposition fluxes throughoutChina from 1990 to 2003 increased slightly on average, with anapparent increase in the Southeast and Southwest, and declineswere marked in the Northeast and in parts of the Midlands (Lü andTian, 2007).

3.3. Simulations of N deposition in China

Zhao et al. (2009a) simulated N deposition using the Models-3/CommunityMultiscale Air Quality (CMAQ) system (V4.4) (Byun andChing, 1999). This model has previously been modified and provento be suitable for Chinese regional and urban-scale air qualitysimulations (Streets et al., 2007; Wang et al., 2008b). The drivingmeteorological inputs are provided by the fifth-generation NCAR/Penn State Mesoscale Model (MM5). To validate the reliability ofCMAQ modeling, the simulated results were compared withobservation values for annual wet S and N deposition from twoprograms: IMPACTS and EANET (http://www.eanet.cc/product/datarep/).

X. Liu et al. / Environmental Pollution 159 (2011) 2251e2264 2253


The model calculated N concentration and deposition in each36 � 36 km2 grid in a domain covering most of East Asia, based onthe latest Chinese emission inventory of NOx, NH3, SO2, and VOCs(Wei et al., 2008; Zhao et al., 2008a, 2009a) and other relatedemissions inventories outside China (e.g., Streets et al., 2003). TotalN flux was calculated as the sum of nitrate, NOx and ammonium,including wet and dry deposition. Northern and eastern Chinawererecognized as the regions receiving the highest N deposition, witha substantial contribution from the high density of energyconsumption and emissions. However, some other modelingresults showed that total N wet and dry deposition rates peakedover the central south China (Lü and Tian, 2007). This is still anopen question for the spatial pattern of N deposition in Chinabecause of the lack of systematic monitoring data in the country.

Holland et al. (1999) compared pre-industrial and contemporaryglobal distribution of N deposition by a global three-dimensionchemical transport model, MOGUNTIA, at a 10� � 10� resolution.The Eastern U.S.A., Western Europe and Southern Asia (includingsoutheastern China) were three hotspots of high N deposition.Dentener et al. (2006b) assessed global N deposition using differentstate-of-the-art global atmospheric chemistry models at 2e3�

resolutions, which further consolidated the high N deposition insoutheast China. Continental scale models of East Asia have hori-zontal resolutions of 0.5e1� or 40e45 km. Most studies havefocused on S deposition or acid deposition, including oxidized S andN (Hao et al., 2001; Halloway et al., 2002; Park et al., 2005). Wanget al. (2008b) and Hayami et al. (2008) carried out modelemodelcomparisons between eight atmospheric transport models (ATMs),and some of these models covered both NHx and NOy deposition.High N deposition often occurred in the eastern part of China (Lüand Tian, 2007). Zhang (2009) constructed N budgets and spatialvariation of N deposition on the North China Plain (NCP) by an ATM,the FRAMEmodel. 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, agreeingwell with directly measured results in this region (e.g. Zhang et al.,2008b; Shen et al., 2009; He et al., 2007, 2010). Deposition of thereduced N species dominated the NCP budget with 2.7 Tg N, 2.5times more N than oxidized N deposition, suggesting a largecontribution from agricultural sources (Zhang, 2009).

4. Impacts of simulated N deposition on different ecosystemsin 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 ecosys-tems have becomemore prominent over the last decade. Numeroussimulated N deposition experiments have been conducted indifferent grassland and forest ecosystems in China (e.g., Mo et al.,2006; Fan et al., 2007b; Xia et al., 2009). Fig. 2 summarizes thelocations of N addition experiments conducted in forest, grasslandand desert ecosystems across China since the late 1990s.

4.1. Forests

China has a large territory with great climatic complexity andspatial variation which sustains a variety of forest ecosystemsranging from boreal forests in the north to tropical rain forests inthe south. These forests play an important role in maintainingbiodiversity and ecological equilibrium and in providing servicesfor social development. High atmospheric N deposition has beenreported in many forest ecosystems (Zhou and Yan, 2001; Zhanget al., 2006; Chen and Mulder, 2007; Hu et al., 2009) with Ndeposition levels commonly exceeding 20 kg N ha�1 yr�1 in centraland east China above which forest health will be seriously

threatened (MacDonald et al., 2002; Bobbink et al., 2010). However,studies on the responses of forest ecosystems to elevated N depo-sition only started in China in the 2000s and the response recordsare far from complete. A Sino-Norwegian project IMPACTS,launched in 1999, has established monitoring sites at five forestecosystems in the south of China to monitor atmospheric N depo-sition and its effects on leaching of N (Chen et al., 2004). To betterunderstand and predict the effects of N deposition on forestecosystems, DingHuShan Forest Ecosystem Long-term NitrogenResearch Project (DHSLTNR) was established. The experimentcovered three tropical/subtropical forest sites (an old-growthmonsoon evergreen broadleaf forest, a mixed pine and broadleafforest and a pine forest) in 2002 (Mo et al., 2006). This experimentaimed to study the responses of forest ecosystems to elevated Ndeposition. Since then, many N addition experiments have beenestablished in different forest types from south to north China(Fig. 2). Here we mainly focus on the effects of N addition on Ndynamics in forest soils, soil acidification, plant growth, biodiver-sity, litterfall decomposition and flux of greenhouse gases.

4.1.1. Effects on N dynamics and soil acidificationIt is well established that elevated N deposition will alter N

cycling greatly in forest ecosystems. Observations in tropical andtemperate forests showed that N inputs to forest floors can affectthe status of N in underlying forest soils and increase soil availableN content (Fan et al., 2007b,c; Hu et al., 2009; Lu et al., 2009; Xuet al., 2009; Fang et al., 2009a, in press). Nitrogen leaching willoccur if N input exceeds soil retention capacity. In five subtropicalforested catchments of South China, Chen et al. (2004) found that Nleaching from soils was high and almost all N leached as NO3

�eN inhigh N deposition sites (35 kg N ha�1 yr�1). Fang et al. (2009a) alsofound that experimental N additions (50e150 kg N ha�1 yr�1)increased dissolved inorganic N (DIN) leaching in all threesubtropical forests, with 25e66% of added N leached over a 3-yearperiod. In addition, leaching of dissolved organic N (DON) increasedunder high N input conditions (Fang et al., 2009b).

Elevated N deposition can increase the acidification potentialof forest ecosystems and decrease soil buffering capacity (Vogt

Fig. 2. Distribution of N addition experiments in forest, grassland and desert ecosys-tems 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; Tuet 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) andpersonal communications (Du E.Z., Lu X.K., Li K.H., Jin G.Z., per. com.).



et al., 2006; Lu et al., 2009). In a mature subtropical forest ofsouthern China, Lu et al. (2009) found that the ecosystem wassensitive to high N addition, and continuous two-year N additions(50e150 kg N ha�1 yr�1) significantly enhanced soil acidification,Al3þ mobilization, and leaching of base cations from soils, thetypical characteristics of N saturation (Aber et al., 1998). Fan et al.(2007b) also found that exchangeable base cations (e.g. Ca2þ andMg2þ) decreased with increasing N addition in a subtropicalChinese fir plantation after three years of N manipulation(60e240 kg N ha�1 yr�1). In addition, the nitrate leached out wasaccompanied by positively charged “counter-ions”, the basecations Kþ, Ca2þ and Mg2þ, resulting in the further acidification ofthe leached soil, or hydrogen and aluminum ions, which maycause the acidification of receiving systems.

4.1.2. Effects on plant growthChanges in N cycling and soil quality induced by N deposition

will affect plant growth. There have been only a few studies on theresponses of plant growth to N addition. These responses have beenfound to be related to plant demand for N. In N-limited forests Ndeposition can meet plant demand for N and improve plantnutrient status, with an increased photosynthetic capacity andstimulation of plant growth. However, excess N will result innutrient imbalance in trees, disturb N metabolism, reduce netphotosynthesis, and restrict plant growth (Li et al., 2005; Lu et al.,2006, 2007; Mo et al., 2008a). By studying the seedling growthresponse of two tropical tree species (Schima superba and Crypto-carya concinna) to N addition, Mo et al. (2008a) found that netphotosynthetic rate and biomass production were increased bylower N addition (50e100 kg N ha�1 yr�1) but both species werenegatively affected by higher N addition. In an N-saturated tropicalforest, Lu et al. (2006, 2007) found that N addition significantlyincreased foliar N in three understory plants and additional Naccumulated as organic N (e.g. free amino acids) and N additioninhibited net photosynthetic capacity in these plants. The negativeeffect of N addition was also found on the belowground biomassproduction in the same forest, which showed that fine root biomassdecreased significantly with increasing levels of N addition (Moet al., 2008b). In a 3-year study, Duan et al. (2009) observed thatelevated CO2 concentration and N deposition increased biomassaccumulation but the responses varied among tree species.Nitrogen deposition stimulated aboveground biomass accumula-tion with decreasing root: shoot ratio and elevated CO2 concen-tration significantly increased biomass allocation to belowgroundor aboveground biomass of different tree species.

4.1.3. Effects on biodiversityBiodiversity can be significantly affected by N addition (Clark

and Tilman, 2008). Although the impacts of N deposition onbiodiversity have generated wide international interest, thereare only a few long-term research projects in China (e.g., theDingHuShan Long-Term Nitrogen Research or DHSLTNR) (Luet al., 2008). These studies show that N deposition can alterspecies diversity, and excessive N can reduce species diversity.Studies in an old-growth tropical forest of southern Chinashowed that high N addition levels (e.g. >100 kg N ha�1 yr�1)significantly reduced understory species diversity and themechanism for change appeared to be N deposition-mediatedsoil acidification rather than fertilization (Lu et al., in press). Inthe same forest, Xu et al. (2006) found that N addition alsosignificantly decreased soil fauna diversity. However, somestudies have shown that species richness was increased by low Naddition levels. For example, Xu et al. (2005) found that 16-months of N addition (50e100 kg N ha�1 yr�1) increased soilfauna diversity in the pine forest of Dinghushan. These changes

may be related to soil N status, vegetation composition and timeof N addition. A few studies have shown that soil microorganismsare very sensitive to N addition. Xue et al. (2007) found thatshort-term N addition significantly increased soil bacterialnumbers but decreased soil fungal counts. A lower soil N statusand different species requirements for N may be attributable toshifts in functional microbial communities.

4.1.4. Effects on litter decompositionNitrogen has long been recognized as an important factor

regulating litter decomposition. Litter decomposition is a key stepin nutrient cycling. Studies on the effects of N addition on litterdecomposition have shown different results, ranging from positiveto negative responses of the decomposition rate and elementalrelease, 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 subtropicalforests, Mo et al. (2006) found that litter decomposition ratesexhibited no significant positive and even some negative responsesto N addition in N-saturated mature forests, but showed significantpositive responses in N-limited disturbed and rehabilitated forests.Nitrogen deposition has significant cumulative effects on litterdecomposition and the initial effects of N addition change overtime (Fang et al., 2007). Effects of N addition on litter decomposi-tion also varied depending on the nutrient status of the litter (Moet al., 2008c).

4.1.5. Effects on the fluxes of greenhouse gasesFew studies have been conducted on the effects of elevated N

deposition on the fluxes of greenhouse gases (GHGs, CO2, CH4, andN2O) from forest soils, and the response pattern depends on foresttype, N status of the soil, and the level of N deposition. In N-satu-rated mature forests, high N addition can reduce soil respiration(CO2 emissions) (Mo et al., 2008b), increase soil N2O emissions(Zhang et al., 2008c), and decrease CH4 uptake rates (Zhang et al.,2008d). However, studies in N-limited pine forests show that Naddition has no significant effects on soil respiration or CH4 uptakein spite of increased N2O emissions as result of high N addition (Moet al., 2007b; Zhang et al., 2008c,d).

4.2. Grasslands

Grasslands in China cover about 40% of the total territory anddiverse grassland types range from vast, continuous areas oftemperate grasslands in arid and semi-arid regions to alpinegrasslands on the Tibetan Plateau, and to small areas of the warmtemperate and tropical regions (Chen and Chen, 2007). The vastarea and wide distribution of Chinese grasslands, however, havebeen largely ignored in terms of global N deposition. Most studieson the responses of grassland ecosystems to N deposition havebeen restricted to the semi-arid temperate steppes (e.g. the InnerMongolia grasslands) in northern China. Current studies includeresponses of plant growth, biodiversity, ecosystem carbonexchange, and N transformation processes.

4.2.1. Effects on plant growthSeveral studies have found that N is amajor limiting factor in the

growth of grasslands (Xia et al., 2009). Nitrogen addition maylessen the N limitation by increasing soil N availability and thusstimulate plant growth (Pan et al., 2004; Bai et al., 2010). In Leymuschinensis communities in typical steppe of Inner Mongolia, Panet al. (2005) found that N addition increased significantly thedensity, height, aboveground biomass, belowground biomass andtotal biomass of Leymus chinensis. With the greater increase ofaboveground biomass than belowground biomass, N addition



significantly reduced the root: shoot ratio of the affected plants(Pan et al., 2005; Zhang et al., 2010b). Therefore, N addition not onlyincreased the biomass of Leymus chinensis population but alsochanged the resource partitioning among plant parts. Plant func-tional traits were greatly changed by N addition at the same time.Observations in a mature typical steppe ecosystem showed that Naddition led to increased foliar N concentration, total chlorophyllcontent, and specific leaf area (SLA) in these plants to some extent,but plant species differed significantly in their responses toincreased N addition rates (Wan et al., 2008; Huang et al., 2009). Inaddition, added N can change productivity and competitive balancebetween C3 (Leymus chinensis) and C4 (Chloris virgata) plants. Ley-mus chinensis accumulated more biomass under N treatments andthe final biomass of Chloris virgata was less impacted than Leymuschinensis (Niu et al., 2008). In a 4-year N addition study, Xia et al.(2009) found that the gross ecosystem productivity increased inspite of the reduction in the biomass of forbs.

4.2.2. Effects on biodiversityPlant diversity commonly responds to N addition as a loss of

richness accompanied by a shift in plant functional group compo-sition (Huang et al., 2009; Bai et al., 2010). N enrichment allowsspecies with acquisitive resource-use strategies to exclude thosewith conservative resource-use strategies without other stresses(e.g. drought). As found by Bai et al. (2010) in a mature communityof Inner Mongolia grasslands, N addition led to a large reduction inspecies richness accompanied by increased dominance of earlysuccessional annuals and loss of perennial grasses and forbs at all Ninput rates. In addition, with increasing N uptake, long-term Naddition would decrease N re-sorption proficiency and make somedominant species less dependent on N re-sorption, while othernutrients (e.g., P) or resources (e.g., water and light) could becomelimiting factors for plant growth (Huang et al., 2009). As a result,interspecific differences in N re-sorptionmay influence the positivefeedback between species dominance and N availability.

Reduced microbial diversity (e.g. functional diversity) has beenobserved in a semi-arid temperate steppe in response to high levelsof N addition but a low level of N addition stimulated microbialdiversity (Zhang et al., 2008e). An N-addition-mediated decline inpH could be largely responsible for the decline in diversity. Zhanget al. (2008e) suggested a critical N loading level between therates of 160 and 320 kg N ha�1 yr�1 in the temperate steppe,beyond which the microbial community would shift its response toN addition. However, this high critical load may only occur inseverely degraded steppe because of soil N deficiency. Furtherstudies are required in other grasslands.

4.2.3. Effects on net ecosystem carbon exchangeNet ecosystem C exchange (NEE) represents the balance

between gross ecosystem productivity (GEP) and ecosystemrespiration (ER). It is addressed well in the changes in NEE responseto N addition in a temperate steppe in northern China. N additionhas been observed to increase NEE in the two hydrologically con-trasting growing seasons because of the larger stimulation in GEPthan in ER (Xia et al., 2009). However, the magnitude of N stimu-lation on NEE declined over time. Niu et al. (2010) considered thatthe temporal decline in the N stimulation of NEE resulted from anN-induced shift in species composition. In any case, the positiveeffects of N on NEE suggest that N deposition under global climatechange may increase C sequestration in the temperate steppe (Niuet al., 2009, 2010). In addition, the interactive effects of water/warming and N addition on NEEwere also studied. The N responsesof NEE under warming were suppressed (Xia et al., 2009), but thereare no additive effects of water and N addition on NEE (Niu et al.,2009).

4.2.4. Effects on N transformation processesAs a part of N cycling, N addition significantly impacts N trans-

formation processes (N mineralization and nitrification). Nitrogenaddition increased soil inorganic N (ammonium and nitrate N) andrates of net N mineralization and nitrification, but reduced soilmicrobial 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 NH4

þeN, NO3�eN, net N

mineralization and nitrification in the early growing season. Ina 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) stimulated Nmineralization but high N addition (280 kg N ha�1 yr�1) inhibited Nmineralization. The mechanism behind these should be clarified inthe future.

4.3. Croplands

Almost all croplands can be regarded as N-saturated ecosystemsdue to continual application of chemical N fertilizers. Nitrogendeposition onto croplands is relatively low compared with Nfertilizer inputs. However, N deposition may make a substantialcontribution to nutrient budgets in some intensively managedcroplands (He et al., 2010) or in some low input agroecosystems.Furthermore, heavy N deposition is normally associated with acidrain and atmospheric Nr pollution, both of which could lead to yielddecline in croplands.

4.3.1. Nutrient contribution of N deposition to croplandsNitrogen is an essential nutrient for plant growth and NH4

þeN,NO3

�eN are two major forms of N taken up by plants. Nitrogendeposition (mainly as NH4

þeN and NO3�eN) is important N source in

agroecosystems (He et al., 2007). Zhang (2006) reported up to58 kg N ha�1 yr�1 wet deposition of NH4

þeN and NO3�eN in the

highly-developed Shanghai region. He et al. (2010) found that totalairborne N inputs to a maizeewheat rotation system on the NCPranged from 99 to 117 kg N ha�1 yr�1, while plant available N fromdeposition 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 uptakeby the two crop species. Based on some earlier measurements, Zhu(1997) estimated approximately 1.5 Tg annual N input from wetdeposition plus irrigation to croplands of China in the 1980s. Liuand Zhang (2009) estimated total N input from wet and drydeposition and found 15 Tg N yr�1 for the whole of China in the2000s, equal to about 50% of N fertilizer use in the country.Assuming that one third of N deposition occurred in the agricul-tural areas, 5.0 Tg annual N deposition may have returned tocroplands of China since the 2000s. This is an important nutrientinput to croplands from the atmosphere.

4.3.2. Negative effects of N deposition on croplandsHigh N addition level of N deposition especially acid deposition

might produce some negative impacts in agroecosystems. Earlierstudies on the effect of acid deposition on croplands were carried outin the late 1970s especially in south China (Zhang et al., 1997; Hu andSu, 1999). High N deposition together with acid rain might decreaseyield and quality of crops by influencing soil properties and soilmicroorganisms (Lin, 2005). The impact of N deposition on soilsdepends on the form of the deposition, the N demands/requirementsof crops, and soil chemistry, mineralogy and moisture status(Hornung,1995). Increased N deposited above background levels willinitially be utilized, resulting in stimulation in plant growth.However, increased N availability can eventually result in induceddeficiencies of other nutrients (Zhou and Ogura, 1996; Qiu et al.,1997). Deposited NH4

þeN can be nitrified to NO3�eN, resulting in



enhanced leaching of base cations and soil acidification (Xu et al.,2002; Guo et al., 2010). Increased N availability can also lead tofrost and drought sensitivity and increased risk of pest and diseaseinfestation (Xu, 1995). Moreover, high NH3 concentrations anddeposition induced by excess N fertilization may cause NH3 toxicityto crops, with symptoms of injury showing in the leaves (Fangmeieret al., 1994). Anthropogenic Nr pollution and deposition may result inyield loss and quality decline of crops. But these negative effects oncroplands have been observed mainly in ‘hotspots’ of N deposition(e.g. areas close to intensive livestock farms, N fertilizer factories orpower plants).

4.4. Aquatic and coastal ecosystems

China has more than 1500 rivers with watersheds larger than1000 km2 andmore than 2800 lakes with surface areas greater than1 km2. The Bohai Sea, Yellow Sea, East China Sea and South ChinaSea cover a total area of about 4.7 million km2. Nitrogen emissionsand deposition are expected to be high in coastal provinces withrapid economic growth. However, only a few studies on N deposi-tion have been reported in the Yellow Sea, East China Sea and BohaiSea and in Taihu Lake (e.g., Zhang and Liu, 1994; Wan et al., 2002;Song et al., 2005). Current studies are mainly focused on theeffects of N deposition on N cycling, phytoplankton productivity,biodiversity and toxic alga blooms.

4.4.1. Effects on N cycling in aquatic and coastal ecosystemsNitrogen deposition has been confirmed to play a significant role

in the N cycles of some marine/coastal and aquatic ecosystems inChina. Total DIN fluxes of atmospheric input through wet and drydeposition were estimated to be 0.93 Tg N yr�1 in the South YellowSea (surface area 2.8 � 105 km2) and 1.75 Tg N yr�1 in the East ChinaSea (surface area 7.7� 105 km2), values of which were as high as 0.94and 1.21 times the riverine N inputs, respectively (Wan et al., 2002).In the surface waters of the West Yellow Sea atmospheric depositionwas themain source of new ‘nutrients’ and 65% of the total DIN inputwas from N deposition (Zhang and Liu, 1994). Bashkin et al. (2002)estimated that 1.06 Tg N yr�1 was deposited onto the Yellow andBohaiSeas (surface area 4.2 � 105 km2), with nearly equal contribu-tions from N wet and dry deposition found. In some riverine andaquatic ecosystems in China, N deposition has been found to bea large component of the total N input, sometimes accounting formost of the N input. For instance, total inorganic N flux fromprecipitation into the Yangtze, Yellow and Pearl River valleys were3.75, 0.64 and 1.29 Tg N yr�1 with NH4eN: NO3eN ratios of 3:1, 3:1and 4:1, respectively (Xing and Zhu, 2002). In Taihu Lake (surface area2338 km2), N fluxof total wet N depositionwas 2.75�10�3 Tg N yr�1

on average during 1987 and 1988, accounting for 6.7% of the total Ninput to the lake area (Sun and Huang, 1993); and these numbersincreased 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 Lakewas 1.67 � 10�3 Tg N yr�1, contributing to about one quarter of thetotal N input from deposition (Song et al., 2005). High atmospheric Ninputs have both positive and negative impacts on biologicalprocesses (e.g. new production, phytoplankton competitive interac-tions, and toxic algal blooms) depending on the nutrient situation inthese ecosystems.

4.4.2. Effects on primary productionNitrogen input through atmospheric deposition has positive

effects on the ‘new production’ in China, especially in oligotrophicmarine and aquatic ecosystems (e.g. central Yellow Sea). In-situincubation experiments showed that phytoplankton species flour-ished in response to nutrient additions and chlorophyll-a increasedsignificantly when rainwater was added (Zou et al., 2000). In the

Yellow Sea, NO3�eN in wet deposition was estimated to account for

about 4.3e9.2% of the NO3�eN requirement for the annual new

production and three times higher productionwould be expected ifdry NO3

�eN deposition, and wet and dry NH4þeN deposition are

included (Chung et al., 1998). In the east coastal Yellow Sea, about115 mg C m�2 of new production would be expected annually fromthe input of nutrients in rainwater (Zou et al., 2000). NH4

þeNappears to be the more important limiting nutrient compared toNO3

�eN in rainwater and chlorophyll-a in NH4þeN incubations were

twice as high as that in NO3�eN incubations (Zou et al., 2000). In-

situ experiments showed that the potential impacts of NO3�eN and

urea on chlorophyll-a concentration, primary productivity andtheir size structures were different and had seasonal variations(Zhu et al., 2008). However, primary production was most likelyphosphorus (P)-limited inmost of the Chinese estuaries and coastalwaters (Huang et al., 1989; Zhang, 1994; Zou et al., 2001). Forinstance, in the coastal Yellow Sea and Jiaozhou Bay, phytoplanktonwas most likely P limited which appears to be a result of both lowabsolute PO4

3e- concentration and a high N:P ratio. In these coastaleutrophic regions, the atmospheric wet deposition had limitedimpacts on production but will further increase the N/P ratio (Zouet al., 2000). Simulation experiments of NO3

�eN addition inseawater showed an increase in the partial pressure of CO2 inseawater, which would weaken the intensity of the carbon sink(Zhang et al., 2006).

4.4.3. Effects on phytoplankton diversityPhytoplankton competitive interactions and composition can be

affected by N inputs including that from N deposition. In-situexperiments showed that NO3

�eN addition had different effects onnetphytoplankton, nanophytoplankton and picophytoplankton atdifferent times of year (Wang and Jiao, 2002), indicating thatadditional N inputs may regulate the composition of the phyto-plankton species. The species of planktonic diatoms and Shannon’sindex decreased (Qu et al., 2000) when N and P concentrations andthe N:P ratios increased artificially.

4.4.4. Relationship to harmful alga bloomHigh atmospheric N and P deposition loads may supply nutri-

ents for harmful algae during blooms in aquatic and marine/coastareas in China. Cyanobacterial blooms in Taihu Lake occurred at theend of April 2007 and had crucial impacts on the livelihoods ofmillions of local people, especially through effects on drinkingwater quality. Annual bulk deposition rates of total N and total Pduring 2007 in Taihu Lake were estimated to be 30 kg N ha�1 and0.84 kg P ha�1, providing substantial nutrient inputs necessary forcyanobacterial blooms in northern Taihu Lake during summer andautumn (Zhai et al., 2009). Nitrogen sources (e.g. NO3

�eN andNH4

þeN) and N:P ratios were reported to have large impacts on thegrowth of toxic bloom-associated algae (e.g. Prorocentrum dong-haiens 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 nutri-ents (N and P) and trace elements necessary to stimulate harmfulblooms (Zhang, 1994). Although nutrient input from atmosphericdeposition (e.g. N, P and S) was much smaller than the riverineinput in Changjiang River estuary, clear seasonal variations in wetdeposition 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 exposureto one or more pollutants below which significant harmful effectson specified sensitive elements of the environment do not occur



according to present knowledge” (UBA, 2004) and this has beenused to negotiate emissions reductions (Hettelingh et al., 1995).Empirical critical loads have been well summarized for Europe andthe United States (Bobbink et al., 2003) but there have been fewstudies in China.

The empirical critical loads of some forests and grasslands areshown in Table 2.We realize that large amount of uncertainty existsas results of the very limited studies and short-term observationsunder high N rates in China. Some studies (e.g., the first three inTable 2) were carried out where historical N deposition has alreadybeen very high, and perhaps even higher than the actual criticalload. The researchersmay have thereforemissed the opportunity toobserve the ecosystem changes that had already occurred. Inaddition, the effects of N deposition might not be distinguishedfrom other anthropogenic impacts such as soil acidification byhigher sulfur deposition and soil degradation by over-grazing.However, the critical loads listed are a first attempt to summarizethe available information in China.

Empirical critical loads were determined by compiling reportedfield observations of detrimental ecological effects and by notingthe deposition levels at which the effects occurred (Bobbink et al.,2002). In recent years N addition experiments have been carriedout in China for some forests and grasslands (Table 2). Similar toexperiments in Europe and North America, several plots wereestablished in forests and/or grasslands and were treated withNO3

�eN, NH4þeN or both (NH4NO3) at different doses (including no

N addition as control). It was assumed that high N input would benecessary for significant impacts in the N addition experiments(Mo et al., 2006; Fan et al., 2007b,c; Lin et al., 2007). Therefore theminimum dose was usually equal to or several times higher thanthe estimated atmospheric N deposition. The experiments havebeen of at least one year’s duration and some that are located inlong-term research stations have already existed for more than fiveyears 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 allhave been included in the present review. A range of critical loadswas assigned to each ecosystem studied, with the higher limit equalto the lowest input level at which a response occurs, and the lowerequal to the highest input level at which no significant effectsoccurred. It should be noted that the total N inputs include both theN added in fertilizer and also the N deposited from the atmosphere(Table 2). The total deposition was simulated through the USEPACMAQ (Models-3/Community Multiscale Air Quality) model (V4.4)(Byun and Ching, 1999) in each 36 � 36 km2 grid in a domaincovering most of East Asia (Zhao et al., 2009a).

As shown in Table 2, the critical loads of forests varied consid-erably from 10 to 30 kg N ha�1 yr�1 for temperate deciduous foreststo 170e300 kg N ha�1 yr�1 for subtropical coniferous plantations;while they were also low (<100 kg N ha�1 yr�1) for temperateconiferous forests and subtropical forests (both coniferous andbroad-leaved). The critical loads of grasslands varied from<50 kgN ha�1 yr�1 for typical temperate steppes and alpine steppesto 150e250 kg N ha�1 yr�1 for subtropical grasslands. Othergrassland types with lower critical loads (<100 kg N ha�1 yr�1)include desert steppes and temperate dry grasslands (Table 2). Thevalues of the critical loads above were based on the biological orchemical response of an ecosystem such as physiological variation,reduced biodiversity, elevated nitrate leaching, and changes in soilmicroorganisms, to varying level of N inputs.

The critical loads of forests and grasslands obtained in China aremuch higher than the values of natural forests and grasslands inEurope, i.e., 10e15 and 10e30 kg N ha�1 yr�1, respectively (Bobbinket al., 2003). One explanation for the difference is that loads wouldbe larger in subtropical ecosystems than in the temperate

ecosystems of Europe. Other factors may include specific ecosys-tems (e.g., warm-humid subtropical forests having more capacityfor cycling N inputs) and the shortcomings of experiment designthat mentioned above.

In a previous study critical loads of nutrient N were calculatedand mapped using the steady state mass balance (SSMB) method(Duan et al., 2001). The SSMB method calculates the critical load ofan ecosystem over the long-term based on defining acceptablevalues for fluxes out of the ecosystem (acceptable/critical Nleaching or NO3

�eN concentration in leachate) (UBA, 2004). Therecommended values for European use were applied (UBA, 2004)because there has been no research on the critical limits of N inChinese ecosystems. The range of values extracted from theprevious map for each vegetation type is also shown in Table 2. Thetwo sets of values showed positive correlation and were compa-rable for temperate deciduous forest, subtropical evergreen broad-leaved forest, typical temperate steppe, and alpine steppe.However, the empirical critical loads were much higher than thosecalculated for other vegetation types.

Although it is too early to doubt the applicability of critical limitson the basis of limited observations, the critical limits of N leachingneed updating according to the empirical critical loads for futureapplication of SSMB in China. One important reason for the largedifference between the calculated critical loads and the empiricalvalues may be underestimation of denitrification during the SSMBcalculation for the subtropical forest ecosystems. As shown in Table2, emissions of nitrous oxide (N2O) increased with enhanced Ninput in desert steppe (Shan, 2008). Preliminary measurementsalso indicated that the forest is a significant source of N2O (Chenand Mulder, 2007), an important greenhouse gas. In a subtropicalforest in southwest China the observed net N uptake by the vege-tation was found to be relatively small and most N was leached tothe groundwater after nitrification of NH4

þeN in the surface soil.However, in the stream only small fluxes of N were recorded in theannual runoff (Larseen et al., submitted for publication). Thissuggests that denitrification is important in removing N from theecosystem in gaseous form. New criteria instead of critical leaching(e.g., critical N2O emission) may be appropriate.

Based on the median of critical load range of each vegetationtype in Table 2, a critical load map for N deposition in China wasdrawn (Fig. 3). The distribution of croplands which were insensitiveto atmospheric N deposition and assigned a very high critical loadof >200 kg N ha�1 yr�1, is also shown in the map. The empiricalcritical loads were lower in northwest China and higher in thesoutheast (Fig. 3). The lowest critical load of N occurred mainly inthe northeast and in some part of the north, followed by northwestChina, especially on the Qinghai-Tibetan Plateau and in the east ofInner Mongolia (Fig. 3). Nitrogen deposition in northeast andnorthwest Chinawas commonly very low (e.g.,<15 kg N ha�1 yr�1).In contrast, the critical loads of N in the southeast, where high Ndeposition existed, were relatively high (subtropical ecosystems insouth 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 the1980s and is expected to continue to increase in the next fewdecades (Liu and Diamond, 2008). As a result, national anthropo-genic Nr emissions mainly from agricultural production and fossilfuel combustion are likely to increase substantially in the nearfuture. On the other hand, the government is making more effort tocontrol environmental pollution by improvement of air quality inmega cities as a result of hosting the 2008 Beijing Summer OlympicGames (Wang et al., 2009b; Chan and Yao, 2008). There are strongresearch needs to forecast future Nr emission trends in China



Table

2Su

mmaryof

Ncritical

load

sforsomeforestsan

dgrasslan

dsin

China.

Veg

etationtype

Site

Dom

inan

tsp

ecies

Ndep

osition

(kgN

ha�

1yr

�1)

Ninput

(kgN

ha�

1yr

�1)

CL

(kgN

ha�

1yr

�1)a

Mainresp

onses

Referen

ces

Forests

Subtropical

coniferou

splantation

Shax

ian,S

anming,

Fujian

(117

� 430E,

26� 310N)

Cunn

ingh

amia

lan

ceolata

5312

0e24

017

0e30

0(70e

140)

Decreasein

litterdecom

position

andnee

dle

K,C

a,an

dMgco

ntent

Fanet

al.,20

07a,

b;Liuet

al.,20

08Su

btropical

mon

soon

evergree

nbroa

d-lea

vedforest

Dingh

ush

anBiosp

hereReserve

,Zh

aoqing,

Guan

gdon

g(1l2

� 330E,

23� l00N)

Schima

supe

rba

3850

e10

090

e14

0(30e

70)

Chan

gein

photosyn

thetic

andphysiologicch

aracteristics

ofdom

inan

tunderstorysp

ecies

Fanget

al.,20

05;

Luet

al.,20

06;20

07;

Xuet

al.,20

05Su

btropical

coniferou

sforest

Tieshan

pingF

orestPa

rk,

Chon

gqing(106

� 410E,

29� 370N)

Pinu

smassonian

a42

<40

40e80

(15e

30)

Nleaching,

biom

assdecrease

ingrou

ndve

getation

Linet

al.,20

07

Subtropical

evergree

nbroa

d-lea

vedforest

Lian

gfen

gaoF

orestPa

rk,

Much

uan

,Sichuan

(103

� 470E,

28� 290N)

Neo

litseaau

rata

18<50

20e70

(30e

70)

Decreasein

nutrientrelease

from

thelitteran

dthe

decom

position

oflig

nin

andcellu

lose

Songet

al.,20

07a,b;

Songet

al.,20

09

Temperateco

niferou

sforest

Chan

gbaish

anFo

rest

Resea

rch

Station,Jilin(127

� 420E,

41� 4l0 N

)Pinu

sko

raiensis

1225

e50

40e60

(15e

30)

Decreasein

soilmicroorga

nism

Zhao

etal.,20

08b,

2009

b

Temperatedeciduou

sforest

Fusong,

Jilin

(127

� 290E,

42� 200N)

Popu

lusalba

,Be

tula

platyp

hyl

70e

2510

e30

(15e

30)

Decreasein

soilmicroorga

nism

Zhao

etal.,20

08b,

2009

b

Grassland

sTy

pical

temperatestep

pe

Inner

Mon

golia

Grassland

Ecosystem

Resea

rchStation(IMGER

S),

XilinRiver

Basin,Inner

Men

gonia

(ll6

� 400E,

43� 320N)

Leym

usch

inen

sis

410

5e17

511

0e18

0for

deg

raded

;<

50for

natural(15e

30)

Peak

valuereached

forsp

ecific

leaf

area

,lea

fN

content,

andtotalch

lorophyllco

ntent

Wan

etal.,20

08;

Bai

etal.,20

10;

Panet

al.,20

04,2

005

Temperatedry

grasslan

dYunwush

anGrassland

NaturalR

eserve

,Ningx

iaTh

ymus

mon

golicus

350

e10

050

e10

0(15e

30)

Thym

usmon

golicus

community

replacedby

Stipahu

ngea

naChen

get

al.,19

96

Subtropical

grasslan

dDon

gchuan

Mudflow

Mon

itoring

Station,X

iaojiangRiver

Basin,

Yunnan

Heterop

ogon

Contortus

415

0e25

015

0e25

0(30e

70)

Graminea

edom

inan

tZh

anget

al.,20

04

Alpinemea

dow

Maq

uGrasslandResarch

Station,

Gan

su(102

� 100E,

34� 010N)

Haibe

iResea

rchStationforAlpine

Mea

dow

,Qingh

ai(37�37

0 N,1

01� 190E)

Kob

resiahu

milis

2<15

0<15

0(15e

30)

Decreasein

biod

iversity

Yao

etal.,20

09;

Shen

etal.,20

02

Desertstep

pe

Siziwan

gqi,W

ulanch

abu,Inner

Men

gonia

(111

� 540E,

41� 470N)

Stipabreviflora

4<10

0<10

0(<

15)

Increa

sein

N2O

emission

Shan

,200

8

Alpinestep

pe

North

toQingh

aiLake

,Qingh

aiStipap

urpu

rea

2<50

<50

(15e

30)

Decreasein

biod

iversity

Zhou

etal.,20

04

aValues

inbracke

tarecritical

load

sof

nutrientN

calculatedby

thestea

dystatemassba

lance

(SSM

B)method

(Duan

etal.,20

01).



considering all forms of trade-offs in N emission increase ordecrease due to expanded intensive agriculture (e.g. improvednutrient management and change of land use) and wider applica-tion of innovative techniques in traffic and industrial emissionreduction.

It is difficult to evaluate the effect of pollution controlmeasures onatmospheric Nr (and other pollutants) concentrations and theirdeposition over the whole country without a national atmosphericdeposition monitoring network. There is an urgent requirement toorganize a long-term national deposition network (like the NADP) tomonitor N wet and dry deposition across the country using uniformmonitoring methods. Cross-site N addition experiments along withtypical forests, grasslands and aquatic ecosystems are urgentlyrequired. Such long-term experiments could provide information onthe impact of elevated N deposition on both terrestrial and marineecosystems against the background of global change.

Modeling tools are very useful for quantifying atmospheric Ndeposition (including both spatial and temporal variations) and itsimpact on natural and semi-natural ecosystems at different scales.For modeling approaches, the most important issue is qualitativeassessment of results which is normally realized by comparing thesimulated data with the measured data in the same region. It iscrucial to reduce gaps between modeled and measured results byimproving the understanding of atmospheric Nr emission, trans-port and deposition processes (e.g., by optimizing model parame-ters) in the future. Critical loads are useful to help propose Nregulation strategies and decrease Nr emissions. However, there areonly limited short-term results on critical loads to soil acidificationand biodiversity. We appeal for systematic and long-term fieldstudies on critical levels and critical loads for grasslands, forests andaquatic ecosystems. International collaboration on N depositionmeasurements, ecosystem responses andmodeling is also required.In addition, education is required to improve public awareness ofenvironmental protection, especially atmospheric Nr pollution anddeposition. Nitrogen regulation tools and strategies should berecommended and taken into account when policy-makersconsider the mitigation of anthropogenic Nr emissions and theirnegative effects.

In summary, N deposition has to some extent become an indi-cator of anthropogenic Nr emissions induced by an expandingChinese economy. Nitrogen deposition can lead to increases inenvironmental nutrient inputs in intensive agricultural ecosystemsand also produce detrimental effects in many natural and semi-natural ecosystems in less populated regions. We believe thatheavy N deposition can be used as an important nutrient resourcein croplands. On the other hand, the potential risk of N depositionon grasslands, forests and aquatic ecosystems should be controlledwithin acceptable levels (below critical loads) by substantiallyreducing the Nr emissions to the environment.

Acknowledgements

We thank Dr. Peter Christie (Belfast, UK) for his linguisticcorrections. This work was supported by One Hundred PersonProject of the Chinese Academy of Sciences (304), the InnovativeGroup Grants from NSFC (30821003), the Program for New CenturyExcellent Talents in University (NCET-06-0111) and the NationalNatural Science Foundation of China (40771188, 41071151,30970521).

Appendix. Supplementary data

Supplementary data associated with this article can be found, inthe online version, at doi:10.1016/j.envpol.2010.08.002.

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