Atmospheric NH3 and NO2 concentration and nitrogen deposition in an agricultural catchment of Eastern China

Science of the Total Environment 408 (2010) 4624–4632

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Science of the Total Environment 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 / s c i t o t e n v

Atmospheric NH3 and NO2 concentration and nitrogen deposition in an agricultural catchment of Eastern China Rong Yang a, Kentaro Hayashi b, Bin Zhu c, Feiyue Li a,d, Xiaoyuan Yan a,⁎ a

State Key Laboratory of Soil and Sustainable Agriculture, Institute of Soil Science, Chinese Academy of Sciences, Nanjing 210008, China Carbon and Nutrient Cycles Division, National Institute for Agro-Environmental Sciences, 3-1-3, Kan-nondai, Tsukuba, Ibaraki 305-8604, Japan Key Laboratory of Atmospheric Physics and Environment, Nanjing University of Information Science and Technology, Nanjing 210044, China d The School of Plant Science, Anhui Science and Technology University, Fengyang 233100, China b c

a r t i c l e

i n f o

Article history: Received 14 December 2009 Received in revised form 18 May 2010 Accepted 4 June 2010 Available online 10 July 2010 Keywords: Dry nitrogen deposition NH3 NO2 Organic nitrogen deposition Rural area

a b s t r a c t To assess the atmospheric environmental impacts of anthropogenic reactive nitrogen in the fast-developing Eastern China region, we measured atmospheric concentrations of nitrogen dioxide (NO2) and ammonia + (NH3) as well as the wet deposition of inorganic nitrogen (NO− 3 and NH4 ) and dissolved organic nitrogen (DON) levels in a typical agricultural catchment in Jiangsu Province, China, from October 2007 to September 2008. The annual average gaseous concentrations of NO2 and NH3 were 42.2 μg m−3 and 4.5 μg m−3 (0 °C, + 760 mm Hg), respectively, whereas those of NO− 3 , NH4 , and DON in the rainwater within the study −1 catchment were 1.3, 1.3, and 0.5 mg N L , respectively. No clear difference in gaseous NO2 concentrations and nitrogen concentrations in collected rainwater was found between the crop field and residential sites, but the average NH3 concentration of 5.4 μg m−3 in residential sites was significantly higher than that in field sites (4.1 μg m−3). Total depositions were 40 kg N ha−1 yr−1 for crop field sites and 30 kg N ha−1 yr−1 for residential sites, in which dry depositions (NO2 and NH3) were 7.6 kg N ha−1 yr−1 for crop field sites and 1.9 kg N ha−1 yr−1 for residential sites. The DON in the rainwater accounted for 16% of the total wet nitrogen deposition. Oxidized N (NO− 3 in the precipitation and gaseous NO2) was the dominant form of nitrogen deposition in the studied region, indicating that reactive forms of nitrogen created from urban areas contribute greatly to N deposition in the rural area evaluated in this study. © 2010 Elsevier B.V. All rights reserved.

1. Introduction Reactive nitrogen (N) levels have increased dramatically worldwide due to anthropogenic activities, such as over-fertilization, high stocking rates of farm animals, and increased combustion of fossil fuels and have resulted in a sharp increase in global N deposition (Galloway et al., 2004). Although N is a critical nutrient for the survival of microorganisms, plants, humans, and animals, excessive N deposition induces a considerable burden on forest, grassland, and aquatic ecosystems, aggravates eutrophication in aquatic systems and soil acidification, and can lead to changes in biodiversity (Vitousek et al., 1997). Deposition monitoring is very important for identifying trends in N emissions, understanding how to control eutrophication in surface water, improving recommendations for the use of fertilizer, and developing deposition models (Smith et al., 2000; Yoshikawa et al., 2008). In the recent two decades, rapid economic growth in China has enhanced the consumption of chemical fertilizer and fossil fuel. China

⁎ Corresponding author. Tel.: + 86 25 8688 1530; fax: + 86 25 8688 1000. E-mail address: [email protected] (X. Yan). 0048-9697/$ – see front matter © 2010 Elsevier B.V. All rights reserved. doi:10.1016/j.scitotenv.2010.06.006

consumes one-third of the world's chemical nitrogen fertilizer, which increased from 12.1 Tg N in 1980 to 25.4 Tg N in 2000 (FAOSTAT, 2006). Nitrogen released from fossil fuel combustion in China increased from 1.1 Tg in 1980 to 3.4 Tg in 2000 (Ohara et al., 2007). As a consequence, nitrogen flow to water bodies and the atmosphere has increased remarkably. The impact of anthropogenic nitrogen cycling on water quality is well recognized in China, as eutrophication of big lakes occurs with an increasing frequency (Xing et al., 2001; Lue et al., 2005; Qin et al., 2007). The impact of nitrogen cycling on the atmospheric environment, however, has not been given sufficient attention in China, probably because it is less visible than that of water pollution. Jiangsu Province, located at the lower reach of the Yangtse River in Eastern China, is a fast-developing province of China. Anthropogenic reactive N has far exceeded the biologically fixed N in natural terrestrial ecosystems in this region (Xing and Zhu, 2002). A few studies have reported a high level of nitrogen deposition in this region (Xie et al., 2008; Zhao et al., 2009). However, these studies primarily focused on the wet deposition of inorganic forms of N (e.g., − ammonium (NH+ 4 ) and nitrate (NO3 )); the contributions of organic N and gaseous dry deposition (e.g., nitrogen dioxide (NO2) and ammonia (NH3)) are not well illustrated. Although dissolved organic

R. Yang et al. / Science of the Total Environment 408 (2010) 4624–4632

nitrogen (DON) is a poorly characterized component of atmospheric precipitation, it may account for 5–88% of the total nitrogen (Cornell et al., 2003). Gaseous NO2 and NH3 not only contribute to N deposition but also play important roles in atmospheric chemistry. Rapid increases in oxidized nitrogen (NOx) and NH3 emissions drastically change the pattern and magnitude of total inorganic nitrogen deposition (Galloway et al., 2004). In this study, we selected a typical agricultural catchment in Jiangsu province and deployed rain gauges to a cropland and a residential area as well as passive samplers to different locations of the catchment to collect dry (NO2 and NH3) and wet (N from precipitation) depositions. This study had three main objectives: (1) to evaluate the atmospheric NO2 and NH3 pollution status; (2) to determine the gaseous NH3 and NO2 deposition levels and relative importance of DON in wet deposition; and (3) to characterize the seasonal pattern of both dry (NO2 and NH3) and wet nitrogen depositions in relation to climate and agricultural activities.

2. Materials and methods 2.1. Study site The study was conducted in a typical intensive agricultural catchment in Jiangsu Province located at the lower reach of the Yangtze River in Eastern China (32°01´N, 119°13´E). This catchment is 3 km northeast of downtown Jurong city and 40 km southeast of Nanjing city. The total permanent population is 18,092. The study catchment has an area of 45.5 km2, of which 54% is cropland, 27.5% is construction and road, 9.3% is tea garden and artificial forest, and the rest (9.2%) is occupied by a reservoir, three small rivers, and thousands of small ponds. The annual mean temperature of the study area is 15 °C, and the annual mean precipitation is 1050 mm. There is no industry or intensive livestock farming in the catchment. Therefore, agriculture is the dominant local source of N contamination. In the residential area, there are a small number of livestock houses for self-consumption. Human and livestock excreta are applied to the farmlands after being managed in the open air in the residential area for a few months. In the field area, rice-wheat and maize-oil rape are the major annual cropping rotations for the paddy fields and upland fields in conventional farming practice, respectively. Synthetic compound fertilizer and urea are the major nitrogen

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nutrient sources for the croplands, and these are usually applied at a rate of 500–600 kg N ha−1 yr−1 for summer and winter crops together. The locations of the study catchment in China and monitoring sites in the study catchment are shown in Fig.1.

2.2. Nitrogen deposition monitoring This study was conducted from October 2007 to September 2008. Automatic samplers are difficult to use due to power limitations in the fields. Therefore, bulk samplers and Ogawa passive samplers (Ogawa & Co., USA, Inc.) were adopted to measure wet and dry (NO2 and NH3) depositions, respectively. Although other reactive nitrogen compo− nents, such as HNO3, particulate NH+ 4 , and particulate NO3 , may have a great contribution to N dry deposition, these are usually measured by pumping ambient air through a filter pack containing three or four filters (EANET; Marner and Harrison, 2004; Aas et al., 2007; Hayashi et al., 2007) or using an annular denuder system (Walker et al., 2004) or other methods (Shen et al., 2009); all of these methods need power to pump the ambient air. Thus, these compounds were not monitored in the current study. To monitor the bulk deposition, two rain gauges were installed at a crop field site and a residential site in the catchment (approximately 1.5 m from the ground with an interval of about 1000 m, at least 50 m away from tree or road). The samplers were made of polyvinyl chloride funnels and bottles, and the funnel had a collection area of 0.028 m2. After each rain event (12 h after a rainfall stops), all collected rainwater was thoroughly mixed and an aliquot of collected rainwater placed in a plastic bottle. The rain gauges were cleaned with deionized water after each collection. The collected rain samples were frozen at − 20 °C until analysis. All of the samples were filtered before − chemical analysis. The inorganic (NH+ 4 and NO3 ) and total N concentrations in the rainwater were measured using a flow injection analyzer (Skalar, Netherlands, analytic error ±3.9%). The automated − procedure for the determination of NH+ 4 , NO3 , and total N was based on the modified Berthelot reaction, hydrazine reduction method, and peroxodisulfate oxidation method, respectively. The concentration of dissolved organic nitrogen (DON) was calculated based on the difference between the total N and inorganic N (Cornell et al., 2003). Xing et al. (unpublished data, private communication) compared N concentrations in rainwater collected using wet and bulk samplers over 2 years in a neighboring county 90 km away from

Fig. 1. Map and location of the study watershed and the monitoring sites.

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+ the study catchment and found that the amounts of NO− 3 , NH4 , and total N collected by the wet sampler were 87 ± 4%, 89 ± 2%, and 87 ± 3%, respectively, of the corresponding N collected by the bulk sampler. The site has similar cultivation customs and climatic conditions as this study's catchment. Therefore, in this study, wet N deposition was estimated from the bulk N deposition using these ratios. Ogawa passive samplers were used to monitor gaseous concentrations of NO2 and NH3 in the catchment. It has been demonstrated that the results provided by these samplers are almost identical to those obtained using continuous active samplers or other passive samplers (Roadman et al., 2003; Sather et al., 2007; Beckerman et al., 2008). Detailed descriptions of the Ogawa passive sampler can be found in the report by Roadman et al. (2003) and at http://www. ogawausa.com/. One sampler has two separate ports that contain filters. In the study, the two ports contained filters designed to collect NO2 and NH3, respectively. Prior to sampling, glass fiber filters (Whatman GF/D, 47-mmdiameter filter) were cut to a suitable size for the sampler (14-mmdiameter filter). The filters were then cleaned by soaking in 5% hypochlorous acid solution for 3 h, after which they were ultrasonically washed three times with deionized water for 20 min each and then dried in an oven. Filters used to collect the NH3 were prepared by soaking them in a solution of 3% (v/v) phosphorous acid and 1% (v/v) glycerol, whereas those used for NO2 were saturated with 25% v/v aqueous triethanolamine/acetone solution. The filters were allowed to dry in airtight desiccators until just damp to prevent the solution from interfering with other components of the sampler, after which all components were placed into the sampler using clean forceps. The prepared samplers were kept in airtight vials in a refrigerator until sampling. Due to the smaller scale of the residential area, the monitoring sites were situated in the center of the village. Therefore, only four passive samplers were installed in the residential area, and fifteen passive samplers were installed in the crop field, at a height of 1.5 m from the ground. Each sampler was replaced after two weeks of exposure to the air. All exposed filters were extracted with 8 ml of deionized water for 20 min in an ultrasonicator. The concentrations of NO2 and NH3 in the extracts were then measured as nitrite and ammonium using an automatic flow analyzer. An automatic weather station (Vantage Pro Plus, Davis Instrument Crop., San Francisco, USA) was also installed 1.5 m from the ground in a crop field (100 m away from the residential area) for meteorological monitoring of surface temperature, relative humidity, wind speed and direction, barometric pressure, and solar radiation. The corrected concentrations of NO2 and NH3 in the air were then calculated using the air temperature and relative humidity recorded by the weather station. The annual mean concentrations of NO2 and NH3 for the whole catchment were taken as the weighted averages of the concentrations of the field and residential areas according to the respective areas in the study catchment. It was estimated that 30% area of the study catchment belong to the residential area, and the rest 70% is crop field area.

2.3. Calculation of the dry deposition velocity For gaseous NO2, the deposition velocity at a height of Z m (Vd (z), m s−1) is expressed as Vd ðzÞ =

1 ðRa ðzÞ + Rb + Rc Þ

where Ra (s m−1) is the aerodynamic resistance (common to all gases) between a specified height (z) and the surface (1.5 m in this study); Rb (s m−1) and Rc (s m−1) are the semi-laminar resistance and the surface resistance, respectively (Wesely, 1989). In contrast to Rc, Ra and Rb can be parameterized from factors describing the meteorological conditions and gas properties by existing techniques. In this study, Ra and Rb were calculated according to the method described by Wesely and Hicks

(1977). The bulk surface resistance Rc can be determined based on the procedures described by Wesely (1989). Rc can be found as Rc =

1 1 rs + rm

1 rlu

+

+

1 rdc + rcl

+

1 rac + rgs

where rs represents the surface bulk resistance component for leaf stomata, rm for leaf mesophyll resistance, rlu for leaf cuticles in healthy vegetation and otherwise (the outer surfaces in the upper canopy), rdc for a gas-phase transfer affected by buoyant convection in canopies, rcl for leaves, twig, bark, or other exposed surfaces in the lower canopy, rac for transfer that depends only on canopy height and density, and rgs for the soil, leaf litter, and so on at the ground surface. The values of related parameters were obtained from the report by Wesely (1989) according to the seasons and surface conditions. For gaseous NH3, the effect of the stomatal compensation point on NH3 deposition was considered due to the counterbalance between deposition and emission. The Vd of NH3 is expressed by: Vd ðzÞ = ð1−

Cs 1 Þ× Ra ðzÞ + Rb + Rc Cair ðzÞ

where Cair(z) (μg m−3) is the concentration of NH3 at 1.5 m. Cs (μg m−3; at 25 °C) is the stomatal compensation point. It results from the thermodynamic and chemical equilibrium between the ammonium concentration in the apoplast (NH+ 4 ) and the gaseous NH3 concentration in the sub-stomatal cavity, which mainly varies with the nitrogen status and the temperature (Loubet et al., 2008). The Cs was calculated as: −12

Cs = 4:79 × 10

 × Γ × exp 10396 ×

 Tleaf −25 298 × ðTleaf + 273Þ

+ where Γ is the ratio of the molar concentrations [NH+ 4 ]/[H ] in the apoplast and Tleaf is the temperature of the plant (°C) (Loubet et al., 2008). Γ is affected by agricultural practices (nitrogen supply) and + plant's developmental stage. It is difficult to measure the NH+ 4 and H concentrations in apoplast. According to the annual fertilization amount of 500–600 kg N ha−1 yr−1 in the study site and the report of Loubet et al. (2008), the Γ value was estimated to be 3000 in this study. The parameter inputs required from meteorological models include values of friction velocity, atmospheric stability, aerodynamic surface roughness, solar radiation, ambient air temperature, humidity, and measures of surface wetness caused by rain and dewfall. The FifthGeneration PSU/NCAR Mesoscale Modeling System, MM5V3.7, was used to generate the required boundary layer meteorology factors that were not available from the weather station (Zhu et al., 2008).

2.4. Calculation of the deposition flux The wet N deposition flux (Fw, kg N ha−1 yr−1) was obtained by multiplying the annual weighted average N concentration by the annual precipitation derived from the volumes collected in the rain samplers. Dry deposition fluxes were calculated as the product of the measured ambient concentration and the modeled deposition velocity. The ambient NO2 and NH3 concentrations were calculated according to Fick's law, using the nitrite and ammonium measured in the extracts and taking into account the local temperature and humidity of the study area in accordance with the Ogawa sampling protocol for NO2 and NH3 (Sampling Protocol Using the Ogawa Sampler, February, 1998). The atmospheric dry deposition flux (Fd, kg N ha−1 yr−1) was calculated by the following equation: Fd = 0:86 × C × Vd where C is the average NO2 or NH3 concentration (μg N m−3), Vd is the dry deposition velocity (m s−1), and 0.86 is a unit conversion factor of μg N m−2 s−1 to kg N ha−1 day−1.

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3. Results and discussion 3.1. Gaseous NO2 and NH3 concentrations The ambient NO2 concentration in the catchment ranged from 16.6 to 68.3 μg m−3, with an annual mean of 42.2 μg m−3 (Table 1). This value is far higher than that observed at several rural sites in south China, which had annual NO2 concentrations ranging from 1.9 to 18.5 μg m−3 (Aas et al., 2007). This value is also higher than that observed at a rural location near Birmingham, UK, which had an annual NO2 concentration of 17 μg m−3 (Marner and Harrison, 2004). However, the value is comparable to NO2 concentrations of 39.5 μg m−3 observed in the suburbs of Beijing (Shen et al., 2008). The background level of atmospheric NO2 was only approximately 0.1 μg m−3 in China (Yu et al., 1997). In the study catchment, the NO2 concentrations were not only close to those observed at typical urban sites in China by the EANET network (EANET, 2007) but also exceeded the mean annual NO2 critical levels of 30 μg m−3 established by the WHO (WHO, 2000). These findings indicate that the rural catchment in this study area is suffering from serious NO2 pollution. The sources of NOx to the atmosphere include lightning and biological fixation of nitrogen, fossil-fuel combustion by power plants and automobiles, biomass burning, and soil emission. It is commonly accepted that NO2 is a typical air pollutant in urban areas caused by combustion of fossil fuels (Li and Lin, 2000). Over the course of a year, east and east– northeast winds prevail in Nanjing city and downtown Jurong city, respectively. The study catchment is located downwind of these two cities. As a result, NO2 produced in urban areas has contributed greatly to the air pollution at the study rural site. Comparatively higher NO2 concentrations were observed in the winter months, particularly December and January, whereas lower concentrations occurred in the summer months (Fig. 2). This seasonal pattern is consistent with previous reports (e.g., Stevenson et al., 2001). The lower NO2 concentration in the summer might be due to greater atmospheric mixing and a higher photochemical reaction, which increase the oxidation of NO2 and its rate of conversion to nitrate by reaction with OH (Atkins and Lee, 1995). The higher concentration in the winter may be attributed to less mixing of air in the lower air boundary and the reduced photochemical reaction in the winter season. In addition, the NO2 generated from the intensive use of fossil fuels in the winter season for heating may also enhance the NO2 concentration. Because of the nature of long-distance transport of NO2, there was no clear difference in the NO2 concentrations between the field site (42.6 μg m−3) and the residential site (41.3 μg m−3). The annual average NH3 concentration for the entire catchment was 4.5 μg m−3 (Table 1), which is higher than the NH3 concentration in most rural sites in East Asia (1.0 μg m−3 to 4.5 μg m−3; EANET, 2007). Although the annual mean NH3 concentration was lower than that in the southeastern United States (5.55 μg m−3, Robarge et al., 2002) and did not exceed the current critical level of 8 μg m−3 (Cape et al., 2009), recent studies have found that the current annual critical level does not protect the most sensitive components of ecosystems, such as lichens, bryophytes, and some sensitive higher plant species. A

Fig. 2. Annual variation of NO2 concentrations in the crop field and residential areas from October 2007 to September 2008. The rainfall is on daily basis.

new critical level of 3 ± 1 μg m−3 was suggested for herbaceous species, whereas 1 μg m−3 has been suggested for lichens and bryophytes in the study of Cape et al. (2009). Therefore, the atmospheric NH3 concentration in the rural catchment poses an environmental risk. The annual average NH3 concentration at the residential sites was 5.4 μg m−3, which was significantly higher than that observed at the field sites (4.1 μg m−3, Table 1). This result is consistent with our emission inventory results estimated in the study catchment, which showed that the NH3 emission per area of 55.0 kg N ha−1 in the residential area is much higher than that of 43.7 kg N ha−1 in the field area (unpublished data). It has been reported that about 90% of all atmospheric NH3 is from a local source and that the ambient NH3 concentration is the most useful parameter for evaluating changes in emission (Erisman and Draaijers, 1995). The higher NH3 concentration in the residential area is likely related to the management of human and livestock excreta. In the catchment, local inhabitants seldom add bedding material to excreta or cover the excreta receptacle; all excreta are usually exposed to the atmosphere directly. In this case, almost 40% of the N in excreta can be emitted to the atmosphere as NH3 and NOx (mainly as NH3 during management) (IPCC guidelines, 2006). Throughout the year, the gaseous NH3 concentrations ranged from 1.3 to 17.2 μg m−3 (Table 1) and were remarkably higher in the summer than winter (Fig.3). A significant positive relationship was also observed between the NH3 concentrations and air temperature in the catchment (P b 0.01, Fig. 4). One explanation of this relationship centers on the fact that the temperature affected the partition equilibrium of NH3 between the gas and liquid phases, as described

Table 1 Comparison of NO-3, NH+ 4 , DON and T-N concentrations in rain and gaseous NO2 and NH3 concentrations in the field and residential areas from October, 2007 to September 2008. Deposition sample

No. of samples

Rain Samples (mg N L-1) 26

Air concentration (μg m-3)

26

Species

NO-3 NH+ 4 DON T-N NO2 NH3

Field area

Residential area

Min

Max

Mean

Min

Max

Mean

0.2 0.2 0 1.2 17.1 1.30

6.2 8.5 9.7 21.8 68.3 16.3

1.3 1.3 0.6 3.2 42.6 ± 2.9 4.1 ± 0.3

0.4 0.1 0 0.8 16.6 1.8

7.6 8.5 3.8 17.7 66.2 17.2

1.4 1.2 0.3 2.9 41.3 ± 3.1 5.4 ± 0.5

t value

Pt N t

Average

-1.44 0.14 1.78 0.37 -1.36 6.54

0.16 0.89 0.09 0.71 0.1 b 0.01**

1.3 1.3 0.5 3.1 42.2 4.5

The t values were determined using SAS (Statistical Analysis System). ** indicates a significant difference between the field and residential areas. Average is weighted average of the concentrations of the field and residential areas.

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by Henry's law (Asman et al., 1998). Further, the higher temperature increased the NH3 emission rate from the soil. In addition, the NH3 concentrations peaked in conjunction with the application of nitrogen fertilizers. In this region, basal fertilizers are usually applied in June and November for summer and winter crop, respectively, whereas topdressing fertilizers are applied in March and July.

+ 3.2. NO− 3 and NH4 concentration in rainwater + The annual average NO− 3 and NH4 concentrations in rainfall were equal (1.3 mg N L−1, Table 1), and they were 4–5 times the levels observed in rural sites in other East Asian countries monitored by EANET (EANET, 2007). Many studies found that NH+ 4 is the dominant N compound in rainwater collected in the countryside (Liu et al., 2006; Xie et al., 2008; Zhao et al., 2009). In the current study, the concentration of NH+ 4 was found to be similar to that in other areas located near the study area (1.35 mg N L−1). However, the NO− 3 concentration was much higher than those in other nearby areas (0.90 mg N L−1, Xie et al., 2008). Although the land in the catchment is used only for agricultural purposes, the surrounding areas are industrially developed. The passage of a highway through the catchment may explain why the level of NO− 3 was much higher in this area than other rural sites in China. Remarkable temporal variations in the concentration of N in rainwater were observed in the catchment (Fig. 5). The highest N concentrations occurred in February and March, which correspond to the lowest monthly rainfall. Conversely, the lowest concentration of N was observed in August, which corresponds to the highest monthly rainfall (Figs. 2, 5). This negative relationship between N concentration and precipitation amount is consistent with the reported results of Liu et al. (2006) and Zhao et al. (2009). In contrast to the gaseous concentration of NH3, no clear difference + between the concentrations of NO− 3 and NH4 in rainwater samples collected at the field site and residential sites were observed (Table 1). N wet deposition occurs via in-cloud scavenging of gases and particles and their deposition in precipitation elements as well as due to belowcloud scavenging occurring by interception of gases and particles (Seinfeld and Pandis, 1998). It is generally agreed that cloudwater concentrations of ammonium dominate the total ammonium measured in rainfall (Goncalves et al., 2003; Nadim et al., 2003). In the case where the ammonia concentration below the cloud base is much

Fig. 4. Relationship between air temperature and NH3 concentrations in the study catchment.

greater than the ammonia concentration that enters the cloud, however, below-cloud scavenging will change the trend of NH+ 4 concentration in rainwater caused by in-cloud scavenging of ammonia (Goncalves et al, 2000; Mizak et al., 2005). In the current catchment, the NH+ 4 in the rainwater from the field and residential areas seems to come mainly from in-cloud ammonia scavenging, where the concentrations of nitrogen species are more homogeneous. 3.3. Organic N concentration in rainwater The annual average DON concentration in the catchment was 0.5 mg N L−1. This value is as high as the inorganic N concentration in the rainwater from several other East Asian countries (e.g., Japan and the Philippines; EANET, 2007). These findings indicate that the DON in rainfall is a non-negligible contributor to N deposition in the study area. No clear differences were observed in the DON between the field and residential sites (Fig. 5c). With the exception of one occasion, no remarkable seasonal variations in DON concentrations were observed in the study sites (Fig. 5c). To date, the sources of DON in precipitation have not been clearly identified. Possible sources of atmospheric organic N include by-products of reactions between NOx and hydrocarbons, marine and terrestrial sources of reduced (amino acid) N, the long-range transport of organic matter (dust, pollen), and bacteria (Neff et al., 2002). Zhang et al. (2008) found that DON may originate from sources similar to inorganic N based on the significant correlation between inorganic and organic N concentrations in the North China Plain; however, this relationship did not exist in the present study or in several other previous studies (Neff et al., 2002). Cornell et al. (2001) reported that urea was a major component of organic N, contributing about half of the organic N and about 15% of total N in both rainwater and aerosols in Hawaii. Therefore, the characteristics of DON likely vary from site to site. Accordingly, future studies should pay greater attention to the contribution of DON, particularly its source and composition, to N deposition. 3.4. Seasonal variation of NO2 and NH3 dry deposition velocity

Fig. 3. Annual variation of NH3 concentrations in the crop field and residential areas from October 2007 to September 2008. The rainfall is on daily basis.

The seasonal variation in the NO2 and NH3 deposition velocity are shown in Fig. 6. The results show that the Vd of NO2 and NH3 in the field area were all higher than that in the residential area, especially in the summer and spring seasons. The values of Vd were determined by the aerodynamic resistance (Ra), quasi-laminar resistance (Rb), and resistance to uptake by the surface elements (Rc). Rc is the dominant resistance, and the contributions of Ra and Rb are normally much smaller (Su et al., 2009). Smith et al. (2000) also reported that the rate

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Fig. 5. Distribution of NO-3 (a), NH+ 4 (b), DON (c), and total N (d) concentrations in rainwater from the crop field and residential sites from October 2007 to September 2008.

of reaction or sorption at the surface has a major influence on the rate of deposition. Thus, the difference in Vd between the field and residential areas was primarily caused by differences in surface conditions. The dry deposition of gases from the atmosphere to a receptor surface is governed by the concentration in the air, turbulent transport processes in the boundary layer, and ability of the surface to capture or absorb gases (Erisman and Draaijers, 1995). Usually, the field area is covered by vegetation, and the gases may react with the epicuticular waxes on leaf surfaces or via diffusion through stomata and subsequent reaction with intercellular fluids; this results in a higher deposition velocity. In the field and residential areas, the calculated annual average deposition velocities (Vd) of NO2 were 0.13 cm s−1 and 0.03 cm s−1, respectively; the Vd values for gaseous NH3 were 0.23 cm s−1 and 0.07 cm s−1, respectively. These findings are comparable to the results of previous studies (Zhang et al., 2004; Marner and Harrison, 2004). 3.5. Wet and dry atmospheric nitrogen deposition in rural regions The seasonal dry NO2 and NH3 deposition fluxes were calculated based on the seasonal average NO2 and NH3 concentrations and the deposition velocity in the field and residential areas. Due to seasonal variation in surface-covering crops in the field area, more gases were deposited in the spring and summer seasons. No seasonal variation was found in the residential area (Table 2). The total gaseous depositions (NO2 plus NH3) were 7.6 kg N ha−1 yr−1 and 1.9 kg N ha−1 yr−1 in the field and residential areas, respectively (Table 2). These values were lower than those (15–16 kg N ha−1 yr−1)

estimated in an arable region of Beijing by Shen et al. (2008) but higher than the average NH3 and NO2 dry deposition flux over South Korea (2.7 kg N ha−1 yr−1; Park et al., 2002). In the study catchment, more gaseous NO2 and NH3 were deposited into the field than residential area (Table 2), as indicated by the gaseous NO2 deposition fluxes of 4.4 kg N ha−1 and 1.1 kg N ha−1 in the field and residential regions, respectively (Table 2). Lü and Tian (2007) reported that NO2 deposition fluxes varied from 0.32 kg N ha−1 yr−1 to 5.80 kg N ha−1 yr−1 in China, averaging 3.03 kg N ha−1 yr−1. Clearly, the NO2 deposition flux in the field area is much higher than the average in China. The NH3 deposition fluxes were only 3.2 kg N ha−1 and 0.83 kg N ha−1 in the field and residential regions, respectively. Overall, the results of this study indicate that dry deposition of NO2 surpasses that of NH3 in the study area, even though it is located in a rural catchment. Wet deposition occurred in each month (Fig. 7). The low deposition in February was mainly due to the sparse rainfall and the high deposition in August was due to heavy fertilization to summer crops in July and August and the heavy rainfall in August. The annual wet deposition levels of inorganic N in the field and residential areas were 27 and 24 kg N ha−1 yr−1, close to the value of 28.0 kg N ha−1 yr−1 observed in neighboring rural sites (Xie et al., 2008). In China, the wet deposition levels of inorganic N were estimated to range from 0.1 kg N ha−1 yr−1 to 62.3 kg N ha−1 yr−1 in different provinces, with an average value of 9.9 kg N ha−1 yr−1(Lü and Tian 2007). Thus, the wet inorganic N deposition flux in the study catchment is a high level in China. Several previous studies showed that NH+ 4 was normally the main component of wet N deposition in rural areas of China (Liu et al., 2006; Zhao et al., 2009). However, the

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nitrogen for sensitive ecosystems (5–10 kg N ha−1 yr−1) (Bobbink and Roelofs, 1995). For dry deposition, however, it has to be noted that we only considered gaseous NO2 and NH3; some N species were missed in this study. Recent studies showed that these two species account for 23– 73% of total inorganic N dry deposition (Zhang et al., 2009; Shen et al., 2009; Park et al., 2002). In that case, the total inorganic N dry deposition should be 10–33 kg N ha−1 in the field area and 2.6–8.1 kg N ha−1 in the residential area. Organic N also exists in the atmosphere. However, its role in atmospheric deposition is unclear, because it is very difficult to obtain organic N's actual deposition rate (Neff et al., 2002). Hauglustaine et al. (1998) reported that peroxyacetyl nitrate and the organic nitrates in many cases are probably the primary sources of N deposition of oxidized N species. Clearly, if all of the N species were considered in the N dry deposition value, then this value could be comparable to of the value describing the N wet deposition fluxes. 3.6. Uncertainties related to the estimated N deposition

Fig. 6. Seasonal variation of NO2 (a) and NH3 (b) deposition velocity in the crop field and residential sites from October 2007 to September 2008. + NO− 3 deposition fluxes reached up to similar level with NH4 deposition fluxes at both the field and residential sites in this catchment (Table 2). This result, together with that of dry deposition, indicates that the high N deposition in the catchment was caused by the higher levels of oxidized N (NO− 3 in precipitation and gaseous NO2) deposition. The DON wet deposition fluxes were 5.3 and 3.2 kg N ha−1 yr−1 in the field and residential areas, respectively (Table 2). Although these values were lower than those observed in the North China Plain (8.6 kg N ha−1 yr−1) (Zhang et al., 2008), they accounted for 16% of the total wet N deposition flux in the study catchment. The total (dry and wet) N deposition values estimated in the current study were 40 and 30 kg N ha−1 yr−1 for the field and residential sites, respectively (Table 2). These values far exceeded the critical load of

In addition to the potentially missing or underreported nitrogen species that may have resulted in underestimation of dry deposition as discussed above, the use of an inferential model may also have resulted in uncertainties in our estimation of dry deposition. Dry deposition models or modules require several types of inputs from observations or from simulations of atmospheric chemistry, meteorology, and surface condition. Surface conditions strongly affect the dry deposition rate through the effect of surface roughness on both deposition velocity and the absorbability of the ground surface to each of the gaseous and particulate species (Loubet et al., 2008). It is indeed difficult to quantify acute N dry deposition to farmlands, however, because the surface conditions of farmlands vary significantly with crop growth and agricultural practices. Another uncertainty source for the dry deposition estimation may arise from the parameters used in the resistance model of dry deposition velocity. These parameters depend largely on the empirical equations derived from early studies in Europe and the USA (Wesely, 1989; Wesely and Hicks, 1977). Wesely and Hicks (2000) reported that uncertainties as large as ±30% are common in deposition velocity estimation. In order to achieve reliable estimates of dry deposition, more experimental parameterizations on site are needed in future research. In this study, we used conversion ratios to exclude dry deposition from the measured bulk deposition. The amount of dry deposition onto bulk precipitation collectors depends on local gas and aerosol concentrations, turbulence intensities, and the collection efficiency of the sampler (Draaijers et al., 1998). The conversion ratios adopted in this study for correcting bulk deposition into wet deposition were obtained by monitoring at a site in a neighboring country. Although the site has similar agricultural practices and climatic conditions as our study catchment, this conversion may have added uncertainty to the estimated amount of wet deposition. There were random observational and systematic errors in precipitation amounts obtained from the rain gauge, as the measurements are sensitive to

Table 2 Dry and wet N deposition fluxes in the study catchment from October, 2007 to September, 2008. sites

Field area

Residential area

Season

Spring (Mar.-May) Summer (Jun.-Aug.) Autumn (Sep.-Nov.) Winter(Dec.-Feb.) Total Spring (Mar.-May) Summer (Jun.-Aug.) Autumn (Sep.-Nov.) Winter(Dec.-Feb.) Total

Wet deposition (kg N ha-1 yr-1)

Dry deposition (kg N ha-1 yr-1)

Total N deposition (kg N ha-1 yr-1)

NO-3

NH+ 4

DON

Total

NO2

NH3

Total

4.7 3.7 2.2 3.0 14 4.2 3.7 2.7 2.7 13

3.2 5.6 1.6 2.9 13 3.3 3.0 2.0 2.9 11

1.1 1.9 1.3 0.93 5.3 0.77 1.7 0.47 0.26 3.2

9.09 11.2 5.03 6.8 32 8.3 8.3 5.1 5.9 28

1.6 ± 0.02 1.9 ± 0.01 0.72 ± 0.01 0.15 ± 0.02 4.4 0.30 ± 0.02 0.22 ± 0.01 0.33 ± 0.01 0.20 ± 0.02 1.1

0.47 ± 0.10 2.3 ± 0.14 0.17 ± 0.06 0.21 ± 0.04 3.2 0.12 ± 0.02 0.31 ± 0.09 0.18 ± 0.04 0.21 ± 0.03 0.82

2.09 4.23 0.89 0.36 7.6 0.42 0.53 0.51 0.41 1.9

11 15 5.9 7.2 40 8.7 8.9 5.6 6.3 30

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Fig. 7. Monthly distribution of wet deposition of NO-3 (a), NH+ 4 (b), DON (c), and total N (d) in the crop field and residential sites from October 2007 to September 2008.

the size, shape and exposure of the rain gauge, as well as wind and topography (Erisman et al., 2003; Ren and Li, 2007). 4. Conclusions Our results showed that a typical rural catchment in a fastdeveloping East China location is suffering from high N deposition. The total nitrogen deposition (wet deposition + dry deposition NO2 and NH3) reached up to 40 kg N ha−1 yr−-1 and 30 kg N ha−1 yr−1 for the field and residential sites, respectively. The DON in the rainwater accounted for 16% of the total wet nitrogen deposition and is therefore a non-negligible component of nitrogen deposition. Oxidized N (NO− 3 in precipitation and gaseous NO2) was the slightly dominant form of N deposited at the catchment. The annual average gaseous concentrations of NO2 and NH3 were 42.2 μg m−3 and 4.5 μg m−3 (0 °C, 760 mm Hg), respectively. This finding indicates that NOx species produced in the surrounding urban area are transported to the study site, where they lead to elevated levels of N deposition. The concentration of NH3 in the air at the study catchment was higher than that in most other rural areas of East Asia and exceeded the critical level for sensitive species. At the catchment, the air concentration of NH3 was higher in residential sites than field sites, indicating that human and livestock excreta are important sources of NH3 and pose an environmental risk. Acknowledgements This research was funded by the Knowledge Innovation Program of the Chinese Academy of Science (No. kzcx2-yw-312), the National Natural Science Foundation of China (No. 40721140018) and the Japan Science and Technology Agency.

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