Water Point and Non-Point Nitrogen Pollution Due to Land-Use Change and Nitrate Deposition in China from 2000 to 2020

Abstract

Water N-NO₃⁻ (mg L⁻¹) pollution is attracting global concern in the face of combating climate change and human health risks. However, there have been comparatively few comprehensively researched studies on water N-NO₃⁻ pollution with respect to N-NO₃⁻ deposition, soil nitrogen, and land-use changes. We collected a total of 7707 published sampling points on N-NO₃⁻ surface and groundwater during flooding and non-flooding seasons during 2000–2020 in China. The types of water N-NO₃⁻ pollution (>20) can be categorized as point pollution (ΔTN ≤ 0 or > 1.5) and non-point pollution (0 < ΔTN ≤ 1.5), which were then assessed with respect to soil nitrogen (ΔTN g kg⁻¹) and water N-NO₃⁻ changes in this study. We found non-point pollution was concentrated in the Huaihe River Basin and Haihe River Basin with higher urbanization (+6%, +4%), cropland (72%, 45%), nitrogen fertilization (g m⁻² yr⁻¹) (>10), and increased wet N-NO₃⁻ deposition (WND) (kg ha⁻¹ yr⁻¹) (+4.6, +3). The Haihe River Basin was found to have the highest N-NO₃⁻ on its surface (306) and in its groundwater (868) and nitrogen fertilization (32). Point pollution was concentrated in the Songhua and Liaohe River Basin with the highest WND (+7.9) but slow urbanization (+1%). N-NO₃⁻ increased during the flooding season compared with the no-flooding season in serious pollution areas. N-NO₃⁻ increased in the Liaohe River and middle and low Yangtze River but was reduced in the Weihe River. Therefore, stringent criteria and management, especially during the flooding season are urgently required to mitigate the degree of N-NO₃⁻ water pollution that occurs due to intensive agriculture and urbanization with increased N-NO₃⁻ deposition.

1. Introduction

Land management, especially fertilization in agroecosystems, is a main controller of nitrogen input into a water system where nitrate (N-NO₃⁻) is the main pollutant [1,2]. The overuse of chemical nitrogen (N) fertilizers in agricultural land results in an increase in nitrate in soil drainage and subsequently in groundwater (GW) [3] and surface water (SW) [4], thereby causing the well-known eutrophication problem [5,6].

The cropland area in China, which is approximately 8.6% of the global cropland area [7], has increased from 130 × 10⁶ to 136 × 10⁶ ha between 2000 and 2020. The Chinese croplands consume approximately 22.8% of global chemical N fertilizers, which is three times the world average level [8]. This value can be compared with the application of N fertilizers in India and the United States, which possess 10.8% and 10.3% of the global cropland area, respectively, but only consume 18% and 10% of the global inorganic N fertilizers [7]. Excluding its croplands, the Chinese urban land area is more than 21 × 10⁶ ha, which is larger than the area of the entire United Kingdom, with a per capita settlement of more than 35% the area of Japan [9]. The urban areas and municipal waste can also provide a significant source of nitrate to the water system. For example, the growth rate of nitrate concentration in the Mississippi River was between 5% and 13% due to urbanization between 2000 and 2008 [10]. Approximately 64% of the nitrate in the water systems of Slovenia came from domestic wastewater and the agricultural sector [11]. As an example, in China, the N content of the urban rivers in Hangzhou were three to five times greater than what was found in the rivers which were not influenced by cities [12]. The extremely rapid urban land area expansions in China, the U.S., and India account for 43% of the world’s total urban sprawl [13]. Therefore, such rapid urbanization in China, which is expected to continue in the future, raises further concerns regarding nitrate pollution.

Nitrogen deposition is affected by land use and results in pollution to water systems [14]. The average annual atmospheric nitrate deposition has increased by 60% from 1980 (13 kg ha⁻²) to 2010 (21 kg ha⁻²) [15] in China. The relatively high average of N-NO³⁻ deposition (7 kg ha⁻²) in China [16] is especially acute in the economic and urbanized regions [17,18], e.g., the Pearl River Delta [19]. Croplands are one of the primary sources of nitrogen deposition [20]. Nitrogen compounds are easily volatilized in gaseous form and are subsequently deposited into soils and water bodies with precipitation following the application of fertilizers and pesticides on farmlands [21], thereby becoming significant contributors to nitrogen deposition. The nitrogen deposition caused by nitrogen overuse accounted for 10% to 20% of the total nitrogen deposition in hotspot regions, which includes China, India, and the United States [22]. It was found that 11.9% of the N surplus in the surface soil of Indian farmlands was because of atmospheric N-NO₃⁻ deposition [23]. The conversion of agricultural land into urban residential or industrial areas in peri-urban regions alters land-use patterns, thereby affecting both the mode and flux of nitrogen deposition. In the Sichuan region of China, the nitrogen deposition flux in underdeveloped areas was only 40% of that which occurred in developed cities [24]. Nitrogen deposition was also found to be a significant factor contributing to water pollution and lake sediment [25,26]. There were 5–8 kg of N ha⁻¹ y⁻¹ transported to the lake sediments in the Colorado Rocky Mountains region of the United States [27], and more than 30% of the nitrate nitrogen pollution in the water bodies of Jincheng came from atmospheric nitrogen deposition [14]. Nitrogen compounds, upon deposition into water bodies, induce eutrophication and threaten the safety of drinking water, thus posing risks to human health [28].

Precipitation has a significant impact on N pollution and concentration in water systems [27], and it also changes total N loading [29]. For example, approximately 10% to 35% of the total nitrogen output from the agricultural basins in the northeastern United States was due to precipitation [30]. The nitrate concentrations in the surface water in Beijing, China, in 2009 and 2010 showed a significant negative correlation (p < 0.05) to the amount of precipitation [31]. Flooding events, in particular, may increase nitrate transport. For example, during the flooding season of 2011, the N-NO₃⁻ content in the urban rivers of Chongqing, China, showed a 14–24% increase [32], and the average nitrate concentration in groundwater of the North China Plain increased by three times during the flooding season compared with the non-flooding season in 2012 [33].

N-NO₃⁻ pollution is a threat to all human health; however, the threat is even more prominent in vulnerable groups such as children and pregnant women. High concentrations of N-NO₃⁻ in drinking water lead to birth defects and induce cancer in infants [34], as well as increase the chances of high hemoglobin cancers, diabetes, thyroid diseases, and spontaneous abortions [35]. In 2017, the World Health Organization (WHO) issued a new specification to adjust an appropriate NO₃⁻ concentration of 10 mg L⁻¹ in drinking water [36]. In China, a nitrate concentration of less than 10 mg L⁻¹ in drinking water is considered appropriate, which can be relaxed to 20 mg L⁻¹ in areas with difficult access to drinking water [33]. However, the regional N-NO₃⁻ concentrations in China are unfortunately higher than the optimum levels. For example, the average N-NO₃⁻ concentration in the groundwater of the middle of the Guanzhong Plain of Shaanxi Province was 29 mg L⁻¹ [37]. The higher N-NO₃⁻ concentration in the groundwaters in Hailun District, Heilongjiang Province, was 123 mg L⁻¹ [38,39], and the average N-NO₃⁻ concentration in the groundwater in Shandong Province was 26 mg L⁻¹ [15].

Although there are numerous, separate studies on the concentration of nitrate in groundwater and surface water, as well as on their involved mechanisms, few studies have tried to provide a comprehensive dataset on the spatial and temporal changes in nitrate concentrations in soil and water systems, particularly with respect to land-use change, N deposition, and regional water regimes. Through using the published data on the N-NO₃⁻ concentrations in water and the changes in soil N content between 2000 and 2020, we have tried to (1) reveal the point and non-point sources of nitrate pollution in groundwater and surface waters; (2) analyze the effects of land use and nitrate nitrogen deposition on the N-NO₃⁻ concentrations in GW and SW; and (3) provide feasible management actions for the purpose of avoiding higher nitrogen pollution in water systems.

2. Materials and Methods

2.1. Study Site

The study site is the whole main Chinese water system. Topographically, China’s elevation gradually decreases from west to east (Figure 1a). Based on precipitation trends, May through to October is considered the flooding season, and November through April is the non-flooding season [40]. In this study, China was divided into the Huaihe River Basin (A), Haihe River Basin (B), Songhua and Liaohe River Basins (C), Yellow River Basin (D), Yangtze River Basin (E), Pearl River Basin (F), Southeast Basin (G), Southwest Basin (H), and Continental Basin (I) according to the nine major water systems (Figure 1b) [41].

2.2. Data

The total soil nitrogen content in 1990–2000 and the basic attribute dataset of a high-resolution national soil information grid with a 1 km spatial resolution (2010–2018) were collected from the National Data Centre for Earth System Science (http://www.geodata.cn/ accessed on 18 September 2022). Information on the nitrogen content (g N kg⁻¹ soil), the soil depth (cm), and the bulk density (g cm⁻³) were collected. The annual changes in the China Land Cover Dataset (CLCD) were produced from Landsat images on the Google Earth Engine at a 30 m spatial resolution from 2000 to 2020 [41].

The point data of the groundwaters and surface waters were retrieved from papers published in the Web of Science and the China National Knowledge Infrastructure (CNKI) (2000–2020) using different keywords, i.e., combinations of “rivers”, “groundwater”, “surface water”, “China”, and “nitrogen”. The N-NO₃⁻ (mg L⁻¹) data were retrieved from 3034 published papers. The collected data for this research were based on the following criteria: (a) a clear determination of concentrations, (b) a clear sampling period, and (c) clear latitude and longitude coordinates. In total, 7707 measured N-NO₃⁻ data points were available from 2000 to 2020.

The population data in China from 2000 to 2020 were from the LandScan Program (https://landscan.ornl.gov/ accessed on 1 September 2023). The N fertilizer use data in 2000–2018 were obtained from the Nation Ecosystem Science Data Center (http://www.geodata.cn/ accessed on 23 July 2023) at a 5 km spatial resolution.

2.3. Data Analysis and Methodology

2.3.1. Calculating Changes in the Soil Nitrogen Content

The variation in the soil N content in the top 0–30 cm layers (ZN) was calculated by the Equations (1) and (2):

$$ZN=\frac{\sum _{i}i\times {Content}_i}{\sum _{i}i}$$

(1)

where i is the soil depth, i.e., 5 (0–5 cm), 10 (5–15 cm), and 15 (15–30 cm); and Content is the N content (g kg⁻¹) at depth I;

$$\mathsf{\Delta }TN={ZN}_{2018}-{ZN}_{1990}$$

(2)

where ΔTN is the change in the soil total N (g kg⁻¹).

2.3.2. Tracing the Point and No-Point Pollution Sources

The increased soil TN in Chinese agricultural fields after years of fertilization was found to be less than 1.5 g N kg⁻¹ [42]. Therefore, when the increased soil TN was higher than 1.5 or less than 0, but the water N-NO₃⁻ was higher than 20 (>20 mg L⁻¹), the sampling points were categorized as point pollution. Accordingly, the water data were categorized as point pollutions when the ΔTN ≤ 0 or ΔTN > 1.5 and as non-point pollutions if 0 < ΔTN ≤ 1.5 [42].

2.3.3. Calculation of the Average Population Density and Nitrogen Fertilizer Application over Multiple Years

$${Per}_i=\frac{\sum _{n}i}{n}$$

(3)

where ${Per}_i$ is the average annual population or fertilization, and n is the number of years.

3. Results

3.1. Spatial and Temporal Distribution of Nitrate

3.1.1. Spatial and Temporal Distribution and Variation of Nitrate during Flooding Seasons

During the flooding seasons of 2006 to 2020, the observed N-NO₃⁻ pollution of >20 mg L⁻¹ was mainly distributed in the Huaihe River Basin (A), Haihe River Basin (B), Songhua and Liaohe River Basin (C), Yellow River Basin (D), Yangtze River Basin (E), and Pearl River Basin (F). Furthermore, 13% of the GW sampled sites in Basin C were detected as point pollutions from 2000 to 2020 (Figure 2). Basin C faced severe point pollutions in 2000–2020, where 10% of the N-NO₃⁻ point pollution sites of GW and SW especially in the Songnen Plain with 16% of the GW point pollutions (Figure 2) had the highest N-NO₃⁻ concentration in GW (624 mg L⁻¹) during the flooding seasons (Figure 3). Basin B had the most severe N-NO₃⁻ pollution, where 64% of the GW sampled sites had N-NO₃⁻ concentration > 20 mg L⁻¹ and approximately 52% of the sites had severe N-NO₃⁻ concentration > 30 mg L⁻¹ (Figure 3). The Basin B area, in particular, experienced the most significant N pollution in 2000–2005, and this was even when 83% of the GW sampled sites had N-NO₃⁻ concentration > 20 mg L⁻¹. The N-NO₃⁻ concentration in GW was found to be the highest (868 mg L⁻¹) in the Ziya River, which is in the center of the Haihe River Basin (B) in 2011–2015 (Figure 2). There were severe GW point pollutions (25 sampled sites) in the southwestern areas of E and in the upper river water system in the northwestern part of F, such as Dianchi Lake, Sanchahe River, and Nanpanjiang River. Even the remote areas in the northwest of China, 6% of the sampling sites, which were near rivers in the western (Hotan River) and central (Shulehe River) regions of Basin I, had point pollutions in 2011–2015 (Figure 2).

During the flooding seasons, the N-NO₃⁻ concentration in surface water (SW) saw the highest concentration at 175 mg L⁻¹ in Heifei city in the Basin E region in 2016–2020 (Figure 3). Basin E experienced increased N-NO₃⁻ pollution from 2000 to 2020, where the percentage of point pollution sites increased from 24% (2006–2010) to 32% (2016–2020). Since 2011, in the southwest of the Basin E region, point pollutions was observed in the Sanchahe River, the upstream of the Nanpanjiang River, and the downstream of Basin E. Approximately 11% of the SW sampled sites from 2011 to 2015 experienced point pollutions in Basin B, with the highest-valued data in the central parts, i.e., the Ziya River (109), and the northeastern parts, including Tianjin city and its surroundings (67). The surface water point pollutions increased from 7% (2011–2015) to 23% (2016–2020) in Basin C areas concentrated in the Liaohe River. The SW point pollution was 31% from 2016 to 2020 in the Li River of Basin F, as well as in nine of the sites from 2011 to 2015 in the Basin B region (the southwest of China). There were no SW sampled sites with severe pollution, i.e., the N-NO₃⁻ concentration > 30 mg L⁻¹, in the Basin G region between 2000 and 2020 (Figure 2).

3.1.2. Spatial and Temporal Distribution and Variation of the Nitrate during the Non-Flooding Seasons

During the non-flooding seasons from 2011 to 2015, the N-NO₃⁻ concentration in the GW was at its highest in the Ziya River at 619 mg L⁻¹, which is at the center of the Haihe River Basin (B) (Figure 4 and Figure 5). Over 38% of the GW sampled sites in Basin B saw N-NO₃⁻ pollutions concentrated in the Ziya River (2000–2011), Luan River (2006–2015), and Yongding River (2016–2020). The GW sampled sites for N-NO₃⁻ pollution were the most severe in the Huaihe River Basin (A), where 68% of the sampled sites had N-NO₃⁻ concentration > 20 mg L⁻¹, and 52% saw N-NO₃⁻ concentration > 30 mg L⁻¹ (Figure 5). The majority of the GW N-NO₃⁻ pollution was concentrated in Shandong Peninsula, which is located in the northeast of Basin A, during 2000–2005 and 2011–2020; however, in 2006–2010, the majority of the GW N-NO₃⁻ pollution was concentrated in the Northern China Plain, which is west of Basin A. The point pollutions were approximately 17% in the sampled sites in Basin C, and these sited were concentrated in the northern (Songhua River) parts in 2006–2010 and the central (Songnen Plain) parts in 2016–2020. The N-NO₃⁻ point pollutions in Basin E (mostly in the eastern and the southwestern parts) and in Basin F were approximately 50% and 41%, respectively, from 2000 to 2020 (Figure 4).

During the non-flooding seasons, the N-NO₃⁻ concentration in the SW had the highest value near Beijing city in Basin B from 2011 to 2015 (306 mg L⁻¹) (Figure 4 and Figure 5). The SW N-NO₃⁻ pollutions increased from 4% (2006–2010) to 40% (2016–2020) in Basin B. The SW N-NO₃⁻ pollution was 14% near the Liaohe River of Basin C from 2011 to 2020, where 12% of the sample sites were found to have been affected by point pollutions from 2016 to 2020. In Basin D, 37% of the SW sampled sites showed N-NO₃⁻ pollution, of which 27% represented severe pollutions (N-NO₃⁻ > 30 mg L⁻¹). However, the average N-NO₃⁻ concentration in the SW was improved in the Weihe River of Basin D, which reduced, on average, from 23 mg L⁻¹ (2011–2015) to 17 mg L⁻¹ (2016–2020) (Figure 4 and Figure 5); however, the average N-NO₃⁻ concentration in the GW did not change much (i.e., a decrease from 35 to 34 mg L⁻¹ between 2011–2015 and 2016–2020). Conversely, the SW N-NO₃⁻ pollutions were found to be deteriorating in the middle and lower regions of Basin E; there, the N-NO₃⁻ pollutions increased from 5% in 2000–2005 to 21% in 2016–2020, where 12% of the sampling sites had severe N-NO₃⁻ pollution (N-NO₃⁻ > 30 mg L⁻¹). In Basin E, the SW point pollutions were over 30% in the sampled sites from 2000 to 2020, and this was mainly observed near the cities in the main stem of the Yangtze River. The SW pollution was concentrated in the Li River of Basin F from 2016 to 2020, where the point pollution was 26%.

3.2. The Land-Use Change and Nitrate Nitrogen Deposition on Nitrogen Pollution in Chinese Water Systems

3.2.1. The Land-Use Change and Urbanization from 2000 to 2020

The percentage of cropland decreased by 5.8%, 4.7%, and 1.5% in Basins A, B, and E, respectively (Figure 6), but the percentage of urban land increased by 5.5%, 4.5%, and 1.3% from 2000 to 2020. The urban land increased by 1.2%, 1.3%, and 2.8% in Basins C, F, and G from 2000 to 2020. Basin A had the highest percentage of cropland area (77.7%) among the nine basins in 2000, where the cropland decreased the most by 5.8%; however, the urban land increased the most by 5.5% from 2000 to 2020 (Figure 6). Basin D saw a decreased percentage of cropland area (–2.4%) with an increase of urban land (+1.3%). Basin H showed the smallest percentage of cropland (4%) in 2011, which increased by 0.7% from 2011 to 2020 (Figure 7). The percentage of urban land increased by more than three times in Basins H (0.02–0.06%) and I (0.06–0.23%) from 2000 to 2020.

The concentrations of nitrate nitrogen (N-NO₃⁻) in the groundwater (GW) in Basins A and B during the flooding season were 25 mg L⁻¹ and 35 mg L⁻¹ higher than that during the non-flooding season, respectively (Figure 8). In the surface water (SW), the concentration of N-NO₃⁻ was 14 and 4 mg L⁻¹ higher during the flooding season compared with the non-flooding season (Figure 8). The concentration of N-NO₃⁻ in the GW in Basin C increased the most during the flooding season, i.e., it was 65 mg L⁻¹ higher than in the non-flooding season (Figure 8). In other basins, the N-NO₃⁻ concentration in the GW during the flooding season was lower than that during the non-flooding season from 2000 to 2020. In this research, the N-NO₃⁻ concentrations in the SW during the flooding season was higher than that during the non-flooding season in Basins A, B, C, and H, with increases of 14.3, 4.5, 3.4, and 10.3 mg L⁻¹, respectively (Figure 8). The research revealed a correlation (p < 0.05) between the concentration of N-NO₃⁻ in the GW and the proportion of cropland and urban land (R = 0.8). When linear fitting was performed on the percentage of SW to cropland and urban land (excluding Basins C and E, which demonstrated evident point pollution), it was found that the concentration of N-NO₃⁻ in the SW in the remaining basins was correlated with the ratio of cropland and urban land (p < 0.05), wherein they yield a correlation coefficient of R = 0.75 (Figure 9).

3.2.2. Population Density and Agricultural Nitrogen Application

The highest average population aggregation (436 persons km⁻²) and the highest average N fertilizer use (175 kg ha⁻¹ yr⁻¹) was found in the Huaihe River Basin (A) (

Table 1. The average population density between 2000 and 2020, as well as the nitrogen fertilization rates between 2000 and 2018 (the data in Table 1), were sourced from the LandScan Population Density Dataset (https://landscan.ornl.gov/ accessed on 1 September 2023) (a). Historical nitrogen fertilizer use in China from 1952 to 2018 (http://www.nesdc.org.cn/ accessed on 23 July 2023) (b).

Basins	Population Density (person km−2)	T Nitrogen Fertilization Rate (kg ha−1 yr−1)
A	436	175.57
B	288	85.10
C	59	25.85
D	106	34.46
E	185	43.54
F	240	44.09
G	276	45.33
H	19	6.08
I	6	2.93

). However, China’s more densely populated areas are concentrated in the middle and lower reaches of the Haihe River Basin, including the Ziya River (>288 persons km⁻²), the downstream areas of the Yellow River Basin, its central areas such as the Weihe River (>106 persons km⁻²), the middle and lower reaches of the Yangtze River Basin, the upstream Sichuan and Chongqing areas (>185 persons km⁻²), the Pearl River Delta region (>240 persons km⁻²), and the Southeast Basin (>276 persons km⁻²) (Figure 10a,b). However, the nitrogen fertilization rate in the Yangtze River Basin (43.5 kg ha⁻¹ yr⁻¹), the Pearl River Basin (44.1 kg ha⁻¹ yr⁻¹), and the Southeast Basin (45.3 kg ha⁻¹ yr⁻¹), which are not predominantly agricultural areas, were on average low (Table 1).

3.2.3. N-NO₃⁻ Dry and Wet Deposition from 2006 to 2015

The main N-NO₃⁻ deposition was wet deposition (0.1–17.8 kg ha⁻¹) from 2006 to 2015, and it was more than three time higher than the dry deposition (0.01–5.8 kg ha⁻¹) (Figure 11). The change in the N-NO₃⁻ wet deposition was much more pronounced during the period of 2006–2010 (0.3–14.6 kg ha⁻¹) to 2011–2015 (0.1–17.8 kg ha⁻¹), when the more serious deposition sites appeared in Basins A, B, C, E, and F. The highest deposition (kg ha⁻¹ yr⁻¹) and the highest increasing trend in N-NO₃⁻ wet deposition (+kg ha⁻¹ yr⁻¹) were concentrated in Basin A (13.4, +4.6), B (16.5, +3.0), the middle and lower reaches of the Yangtze River of Basin E (17.8, +4.9), the south of Basin C (16.4, +7.9), and the Pearl River Delta of Basin F (17.4, +2.7) from 2006–2010 to 2011–2015. These areas also had the highest N-NO₃⁻ dry deposition (Figure 11). The N-NO₃⁻ dry deposition stayed relatively constant from 2006–2010 (0.01–5.8 kg ha⁻¹) to 2011–2015 (0.01–5.5 kg ha⁻¹). The N-NO₃⁻ dry deposition in the Pearl River Delta (F) appeared to decrease (–0.9 kg ha⁻¹ yr⁻¹) during 2006–2010 to 2011–2015.

4. Discussion

4.1. The Nitrate Pollution Sources in the Nine Water Basins

The studied N-NO₃⁻ pollution in the water systems included the main non-point pollutions (e.g., A and B) and point pollutions (e.g., C and the southwest of China) (Figure 2 and Figure 4). Basin B had the most severe N-NO₃⁻ pollution in the GW and SW among the nine basins from 2000 to 2020 and also the highest proportion of pollution (40% N-NO₃⁻ pollution), where the highest N-NO₃⁻ concentration in the GW was 868 mg L⁻¹ in the flooding seasons and 306 mg L⁻¹ in the SW in the non-flooding seasons (Figure 2 and Figure 4). Basin B had had the highest nitrate pollution among the other basins mainly because of its urban land cover [45], which increased from 9% to 13% in 2000–2020 and was accompanied by the highest increase in N-NO₃⁻ wet deposition (+3 kg ha⁻¹ yr⁻¹) (Figure 9) and a high population density (288 persons km⁻²) (Table 1). In particularly, the Ziya River in the central part of Basin B featured extensive cropland cover (Figure 6), which is also where the highest nitrogen fertilizer usage took place (320 kg km⁻² yr⁻¹) (Figure 10b). Basin A had the most severe non-point nitrogen pollution because more than 80% of the region was utilized for urban areas and croplands in 2000. The nitrogen pollution in the water systems made it extremely easy for lake sediment deposits. Thus, the lake sediment pollution index of TN (STN), which is affected by the amount of urban land, was 60% higher than that in the unaffected regions [46]. However, in the middle and lower reaches of the Songhua River, the STN values ranged from 0.2 to 10.1 (Figure 12) with a high N-NO₃⁻ concentration in the GW; this region also contained a lower degree of population (59 persons km⁻²) and a lower average N fertilization rate (26 kg km⁻² yr⁻¹) (Figure 10a,b; Table 1). These conditions suggest that the industry and urban areas in the region greatly contributed to the point pollution in Basin C, especially during the time the government relocated the polluting industries to the southern parts of Basin C [47]. Therefore, the percentage of SW N-NO₃⁻ pollution (>20 mg L⁻¹) in the Liaohe River of Basin C exhibited a severe increasing trend of 9% in 2011–2015 to 24% in 2016–2020 (Figure 3). The urban land in the Yangtze River Basin had a huge impact on the pollution of the water system; the concentration of urban areas and dense population (Figure 6 and Figure 10) were the main reasons for the continuous rise of surface water from 2000 to 2020. Specifically, the lake sediment pollution index of TN (STN) in the downstream regions of Basin E was higher than Basin C, which had STN values ranging from 0.5 to 12.2 (Figure 12). The N-NO₃⁻ pollution of the GW and SW became worse in Basins C and E from 2000 to 2020; however, they were mitigated in the Weihe River of Basin D (Figure 2 and Figure 4). The improved N-NO₃⁻ pollution in the SW in the Weihe River of Basin D during the non-flooding season was in contrast to the polluted GW (Figure 4), which implied a potentially higher point pollution. The mitigation of point pollution in the Weihe River was attributed to the governmental policies that were implemented along the Weihe River from 2008, including the shutting down of highly polluting factories and the implementation of environmental protection projects (e.g., beach remediation, river dredging, ecological restoration, and wetland construction) [48].

4.2. The Risk of Nitrate Pollution in the Groundwater and Surface Water during the Flooding and Non-Flooding Seasons

Precipitation greatly increased the water N-NO₃⁻ concentration because of the dilution that occurs during the flooding season [27]. In Basins A and B, which have high urban land and cropland areas, increased precipitation increased the N-NO₃⁻ concentrations in the groundwater (+25 mg L⁻¹; +35 mg L⁻¹) and surface water (+14 mg L⁻¹; +4 mg L⁻¹). The highest GW N-NO₃⁻ concentrations increased by 65 mg L⁻¹ in the point pollution of Basin C during the flooding season in comparison with the non-flooding season. Therefore, in the heavily polluted areas, the increased precipitation made the GW N-NO₃⁻ pollution worse. This was also demonstrated in other regions because of the intensified land use of urban land and croplands. For example, the GW N-NO₃⁻ concentration in the UK increased by 30% from 1976 to 2006, which peaked after precipitation [50]; in addition, GW N-NO₃⁻ concentrations also increased in South Korea by up to 22% after precipitation [51]. In these examples, the large areas of urban land, high levels of fertilizer application, and severe urban and industrial pollution were understood as major factors [52]. In addition to these major factors, however, the development of tourism, particularly in Erhai Lake in Basin H, drove land-use changes, thereby resulting in a decreased soil conservation rate and an increased SW nitrogen rate [53]. Therefore, the SW N-NO₃⁻ concentration was 8 mg L⁻¹ higher during the flooding season than in the non-flooding season in Basin H (Figure 2 and Figure 4), and these particular regions were predominantly concentrated in Erhai Lake [54].

The groundwater nitrogen pollution in China was found to be positively correlated with the proportion of land use (specifically cropland and urban land) (r = 0.8). The pollution in surface water was also found to be correlated in areas where non-point pollution was not the main source, except in Basin C and E (r = 0.75) (Figure 9). Non-point pollution has a lag effect [55], and the N retention capacity of groundwater is prominent [56]. Therefore, compared with surface water with its stronger mobility, the groundwater was more severely affected by anthropogenic nitrogen input.

4.3. Water Quality and Pollution Control Strategies

The N-NO₃⁻ pollution in China water system originates from human activities mainly because of distribution of cropland and urban land (Figure 9) [57]. A (88%) and B (58%) with highest cropland and urban land areas have the worse N-NO₃⁻ concentrations in GW than in SW, where the percentage of GW N-NO₃⁻ pollution is higher than SW. Excessive N fertilization increases soil N content [58] due to stimulated nitrification [59], which consequently exacerbates the concentration of N-NO₃⁻ in surrounding water systems, and it is not only China but also the United States and India that are facing the environmental problems caused by massive fertilizer application [7,60]. Therefore, it is necessary to reduce the amount of N fertilization [61] by adopting reasonable field management practices, such as increasing the use of organic fertilizers, e.g., manure and returning straw [62], as well as incorporating biochar [63] and implementing reduced and precision fertilization [64], especially in Basins A, B, and C.

In densely populated urban land areas, such as fast urbanization areas, N-NO₃⁻ pollution originates from the discharges of industrial pollutants and home sewage [65]. This occurred in Basins A and B, the middle and lower reaches of Basins C and E, and the Pear River Delta of Basin F, as well as in southwestern China (Figure 2 and Figure 4). Therefore, it is imperative to manage industrial wastewater and improve standards in the flooding and non-flooding seasons, which should include optimizing urban layouts such as the relocation of industrial areas away from the city, as well as the establishment of wooden shelterbelts [66] and NOx emission indicators [67]. The results of such proper management strategies can be seen, for example, in Basin G. Despite the fast urbanization in Basin G (which has seen the area of urban land during 2000 to 2020 doubled from 2.5% to 5.2% and where the population density (288 persons km⁻²) is among the highest in China), the GW and SW N-NO₃⁻ concentration rates have remained constant and have shown a less significant degree of N-NO₃⁻ pollution. This appears to be mainly because of the relatively higher percentage of forests and grasslands in Basin G (75%), which also provide a good example for the control of N-NO₃⁻ pollution. These findings are consistent with the management plans utilized in the United States, where net nitrogen input is more easily retained, such as in the case of the Mississippi River basin, which has reached an 80% [60] net nitrogen iuput. From 2000 to 2020, the nitrogen fertilizer application rate in India increased from 13% worldwide to 18% [7], and this type of management should be followed even more in regions where nitrogen fertilizer usage is increasing.

5. Conclusions

The rapid agriculture and economic development in China has led to the accumulation of a substantial amount of nitrogen in groundwater (GW) and surface water (SW). Based on 7707 published studies on the N-NO₃⁻ concentration of surface and groundwater during the flooding and non-flooding season from 2000 to 2020 in China, we found that non-point pollution was concentrated in the Huaihe River Basin (A) and Haihe River Basin (B), and that point pollution was concentrated in the Songhua and Liaohe River Basin (C). Basin B saw the most severe N-NO₃⁻ pollution in more than 40% of the GW and SW studied sites. The N-NO₃⁻ concentrations in Basin B were even as high as 868 mg L⁻¹ in the GW during the flooding seasons and as high as 306 mg L⁻¹ in the SW during the non-flooding seasons. Industrial emissions have the most significant impact on N-NO₃⁻ wet deposition at +7.9 kg N ha⁻¹ yr⁻¹ in Basin C, which lead to the highest increase in the GW N-NO₃⁻ during the flooding season (+66 mg L⁻¹). In contrast, the high forest and grassland coverage (75%) in the Southeast Basin caused a small change in N-NO₃⁻ wet deposition (+1 kg ha⁻¹ yr⁻¹), even when subjected to intensive urbanization (+2.8%). Facing serious water N-NO₃⁻ pollution necessitates taking on stricter criteria for the management of sewage discharge, nitrogen fertilizer application, and control of the industrial emissions of ammonia and nitrous oxides.