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# Managing nitrogen to restore water quality in China

Chaoqing Yu, Xiao Huang, Han Chen, H. Charles J. Godfray, Jonathon S. Wright, Jim W. Hall, Peng Gong, Shaoqiang Ni, Shengchao Qiao, Guorui Huang, et al.

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Chaoqing Yu, Xiao Huang, Han Chen, H. Charles J. Godfray, Jonathon S. Wright, et al. Managing nitrogen to restore water quality in China. Nature, 2019, 567 (7749), pp.516-520. 10.1038/s41586-019-1001-1 . cea-02186561

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Distributed under a Creative Commons Attribution 4.0 International License
1 Managing nitrogen to restore water quality in China

2

3 Chaoqing Yu 1,2,* , Xiao Huang1,3, Han Chen1, H. Charles J. Godfray4, Jonathon S. Wright1, Jim

4 Hall3, Peng Gong1, Shaoqiang Ni1, Shengchao Qiao1, Guorui Huang1, Yuchen Xiao1, Jie Zhang1,

5 Zhao Feng1, Xiaotang Ju6, Philippe Ciais7, Nils Chr. Stenseth1,8, Dag O. Hessen7, Zhanli Sun9, Le

6 Yu1, Wenjia Cai1, Haohuan Fu1, Xiaomeng Huang1, Chi Zhang9, Hongbin Liu11, James Taylor12

7

8 ¹Ministry of Education Key Laboratory for Earth System Modeling, Department of Earth System

9 Science, Tsinghua University, China

10 ²AI for Earth Laboratory, Cross-Strait Tsinghua Research Institute, Beijing, China

11 ³Norwegian Institute of Bioeconomy Research, Saerheim, Norway

12 ⁴Oxford Martin School & Department of Zoology, University of Oxford, UK

13 ⁵Environmental Change Institute, Oxford University, UK

14 ⁶College of Resources and Environmental Sciences, China Agricultural University, China

15 ⁷Laboratoire des Sciences du Climat et de l'Environnement (LSCE), France

16 ⁸Centre for Ecological and Evolutionary Synthesis (CEES), University of Oslo, Norway

17 ⁹Leibniz Institute of Agricultural Development in Transition Economies (IAMO), Germany

18 ¹⁰School of Chemical Science and Engineering, Royal Institute of Technology, Sweden

19 ¹¹Institute of Agricultural Resources and Regional Planning, Chinese Academy of Agricultural

20 Sciences, China

21 ¹²School of Agriculture, Food & Rural Development, Newcastle University, UK

22 *Correspondence to: chaoqingyu@yahoo.com (Chaoqing Yu)

23

24
**Abstract (290 words)**

The nitrogen (N) cycle has been radically changed by human activities [1]. China consumes nearly 1/3 of the world's N fertilizers, and widespread excessive application of N fertilizers [2, 3] and increased discharge from livestock, domestic and industrial sources have resulted in pervasive pollution of water bodies. Yet challenges in monitoring and quantifying diffusion of N pollution from heterogeneous sources inhibit understanding of safe "boundaries" [4] of N use and effective N management for meeting local water-quality standards. We use a combination of water quality observations and simulated N discharge from agricultural and other sources to estimate spatial patterns of N discharge into water bodies across China for 1955-2014. We find that the critical surface-water quality standard (1.0 mgN·L⁻¹) was exceeded in most provinces by the mid-1980s, and that current rates of anthropogenic N discharge (14.5±3.1 MtN·yr⁻¹) to freshwaters are about 2.7 times the estimated ‘safe’ N discharge threshold (5.3±0.7 MtN·yr⁻¹). Current efforts to reduce pollution through wastewater treatment (WWT) and improving cropland N management can partially remedy this situation. Domestic WWT has already helped to reduce net discharge by 0.7±0.1 MtN in 2014, but at high monetary and energy costs. Improved cropland N management could cut another 2.3±0.3 MtN·y⁻¹, about 25% of the excess discharge to freshwaters. Successfully restoring a clean water environment in China will further require transformational changes to boost the national nutrient recycling rate from its current average of 36% to ~87%, a level typical of traditional Chinese agriculture. Though ambitious, such a high level of N recycling is technologically achievable at a capital cost of approximately $100 billion and operating costs of $17-25 billion·yr⁻¹, and would have co-benefits including recycled wastewater for crop irrigation, improved environmental quality and ecosystem services, and new economic opportunities for farmers and fertilizer industries.

Keywords: Sustainability, nutrient recycling, wastewater treatment, thresholds, water quality
**Main text** (1840 words)

The Earth's biogeochemical cycles have been strongly affected by human activity. For example, the amount of reactive nitrogen (N) entering the global environment increased from ~15 million tons (Mt) in 1860 to 185 Mt in 2010 [1, 5], while agricultural use of N fertilizers has expanded from 12 Mt in 1961 to 110 Mt in 2014 [6]. Though critical for crop yields and food production, human inputs of reactive N to terrestrial and freshwater ecosystems cause water pollution (e.g. NO₃-N, NH₄-N) and air pollution (NH₃, NOₓ), as well as global warming and stratospheric ozone depletion (N₂O) [7].

China has dramatically increased its food production over the past four decades. Domestic grain production has risen from 132 Mt in 1950 to 607 Mt in 2014 without expanding total planting area (Fig.S1). This has been achieved through massive increases in synthetic fertilizer use (Fig.1), along with increasingly efficient agricultural technologies, practices and expanded irrigation infrastructure. China now accounts for about 32% [6] of the world's consumption of N fertilizers. To illustrate the critical role of N inputs on crop production, we use the nitrogen-based DeNitrification and DeComposition (DNDC) biogeochemical model [8] and agricultural data from 2403 counties to simulate major grain crops (rice, wheat and maize) during the period 1955-2014 (Fig.S2, S3; Methods). Fig.1b shows that 45±3% of current grain yields can be attributed to use of synthetic N fertilizers. Similar conclusions have been reached following controlled long-term experiments in China [9] and other locations [10, 11] (Table S1).

Meanwhile, traditional practices of recycling organic wastes as fertilizers have been largely abandoned. The availability of subsidized synthetic N fertilizers and new sewage infrastructure have resulted in nutrient-recycling rates dropping sharply from >90% in the late 1970s to 36% in 2014 (Fig.1c, d). Changing N management practices have caused average nitrogen use efficiency (NUE) in China's croplands to decrease significantly from the 1960s to the 2010s [2, 3]. Increased N discharges from croplands and non-recycled organic wastes have resulted in pervasive water pollution.

Quantifying safe levels [4] of N discharge to the environment is a prerequisite for effective N management. Rockstrom et al. (2009) proposed 25% (35 MtN·yr⁻¹) of current human N fixation from the atmosphere as a global “safe boundary”, but qualified this estimate as “a first guess” [12]. Recent studies have attempted to clarify the global safe boundary for N released by human activities using simplified N-budget models [5, 13], but relationships among N discharge, N concentrations in freshwater bodies and biotic responses to N in water bodies are complex and often differ by region. Thus, large uncertainties remain regarding regional ‘safe’ N discharge thresholds, owing to insufficient understanding of spatiotemporal heterogeneity in biogeochemical and hydrological processes. The concept of a safe boundary is most useful when
applied to regional and local scales, where practical management options are available. Yet, local to regional safe N boundaries have not been established across representative regions. Here we use observations of water quality in representative rivers and lakes across China to characterize regional thresholds of total N discharge to the water environment, assess excess N, and evaluate the potential for reducing N pollution.

We collected observational data on concentrations of total N (TN) in water bodies and estimate provincial N discharge into the inland aquatic environment for 1955-2014. Here, N discharge comprises anthropogenic sources from croplands, human and livestock excrement, organic garbage, and industrial waste (Fig.2,3; Table S1; Fig.S4-6; Methods). N concentrations in water bodies were low before the 1980s (typically <1.0 mgN·L⁻¹), but increased rapidly to levels exceeding 15 mgN·L⁻¹ in many catchments after the 1990s. Groundwater concentrations of NO₃-N also increased sharply after the 1980s, with stable isotope analysis implicating agricultural and domestic sources [14]. The Ministry of Water Resources classified 80.2% of groundwater samples (from 2103 wells; Fig.S7) as polluted in 2015 [15].

We then identified empirically the critical N discharge threshold for each province as total anthropogenic N discharge during the year when pollution levels first breached the water-quality standard (1.0 mgN·L⁻¹ [16, 17], Methods; Fig.2,3; Table S3) in representative catchments. Aggregating N-discharge thresholds from catchment to national scale leads to a novel estimate of the national N discharge threshold (5.3±0.7 MtN·yr⁻¹; Fig.3b). This input can be conceptualised as China's share of the total planetary safe N boundary. The current national anthropogenic N discharge rate is 14.5±3.1 MtN·yr⁻¹ (2010-2014 average), well above the threshold. Agricultural systems are responsible for 59% of current N discharge (35% croplands, 24% livestock). Another 39% is attributed to domestic waste (13% urban sewage, 8% rural sewage, 18% organic garbage) and the remaining 2% to industrial waste.

Although China's declared target of zero growth in chemical fertilizer use by 2020 [18] has already been achieved, synthetic N use was still 30.5 Mt in 2016. Proposed remedies for maintaining agricultural productivity and reducing negative environmental impacts have centered on improved N management (INM) in croplands: applying the right fertilizer products, at the right rate, at the right time, in the right place [2]. To assess the potential benefits of INM in China's croplands, we use the DNDC model to estimate the minimum synthetic N input for maintaining historical yields (Methods) through broadcasting of fertilizers (BF) or the use of controlled-release fertilizers (CRF). Fig.3 shows that INM could reduce cropland N discharge from current (2010-2014 average) 5.1±0.3 MtN·y⁻¹ to 2.8-3.0 MtN·y⁻¹ (range encompasses the CRF and BF scenarios), and also fluxes of NH₃ (by 39~72%) and N₂O emissions (by 47~55%). INM could thus reduce current total cropland N inputs from 36.9±0.4 Mt to 24.1~26.9 MtN·yr⁻¹, or the use of synthetic N from 29.8±0.3 MtN·yr⁻¹ to 17.0~19.8 MtN·yr⁻¹. If all smallholder farmers adopted INM practices [19], this alone could reduce the excess anthropogenic N discharge (9.2±2.0 MtN·y⁻¹,
146 Methods) by 23-25%.

147

148 Quantifying safe boundaries at small geographical scales is helpful for clarifying
149 more detailed environmental issues and local management options. Fig.4a-c
150 summarizes the critical challenges for N management at the provincial level. Under
151 current N management (CNM), cropland N discharge by itself already exceeds critical
152 pollution thresholds in 14 of 31 provinces. INM-CRF could lower cropland pollution
153 to below the critical thresholds in 29 provinces, but two provinces cannot reach safe
154 thresholds without lowering food production (Inner Mongolia by 36%, Shaanxi by
155 12%).

156

157 Wastewater treatment (WWT) is another primary solution for reducing point-source
158 water pollution, not only by N, but also with by phosphorus and enteric bacteria in
159 inland water bodies. Domestic WWT has expanded considerably since the 1980s. In
160 2014, about 49.43 billion m³ (Bm³) of municipal wastewater (75% of urban water use)
161 was treated, with energy consumption of 14.8 billion kWh. The total economic cost
162 (infrastructure and operations) of WWT was about $20.8 billion ($0.42/m³ [20]),
163 equivalent to 2.2% of national rural household income (population 619 million) or
164 0.2% of GDP in 2014. But the amount of N removed by WWT was only 0.70±0.1 MtN
165 (~26% of total municipal sewage N), because only 56% of WWT plants had
166 N-removal facilities [21], with an average N removal rate of 55% [22]. Average efflux
167 concentrations in treated water released to the environment were 14.3-16.5 mgN·L⁻¹ [21,
168 23]. Further improvements in N-removal efficiency (e.g. tertiary WWT) are feasible,
169 and have been achieved in some wastewater treatment works, although at additional
170 capital [24] and operational costs [25].

171

172 China needs a holistic strategy to mobilise and integrate all relevant
173 socio-economic sectors to cut effectively N pollution, not just from croplands but also
174 from livestock, domestic and industrial wastes. As 63% of current N discharge to
175 freshwaters is from non-recycled livestock and domestic waste, future policies should
176 pursue a transformational expansion of nutrient-recycling systems, together with water,
177 sanitation and hygiene (WASH) programmes. Results from long-term (>20 years)
178 fertilisation experiments in China indicate that combining synthetic fertilizers with
179 manure can improve soil quality and generate 8.2-9.9% larger yields of rice, maize and
180 wheat compared to synthetic fertilizers alone [9]. Fig.4c, d demonstrate that reducing N
181 discharge by enough to return to provincial safe thresholds, though daunting, could
182 nearly be achieved by implementing INM and increasing the national
183 nutrient-recycling rate to 86-88%. Rates in nine provinces would need to exceed 95%
184 under INM-CRF. Even with a recycling rate of 100%, Shanghai would still need to cut
185 industrial N discharge by 37%, and Shanxi must either reduce food production by 3%
186 or cut industrial N discharge by 80%. Raising national nutrient-recycling rates to
187 86-88% could allow food productivity to be maintained at current levels while further
188 reducing synthetic N requirements under INM from 17-19.8 MtN·yr⁻¹ to 10.8~13.6
189 MtN·yr⁻¹ (assuming equivalent NUE for organic and synthetic N). Side effects of
increased recycling might include a rise in N₂O emissions [26].

191

We evaluate the costs of three strategies to achieve our proposed increase in nutrient recycling: 1) traditional wet manure recycling, 2) dry compost recycling, and 3) direct wastewater recycling, with consideration of the potential impacts of new toilet technologies and infrastructure changes (Methods). We estimate that the operational costs of recycling organic waste would range from $8-26 billion·yr⁻¹ based on 2010 prices (Table.S4). Traditional recycling is the most expensive option and carries WASH risks. A more practical approach is to deliver domestic wastewater for irrigation by separating industrial wastewater from domestic wastewater and connecting household sewage systems with existing irrigation systems. This approach would cost ~$100 billion for infrastructure, with operational costs of $17-25 billion·yr⁻¹.

202

Overall, the costs of building nutrient recycling systems are small relative to the annual cost of water pollution at current levels, estimated at 1.5% of national GDP in 2010 ($91 billion; $20 billion from treatment and $71 billion from impacts of environmental degradation) [27]. Farmers would benefit more from nutrient recycling than from WWT because the former offers potential increases in job opportunities and income in rural areas. The human health risks of enhanced nutrient recycling could be reduced by sealing domestic wastewater systems to minimize physical contact and by enhancing wastewater disinfection (e.g., chlorination, aerobic treatment, ozonation and/or UV light). Further improvements in livestock-waste management (~14.3 MtN·yr⁻¹ in excretion), such as manipulating animal diets, trapping particulate emissions and applying methane tanks, could also reduce air and water pollution [28], increase manure-N recycling rates and improve sanitation. These initiatives would create new market opportunities for the fertiliser industry [29] and farmers.

216

The massive challenges that China must overcome to restore safe and sustainable levels of N in water bodies are also shared by other regions where water pollution is increasing because of excess nutrient discharge. This includes many parts of Asia, South and Central America, and sub-Saharan Africa [7]. Well-defined targets for N releases into the local environment are essential for formulating effective regional policies to reduce pollution, which are in turn essential for progress at the global level (for example staying within “planetary boundaries”). Given the many environmental problems related to different types of N compounds and the complexity of the N cycle, coordinated approaches to nutrient management contribute to remaining within multiple planetary boundaries (climate, air quality, biodiversity) and have widespread co-benefits, including conserving water resources, lowering both air and water pollution, reducing N₂O-induced stratospheric ozone depletion, and increasing rural incomes.

230

231

**Acknowledgments:** This work was financially supported by the Chinese National Basic Research Program (2017YFA0603602 and 2014CB953803). HCJG and JH

233
234 acknowledge support from the Wellcome Trust, Our Planet Our Health (Livestock,

235 Environment and People project 205212/Z/16/Z).

236
<figure><img src="image_3.png" /></figure>

<caption>Fig.1 The changing nitrogen cycle and its contributions to food production in China during 1955-2014. (a) Nitrogen application in croplands (stacked bars); the yellow and purple lines mark our model-based estimates of the minimum nitrogen required to maintain annual food production under the two INM scenarios without changes in nutrient recycling rates (BF: broadcasting of fertilizer; CRF: controlled-release fertilizers). Error bars indicate upper and lower bounds on total N applied to croplands (see Methods). (b) Contribution of synthetic nitrogen (~45%, red line) to grain productivity. The blue line shows climate-driven variations in modeled grain yields (rice, maize and wheat) under the baseline scenario (current cropping with average agricultural inputs for 2007-2011; see Methods). The red line shows variations in modeled grain yields due to variations in climate and historical synthetic nitrogen inputs. The black line shows actual yields as recorded in the statistical yearbooks. The unexplained 25% growth in grain yields between 1955 and 2010 is attributed to additional contributions from irrigation expansion and technological advances that coupled with increased N inputs. (c) The broken nitrogen cycle: increases in synthetic nitrogen application are associated with declining rates of nutrient recycling since the late 1970s. The recycling rate is defined as the ratio of nitrogen from organic waste returned to croplands against the amount of nitrogen in organic waste, excluding emissions to the atmosphere. (d) The major anthropogenic N flows to croplands (green arrows: organic N, blue line: synthetic N), water (red arrows) and the atmospheric</caption>
261
262

(purple arrows) environment (2010-2014 average). N removal via WWT is based on
data from 2014. Imported N is counted in our assessment of excretion (Fig.S8).

<figure><img src="image_4.png" /></figure>

263

264 Fig.2 Reconstruction of N discharge during 1955-2014 and the associated evolution of
265 surface-water quality in terms of TN (data sources in Methods). Raster maps of
266 cropland N are based on leaching and runoff simulated by the DNDC model. Urban N
267 includes discharges of human sewage, industrial waste and garbage at provincial levels.
268 Rural N includes discharges of rural human sewage, livestock manure and garbage at
269 provincial levels. Observed national grain yields, synthetic N use, and estimated
270 nutrient recycling and discharge rates are shown in the panels to the right of the map.
271 To view an animation of this figure, open the attached powerpoint file, and initiate the
272 slide show. The animation will begin automatically and step through successive years.
<figure><img src="image_5.png" /><figcaption>Fig.3 Thresholds of N discharge to the water environment and the potential contributions of INM in reducing N pollution. (a) The years in which N pollution first exceeded the national standard in each province (see Table S3 for representative catchments and references). (b) 60-year records of nitrogen discharge (runoff and leaching) to the water environment in mainland China from anthropogenic sources. The yellow and purple lines mark simulated minimum N discharge amounts through INM under the BF and CRF scenarios, respectively. The critical threshold is estimated by aggregating provincial-level N discharge amounts from the years marked in (a). Error bars indicate estimated upper and lower bounds of total N discharge into the water environment (see Methods). (c) Predicted reductions in emissions of air pollution (NH₃ and N₂O) from croplands are estimated by substituting CNM with INM. Controlled-release fertilizers (CRF) cannot significantly reduce N discharge to the aquatic environment relative to broadcasting of fertilizer (BF) under INM, but do reduce N loss in the form of NH₃.</figcaption></figure>
<figure><img src="image_6.png" /><figcaption>Fig.4 Anthropogenic nitrogen discharge and requirements to meet the critical threshold in each province of mainland China. Upper panels show provincial-level N discharge (see Fig.3a for locations) into the water environment under (a) current nitrogen management (CNM) and (b) improved nitrogen management (INM-CRF) in croplands based on 2010-2014 mean values. Percentages listed in (a) and (b) are current nutrient recycling rates. Lower panels list the minimum N recycling rates (in percentage of recyclable solid-liquid organic wastes) required for each province to meet its critical pollution threshold under (c) CNM and (d) INM-CRF. All discharge and recycling rates are normalized against the critical pollution threshold (marked as a red circle).</figcaption></figure>
This Figure will appear in SI, for enhancing Fig.1d. It was required by the reviewer #4.

<figure><img src="image_7.png" /></figure>

Fig.S8 An approximate estimate of China's soybean consumption (biomass values in black text) and N budget (in blue text) in 2016. In recent years imported grain has increased substantially in China, and reached 114.6 Mt in 2016. Most of the imported grain was in the form of soybeans (73%). This imported N has been considered in our study in estimates of both human diet and livestock feed.

(Sources: (1) Imported grain data, http://www.yuanbangtouzi.com/a/job/shichanggenzong/2017/0125/448.html)

(2) fraction of soybean consumption, http://www.chyxx.com/industry/201805/638760.html, in Chinese, access on Sep.17, 2018,

(3) fraction of oil cake consumption, http://www feedtrade.com.cn/technology/nutrition/ingredient/2012-08-22/2007111_2.html, access on Sep.17, 2018)
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