Observation of the nitrogen deposition in the lower Liaohe River Plain, Northeast China and assessing its ecological risk

Atmospheric Research 101 (2011) 460–468

Contents lists available at ScienceDirect

Atmospheric Research j o u r n a l h o m e p a g e : w w w. e l s ev i e r. c o m / l o c a t e / a t m o s

Observation of the nitrogen deposition in the lower Liaohe River Plain, Northeast China and assessing its ecological risk W.T. Yu a, C.M. Jiang a,⁎, Q. Ma a, Y.G. Xu a, H. Zou a, S.C. Zhang b a b

Laboratory of Nutrients Recycling, Institute of Applied Ecology, Chinese Academy of Sciences, Shenyang 110016, PR China Northeast Institute of Geography and Agricultural Ecology, Chinese Academy of Sciences, Changchun, 130012, PR China

a r t i c l e

i n f o

Article history: Received 19 November 2010 Received in revised form 12 April 2011 Accepted 19 April 2011 Keywords: N deposition NH4+/NO3− ratio Precipitation Agroecosystems Ecological disturbance Critical loads Lower Liaohe River Plain, Northeast China

a b s t r a c t Little information about the nitrogen (N) deposition in the lower Liaohe River Plain (LRP) of Northeast China was available. A continuous 5-year observation (from 2004 to 2008) was conducted to determine the nitrogen deposition in precipitation collected by a standard rain gage, and then we assessed its ecological consequences. The mean inorganic nitrogen concentrations were 1.82 mg N L−1 for NH4+–N and 0.89 mg N L−1 for NO3−–N. On an annual basis, the averaged amounts of N deposition were 14.5 kg N ha−1 year−1, which were much lower than that of hotspots in China. However, the amount of deposited N should still be taken account into the N fertilizer management of the agroecosystems. Compared with the critical loads (CL) of N deposition, the deposited N would likely threaten the natural ecosystems in LRP. The N concentrations in rainwater and the amount of N deposition showed a clear seasonal pattern, and precipitation played an important role in regulating the N concentration. Furthermore, the seasonal fluctuation of NH4+/NO3− ratio reflected that the deposited N originated from both fertilizer application and fossil fuel combustion, and environmental factor (soil temperature, lightning and sunshine) might also contribute to the seasonal cycles of NH4+/NO3− ratio. The high annual NH4+/NO3− ratio (2.05) compared with the more developed region suggested that N deposition in this region was mostly affected by agricultural activities rather than industrial activities. Crown Copyright © 2011 Published by Elsevier B.V. All rights reserved.

1. Introduction Humans have intensively altered the global nitrogen (N) cycle due to intensification of agriculture activity and increase of fossil-fuel consumption (Galloway et al., 2004; Galloway et al., 2008). These activities would lead to an acceleration of reactive N emission, and subsequently increase global N deposition (Vitousek et al., 1997). Reactive N creation increased from 15 Tg N in 1860 to 156 Tg N in 1995 and it increased further to 187 Tg N in 2005 (Galloway et al., 2008). The deposition of reactive N has also more than doubled over the last 100 years and the increasing trend is considered to be enhanced in the next few decades (Galloway et al., 2004). Although N is an essential plant nutrient, excessive reactive N input to terrestrial ecosystems could cause a lot of ecological risk (Matson et al., 2002), such as ⁎ Corresponding author. Tel.: + 86 24 8937 0421; fax: + 86 24 8397 0366. E-mail addresses: [email protected], [email protected] (C.M. Jiang).

eutrophication of water bodies (Gao et al., 2007), decrease of plant diversity (Bobbink et al., 2010; Stevens et al., 2010), alteration of greenhouse gas flux (Jiang et al., 2010; Lund et al., 2009) and so on. Consequently, observation of N deposition is an important issue for minimizing its influences on sensitive natural ecosystems and improving N use efficiency in agricultural ecosystems (Shen et al., 2009). Determination of inorganic N (NH4+–N and NO3−–N) in precipitation is an effective approach to evaluate the N deposition status, which has been widely adopted in Europe and America (Holland et al., 2005). NH4+ in rainwater originates from NH3, the major sources of which are animal waste and fertilizer application; while the formation of NO3− is considerably more complex. NO3− is an end product of a series of gas-phase photochemical and heterogeneous reactions involving N oxides which are primarily derived from the combustion of fossil fuels (Gao et al., 2007; Hertel et al., 2006; Rojas and Venegas, 2010). Investigation of N deposition has been carried out in many areas

0169-8095/$ – see front matter. Crown Copyright © 2011 Published by Elsevier B.V. All rights reserved. doi:10.1016/j.atmosres.2011.04.011

W.T. Yu et al. / Atmospheric Research 101 (2011) 460–468

of China (Lu and Tian, 2007). However, only limited information about N deposition in precipitation of the lower Liaohe River Plain (LRP) in Northeast China has been reported according to the best of our knowledge. The LRP is located at central part of Liaoning Province, with relatively large population density, highly intensive agricultural activity and high level of economic development in Northeast China. This research could provide great insights to reasonable application of N fertilizer and ecological protection in LRP. The specific objectives of this study were: (1) to quantify the concentrations of NH4+–N and NO3−–N in precipitation and determine the amount of N deposition; (2) to characterize the temporal variation of NH4+/NO3− ratio and elucidate its ecological implication; (3) to evaluate potential ecological consequences of the N loads. 2. Materials and methods 2.1. Research site The monitoring site was established at the central part of LRP (Fig. 1): Shenyang Experimental Station (41°32′ N latitude, 123°23′ E longitude and at an altitude of 31 m above sea level), Chinese Academy of Sciences. Geographically, the station is located at the intersection of the two transects driven by hydrologic factor from east to west and by thermic factors from north to south in China, and is the regional representative site in Chinese Ecosystem Research Network (CERN). This site was mainly affected by warm-temperate continental monsoon climate featuring four distinct seasons. The mean annual temperature of 7.5 °C (maximum, 39.3 °C; minimum, −33.1 °C), the frost-free period is 147–164 days, and the average

461

annual precipitation is about 520 mm. The monitoring site is surrounded by paddy field and dry farmland with corn or soybean. The cropping system is single harvest per year. This site is about 35 km south of Shenyang and 40 km north of Liaoyang, Benxi and Anshan, which are all typical heavy industry cities in Northeast China. The northwest of the site is plain, and the southeast is mountainous for 50 km around (Fig. 1). 2.2. Sampling and chemical analysis Due to the severe weather condition in winter, we did not collect precipitation sample in January, February and December of every year. Because the site was in typical monsoon climate zone, there was almost no precipitation in the above dry period. This scheme of sampling collection had very small influence on the estimation of whole annual N deposition from precipitation. At the monitoring site, a rain gage (SDM6, Tianjin Weather Equipment Inc., China) was installed to collect rainwater samples. After each rain event, the amount of rainwater was recorded, and the samples were thoroughly mixed and stored in plastic bottles then immediately stored at the refrigerator of −20 °C until analysis of NH4+–N and NO3−–N concentrations, which could avoid transformation of inorganic N to organic N by microorganisms (Cornell et al., 2003). The rain gage consists of a 20 cm diameter, funnel-shaped top connected to an inner collecting tube. The gage stands 60 cm from the ground to avoid splash from the ground surface and to minimize the impact of wind turbulence. The plastic bottles were cleaned with acid wash of 10% HCl solution, then cleaned with deionized water to avoid contamination (Anderson and Downing, 2006), and kept in plastic bags until they were used for sample collection. For the rain gage, the funnel-shaped top,

Sampling site

Fig. 1. Location of the sampling site.

462

W.T. Yu et al. / Atmospheric Research 101 (2011) 460–468

March of 2004, 2005 and 2006, in November of 2005 and 2008, and in April and October of 2004. We thus have not collected rain samples in these periods.

the inner collecting tube and all equipment in contact with the samples were cleaned using the same above method just after each sample collection. Because the rain collector is a standard rain gage rather than a wet-only, event based system, the inorganic N deposition we quantified was the bulk deposition (wet deposition plus sedimenting dry deposition such as dust during dry periods). The NH4+–N and NO3−–N concentrations of rainwater was colorimetrically analyzed using a continuous flow analyzer (TRAACS 2000, Bran-Luebbe Inc., Germany) (Kamphake et al., 1967; Searle, 1984).

3.2. Seasonal variation of NH4+–N and NO3−–N concentrations and deposition Fig. 2b showed that the monthly volume-weighted concentrations of inorganic N varied dramatically with time, and the lower values occurred between June and September, which was opposite to the trend of precipitation. This phenomenon indicates that there is a dilution effect of rainwater on inorganic N concentration. Furthermore, if the mean NH4+–N or NO3−–N concentrations of each rainfall event were grouped by the precipitation amount, we found inorganic N concentration gradually increased with the decrease of precipitation amount and the increasing degree was more dramatic when rainfall was below 20 mm (Fig. 3a, b). Therefore, lower rainfall is an efficient removal process for scavenging NH4+ and NO3− in the atmosphere. Additionally, we divided one year's period into dry (October to May of next year) and wet season (June to September), and found that the proportion of light precipitation event (b20 mm) to the total precipitation event in dry season was much greater than that in wet season (92% for dry, 57% for wet) (Fig. 3c). This distribution pattern of precipitation implies that suspended particle in the air is an important contribution to NH4+–N and NO3−–N in rainwater, which may mostly occur in dry season, because the residence time of the suspended particles in the air tend to be longer during the dry season and substantial atmospheric particles carrying the N compounds are accumulated (Huang et al., 2010), and the higher N concentrations in rainwater are subsequently formed by both the in-cloud and below-cloud scavenging mechanisms (Hertel et al., 2006; Zhang et al., 2008). In our observation period, N deposition exhibited clear seasonal pattern, and more than 70% of deposited N we quantified occurred between June and September (Fig. 2c), which was in accordance with the seasonal variation pattern of rainfall (Fig. 2a). In addition, we found that the yearly amount of N deposition tended to correspond with that of precipitation (with the exception of the years 2006 and 2007) (Table 1). These results suggest that precipitation is an important factor influencing the wet inorganic N deposition (Anderson and Downing, 2006; Zhao et al., 2009). Because the precipitation fluctuates much more between months than between years and wet deposition is strongly dependent on rain, monthly variability of N wet deposition is usually larger than its yearly variability in our observation (CV for moth was

2.3. Calculation of N deposition The monthly and annually volume-weighted concentrations (C′) of NH4+–N or NO3−–N (mg N L −1) were calculated according to the data from the corresponding month or year. The volume-weighted concentrations were calculated using the equation: 0

C =



  n  n ∑ Ci Pi = ∑ Pi

i=1

i=1

where Pi is the rainwater amount (mm) in the precipitation event i, Ci is the measured concentration (mg N L−1) of NH4+–N or NO3−–N, and n is the number of precipitation events. The monthly and yearly wet deposition (WD) of NH4+–N or NO3−–N (kg N ha −1) was according to the following formula: WD = 10

−2

n

∑ Ci Pi

i=1

where n is the number of precipitation events in the monthly and yearly period, Pi is the rainwater amount (mm) in the precipitation event i, Ci is the corresponding concentration (mg N L −1) of NH4+–N or NO3−–N in rainwater, and 10 −2 is a unit conversion factor. 3. Results and discussion 3.1. Seasonal variation of precipitation In the observation period, the yearly precipitation ranged from 446.8 mm to 612.2 mm, with the average of 532.1 mm (Table 1). The precipitation showed clear seasonal pattern, and most of the precipitation concentrated between June and September (Fig. 2a), because this site was in typical monsoon climate zone. The largest monthly precipitation was 270 mm in July of 2004. In contrast, there was no precipitation in

Table 1 Concentrations of NH4+–N and NO3−–N in precipitation and N deposition from 2004 to 2008 in LRP. Year

2004 2005 2006 2007 2008 Average n number of sample.

Precipitation

446.8 612.2 576.1 494.4 530.9 532.1

n

22 29 38 27 28

N concentration (mg N L−1)

N deposition (kg N ha−1 year−1)

NH4+–N

NO3−–N

NH4+/NO3−

NH4+–N

NO3−–N

Total

1.21 2.14 1.35 2.32 2.10 1.82

0.62 0.85 0.78 1.03 1.16 0.89

1.93 2.52 1.73 2.25 1.82 2.05

5.4 13.1 7.7 11.5 11.2 9.8

2.8 5.2 4.5 5.1 6.1 4.7

8.2 18.3 12.2 16.6 17.3 14.5

W.T. Yu et al. / Atmospheric Research 101 (2011) 460–468 300

463

a

Precipitation

Precipitation (mm)

250

200

150

100

50

0

b

NO3-N

N concentration (mg N L-1)

Volume-weighted

NH4-N NH4/NO3

10

8

15 6 10

4

NH4/NO3 ratio

20

2

5

0 0 6

c

NO3-N

-1

N deposition (kg N ha )

NH4 _N 5 4 3 2 1 0 2004-1-1 2004-7-1 2005-1-1 2005-7-1 2006-1-1 2006-7-1 2007-1-1 2007-7-1 2008-1-1 2008-7-1

Date Fig. 2. Monthly variation of precipitation (a), volume-based concentration of inorganic N and NH4+/NO3− ratio (b), and amount of inorganic N bulk deposition (c) from 2004 to 2008.

70%; CV for year was 29%) (CV, coefficient of variation: defined as Standard Deviation / Mean × 100%). This result demonstrates that the monthly information on N deposition is more important than the yearly information for clarifying the deposition status and evaluates the subsequent ecological effects (Lu and Tian, 2007). 3.3. Yearly NH4+–N and NO3−–N concentrations and N deposition Yearly average concentrations were 1.82 mg N L −1 for NH4+–N (ranged from 1.21 mg N L−1 to 2.32 mg N L−1) and 0.89 mg N L−1 for NO3−–N (ranged from 0.62 mg N L −1 to 1.16 mg N L−1) (Table 1). NH4+–N and NO3−–N concentrations in precipitation in LRP are much higher than those (0.02– 0.05 mg N L−1 for NH4+–N and 0.02–0.08 mg N L−1 for NO3−–N)

at remote areas across the world (Galloway et al., 1982) and those (0.36 mg N L−1 for NH4+–N and 0.10 mg N L−1 for NO3−–N) in Linzhi, a pristine area on the Qinghai–Tibetan Plateau (Jia et al., 2009). The magnitudes of inorganic N concentration are comparable to the recently reported values (1.20 mg N L−1 for NH4+–N and 0.97 mg N L−1 for NO3−–N) in the Yangtze River Delta region, a rapidly developing area located in Southern China (Zhao et al., 2009). Nevertheless, because of smaller rainfall in LRP, the amount of N deposition was only about half of that in the Yangtze River Delta region (26.8 kg N ha−1 year−1) (Zhao et al., 2009), indicating less serious N pollution than that in Southern China with high level of economic development. Generally, NH4+/NO3− ratio in rainwater can be used as an evaluation of industrialization degree (Zhao et al., 2009). Yearly mean NH4+/NO3− ratio varied from 1.73 to 2.52, with an average value of 2.05 in our observation periods (Table 1). In

W.T. Yu et al. / Atmospheric Research 101 (2011) 460–468

Mean NH4-Nconcentration (mg N L-1)

464

a 3

2

1

0 0-20

20-40

40-60

60-80

Mean NO3-Nconcentration (mg N L-1)

Precipitation (mm)

b 1.5

1.0

0.5

0.0 0-20

20-40

40-60

60-80

Precipitation (mm)

Number of events

100

c

Heavy precipitation (>20mm) Light precipitation (<20mm)

75

50

Fenn et al., 2003), across European countries (about 1.6) (Holland et al., 2005) and the more developed regions of East China (Table 2). Therefore, we could speculate that with rapid growth of the economy and continuing industrialization and urbanization in LRP, the NH4+/NO3− ratio might exhibit a clear decreasing trend in near future (Akimoto, 2003; Jia and Chen, 2009). Based on N concentrations and precipitation, we could estimate the amount of N deposition. From 2004 to 2008, yearly N deposition varied from 5.4 kg N ha−1 to 13.1 kg N ha−1 for NH4+–N (with the average of 9.8 kg N ha −1), from 2.8 kg N ha−1 to 6.1 kg N ha−1 for NO3−–N (with the average of 4.7 kg N ha−1) and from 8.2 kg N ha−1 to 18.3 kg N ha−1 for total inorganic N (with the average of 14.5 kg N ha−1) (Table 1). Holland et al. (1999) estimated pre-industrial inorganic-N deposition in Northern Hemisphere, ranging from only 0.4 to 1.0 kg N ha−1 year−1. The Qinghai–Tibetan Plateau, which is very far away from the industrial area and remains relatively undisturbed by humans' activities, receives atmospheric N deposition as low as 2.4 kg N ha−1 year−1 (Jia et al., 2009). Compared with the above data, the present results indicate that the N deposition in LRP has been intensively influenced by humans' activities. Across Europe and the United States, the networks monitoring regional precipitation chemistry have been carried out for more than 30 years. During 2003 to 2005, the magnitude of inorganic-N deposition in rainwater quantified by the European Monitoring and Evaluation Programme (EMEP) (2003–2005) and the National Atmospheric Deposition Program/National Trends Network (NADP/NTN) (2003–2005) varied from 1.04 to 18.4 kg N ha−1 year−1 and 0.04 to 8.53 kg N ha−1 year−1, respectively. The values of LRP fell in high end of the ranges. Across Chinese province, Lu and Tian (2007) reported that the wet inorganic-N deposition ranged from 0.10 to 62.25 kg N ha−1 year−1, with an average value of 9.88 kg N ha−1 year−1. The amount of N deposition in LRP was much smaller than that of hotspots in China, especially in the more developed area of East China (Table 2). 3.4. Seasonal variation of NH4+/NO3− ratio

25

0

Dry season

Wet season

Fig. 3. Mean NH4+–N (a) or NO3−–N (b) concentration in rainwater grouped by the precipitation amount during 5-year period, and the distribution of heavy and light precipitation events in dry and wet season, respectively (c).

the Western China with much lower economic development degree, NH4+/NO3− ratio is higher (Table 2). Li and Li (1999) found that NH4+/NO3− ratio for N deposition averaged 4.1 in the Guanzhong Plain, Northwest China, and Jia et al. (2009) reported that NH4+/NO3− ratio equaled 3.6 in Linzhi, on the Qinghai–Tibetan Plateau. Historically, present NH4+/NO3− ratio of LRP was much smaller than the value (about 4) reported in the 1980s (Wang et al., 1986), suggesting the quick development of the industrialization. However, the present NH4+/NO3− ratio was still much larger than highly industrialized North America (less than 1) (Fahey et al., 1999;

Anthropogenic source of NH4+ in rain is NH3, mainly deriving from volatilizing of fertilizer and excrements of human and animal, which are closely related with agricultural activities; while the major anthropogenic sources of NO3− are nitrogen oxides, originating from fossil fuel combustion from power plants and automobiles, which is an indicator of industry activity. NH4+/NO3− ratio showed marked seasonal fluctuation with the extent from 1.1 to 8.9 (Fig. 2b). Seasonal variation pattern of NH4+/NO3− ratio was relatively complex. However, by synthesizing the 5-year data, we could still achieve the general fluctuation pattern. There were two peaks in one year's observation (Fig. 2b). The first peak occurred in the early spring (March or April) (if we could collect rainwater in this period) and the second peak emerged in summer (June or July). The first peak was higher than the second one, and this phenomenon was much more obvious in 2008. In autumn, the NH4+/NO3− ratio often showed a slight decreasing trend (from September to October). The second peak in summer and decreasing trend in autumn were easy to explain. The increasing value of NH4+/NO3− ratio in June or July coincided very well with the NH3 volatilization caused by N topdressing in the field of LRP area

W.T. Yu et al. / Atmospheric Research 101 (2011) 460–468

465

Table 2 Comparison of N concentrations and N depositions in this study with those of other researches in China. Location

Lijiang, Yunnan Linzhi, Tibet Yangling, Shaanxi Lhasa, Tibet Chongqing Zhengzhou Tai lake region Rural area, Beijing Suburb, Beijing Beijing Guangzhou Shanghai LRP, Liaoning

Sampling period

1987–1989 2005–2006 1994 1998–2000 1996–2002 1999 2003–2005 2004 1998–2004 2004 2005–2006 1998–2003 2004–2008

Character

RE/L RE/L R/L U/L U/M U/M S/H R/H S/H U/H U/H U/H S/M

N concentration (mg N L−1)

N deposition (kg N ha−1 year−1)

References

NH4+–N

NO3−–N

NH4+/NO3−

NH4+–N

NO3−–N

Total

0.079 0.36 3.23 0.12 2.38 2.64 1.35 n.d. 4.76 n.d. 1.27 2.16 1.82

0.026 0.10 1.11 0.1 0.66 0.65 0.90 n.d. 2.22 n.d. 1.06 2.59 0.89

3.0 3.6 3 1.2 3.6 4.1 1.5 1.7 2.1 0.8 1.2 0.8 2.05

n.d. 1.9 12.3 n.d. n.d. n.d. 16.3 11.6 20.9 9.2 24.3 26.7 9.8

n.d. 0.5 4.2 n.d. n.d. n.d. 10.9 6.7 9.7 11.7 16.2 31.4 4.7

n.d. 2.4 16.5 n.d. n.d. n.d. 27.1 18.3 30.6 20.9 40.5 58.1 14.5

(Liu et al., 1993) (Jia et al., 2009) (Li and Li, 1999) (Zhang et al., 2002) (Zhou et al., 2003) (Zhao et al., 2001) (Xie et al., 2008) (Liu et al., 2006) (Liu et al., 2006) (Liu et al., 2006) (Huang et al., 2010) (Zhang, 2006) This study

U—urban; S—suburban; R—rural; RE—remote; H—high-level economy; M—mid-level economy; L—low-level economy; n.d.—no data.

(Chen et al., 2007), indicating that agricultural NH3 emissions were the main source of N deposition in this period (Aneja et al., 2003; Su et al., 2003). The decline of NH4+/NO3− ratio in autumn is due to larger amount of nitrogen oxides emission during the cold season when high amount of coal was consumed for heating in Northern China. While in the Tai lake region, Southern China, Xie et al. (2008) found that the NO3− deposition rate was steady throughout the year because fossil fuel combustion in industry and motor vehicles was relatively constant in this region. The relatively high NH4+/NO3− ratio in March or April was detected in Beijing area, Northern China (Liu et al., 2006). Shen et al. (2009) also found a peak of atmospheric NH3 concentration in March or April in the same area. They both attributed the phenomenon to the NH3 volatilization induced by application of N fertilizer. However, this explanation could not totally fit the case in our results. Due to colder weather condition than Beijing area, the basal N fertilizer was applied in late April and early May (Liu et al., 1996). Surprisingly, we found high NH4+/NO3− ratio in March, ahead of N fertilizer application. In the area of LRP, spring soil thawing is initiated from March (Institute of Forest Soil, 1980). Edwards and Killham (1986) found that soil freeze/thaw cycle could promote NH3 volatilization. Consequently, we inferred that this peak of NH4+/NO3− ratio could have resulted from the soil freeze/thaw cycle in early spring. More speculatively, the higher NH4+/NO3− ratio in spring than in summer might be largely caused by the meteorology factor. For example, higher lightning frequency and sun radiation in summer could favor the transformation from nitrogen oxides to nitrate and nitric acid by photochemical reaction in atmosphere (Hastings et al., 2003; Kadowaki, 1986; Khoder, 2002; Price et al., 1997; Schumann and Huntrieser, 2007). The plausibility of above scenario should be corroborated in further study using a variety of approaches, including determining the N isotopic composition in rainwater. 3.5. Effects of N deposition on agricultural and natural ecosystem In the farmland of LRP, the mean annual N fertilizer input was 150 kg N ha −1 (Liaoning Statistics Bureau, 2006). This

high N deposition indicated that N deposition should be taken into account when calculating the N fertilizer requirements of crop in this region, which could avoid the excessive application of N fertilizer, help to increase N use efficiency and thus reduce the risk of higher N losses to the environment (He et al., 2010; Zhu and Chen, 2002). The deposited N calculated in this study would have potential negative effects on the natural ecosystem in LRP and its vicinity. Before the analysis, we must know the term, critical loads (CL) of a pollutant, which has been defined as “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 according to present knowledge” (Fenn et al., 2010). The typical zonal vegetation in the study site is dominated by temperate mixed coniferous and broad-leaf forest (Institute of Forest Soil, 1980). To our knowledge, the adverse effects of N deposition on the vegetation included the following aspects: (1) Lichen is considered as the most sensitive terrestrial responders to air pollution, such as increased N deposition (Conti and Cecchetti, 2001; Fenn et al., 2007; Van Herk et al., 2007). At a mixed conifer forest ecosystem, Fenn et al. (2008) concluded that biological impacts on lichen communities were clearly occurring at N deposition levels as low as 3–5 kg N ha−1 year−1. Therefore, the dominance of lichen communities might have shifted from oligotrophs (adapted to low nutrient environments) to eutrophs (thrive in nutrient-rich environments) around our monitoring site like in Northwestern America (Geiser et al., 2010). (2) The effects of increasing N deposition on the biodiversity in herbaceous layer of forest ecosystem have caused great concern because most species diversity of forests occurs in herb layer. Previous studies showed varying response and did not get consistent CL (Gilliam, 2006; Lovett et al., 2009). Available evidence suggests that the threshold for N deposition effects on understory biodiversity is 20 kg N ha−1 year−1, and may be as low as 10–15 kg N ha−1 year−1 (Bobbink et al., 2010). Thus, the diversity of native vascular plant in natural ecosystem of LRP is expected to decline, because N deposition could favor the invasion of nitrophilous plants caused by the decrease of N spatial heterogeneity (Cassidy et al., 2004; Gilliam, 2006). (3) Soil chemical response is less sensitive to chronic N deposition than

466

W.T. Yu et al. / Atmospheric Research 101 (2011) 460–468

species composition (Bowman et al., 2006). However, long-term N deposition can lead to loss of soil fertility when the ecosystems reach the “nitrogen saturation” status (Aber et al., 1998). Aber et al. (2003) found increased nitrate leaching in some forested ecosystems in the Northeast of America as N deposition levels increase. Thus, there might be a potential for nitrate leaching in this area, if we compare our N deposition estimation (see the discussion in Section 3.6) with the CL of nitrate leaching in temperate forest (17 kg N ha−1 year−1) (Fenn et al., 2008). These nitrates could affect water quality and cause eutrophication of natural surface water bodies. (4) If deposited N is above the CL of 15–25 kg N ha −1 year −1, it would increase the susceptibility of boreal forest ecosystems to frost, drought, pathogens and pests (De Vries et al., 2007). Consequently, we speculate that the natural ecosystems in LRP and its vicinity may be predisposed to be affected by these stresses. 3.6. Further research needs In the present study, we just quantify the inorganic N (NH4+–N and NO3−–N) in precipitation collected by a standard rain gage. However, some studies have found that organic N deposition can account for 10–30% of wet N deposition (Keene et al., 2002; Neff et al., 2002; Violaki et al., 2010). Moreover, dry N deposition, including gaseous HNO3, NH3, and NO2, particulate NH4+ and NO3−, can provide as much amount of N as the wet deposition or even larger (Anatolaki and Tsitouridou, 2007; Anderson and Downing, 2006; Lu and Tian, 2007; Shen et al., 2009), although the wet/dry ratio is not fixed, and would vary as a result of changing emissions sources and N compounds (Baron, 2006). Consequently, the magnitude of N deposition we quantified (Table 1) could underestimate the total N deposition up to 50%. The total N deposition might be greater than 29 kg N ha−1 year−1 in LRP. Exact investigations are needed to determine these N depositions in future study. The micrometeorological techniques, such as eddy covariance and gradient methods (Fowler et al., 2001), are ideal methods to quantify the dry N flux because these methods could determine bidirectional N fluxes (e.g. NH3) of the atmosphere–land exchange (Milford et al., 2001; Sutton et al., 2009; Walker et al., 2006; Wolff et al., 2010). Admittedly, because the rain collector we used was a standard rain gage rather than a wet-only, event based system, it was inevitable to collect some portion of dry N deposition (such as dust during dry periods) other than the wet N deposition. The wet-only rain collector is needed in future study to quantify the N wet deposition more exactly. Stable N isotope in rainwater could provide novel information that which sources contributed to the atmospheric N (Jia and Chen, 2009; Xie et al., 2008; Zhang et al., 2008); meanwhile, back trajectory analysis could track the geographical location of the air mass from which the precipitation was produced (Chiwa, 2010; Occhipinti et al., 2008; Russell et al., 1998). Integrating the two methods, we would detect the potential sources of N in rainwater more accurately and then we could take more effective measures to decrease the emission of reactive N to atmosphere. Additionally, the seasonal variation of stable O (oxygen) isotope observed in NO3− would reflect a shift in the predominant oxidation pathway for conversion from nitrogen oxides to nitrate (nitrogen oxides being oxidized by ozone or hydroxyl radical) (Buda and DeWalle, 2009; Hastings et al., 2003).

4. Conclusions In summary, a continuous 5-year observation (from 2004 to 2008) in LRP was carried out to analyze the inorganic N (NH4+–N and NO3−–N) concentration in precipitation and determine the amount of N deposition in an intensive agricultural and industrial activities region of Northeast China. The mean inorganic N concentrations were 1.82 mg N L−1 for NH4+–N and 0.89 mg N L −1 for NO3−–N. On an annual basis, the averaged amounts of total inorganic N, NH4+–N and NO3−–N depositions were 14.5 kg N ha −1 year −1 , 9.8 kg N ha −1 year −1 and 4.7 kg N ha−1 year −1, respectively. Although the amount of N deposition was smaller than that of hotspots in China, serious ecological perturbations in this region are expected. Monthly and yearly N depositions coincided very well with the precipitations, suggesting that precipitation was the determining factor that controlled the wet deposition. We also found that precipitation played an important role in regulating the N concentration in rainwater, particularly when the rainfall was below 20 mm. In addition, the seasonal fluctuation of NH4+/ NO3− ratio reflected that the deposited N originated from both fertilizer application and fossil fuel combustion, and environmental factor (soil temperature, lightning and sunshine) might also contribute to the seasonal cycles of NH4+/NO3− ratio. The high annual NH4+/NO3− ratio (2.05) compared with the more developed region suggested that N deposition in this region was mostly affected by agricultural activities rather than industrial activities. Finally, additional methods are needed to investigate the total N deposition and to distinguish the source of reactive N deposition in the region. Acknowledgments This study was financially supported by the National Natural Science Foundation of China (nos. 31070547; 31000206) and the National Key Technology R & D Program (no. 2007BAD89B02). We are grateful to everyone who assisted with field and laboratory work. We are also grateful to anonymous reviewers for their valuable comments on earlier versions of the manuscript. References Aber, J., 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, 921–934. Aber, J.D., Goodale, C.L., Ollinger, S.V., Smith, M.L., Magill, A.H., Martin, M.E., Hallett, R.A., Stoddard, J.L., 2003. Is nitrogen deposition altering the nitrogen status of northeastern forests? Bioscience 53, 375–389. Akimoto, H., 2003. Global air quality and pollution. Science 302, 1716–1719. Anatolaki, C., Tsitouridou, R., 2007. Atmospheric deposition of nitrogen, sulfur and chloride in Thessaloniki, Greece. Atmospheric Research 85, 413–428. Anderson, K.A., Downing, J.A., 2006. Dry and wet atmospheric deposition of nitrogen, phosphorus and silicon in an agricultural region. Water Air and Soil Pollution 176, 351–374. Aneja, V.P., Nelson, D.R., Roelle, P.A., Walker, J.T., Battye, W., 2003. Agricultural ammonia emissions and ammonium concentrations associated with aerosols and precipitation in the southeast United States. Journal of Geophysical Research 108, 1–12. Baron, J.S., 2006. Hindcasting nitrogen deposition to determine an ecological critical load. Ecological Applications 16, 433–439. Bobbink, R., Hicks, K., Galloway, J., Spranger, T., Alkemade, R., Ashmore, M., Bustamante, M., Cinderby, S., Davidson, E., Dentener, F., Emmett, B.,

W.T. Yu et al. / Atmospheric Research 101 (2011) 460–468 Erisman, J.W., Fenn, M., Gilliam, F., Nordin, A., Pardo, L., De Vries, W., 2010. Global assessment of nitrogen deposition effects on terrestrial plant diversity: a synthesis. Ecological Applications 20, 30–59. Bowman, W.D., Gartner, J.R., Holland, K., Wiedermann, M., 2006. Nitrogen critical loads for alpine vegetation and terrestrial ecosystem response: are we there yet? Ecological Applications 16, 1183–1193. Buda, A.R., DeWalle, D.R., 2009. Using atmospheric chemistry and storm track information to explain the variation of nitrate stable isotopes in precipitation at a site in central Pennsylvania, USA. Atmospheric Environment 43, 4453–4464. Cassidy, T.M., Fownes, J.H., Harrington, R.A., 2004. Nitrogen limits an invasive perennial shrub in forest understory. Biological Invasions 6, 113–121. Chen, Z., Chen, L., Wu, Z., Zhang, Y., Juan, Y., 2007. Ammonia volatilization from rice field under different water conditions in lower Liaohe River Plain. Chinese Journal of Applied Ecology 18, 2771–2776 (in Chinese). Chiwa, M., 2010. Characteristics of atmospheric nitrogen and sulfur containing compounds in an inland suburban-forested site in northern Kyushu, western Japan. Atmospheric Research 96, 531–543. Conti, M.E., Cecchetti, G., 2001. Biological monitoring: lichens as bioindicators of air pollution assessment — a review. Environmental Pollution 114, 471–492. Cornell, S.E., Jickells, T.D., Cape, J.N., Rowland, A.P., Duce, R.A., 2003. Organic nitrogen deposition on land and coastal environments: a review of methods and data. Atmospheric Environment 37, 2173–2191. De Vries, W., Kros, H., Reinds, G.J., Wamelink, W., Mol, J., van Dobben, H., Bobbink, R., Emmett, B., Smart, S., Evans, C., 2007. Developments in deriving critical limits and modelling critical loads of nitrogen for terrestrial ecosystems in Europe. Alterra-rapport 1382: Alterra, Wageningen. Edwards, A.C., Killham, K., 1986. The effect of freeze thaw on gaseous nitrogen loss from upland soils. Soil Use and Management 2, 86–91. European Monitoring and Evaluation Programme (EMEP), 2003–2005. EMEP Measurement Data OnlineWebsite at http://www.nilu.no/projects/ccc/ emepdata.html 2003–2005. Fahey, T.J., Williams, C.J., Rooney-Varga, J.N., Cleveland, C.C., Postek, K.M., Smith, S.D., Bouldin, D.R., 1999. Nitrogen deposition in and around an intensive agricultural district in central New York. Journal of Environmental Quality 28, 1585–1600. Fenn, M.E., Haeuber, R., Tonnesen, G.S., Baron, J.S., Grossman-Clarke, S., Hope, D., Jaffe, D.A., Copeland, S., Geiser, L., Rueth, H.M., Sickman, J.O., 2003. Nitrogen emissions, deposition, and monitoring in the western United States. Bioscience 53, 391–403. Fenn, M.E., Geiser, L., Bachman, R., Blubaugh, T.J., Bytnerowicz, A., 2007. Atmospheric deposition inputs and effects on lichen chemistry and indicator species in the Columbia River Gorge, USA. Environmental Pollution 146, 77–91. Fenn, M.E., Jovan, S., Yuan, F., Geiser, L., Meixner, T., Gimeno, B.S., 2008. Empirical and simulated critical loads for nitrogen deposition in California mixed conifer forests. Environmental Pollution 155, 492–511. Fenn, M.E., Allen, E.B., Weiss, S.B., Jovan, S., Geiser, L.H., Tonnesen, G.S., Johnson, R.F., Rao, L.E., Gimeno, B.S., Yuan, F., Meixner, T., Bytnerowicz, A., 2010. Nitrogen critical loads and management alternatives for N-impacted ecosystems in California. Journal of Environmental Management 91, 2404–2423. Fowler, D., Coyle, M., Flechard, C., Hargreaves, K., Nemitz, E., Storeton-West, R., Sutton, M., Erisman, J.W., 2001. Advances in micrometeorological methods for the measurement and interpretation of gas and particle nitrogen fluxes. Plant and Soil 228, 117–129. Galloway, J.N., Likens, G.E., Keene, W.C., Miller, J.M., 1982. The composition of precipitation in remote areas of the world. Journal of Geophysical Research 87, 8771–8786. Galloway, J.N., Dentener, F.J., Capone, D.G., Boyer, E.W., Howarth, R.W., Seitzinger, S.P., Asner, G.P., Cleveland, C.C., Green, P.A., Holland, E.A., Karl, D.M., Michaels, A.F., Porter, J.H., Townsend, A.R., Vorosmarty, C.J., 2004. Nitrogen cycles: past, present, and future. Biogeochemistry 70, 153–226. Galloway, J.N., Townsend, A.R., Erisman, J.W., Bekunda, M., Cai, Z.C., Freney, J.R., Martinelli, L.A., Seitzinger, S.P., Sutton, M.A., 2008. Transformation of the nitrogen cycle: recent trends, questions, and potential solutions. Science 320, 889–892. Gao, Y., Kennish, M.J., Flynn, A.M., 2007. Atmospheric nitrogen deposition to the New Jersey coastal waters and its implications Supplement Ecological Applications 17 S31–S41. Geiser, L.H., Jovan, S.E., Glavich, D.A., Porter, M.K., 2010. Lichen-based critical loads for atmospheric nitrogen deposition in Western Oregon and Washington Forests, USA. Environmental Pollution 158, 2412–2421. Gilliam, F.S., 2006. Response of the herbaceous layer of forest ecosystems to excess nitrogen deposition. Journal of Ecology 94, 1176–1191. Hastings, M.G., Sigman, D.M., Lipschultz, F., 2003. Isotopic evidence for source changes of nitrate in rain at Bermuda. Journal of Geophysical Research 108, 4790.

467

He, C.E., Wang, X., Liu, X., Fangmeier, A., Christie, P., Zhang, F., 2010. Nitrogen deposition and its contribution to nutrient inputs to intensively managed agricultural ecosystems. Ecological Applications 20, 80–90. Hertel, O., Skjoth, C.A., Lofstrom, P., Geels, C., Frohn, L.M., Ellermann, T., Madsen, P.V., 2006. Modelling nitrogen deposition on a local scale — a review of the current state of the art. Environmental Chemistry 3, 317–337. Holland, E.A., Dentener, F.J., Braswell, B.H., Sulzman, J.M., 1999. Contemporary and pre-industrial global reactive nitrogen budgets. Biogeochemistry 46, 7–43. Holland, E.A., Braswell, B.H., Sulzman, J., Lamarque, J.F., 2005. Nitrogen deposition onto the United States and western Europe: synthesis of observations and models. Ecological Applications 15, 38–57. Huang, D., Xu, Y., Zhou, B., Zhang, H., Lan, J., 2010. Wet deposition of nitrogen and sulfur in Guangzhou, a subtropical area in South China. Environmental Monitoring and Assessment 171, 429–439. Institute of Forest Soil, 1980. Soil of Northeast China. Science Press, Beijing (in Chinese). Jia, G.D., Chen, F.J., 2009. Monthly variations in nitrogen isotopes of ammonium and nitrate in wet deposition at Guangzhou, south China. Atmospheric Environment 44, 2309–2315. Jia, J.Y., Zhang, Y., Cai, X.B., Liu, X.J., 2009. A dynamic changes of wet deposition of nitrogen in southeast Tibet: taking Linzhi experiment station as an example. Acta Ecologica Sinica 29, 1907–1913 (in Chinese). Jiang, C.M., Yu, G.R., Fang, H.J., Cao, G.M., Li, Y.N., 2010. Short-term effect of increasing nitrogen deposition on CO2, CH4 and N2O fluxes in an alpine meadow on the Qinghai–Tibetan Plateau, China. Atmospheric Environment 44, 2920–2926. Kadowaki, S., 1986. On the nature of atmospheric oxidation processes of sulfur dioxide to sulfate and of nitrogen dioxide to nitrate on the basis of diurnal variations of sulfate, nitrate, and other pollutants in an urban area. Environmental Science & Technology 20, 1249–1253. Kamphake, L.J., Hannah, S.A., Cohen, J.M., 1967. Automated analysis for nitrate by hydrazine reduction. Water Research 1, 205–216. Keene, W.C., Montag, J.A., Maben, J.R., Southwell, M., Leonard, J., Church, T.M., Moody, J.L., Galloway, J.N., 2002. Organic nitrogen in precipitation over Eastern North America. Atmospheric Environment 36, 4529–4540. Khoder, M.I., 2002. Atmospheric conversion of sulfur dioxide to particulate sulfate and nitrogen dioxide to particulate nitrate and gaseous nitric acid in an urban area. Chemosphere 49, 675–684. Li, S.Q., Li, S.X., 1999. Nitrogen added to ecosystems by wet deposition in Guanzhong Area in Shaanxi. Agricultural Environment Protection 18, 97–101 (in Chinese). Liaoning Statistics Bureau, 2006. Liaoning Statistical Yearbook 2006. China Statistics Press, Beijing (in Chinese). Liu, J.Q., Keene, W.C., Wu, G.P., 1993. Study of precipitation background value in Lijiang. China. China Environmental Science 13, 246–251 (in Chinese). Liu, X.H., Li, F.C., Zhou, C.Y., 1996. Geoponics. China Agricultural Press, Beijing (in Chinese). Liu, X.J., Ju, X.T., Zhang, Y., He, C., Kopsch, J., Fusuo, Z., 2006. Nitrogen deposition in agroecosystems in the Beijing area. Agriculture Ecosystems & Environment 113, 370–377. Lovett, G.M., Tear, T.H., Evers, D.C., Findlay, S.E.G., Cosby, B.J., Dunscomb, J.K., Driscoll, C.T., Weathers, K.C., 2009. Effects of air pollution on ecosystems and biological diversity in the eastern United States. Annals of the New York Academy of Sciences 2009 (1162), 99–135. Lu, C.Q., Tian, H.Q., 2007. Spatial and temporal patterns of nitrogen deposition in China: synthesis of observational data D22S05 Journal of Geophysical Research-Atmospheres 112. doi:10.1029/2006JD007990. Lund, M., Christensen, T.R., Mastepanov, M., Lindroth, A., Strom, L., 2009. Effects of N and P fertilization on the greenhouse gas exchange in two northern peatlands with contrasting N deposition rates. Biogeosciences 6, 2135–2144. Matson, P., Lohse, K.A., Hall, S.J., 2002. The globalization of nitrogen deposition: consequences for terrestrial ecosystems. Ambio 31, 113–119. Milford, C., Hargreaves, K.J., Sutton, M.A., Loubet, B., Cellier, P., 2001. Fluxes of NH3 and CO2 over upland moorland in the vicinity of agricultural land. Journal of Geophysical Research-Atmospheres 106, 24169–24181. National Atmospheric Deposition Program/National Trends Network (NADP/ NTN), 2003–2005. Annual Data for Custom Site ListWebsite at http://nadp. sws.uiuc.edu/nadpdata/annualReq.asp?site=Custom156 2003–2005. Neff, J.C., Holland, E.A., Dentener, F.J., McDowell, W.H., Russell, K.M., 2002. The origin, composition and rates of organic nitrogen deposition: a missing piece of the nitrogen cycle? Biogeochemistry 57, 99–136. Occhipinti, C., Aneja, V.P., Showers, W., Niyogi, D., 2008. Back-trajectory analysis and source-receptor relationships: particulate matter and nitrogen isotopic composition in rainwater. Journal of the Air & Waste Management Association 58, 1215–1222. Price, C., Penner, J., Prather, M., 1997. NOx from lightning.1. Global distribution based on lightning physics. Journal of Geophysical Research—Atmospheres 102, 5929–5941.

468

W.T. Yu et al. / Atmospheric Research 101 (2011) 460–468

Rojas, A.L.P., Venegas, L.E., 2010. Interannual variability of estimated monthly nitrogen deposition to coastal waters due to variations of atmospheric variables model input. Atmospheric Research 96, 88–102. Russell, K.M., Galloway, J.N., Macko, S.A., Moody, J.L., Scudlark, J.R., 1998. Sources of nitrogen in wet deposition to the Chesapeake Bay region. Atmospheric Environment 32, 2453–2465. Schumann, U., Huntrieser, H., 2007. The global lightning-induced nitrogen oxides source. Atmospheric Chemistry and Physics 7, 3823–3907. Searle, P.L., 1984. The Berthlot or indophenol reaction and its use in the analysis chemistry of nitrogen: a review. Analysis 109, 549–565. Shen, J.L., Tang, A.H., Liu, X.J., Fangmeier, A., Goulding, K.T.W., Zhang, F.S., 2009. High concentrations and dry deposition of reactive nitrogen species at two sites in the North China Plain. Environmental Pollution 157, 3106–3113. Stevens, C.J., Dupre, C., Dorland, E., Gaudnik, C., Gowing, D.J.G., Bleeker, A., Diekmann, M., Alard, D., Bobbink, R., Fowler, D., Corcket, E., Mountford, J.O., Vandvik, V., Aarrestad, P.A., Muller, S., Dise, N.B., 2010. Nitrogen deposition threatens species richness of grasslands across Europe. Environmental Pollution 158, 2940–2945. Su, C., Yin, B., Zhu, Z.L., Sheng, Q., 2003. Ammonia volatilization loss of nitrogen fertilizer from rice field and wet deposition of atmospheric nitrogen in rice growing season. Chinese Journal of Applied Ecology 14, 1884–1888 (in Chinese). Sutton, M.A., Nemitz, E., Milford, C., Campbell, C., Erisman, J.W., Hensen, A., Cellier, P., David, M., Loubet, B., Personne, E., Schjoerring, J.K., Mattsson, M., Dorsey, J.R., Gallagher, M.W., Horvath, L., Weidinger, T., Meszaros, R., Dammgen, U., Neftel, A., Herrmann, B., Lehman, B.E., Flechard, C., Burkhardt, J., 2009. Dynamics of ammonia exchange with cut grassland: synthesis of results and conclusions of the GRAMINAE Integrated Experiment. Biogeosciences 6, 2907–2934. Van Herk, C.M., Mathijssen-Spiekman, E.A.M., De Zwart, D., 2007. Long distance nitrogen air pollution effects on lichens in Europe. The Lichenologist 35, 347–359. Violaki, K., Zarbas, P., Mihalopoulos, N., 2010. Long-term measurements of dissolved organic nitrogen (DON) in atmospheric deposition in the Eastern Mediterranean: Fluxes, origin and biogeochemical implications. Marine Chemistry 120, 179–186. Vitousek, P.M., Aber, J.D., Howarth, R.W., Likens, G.E., Matson, P.A., Schindler, D.W., Schlesinger, W.H., Tilman, G.D., 1997. Human alteration of the

global nitrogen cycle: sources and consequences. Ecological Applications 7, 737–750. Walker, J.T., Robarge, W.P., Wu, Y., Meyers, T.P., 2006. Measurement of bidirectional ammonia fluxes over soybean using the modified Bowenratio technique. Agricultural and Forest Meteorology 138, 54–68. Wang, W., Wang, D., Liu, X., 1986. Observation nitrogen concentration in rainwater of Shenyang. The Balance of Organic Matter in Cultivated Land and Improvement by Application of Organic Fertilizers. Institute of soil fertility, Shenyang Agricultural University Press, Shenyang, pp. 210–212. in Chinese. Wolff, V., Trebs, I., Foken, T., Meixner, F.X., 2010. Exchange of reactive nitrogen compounds: concentrations and fluxes of total ammonium and total nitrate above a spruce canopy. Biogeosciences 7, 1729–1744. Xie, Y.X., Xiong, Z.Q., Xing, G.X., Yan, X.Y., Shi, S.L., Sun, G.Q., Zhu, Z.L., 2008. Source of nitrogen in wet deposition to a rice agroecosystem at Tai lake region. Atmospheric Environment 42, 5182–5192. Zhang, X.F., 2006. Atmospheric nitrogen wet deposition and its effects on wetland water environment of Shanghai area. Chinese Journal of Applied Ecology 17, 1099–1102 (in Chinese). Zhang, D.D., Peart, M.R., Jim, C.Y., La, J., 2002. Alkaline rains on the Tibetan Plateau and their implication for the original pH of natural rainfall. Journal of Geophysical Research-Atmospheres 107, 4198–4203. Zhang, Y., Liu, X.J., Fangmeier, A., Goulding, K.T.W., Zhang, F.S., 2008. Nitrogen inputs and isotopes in precipitation in the North China Plain. Atmospheric Environment 42, 1436–1448. Zhao, Y., Sun, Z.D., Wang, F., Zhao, Y.G., 2001. The correlation between acid material in atmosphere and precipitation acidity. Research of Environmental Sciences 14, 20–23 (in Chinese). Zhao, X., Yan, X.Y., Xiong, Z.Q., Xie, Y.X., Xing, G.X., Shi, S.L., Zhu, Z.L., 2009. Spatial and temporal variation of inorganic nitrogen wet deposition to the Yangtze River Delta Region, China. Water Air and Soil Pollution 203, 277–289. Zhou, Z.Y., Chen, D.R., Ying, J., Zhang, W.L., 2003. The chemical composition of precipitation analysis in Chongqing city. Chongqing Environmental Science 25, 112–114 (in Chinese). Zhu, Z.L., Chen, D.L., 2002. Nitrogen fertilizer use in China — contributions to food production, impacts on the environment and best management strategies. Nutrient Cycling in Agroecosystems 63, 117–127.