Emissions of reactive nitrogen
Double cropping of rice and other crops is a general practice in the warm regions of Japan, particularly western Japan. Double or triple cropping of leaf vegetables is also practiced in eastern Japan. However, single cropping is the general practice in most of Japan because of the cold winters and snow. Therefore, the fallow period tends to be long. In the case of rice cropping in eastern Japan, the fallow period is approximately 6 months. The application rate of N fertilizers is usually small, <100 kg N ha−1 per cropping for paddy rice. This small application rate is an environmentally friendly practice and simultaneously optimizes the palatability of rice. In contrast, double cropping with a short fallow period or no fallow period at all is practiced in many parts of China. Furthermore, the application rate of N fertilizers is large, and is typically approximately 200–300 kg N ha−1 per cropping for paddy rice and second crops (e.g. maize, wheat, oil seed rape and cotton). Therefore, there is a large difference in the annual application rates of N fertilizers in rice cropping areas, with <100 kg N ha−1 year−1 in Japan and approximately 400–600 kg N ha−1 year−1 in China. The application rate of N fertilizers is larger for leaf vegetable cropping. For example, the mean application rates of N fertilizers to vegetable fields in Nanjing and Wuxi in China reached 1,862 and 2,749 kg N ha−1 for 2 years, respectively (Wang et al. 2008a). Therefore, Chinese agricultural areas have a high potential for N pollution, including NH3 volatilization loss.
Although Japan strongly depends on imports for food and feed, the self-sufficiency ratio of livestock products in 2006 accounted for 67% on a caloric basis (Ministry of Agriculture, Forestry and Fisheries 2008); however, imported feed accounted for 76% of the domestic production. An important feature of Japanese livestock production is its intensity. Intensive livestock production is accompanied by vast amounts of livestock waste within a narrow area, which may become a strong NH3 emitter. The N load accompanying the treatment and disposal of livestock waste is a serious problem in Japan. In contrast, the N load per area of livestock production in China remains smaller than that in Japan (e.g. Shindo et al. 2006).
The amount of Nr emitted from agricultural activities in Japan and China must be considered. Yan et al. (2003) compiled an inventory of NH3 and NO emissions from croplands in East Asian countries in 1995. The estimated NH3 emissions in China from chemical fertilizers and organic fertilizers, such as animal excreta, were 3.56 and 2.04 Tg N, respectively, whereas those in Japan were 0.059 and 0.069 Tg N, respectively, in which Yan et al. (2003) determined the emission factors for urea and ammonium bicarbonate, the two most consumed N fertilizers in China, through experiments conducted in China. In contrast, the estimated NO emissions in China from chemical fertilizers, crop residues and organic fertilizers were 0.106, 0.005 and 0.058 Tg N, respectively, whereas those in Japan were 0.003, 0.0002 and 0.001 Tg, respectively (Yan et al. 2003). Several studies have examined NH3 emission in Japan. Kannari et al. (2001) showed that the total NH3 emission in Japan in 1994 was 0.43 Tg N, of which 0.26 and 0.023 Tg were emitted from livestock waste and chemical N fertilizers, respectively. Furthermore, Murano and Oishi (2000) showed, regarding NH3 emissions in Japan in 1991, that 0.19 and 0.050 Tg N of NH3 were emitted from livestock waste and chemical N fertilizers, respectively. However, these two studies used NH3 emission factors determined in Europe. In all the three studies, the authors noted that the emission factors were insufficient and did not completely reflect the status in Japan and China. Attention should be given to the development of Nr emission factors specific to Japanese and Chinese agriculture. Yan et al. (2006) also compiled an emission inventory of various substances in relation to biomass burning in China in 2000, in which the total amount of field-burned crop residues was 122 Tg, accounting for 19.4% of the total crop residues. The emissions of NOX, NH3 and PM2.5 from crop residue used as fuel were estimated to be 0.043 and 0.061 Tg N and 1.1 Tg, respectively, and those from field-burned crop residues were estimated to 0.14 and 0.049 Tg N and 0.48 Tg, respectively (Yan et al. 2006). However, Freney et al. (2008) noticed that the estimated annual emission of NH3 in Asia from agricultural practices varied greatly among research groups, ranging from 11.7 to 19.1 Tg N year−1 from 1990 to 2001, whereas the values for NOX emission fell within a narrow range of 0.58–0.66 Tg N year−1. The emission inventory of NH3 has room for improvement.
The actual rate of Nr emission from agroecosystems varies largely with different agricultural practices and their magnitude, including the application rate of N fertilizer. Therefore, a ‘ratio’, for example, NH3 volatilization loss to applied N amount, is a convenient scale to explain the effects of agricultural activities on Nr emissions from agroecosystems. It is, however, noteworthy that the ratio varies depending on the conditions.
Case studies of NH3 emission from farmlands in Japan and China are summarized in Table 1. Case studies at a paddy field in central Japan (Hayashi et al. 2006a, 2008a) showed that NH3 volatilization loss as a result of the application of urea from the surface of the paddy field, excluding rice plants, averaged 1.4%, whereas loss from the whole paddy field was 8.2% (Table 1). In particular, the first supplemental fertilization with a rate of 30 kg N ha−1 resulted in high NH3 volatilization loss (21%), which indicated that rice plants also acted as emitters of NH3 (Hayashi et al. 2008a). Furthermore, Hayashi et al. (2008b) found temporal and rapid increases in the NH4+ concentration in the xylem sap of rice with the application of urea. Thus, NH3 emission from rice plants with excessive N nutrition is highly possible. There are a number of case studies of NH3 emission from Chinese paddy fields. The application of urea by surface incorporation at puddling results in relatively small volatilization losses, 8.8% (Cai et al. 1986) and 10.6% (Cai et al. 1992a) (Table 1). However, supplemental fertilization with urea by surface application boosted the total NH3 volatilization loss to 28.7 and 29.9% (Fan et al. 2006) (Table 1). Furthermore, alkaline soils with a high pH (8.8) showed large volatilization losses of 30 and 39% from applied urea and ammonium bicarbonate, respectively (Zhu et al. 1989) (Table 1). Although acidic conditions generally inhibit NH3 volatilization, a paddy field with a pH of 5.5 showed a large volatilization loss of 40% from applied urea (Cai et al. 1992b) (Table 1). Cai et al. (1992b) attributed this rather extreme result to the high temperature, more than 30°C in many cases, and the shallow floodwater with a high pH (>7), despite the low soil pH.
Table 1. Case studies of ammonia volatilization loss from farmlands in Japan and China
| Tokyo||Fluvisol||5.8||Forage rice||AP||In/B/B||100 × 3||300||300||DC†||12.2||12.2||3|
|ADS||In/B/B||100 × 3||300||103||DC†||51.7||150.8||3|
|ADS||In/B/B||150 × 3||450||154||DC†||37.7||110.0||3|
|ADS||In/B × 5||75 × 6||450||154||DC†||29.6||86.3||3|
| Jiangsu||Lacustrine clay||5.4||Rice||U||In||90||90||90||MM||8.8||8.8||4|
| Jiangsu||Anthrosol||7.2||Rice||U||B × 3||41/53/41||135||135||MM||28.7||28.7||5|
|U||B × 3||81/108/81||270||270||MM||29.9||29.9||5|
| Hubei||Yellow sediment||6.8||Rice||U||B||225||225||225||AT†||20.6||20.6||6|
| Zhejiang||Alluvial- lacustrine deposit||6.0||Rice||U||In||60||60||60||MM||10.6||10.6||7|
| Jiangxi||Red clay||5.5||Rice||U||In||90||90||90||MM||40||40||8|
| Henan||Calcareous sandy loam||8.8||Rice||U||B||90||90||90||MM||30.1||30.1||9|
| Ibaraki||Andosol||5.9||Pac choi||PM||In||120||120||7.3||DC†||0.0||0.7||11|
| Henan||Calcareous sandy loam||8.4–8.8||Maize||U||B+Ir||75||75||75||MM||18||18||12|
| Henan||Calcareous sandy loam||8.8||Maize||U||Pl||80||80||80||MM||12||12||13|
| Hokkaido||Gray upland soil||6.4||Timothy||CS||B||276||276||120||DC||8||18||15|
Volcanic ash soils are typical in Japanese upland fields. A relatively low pH and a high cation exchange capacity are key features of volcanic ash soils, which effectively inhibit NH3 volatilization. For example, a case study at an upland field of volcanic ash soils in northern Japan showed that the application of composted cattle manure by surface incorporation to bare soil surface resulted in a very small NH3 volatilization loss (0.8% to total N), and the subsequent application of chemical fertilizers as a mixture of ammonium sulfate (AS) and ammonium phosphates by placement (3 cm depth) resulted in negligible loss (Hayashi et al. 2009a) (Table 1). Moreover, case studies at an upland field of light-colored volcanic ash soils in central Japan also showed negligible NH3 volatilization losses from surface-incorporated manure and AS (Hayashi et al. 2009b) (Table 1). However, Matsunaka et al. (2008) reported a relatively large NH3 loss in grassland on grey upland soils in northern Japan, where surface applications of dairy cattle slurry resulted in the loss of 8–17% to total N (or 18–35% to NH4-N) (Table 1). Comprehensive studies in an upland field of calcareous soils with a high pH of 8.8 used for rotations of maize and wheat in the North China Plain showed that the NH3 volatilization losses from applied urea to maize as summer cropping were 26–48%, 18% and 11–12% for surface application, broadcast followed by irrigation and deep placement, respectively (Cai et al. 2002; Zhang et al. 1992) (Table 1). Despite the strong inhibition effect of deep placement, 11–12% of volatilization loss still occurred. To a certain extent this may be ascribed to the high soil pH of 8.8. Fertilizer application at farmlands with a high soil pH presents the problem of N use efficiency as well as the emission of airborne N. In contrast, growing wheat as a winter crop in the same field resulted in relatively small volatilization losses of 2.0–15%, 0.6% and 2.3% for surface application, broadcast followed by irrigation and deep placement, respectively (Cai et al. 2002) (Table 1). The small losses after wheat cropping may result from the low temperature at fertilization in autumn and the lack of canopy emission without plants at fertilization.
Emission of NO from farmlands results from denitrification. Therefore, soil conditions appear to be the strongest factors controlling NO emission. Unfortunately, only a limited number of studies have been conducted, and these studies form part of N2O related studies (Table 2). Case studies at an upland field of light-colored volcanic ash soils in central Japan showed a range of NO emissions to applied total N (0.02–1.19%) (Akiyama and Tsuruta 2002, 2003; Akiyama et al. 2000) (Table 2), with a mean value of 0.48%. Katayanagi et al. (2008) reported a similar result in maize fields in northern Japan (0.44%) (Table 2). They also reported 0.24–0.28% of emissions from grassland and pasture (Table 2). However, Kusa et al. (2006) recorded an extremely high emission of NO (1.64–3.98%) at an upland field of volcanic ash soils with poor drainage in northern Japan (Table 2). Highly activated denitrification may result in a high NO emission in addition to N2O.
Table 2. Case studies of nitrogen oxide emission from farmlands in Japan
| Hokkaido||Andosol (poor drainage)||5.6||Maize||CF+M 1||128||CC||1.64||1|
| Hokkaido||Andosol and Histosol||5.2–5.3||Maize||CF+M||220||CC||0.44||2|
|Andosol (partly Histosol)||5.0–6.1||Grassland||CF+M||143||CC||0.28||2|
| Ibaraki||Andosol||5.9||Pac choi||CN 1||150||CC||0.32||4|
| Ibaraki||Andosol||5.9||Pac choi||PM||150||CC||0.07||5|
Atmospheric deposition of reactive nitrogen
The air concentration provides important information for identifying the degree of air pollution and for evaluating the rate of dry deposition. However, the rate of wet deposition also increases at a high air concentration owing to below-cloud scavenging (e.g. Hojito et al. 2006b). The annual means of the air concentration of Nr in Japan and China are shown in Table 3. Particularly high concentrations of NH3 are found at several sites in Japan and China (Table 3); these sites were affected by neighboring livestock facilities (Hojito et al. 2006a; Hu et al. 2007). Freney et al. (2008) also pointed out that the background concentrations of NH3 in Chinese agroecosystems were high, that is, 20–25 μg N m−3, with extreme values exceeding 45 μg N m−3 and no values below 5 μg N m−3. Although the amount of data is limited, the air concentrations of NOX and particles tend to be higher in China than in Japan (Table 3).
Table 3. Annual mean values of air concentration of reactive nitrogen in Japan and China
| Hokkaido||Rishiri||45°07′N||141°12′E||40||Re||B||FP, AT||2002–2007||0.3||0.0||–||0.0||–||0.6||0.3||0.1||1|
| Aomori||Tappi||41°15′N||141°21′E||105||Re||B||FP, AT||2003–2007||0.3||0.1||–||0.0||–||0.9||0.5||0.3||1|
| Niigata||Sado-seki||38°14′N||138°24′E||136||Re||B||FP, AT||2003–2007||0.4||0.2||–||0.0||–||0.9||0.5||0.2||1|
| Tochigi||NILGS 1||36°55′N||139°57′E||350||Ru||M||PS||2004||3.2||–||–||–||–||–||–||–||3|
| Tochigi||NILGS 2||36°55′N||139°56′E||310||Ru||M||PS||2004||9.2||–||–||–||–||–||–||–||3|
| Tokyo||Ogasawara||27°05′N||142°13′E||230||Re||B||FP, AT||2003–2007||0.3||0.0||–||0.0||–||0.3||0.2||0.1||1|
| Nagano||Happo||36°42′N||137°48′E||1850||Re||B||FP, AT||2003–2007||0.3||0.2||–||0.0||–||1.4||0.5||0.1||1|
| Gifu||Ijira||35°34′N||136°41′E||140||Ru||B||FP, AT||2003–2007||0.8||0.2||–||0.2||–||2.0||1.1||0.1||1|
| Shimane||Oki||36°17′N||133°11′E||90||Re||B||FP, AT||2002–2007||0.5||0.1||–||0.0||–||0.9||0.8||0.3||1|
| Shimane||Banryu||34°41′N||131°48′E||53||U||B||FP, AT||2003–2007||0.7||0.2||–||0.2||2.1||2.3||1.0||0.3||1|
| Kochi||Yusuhara||32°22′N||132°56′E||790||Re||B||FP, AT||2003–2007||0.3||0.3||–||0.1||–||1.0||1.1||0.1||1|
| Okinawa||Hedo||26°52′N||128°15′E||60||Re||B||FP, AT||2003–2007||0.5||0.1||–||0.0||–||0.5||0.6||0.3||1|
| Jinagxi||Yingtan (Forest Micro-Met. Sub-Station)||Re||NS||AT, GT, FT||2003–2004||67.6||–||–||–||5.2||–||6.5||12.3||6|
| Xiamen||Hongwen||24°28′N||118°08′E||50||U||B||FP||2007||4.9||0.6||–||–|| ||–||3.1||1.4||1|
| Hongkong||Cape D’Aequier||22°12′N||114°15′E||1||Re||M||PS||1999–2000||0.6||–||–||–||–||–||–||–||4|
The atmospheric deposition of Nr as the sum of the wet and dry deposition at 10 sites in Japan, eight of which were remote sites, ranged from 3.1 to 18.2 kg N ha−1 year−1 as a 5-year mean from 2003 to 2007 (Ministry of the Environment 2009). The largest deposition was recorded at a rural site, Ijira, in central Japan, and the smallest depositions were recorded at clean remote islands, such as Rishiri in northernmost Japan and Ogasawara in the Pacific Ocean. According to earlier nationwide monitoring implemented by the Ministry of the Environment, Japan, during 1983–2002, wet deposition of Nr (as bulk precipitation during 1983–1988 and wet-only precipitation after that) at 23, 29 and 29 remote, rural and urban sites was 7.0, 7.9 and 8.2 kg N ha−1 year−1, respectively. Wet deposition of Nr was greatest at the urban sites along the central part of the Japan Sea coastal regions, with a mean value of 11.4 kg N ha−1 year−1; however, in addition to this region, wet deposition of NO3− was also higher on the Pacific coast and the Southwestern Islands regions, and wet deposition of NH4+ was also higher on the Pacific coast and the East China Sea coastal regions (Hayashi et al. 2006b). In contrast, Lü and Tian (2007), who attempted to evaluate the wet deposition of NH4+ and NO3− and the dry deposition of NO2 across China from 1990 to 2003, demonstrated that atmospheric deposition peaked over south central China, with a mean value of 12.9 kg N ha−1 year−1 and a maximum value of 63.5 kg N ha−1 year−1.
Table 4 is a summary of the recent data of annual atmospheric deposition in Japan and China. The rates of wet deposition of NH4+ were on par with those of NO3− in Japan in many cases; however, the contribution of NH4+ was greater than that of NO3− at some sites. In particular, the rates of wet deposition of NH4+ outweighed those of NO3− at the Hinoki and Aoki sites in an intensive dairy farming area in Tochigi Prefecture, central Japan (Hojito et al. 2006b) (Table 4); however, these data were obtained by open bulk samplers that involved dry deposition to some extent. In contrast, the maximum level of wet deposition of inorganic N (17.7 kg N ha−1 year−1) was observed at the Kajikawa site in Niigata Prefecture, central Japan, a remote forest site in the central part of the Japan Sea coastal region (Kamisako et al. 2008) (Table 4). In addition, Sase et al. (2008) used filtering-type bulk samplers to prevent evaporation of the collected samples or deterioration as a result of reactions with gases or aerosols in the atmosphere. Although limited data are available, a range of 1.3–5.5 kg N ha−1 year−1 of wet deposition of organic N was reported in central Japan (Hayashi et al. 2007; Hojito et al. 2006b). The rates of wet deposition of NH4+ were also on par with those of NO3− in China; however, NH4+ showed larger contributions in many cases (Table 4). The larger NH4+ wet deposition was mainly ascribed to the urban and agricultural sources of NH3 in the urban and non-urban sites, respectively. The rates of wet deposition of inorganic N at remote sites in China were similar to those in Japan, whereas those at other sites in China (22.7 kg N ha−1 year−1 as a mean value of Table 4 involving open bulk data) were several-fold greater than those in Japan. The rates of wet deposition of organic N in China ranged from 0.7 to 26.4 kg N ha−1 year−1 (Table 4) and these rates are also likely to be greater than the rates in Japan, excluding inland remote sites.
Table 4. Annual wet and dry deposition of reactive nitrogen in Japan and China
| All Japan||23 remote sites|| || || ||Re||D,W,B,M||WOB||–||1983–2002||–||3.7||3.4||–||–||–||–||–||–||–||1|
| All Japan||29 rural sites|| || || ||Ru||D,W,B,M||WOB||–||1983–2002||–||4.5||3.4||–||–||–||–||–||–||–||1|
| All Japan||29 urban sites|| || || ||U||D,W,B,M||WOB||–||1983–2002||–||4.7||3.6||–||–||–||–||–||–||–||1|
| Tochigi||NILGS 1||36°55′N||139°57′E||350||Ru||M||OB||–||2004||–||7.0||7.1||1.8||–||–||–||–||–||–||7|
| Tochigi||NILGS 2||36°55′N||139°56′E||310||Ru||M||OB||–||2004||–||7.8||7.1||2.2||–||–||–||–||–||–||7|
| Liaoning||Dalian|| || || ||Ru||D||OB||–||2006||–||8.1||10.3||10.9||–||–||–||–||–||–||8|
| Beijing||China Agric. Univ.||39°57′N||116°18′E|| ||SU||E||OB||–||1999–2004||–||19.8||9.5||–||–||–||–||–||–||–||9|
| Beijing||Dongbeiwang||40°03′N||116°17′E|| ||SU||E||OB||–||2003–2004||–||24.1||10.3||–||–||–||–||–||–||–||9|
| Beijing||Changping|| || || ||Ru||D||OB||–||2005–2006||–||20.1||13.3||4.6||–||–||–||–||–||–||8|
| Beijing||China Agric. Univ.|| || || ||SU||D||OB||–||2005–2006||–||17.4||10.5||10.3||–||–||–||–||–||–||8|
| Beijing||Dongbeiwang|| || || ||SU||D||OB||–||2005–2006||–||27.0||12.5||9.5||–||–||–||–||–||–||8|
| Beijing||Shunyi (Yanhe)|| || || ||Ru||D||OB||–||2005–2006||–||16.1||8.5||3.1||–||–||–||–||–||–||8|
| Beijing||Fangshan|| || || ||Ru||E||OB||–||2005–2006||–||18.3||11.3||–||–||–||–||–||–||–||10|
| Beijing||Beijing Academy of Agro-Forestry Sciences|| || || ||U||E||OB||–||2005–2006||–||18.7||10.3||–||–||–||–||–||–||–||10|
| Beijing||Shunyi (Liangshan)|| || || ||Ru||E||OB||–||2005–2006||–||19.4||8.7||–||–||–||–||–||–||–||10|
| Hebei||Baoding|| || || ||Ru||D||OB||–||2005–2006||–||14.1||8.6||4.6||–||–||–||–||–||–||8|
| Hebei||Quzhou|| || || ||Ru||D||OB||–||2005–2006||–||10.5||5.6||5.5||–||–||–||–||–||–||8|
| Hebei||Wuqiao|| || || ||Ru||D||OB||–||2005–2006||–||10.1||5.9||6.6||–||–||–||–||–||–||8|
| Shandong||Huimin|| || || ||Ru||D||OB||–||2005–2006||–||19.1||11.0||8.6||–||–||–||–||–||–||8|
| Shandong||Qingdao|| || || ||Ru||D||OB||–||2006||–||7.9||5.5||7.4||–||–||–||–||–||–||8|
| Shanghai||44 urban, suburban, and rural sites|| || || || ||NS||NS||–||1998–2003||–||26.6||31.5||–||–||–||–||–||–||–||11|
| Changshu||Changshu||31°33′N||120°42′E|| ||Ru||E||OB||–||2001–2003||–||12.9||9.5||4.7||–||–||–||–||–||–||12|
| Changshu||Changshu||31°32′N||120°42′E|| ||Ru||E||WO||–||2003–2005||–||16.9||11.0||–||–||–||–||–||–||–||13|
| Changshu||Wuxi||31°36′N||120°28′E|| ||Ru||E||WO||–||2003–2005||–||15.6||10.7||–||–||–||–||–||–||–||13|
| Sichuan||Gonggashan|| || || ||Re||D||OB||–||2005–2006||–||4.7||2.6||26.4||–||–||–||–||–||–||8|
| Zhejiang||Fenghua|| || || ||Ru||D||OB||–||2004–2005||–||20.0||10.7||23.1||–||–||–||–||–||–||8|
| Jiangxi||Yingtan (Forest Micro-Met. Sub-Station)|| || || ||Re||M||OB||Inf||2003–2004||2003–2004||5.2||12.4||–||39.0||–||–||6.0||2.8||5.3||15|
| Jiangxi||Yingtan||28°12′N||117°00′E|| ||Ru||M||WO||Inf||2004–2005||2004–2005||12.3||14.1||4.2||22.0||–||–||1.3||2.5||2.5||16|
| Zhuhai||Xiang Zhou||22°16′N||113°34′E||40||U||D||WO||–||2000–2007||–||10.8||4.1||–||–||–||–||–||–||–||2|
| Zhuhai||Zhuxian Cavern||22°12′N||113°31′E||45||U||D||WO||–||2000–2005||–||9.4||5.2||–||–||–||–||–||–||–||2|
| Guangdong||Dinghushan||23°10′N||112°10′E|| ||Re||NS||OB||–||2003||–||23.2||10.9||–||–||–||–||–||–||–||18|
| Inner Mongolia||Duolun|| || || ||Re||D||OB||–||2006||–||13.2||3.5||0.7||–||–||–||–||–|| ||8|
| Xingjiang||Urumchi|| || || ||Ru||D||OB||–||2006||–||3.6||1.3||1.5||–||–||–||–||–||–||8|
| Tibet||Linzhi|| || || ||Re||D||OB||–||2005–2006||–||2.4||0.0||5.6||–||–||–||–||–||–||8|
| Yellow Sea||Qianliyan Island|| || || ||Re||E||WO||–||2000–2003||–||0.0||3.4||–||–||–||–||–||0.0||1.4||19|
| East China Sea||Shengsi Archipelago|| || || ||Re||E||WO||–||2000–2003||–||4.6||3.0||–||–||–||–||–||1.0||1.0||19|
In contrast to the large quantity of data for wet deposition, limited information has been available for dry deposition in Japan and China. Case studies in Japan have shown smaller rates of dry than wet deposition (Table 4). However, dry deposition accounted for 40% at maximum (Ministry of the Environment 2009) and 51% (Hayashi et al. 2007) of the atmospheric deposition of inorganic N. The actual contribution of dry deposition to total deposition appears greater because these studies excluded the dry deposition of NO2. Two case studies in China have shown an extremely large rate of dry deposition of NH3, 22.0 and 39.0 kg N ha−1 year−1 (Hu et al. 2007; Wang et al. 2008b) (Table 4). The large deposition resulted from the high air concentrations of NH3 arising from neighboring livestock facilities, such as piggeries. Local sources strongly affect dry deposition of NH3. All case studies shown in Table 4 adopted the inferential method (or resistance model) (e.g. Hicks et al. 1987) to estimate the dry deposition flux. By definition, the resistance model infers a deposition velocity on the basis of micrometeorological data and parameterizations of the effects of surface conditions, such as vegetation and soil. However, the parameterizations depend largely on the empirical equations derived from early studies in Europe and the USA. Although methodological improvements have been progressing (e.g. Matsuda 2008; Zhang et al. 2003), the development of an improved inferential method that takes the key features of Japanese and Chinese agroecosystems into consideration is still required.
Shindo et al. (2005) revealed that long-term N deposition is an important factor to determine the N concentration in stream water in natural ecosystems on the basis of 1 km × 1 km grid estimation for the whole of Japan. Xie et al. (2007) also recognized the importance of atmospheric deposition as a source of N pollution of water bodies, although these researchers held to the view that N pollution in the Taihu Lake region in south-eastern China should be primarily ascribed to the direct discharges of urban sewage and rural human and animal excreta without treatment.
In 2002–2007, in a mountainous forested catchment in Niigata, central Japan, which received the highest level of N deposition, leaching of N to stream water occurred even during the growing season, and the rate of N deposition certainly contributed to the high NO3− concentrations in the stream water (Kamisako et al. 2008). Matsubara et al. (2009) reported the current serious status of the river water chemistry in Niigata and Gifu, central Japan, which is characterized by a significant long-term decline in pH (Niigata, 1986–2003; Gifu, 1988–2003). These researchers pointed out the possible impact of acid deposition, including inorganic N species, on the acidification of river water because of the higher levels of deposition load than those in Europe and the USA.
Earnest studies have also been conducted in China in relation to the impact of N deposition on forested catchments. Du et al. (2008) concluded that the Shaoshan Forest, a subtropical evergreen mixed forest in south-central China, was still far from N saturation, despite the high N deposition. In contrast, Fang et al. (2009) found large and rapid N leaching from a mature monsoon broadleaf forest in southern China, which accounted for 80% of N deposition, perhaps as a result of long-term N accumulation by the high N deposition of >30 kg N ha−1 year−1 over the course of 15 years. An increase in N deposition generally enhances the decomposition of organic matter because of the improvement in N nutrition (e.g. Månsson and Falkengren-Grerup 2003). However, further additions of N to subtropical and tropical forests in southern China, which have already been exposed to high N deposition, have resulted in significant reductions in litter decomposition and soil respiration (Fang et al. 2007; Mo et al. 2006, 2007, 2008). Mo et al. (2007) also reported that the reduction occurred primarily in the hot and wet growing season. More attention should be given to the diverse ecosystems in East Asia and to their various responses against N load.