Chinese coastal seas are facing heavy atmospheric nitrogen deposition

The study included six sampling sites and all the sampling sites were located in the coastal region (near the sea) or on islands. The sampling sites were set up in Liaoning, Shandong, Zhejiang, Fujian, and Guangdong provinces, respectively (figure 1). The specific sites are Dalian (DL, 38°44'N, 120°31'E), Changdao island (CD, 37°55'N, 124°44'E), Linshan island (LS, 35°45'N, 120°08'E), Fenghua (FH, 29°36'N, 121°31'E), Fuzhou (FZ, 26°10'N, 119°21'E), and Zhanjiang (ZJ, 21°12'N, 110°17'E), respectively. Dalian is a coastal city and is famous for business development. DL is an urban site where the mean annual temperature is 8–11 °C and the annual rainfall varies between 550 and 1000 mm. DL belongs to the warm temperate and semi-humid monsoon and marine climatic zone. CD is largest island in Shandong province, located in the joint of the Bohai Sea and the Yellow Sea with about 40 000 inhabitants. CD island is remote and does not have industry or agricultural activities near the sampling site. LS is located in the Yellow Sea belonging to Jiaonan city (region), Shandong province. There are about 2700 inhabitants on the island. FH is located in Chunhu town, Fenghua city, Zhejiang province. It is a typical coastal area of the East China Sea and has an advanced fishery. FZ is surrounded by hills and has low population density. The FZ site is surrounded by some agricultural land growing tobacco and paddy rice without large livestock farms or industry. ZJ is located in Leizhou Peninsula, Zhanjiang city, Guangdong province, close to the South China Sea. The climate is tropical with relatively high temperatures, high rainfall, and high plant cover. This area is rich in tropical fruits and farmers use large quantities of fertilizers for fruit production. Although only the DL is located at real urban site, all sampling sites (except FZ) are affected by rapid urbanization (e.g., increased traffic vehicles and population intensities) in coastal area of China. Besides the local N_r sources, all coastal sites receive atmospheric N_r transport from both inland (mainly agricultural sources) and marine (including international ship emissions) ecosystems.

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Reactive N components were collected by the DELTA (DEnuder for Long-Term Atmospheric Sampling) system which is an active sampling equipment, fulfilling long-time sampling and thus cost saving and it is widely used in the European N deposition monitoring network (Flechard et al 2011). Atmospheric HNO₃, NH₃, particulate NO₃⁻ and NH₄⁺ can be captured by the coated wall of the denuders and the subsequent filters. The sampling time is one month. The coating solution is 5% m/v citric acid in methanol for absorbing NH₃, and 1%(m/v) K₂CO₃ + 1%(m/v) glycerol in methanol for collection of HNO₃. The coating solution was 13%(m/v) citric acid in methanol added to the filters absorbing pNH₄⁺, and the coating solution which absorbs pNO₃⁻ is 5%(m/v) K₂CO₃ + 10%(m/v) glycerol in methanol. The extraction solution is 0.05% H₂O₂ for the denuders and filters absorbing NO₃⁻ (10 ml) and the extraction solution for NH₃ and pNH₄⁺ is high purity water (denuders 6 ml and filters 10 ml). The NH₄⁺–N and NO₃⁻–N concentrations in the samples were analyzed by an AA₃ continuous-flow analyzer (Bran + Luebbe GmbH, Norderstedt, Germany). The DELTA system also absorbs ${\rm S}{{{\rm O}}_{2}}$ and ${\rm SO}_{4}^{2-}$ by the denuders and filters. ${\rm SO}_{4}^{2-}$ concentration was measured by ion chromatography (761 Compact, Switzerland). Atmospheric concentrations of NO₂ were sampled by passive samplers (UK Environmental Change Network Diffusion tubes) extensively described in Luo et al (2013). Monthly reactive N concentrations were collected mostly starting from August or September 2010 and ending in May or June 2012. The detailed sampling periods and the numbers of reactive N and SO₄²⁻ samples at the six sites are listed in table 1.

Rain water samples were collected by precipitation collectors directly after every rainfall event. Wet deposition of N was collected in situ by precipitation collectors (SDM6, Tianjin Weather Equipment, China). The sampling period was one complete year (2011). The samples were analyzed for concentration of NH⁺₄–N and NO⁻₃–N using the AA3 continuous-flow analyzer mentioned above. Wet deposition of NH₄⁺–N and NO₃⁻–N was then calculated by the following two equations:

N deposition per event (kg N ha⁻¹) = precipitation(mm) × NH₄⁺–N or NO₃⁻–N concentration in rainwater (mg N L⁻¹) × 0.01.

N deposition per month or year (kg N ha⁻¹) = 0.001 × ∑N deposition per event in a month or year.

Compared with wet deposition N dry deposition is relatively complicated because it changes with the meteorological conditions and the surface characteristics. In the inferential model the dry deposition flux can be expressed by

$$\begin{eqnarray}&&F=C\times {{V}_{{\rm d}}}\end{eqnarray} \tag{ 1 }$$

and V_d can be expressed by

$$\begin{eqnarray}&&{{V}_{{\rm d}}}={{\left( {{R}_{{\rm a}}}+{{R}_{{\rm b}}}+{{R}_{{\rm c}}} \right)}^{-1}}.\end{eqnarray} \tag{ 2 }$$

Here, R_a is the aerodynamic resistance, R_b is the quasi-laminar boundary layer resistance, and R_c is the surface or canopy resistance. In the three resistances, R_a is dependent on the meteorological conditions and it should be calculated for a range of atmospheric stabilities, and R_b is mainly about the gas own characteristic. R_c is quite complicated in inferential model and difficult for dry deposition calculation due to the different surface features and gas or particles species (Erisman and Pul Van 1994).

In this study we did not have enough equipment or suitable methods to calculate the dry deposition velocities at the ocean surface. In addition, studies on N dry deposition velocities above the ocean surface are very limited, so we decided to use the deposition velocities from the Chinese literature on the ocean. We therefore cited deposition velocities of the gases NH₃, HNO₃, and NO₂ from Zhang et al (2010) and those of particulate pNH₄⁺ and pNO₃⁻ from Zhang et al (2004).

Atmospheric concentrations of reactive N components are shown in figure 2. Annual average NH₃ concentrations were 2.73, 3.50, 3.97, 7.34, 2.13, and 5.62 μg N m⁻³ at sites DL, CD, LS, FH, FZ, and ZJ, respectively. The concentrations of NH₃ were higher at FH and ZJ than at the other sites because of the high level of agricultural activity. HNO₃ and NO₂ concentrations ranged from 0.46 to 1.22 μg N m⁻³ and 3.10 to 7.19 μg N m⁻³ across all six sites. The lowest concentrations of HNO₃ and NO₂ were observed at FZ and the highest were found at FH and CD, reflecting the weak effects of power plants, industry and transportation on the concentrations measured at FZ but strong effects on FH and CD because of the large population densities and advanced economic activities. Average pNO₃⁻ concentrations were 1.14–2.68 μg N m⁻³ at the six sampling sites. The highest value was measured at CD and the lowest at FZ. Average concentrations of pNH₄⁺ were 4.16, 4.45, 4.78, 4.07, 2.29, 3.30 μg N m⁻³ at sites DL, CD, LS, FH, FZ, ZJ, respectively. Air concentrations of pNH₄⁺ at the North coastal sites (DL, CD and LS) were almost twice of those at the Southern FZ site, whereas the concentrations of pNH₄⁺ at FH and ZJ were close to the North coastal sites. Three North sampling sites located close to the North China Plain, an intensively managed agricultural region, and the higher pNH₄⁺ concentrations in this area showed agricultural source impacts on secondary aerosols (e.g., NH₄⁺-based PM_2.5).

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Table 2 shows the seasonal variation in reactive N concentrations at the six sampling sites. A peak NH₃ concentration was found in summer at all sites (except DL) because high temperatures together with N fertilization in summer induce high NH₃ emissions from various sources. The lowest concentration of NH₃ was observed in winter at all sampling sites. The air concentration of pNH₄⁺ was highest in spring or winter. The results may be explained by low rainfall in the spring and winter, and in addition, coal consumption for heating in the winter and spring can also stimulate the formation of particulate NH₄⁺. HNO₃ concentrations did not show notable seasonal variation across the six sampling sites. Peak concentrations of pNO₃⁻ were observed in winter or spring and lowest in the summer. Seasonal variation in pNO₃⁻ is similar to that of pNH₄⁺ in air as mentioned above. The meteorology and heating stimulate higher pNO₃⁻ pollution especially in North China in winter. The pNO₃⁻ concentrations were 24.2% higher at coastal sites of North China (DL, CD and LS) than at sites of South China (FH, FZ and ZJ) in winter. Concentrations of NO₂ mainly peaked in winter or spring and had a close relationship with heating, especially in North China.

Rainfall distribution showed seasonal and regional variations. South China sampling sites had more rainfall than the Northern sites. The latter had lesser or even no rain in winter and spring but the Southern sites experienced more rain in these two seasons. The inorganic N concentrations in rainwater were 0.46–1.67 mg N L⁻¹ for NO₃⁻–N and 0.47–1.31 mg N L⁻¹ for NH₄⁺–N (table 3). Nitrogen concentrations in rainwater followed the pattern in atmospheric reactive N concentrations. For example, CD and FZ had the highest and lowest inorganic N concentrations in rainwater in relation to the highest and lowest concentrations of atmospheric reactive N at the two sites, respectively. N wet deposition ranged from 14.2 to 25.2 kg N ha⁻¹ yr⁻¹. N wet deposition rate was controlled by the N concentration in rainwater and the amount of annual rainfall. The NH₄⁺/NO₃⁻ ratio was lower than 1.0 except at ZJ, so wet deposition of NO₃⁻–N was slightly higher than that of NH₄⁺–N in our study.

Annual N dry deposition was 12.9, 16.9, 15.2, 23.1, 7.8 and 18.0 kg N ha⁻¹ at DL, CD, LS, FH, FZ, and ZJ sampling sites, respectively (figure 3(a)). The estimated dry deposition velocities of NH₃, HNO₃, NO₂, pNH₄⁺ and pNO₃⁻ averaged 0.626, 0.63, 0.005 34, 0.27 and 0.27 cm s⁻¹, respectively, which were taken directly from (Zhang et al 2004, 2010). NH₃ and pNH₄⁺ were the major parts in the N dry deposition because they have both higher concentrations and higher deposition velocities, and the total NH₃ and pNH₄⁺ contribution to the dry deposition was 72.1–81.7%. FH had the highest N dry deposition rate because of the highest reactive N concentrations (especially NH₃ and NO₂). FZ had the lowest dry deposition rate because all reactive N components were very low at this site. In general, the N dry deposition rate is comparable to the N wet deposition in our study (except FZ). Given the wet N deposition, the annual total N deposition rates were 29.9, 32.8, 31.1, 44.6, 22.0, 43.2 kg ha⁻¹ yr⁻¹ at DL, CD, LS, FH, FZ, and ZJ (figure 3(b)), respectively.

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NH₃ and NO_x are two primary reactive N species in air which are emitted mainly from human activity. Rapid economic development in China has led to continuing increases in NH₃ and NO_x emissions since the 1980s (Liu et al 2013). However, the N emissions varied from region to region and also from season to season due to the large differences in population density, economic function, land use types and weather conditions. The NH₃ emission 'hot-spots' were in Hebei, Henan, Shandong and Jiangsu provinces (Clarisse et al 2009, Zhang et al 2010), while the NO_x emission 'hotspots' were in some well developed regions such as Jing-Jin-Ji (Beijing-Tianjin-Hebei) Region, the Yangtze River Delta and the Pearl River Delta (Richter et al 2005, Gu et al 2012). In this study the NH₃, HNO₃ and NO₂ concentrations were much lower than our previous results on the North China Plain (Shen et al 2009, Luo et al 2013), but are comparable to the results from three South China sampling sites in Hunan province (Shen et al 2013).

Atmospheric pNH₄⁺, pNO₃⁻, and pSO₄²⁻ are important components of secondary aerosols. Particulate pollution is a major public concern and there are many studies focusing on urban regions (Chan and Yao 2008). Our data show that atmospheric concentrations of pNH₄⁺ and pNO₃⁻ or pSO₄²⁻ were significantly correlated (figures 4(a) and (b)). The ${\rm pNO}_{3}^{-},\;{\rm pSO}_{4}^{2-}\;{\rm and}\;{\rm pNH}_{{\rm 4}}^{+}$ made an important contribution to the formation of secondary particulates (e.g., NH₄NO₃ and (NH₄)₂SO₄) and their transportation to the coastal sea. China has made great progress in controlling particulate matter but concentrations remain high because industrial development has offset the PM control technology and the conditions of particulate pollution are even more problematic in the cities of North China (Wang and Hao 2012). Compared with particulate NH₄⁺ and NO₃⁻ concentrations on the North China Plain (Luo et al 2013), the concentrations of pNH₄⁺ and pNO₃⁻ at the six coastal sites were lower especially at one Southeast sampling site (FZ). China uses large amounts of coal which accounts for about 70% of its energy consumption. SO₄²⁻ is therefore one of the most important components in particulate matter (Chan and Yao 2008). The SO₄²⁻/NO₃⁻ ratio is very low in this study (SO₄²⁻/NO₃⁻ ratios were 0.48, 0.49, 0.69, 0.49, 0.87, 0.50 at DL, CD, LS, FH, FZ, and ZJ, respectively), reflecting the possibly higher contribution of the transportation sector and rapid urbanization to the present-day air pollution. Our results are consistent with recent reports of decreasing SO₄²⁻/NO₃⁻ trends in Beijing and Shanghai (Shen et al 2011b, Huang et al 2012). Natural processes can also produce aerosols. Cusack et al (2012) found crustal material and marine aerosol comprising 9% and 2% respectively of the total sum of the chemical components at a sampling site in Spain. In addition, forest fires and volcanic eruption form large amounts of natural-source aerosols.

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Regarding seasonal variation and taking NH₃ as an example, we know that fertilized soils, animal/human manures and biomass burning are major NH₃ emission sources worldwide. These emission sources have a close relationship with temperature and high temperatures stimulate NH₃ emission in summer (Shen et al 2011a). The oceans are important natural sources of NH₃ emissions Olivier et al (1998) and high temperatures also stimulate NH₃ emissions from the oceans in summer. Meteorological conditions may affect the air quality in the coastal sea in different seasons. The wind comes mainly from the sea in the summer and brings relatively clear air from the ocean which is also an explanation for the lower pNH₄⁺ and pNO₃⁻ concentrations in summer. Wind comes mainly from the West or North in winter and the sampling sites were affected by the land sources. As a result, particulate pollution is usually higher in winter and spring. There are always sand storms from Inner Mongolia in spring and there can be long-range transport to the China seas affecting the air quality in this region and to a lesser extent the Yellow Sea followed by the Bohai Sea, the East China Sea, and the Northern South China Sea (Tan et al 2012).

Precipitation can remove various components from the atmosphere, so the precipitation N concentrations have a close relationship with the atmospheric concentrations of NH₃ and NH₄⁺ (Xiao et al 2012). As reported by Liu et al (2013), inorganic N concentrations in precipitation show remarkable differences in different regions of China. The average concentration was 1.20 mg N L⁻¹ for NH₄⁺–N, and 0.97 mg N L⁻¹ for NO₃⁻–N at a South China site in Guangzhou city (Jia and Chen 2010). Chen and Mulder (2007) reported lower inorganic N concentrations in precipitation (NH₄⁺–N 0.29–2.33 mg N L⁻¹ and NO₃⁻–N 0.20–1.11 mg N L⁻¹) at four South China sites in Hunan, Chongqing and Guizhou. Our study reflects the consistency of rainwater N concentrations with rainfall distribution in different regions. The North China sites had higher N concentrations in precipitation than the South China sites due to the lower rainfall and higher atmospheric reactive N concentrations (Zhang et al 2008). This study confirms the higher inorganic N concentrations in rainwater at coastal sites of North China than South China. The lower inorganic N concentration in rainwater at FZ in our study compares well with the inorganic N concentration in rainwater (1.74 mg N L⁻¹) in Xiamen (Chen et al 2011).

Nitrogen wet and dry deposition in coastal areas in our study was overall lower than that in intensive agricultural areas of China. For example, on the North China Plain, N wet and dry deposition ranged from 60–102 kg N ha⁻¹ yr⁻¹ due to high air pollution and the inter-annual variability of rainfall (Pan et al 2012, Luo et al 2013). Nitrogen wet deposition in the Taihu Lake region and Shanghai area in East China was estimated to be about 30 kg N ha⁻¹ yr⁻¹ (Xie et al 2008) and 58 kg N ha⁻¹ yr⁻¹ (Mei and Zhang 2007), respectively. Jia and Chen (2010) reported an N wet deposition of 52.8 kg N ha⁻¹ yr⁻¹ at a South urban sampling site of Guangzhou. However, N wet deposition on the lower Liaohe River Plain in Northeast China was only 14.5 kg N ha⁻¹ yr⁻¹ (Yu et al 2011) and close to our results.

For the coastal sea area, streams and rivers are additional nutrient sources to estuarine and coastal marine environments. The inputs of N and P increased rapidly by erosion from arable lands or carrying human wastewater, so the coastal sea is more affected by human activity than only by air pollution (Smith et al 1999). Atmospheric N deposition represents a long-term low level fertilization effect for the ocean and has consequences for the natural biogeochemical cycles of N and increases the productivity of the oceans (Duce et al 2008). Compared with the riverine N input into the coastal ocean, N deposition may have long range impacts on the open ocean downwind of the primary source regions, especially in Eastern North America, Europe, and South and East Asia (Doney et al 2007). China is one of the hotspots of global reactive N emissions and deposition (Galloway et al 2008, Liu et al 2013). Several studies have quantified potential reactive N long range transport. For example, a modeling study showed that the high emissions of reactive N in China made a significant contribution to Japan's HNO₃ deposition through long-range transport (Holloway et al 2002). The Yellow Sea regions showed high N deposition through long-range transport from East Asia (Kim et al 2010). Total N deposition averaged 33.9 kg N ha⁻¹ yr⁻¹ for all the sampling sites in this study. If we consider this to be representative of N deposition for the sea area, it makes a significant contribution in addition to the river or the groundwater N entering the oceans. Assuming an annual mean N deposition of 33.9 kg N ha⁻¹, the total N deposition into the whole Chinese coastal seas (total area approx. 0.4 million km², calculating from 18 000 km length of boundary line of territorial waters with 22 km width of territorial waters) can amount to 1.36 Tg N yr⁻¹. It is a large N nutrient input from atmospheric deposition, which is comparable to inorganic N output of 1.61 Tg N yr⁻¹ from the Yangtze River to the East China Sea (Yan et al 2010) and estimated feed N input (∼1.5 Tg N yr⁻¹) from coastal aquaculture (www.zg3n.com.cn/2012/0904/15059.html). Although no systematic studies show the total N input from riverine source, aquaculture and deposition (three major N sources in coastal zones), the contribution of N deposition to Chinese coastal N input (amounting to about 30% of total N input) should not be neglected. Such high levels of N deposition in particular NH₄⁺/NH₃ deposition may lead to secondary N₂O emissions either from soils or marine ecosystems (Matson et al 2002, Duce et al 2008), due to nitrification and denitrification. Our monitoring results are 3–8 times higher than the modeled N deposition rates (2.5–11 kg N ha⁻¹ yr⁻¹) in the East China Sea (Zhang et al 2010). One possible explanation may be that we focus on the coastal sea and atmospheric reactive N concentrations and deposition show a decreasing gradient downward from the shore and are therefore significantly lower for the whole sea area. Another possibility is that the modeling results underestimate N wet and dry deposition in the coastal areas.

In addition, the coastal ocean environment is also affected by P and iron inputs and other complex chemical process (Deutsch et al 2007), so it is quite complicated to quantify the environmental impact of N deposition to the coastal ocean and further research is also needed to quantify all the nutrient fluxes from mainland China to the ocean. Nitrogen dry deposition velocities were taken directly from the literature (e.g., Zhang et al 2004, 2010) and are therefore highly uncertain, leading to uncertainties for the dry deposition rates in this study, and this needs more accurate estimation in future studies.