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ABSTRACT Mitigating agricultural ammonia (NH3) emissions in China is urgently needed to avoid further damage to human and ecosystem health. Effective and feasible mitigation strategies hinge

on integrated knowledge of the mitigation potential of NH3 emissions and the associated economic costs and societal benefits. Here we present a comprehensive analysis of marginal abatement

costs and societal benefits for NH3 mitigation in China. The technical mitigation potential of agricultural NH3 emissions is 38–67% (4.0–7.1 Tg N) with implementation costs estimated at US$

6–11 billion. These costs are much lower than estimates of the overall societal benefits at US$ 18–42 billion. Avoiding unnecessary fertilizer use and protein-rich animal feed could provide

30% of this mitigation potential without additional abatement costs or decreases in agricultural productivity. Optimizing human diets with less animal-derived products offers further

potential for NH3 reduction of 12% by 2050. SIMILAR CONTENT BEING VIEWED BY OTHERS AIR QUALITY, NITROGEN USE EFFICIENCY AND FOOD SECURITY IN CHINA ARE IMPROVED BY COST-EFFECTIVE AGRICULTURAL

NITROGEN MANAGEMENT Article 14 October 2020 COST-EFFECTIVE MITIGATION OF NITROGEN POLLUTION FROM GLOBAL CROPLANDS Article Open access 04 January 2023 GREENHOUSE GAS EMISSIONS FROM NITROGEN

FERTILIZERS COULD BE REDUCED BY UP TO ONE-FIFTH OF CURRENT LEVELS BY 2050 WITH COMBINED INTERVENTIONS Article 09 February 2023 INTRODUCTION Anthropogenic ammonia (NH3) emissions, primarily

from agriculture, have adversely affected environmental quality, including air pollution, soil acidification, eutrophication of water bodies, and led to tremendous damage to human health and

ecosystem health1,2. The cost of damage associated with agricultural NH3 emissions was estimated at US dollars (US$)55–114 billion in the European Union (EU) in 2008, with the largest

contribution due to increased human mortality from exposure to NH3-containing aerosols3,4. In the United States (US), annual health costs due to NH3 emissions were estimated at US$69–180

billion in 20115. Mitigating NH3 emissions has attracted much attention in high-income countries. For example, the Gothenburg Protocol was signed in 1999 to control long-range transboundary

transport of air pollutants among member countries within the United Nations Economic Commission for Europe. Following the Gothenburg protocol, the EU adopted the first National NH3 Emission

Ceilings directive (2001/81/EC) in 20016. The efficacy and costs of NH3 abatement and their climate co-benefits were evaluated in 20157, focusing on European countries. To date, only a few

countries have estimated their national NH3 mitigation potential and associated costs and benefits (Table 1). China is the world’s largest emitter of NH3 (9–13 Tg N year−1 in the 2010s),

with over 80% contributed by agriculture8,9. Low fertilizer nitrogen (N) use efficiency (NUE) and poor animal waste management have resulted in enormous NH3 emissions in China8,10. Worse

still, regional NH3-related pollution is enhanced due to the increasing decoupling between crop and livestock production systems11. In recent years, frequent smog events with high

concentrations of PM2.5 (fine particulate matter < 2.5 µm) in China have triggered both public anxiety and concerns of the Chinese government12. A substantial proportion of PM2.5

pollution was caused by aerosol formation driven by NH3 emissions13,14. Studies have suggested that the current clean air policy for reductions in primary PM2.5, sulfur dioxide (SO2), and

nitrogen oxides (NOx) has limitations, and that PM2.5 pollution can be cost-effectively controlled only if NH3 emissions are abated as well as those of SO2 and NO_x_15,16,17. Studies have

also suggested that many NH3 abatement techniques may simultaneously reduce agricultural methane (CH4) and nitrous oxide (N2O) emissions, bringing co-benefits for agricultural greenhouse gas

(GHG) mitigation7,18,19,20. However, NH3 emission reduction in China may worsen the adverse impact of acid rain on crops and forests by increasing rainfall acidity21,22. To date, China has

not yet formulated or implemented policies to reduce NH3 emissions23, although there are many available measures to reduce NH3 emissions from agriculture, most of which have been validated

and adopted in the EU and North America7,24. Many NH3 abatement measures have not been widely practiced in China and their implementation costs and the impacts on agricultural GHG emissions

have not been assessed. Given that poor smallholder farmers still dominate China’s agricultural production and that agricultural N pollution is severe25, it is crucial to identify feasible

and cost-effective NH3 abatement measures for Chinese agriculture. A national systematic assessment of NH3 mitigation potential, and the associated costs and societal benefits, is urgently

needed for China to establish cost-effective mitigation strategies and targets. To fill the knowledge gap, this study builds an integrated NH3 mitigation assessment framework (Supplementary

Fig. 1) with the combination of Coupled Human And Natural Systems (CHANS), GAINS, Weather Research and Forecasting-Community Multiscale Air Quality (WRF-CMAQ), and exposure–response models

to: (1) identify feasible NH3 abatement options and to estimate the agricultural NH3 mitigation potential and the associated implementation costs and societal benefits; (2) determine

mitigation priorities and strategies for China; and (3) to explore optimal NH3 mitigation pathways for the next 30 years (2020–2050) using scenario analysis and cost-benefit assessment. We

find that the relative NH3 mitigation potential in China is twice that in Europe. The overall societal benefits of agricultural NH3 mitigations in China far exceed the abatement cost and

increase when including the synergy with reduction of GHG emissions. RESULTS AND DISCUSSION NH3 MITIGATION POTENTIAL, ABATEMENT COSTS, AND SOCIETAL BENEFITS For cropping systems NH3

abatement measures include reductions of urea-based fertilizer, promotion of enhanced efficiency N fertilizer (EENF), and deep placement of fertilizer (Supplementary Table 1). The NH3

mitigation potential of crop production is around 2.0–3.4 Tg N year−1 at an abatement cost of US$1.9–3.4 billion. The three major staple crops in China have the largest NH3 mitigation

potential at 460–954 Gg N for maize, 516–684 Gg N for rice, and 446–731 Gg N for wheat. The large reduction potential is mainly due to large sowing areas and poor fertilization practices.

The production of vegetables and fruits consumes one-third of total synthetic N fertilizer use in China, and their NH3 mitigation potentials are estimated at 30–55% (269–493 Gg N) and 20–40%

(118–235 Gg N), respectively. Unit abatement costs (US$ ha−1 year−1, Table 2) for cash crops (sugar, fruits and, vegetables) are higher than those for staple crops because the production of

cash crops is more intensive, requiring higher inputs of manpower, fertilizer, and financial resources26. For livestock production systems NH3 abatement measures include manipulation of

feed rations, improved housing facilities and manure management practices (Supplementary Table 2). The NH3 mitigation potential of livestock production is around 2.1–3.6 Tg N year−1 at an

abatement cost of US$4.2–7.8 billion. The pig industry in China has the largest NH3 mitigation potential at 838–1408 Gg N, followed by poultry farming (421–843 Gg N) and cattle production

(448–751 Gg N). Livestock units (LUs), a metric used in this study to compare the unit abatement costs between different animal types (Table 2) on the basis of the feed requirement of each

type of animal27. Generally, the unit abatement costs differ notably among animal types. For instance, the unit abatement cost is the highest for beef cattle (US$44–86 LU−1 year−1) while it

is only US$8–14 LU−1 year−1 for hogs. Differences are due to inherent disparities in animal feed, digestibility and farming practices28. The total abatement cost for the cattle industry is

the highest (US$1.7–3.3 billion), followed by pig (US$1.6–2.9 billion) and poultry (US$0.6–1.2 billion) farming. For other types of livestock abatement costs are relatively low owing to

their smaller NH3 emission rates and smaller animal populations. When mitigation options are combined for different crops and animal types, total agricultural NH3 mitigation potential is

estimated at 38–67% (4.0–7.1 Tg N) of total NH3 emissions, with implementation costs estimated at US$6–11 billion, equivalent to 0.04–0.08% of the national GDP of China. The societal

benefits of NH3 emission reduction were also quantified for comparison with the implementation costs of abatement measures. The mitigation of NH3 emissions by 38–67% could reduce PM2.5

concentrations by 8–20%, and avoid premature mortalities by 90–240 thousand people with health benefits at US$10–26 billion. Ecosystem benefits due to NH3 mitigation in terms of avoided soil

acidification and water eutrophication are estimated at US$10–17 billion. Further, NH3 mitigation could simultaneously reduce agricultural GHG emissions by 9–35% (101–385 Tg CO2 equivalents

(CO2-eq)), and generate climate benefits of US$1–3 billion. However, reductions of NH3 emissions might increase the acidity of precipitation and cause an economic loss of US$4–7 billion.

Although acid rain damage partly offsets the benefit of NH3 mitigation, the overall societal benefits of NH3 control (US$18–42 billion) still far outweigh the abatement costs (US$6–11

billion), suggesting that mitigation of agricultural NH3 emissions could generate net economic benefits (NEBs) on a national scale. NH3 MITIGATION PRIORITY AND STRATEGIES FOR CHINA The

marginal abatement cost curve (MACC) can be used to support decision making in prioritizing strategies for NH3 mitigation15,29. Using data presented in Table 2, a bottom-up MACC was

constructed (Fig. 1), which plots the cumulative NH3 emission-reducing potential of measures with increasing implementation cost per unit of NH3 emission reduction. The MACC illustrates that

the agricultural sector offers an average reduction potential of 5.5 Tg NH3–N at a total cost of US$8.3 billion. A reduction of 1.6 Tg NH3–N (30% of the total reduction) is potentially

available at a negative cost (cost-saving) for the agricultural sector by preventing unnecessary use of N fertilizer and protein-rich animal feed. These savings can be used to cover the

implementation cost of the next incremental reduction of 1.7 Tg NH3–N. As a result, 3.3 Tg NH3–N (60% of the total reduction) can be abated at zero cost. Overall, a 90% reduction could be

achieved at an average cost below US$1.2 per kg NH3–N, which is the estimated threshold of abatement cost in the EU27 to meet the targets of the Thematic Strategy on Air Pollution30. The

MACC highlights the importance of reducing synthetic N fertilizer use, coupled with improved animal feeding practices, as potential foci for mitigating China’s agricultural NH3 emissions. We

found that mitigating China’s agricultural NH3 emissions should start with the reduction of urea-based N fertilizer (RNF) in cropland by 20%. This option could offer 27% of mitigation

potential for cropland and save fertilizer costs of US$2.0 billion without yield loss. In general, fertilizer N application in China has far exceeded the crop demand31. Studies have proven

that reducing current N applications by 30–60% with optimum N management would still maintain crop yields while saving unnecessary economic expenditure for farmers19,31. Also, the reduction

of protein (N) rich feed (LNF) could reduce N losses from excreta and result in a decrease in 0.5 Tg of NH3–N emission without undermining animal productivity. The next priority mitigation

measures are covered storage (CS) of pig and poultry manures, followed by improved application (IA) of animal manure to cropland. These measures are able to reduce losses of 1.2 Tg of NH3–N

with a cost of US$1.1 billion for manure processing. In addition, housing adaptation (HA) for poultry can reduce another 0.3 Tg of NH3–N emission at a cost of US$0.6 billion. Improved N

fertilizer management (IFM) for three major staple crops can reduce NH3–N emission by 1.2 Tg of at a cost of US$2.6 billion. In practice, the net cost of IFM may be smaller because the

increased crop quality and yield may partly compensate for the investment costs (e.g., equipment such as subsurface injectors) and operational costs (e.g., material, labor, and energy).

Advanced housing systems for pigs and cattle (e.g., air-scrubbing techniques) can be very effective for reducing NH3 emissions (up to 80%) but are also costly (US$10–30 per kg NH3–N

mitigated). NH3 MITIGATION PATHWAYS IN 2020–2050 Scenario analysis and cost-benefit assessment guide the optimization of NH3 mitigation strategies and pathways in the future. In this study,

one baseline scenario of business as usual (BAU) and four mitigation scenarios (DIET, NUE, REC, and ALL) toward 2050 are analyzed, which comprise a range of packages of mitigation options

(see Table 3 for details) to explore optimal mitigation pathways. The simulated NH3 emission trends (Fig. 2) for the next 30 years (2020–2050) reveal that there would be substantial NH3

mitigation potential with broad welfare benefits (Fig. 3). Under the BAU scenario total agricultural NH3 emissions in China are estimated to first increase from 10.9 Tg N in 2015 to 12.1 Tg

N in 2035 because of a growing and changing food demand for China’s increasing and wealthier population27. The emission would then slightly decrease to 11.9 Tg N following a decrease in

China’s population toward 205032. NH3 emission from synthetic N fertilization is expected to remain stable during this period considering the national “Zero-growth Action Plan” for chemical

fertilizer use33. The major cause of increased NH3 emission is the rising livestock production to meet the growing demand for animal products both in total and per capita terms27,34. In

contrast, the DIET scenario assumes optimizing human dietary structure (transitioning toward more plant-based diets) to reduce the animal-based food N to 40%, which is in line with Chinese

dietary guidelines35. The increased human consumption of plant-based food N will shorten the food chain and decrease food-feed competition from decreased livestock farming. Decreased

livestock production (meat, eggs, and milk) in DIET reduces the demand for crop production by 20–30% relative to the BAU scenario, which could reduce NH3 emission by 21% by 2050 (Fig. 2).

Based on the proposed improvement in NUE in crop and livestock production systems by adopting advanced farming practices, or techniques as identified in this study, agricultural NH3

emissions are projected to decline from 11.9 to 8.8 Tg N in 2050 under the NUE scenario (Fig. 3). This scenario could decrease synthetic N fertilizer use by 13 Tg N, decreasing NH3 emission

from cropping systems by 39%. In addition, NH3 emission from livestock systems could be reduced by 1.9 Tg N through improved animal feed and waste management. Cropland in China is poorly

coupled with its livestock production systems11. The REC scenario aims to reconnect the two agricultural subsystems by increasing the recycling of manure to croplands. The total excretion N

generated by livestock production was 13.4 Tg N in 2015 and is estimated to reach 18.2 Tg N in 2050 under the REC scenario. Simultaneously, the total N demand of crops in China is estimated

to be 25.0 Tg N in 2050, suggesting that animal excretion N would be within the carrying capacity of cropland in China. Under the assumption that 60% of manure N is recycled to croplands,

the REC scenario could save 10.9 Tg chemical N fertilizer and reduce NH3 emission by 24% (3.3 Tg N) in 2050. The abatement cost is estimated at US$3.8 billion, acknowledging the considerable

socio-economic barriers of relocation and adaptations of the livestock supply chain and transport infrastructure3. To achieve the most ambitious mitigation target, the ALL scenario combines

all the mitigation options identified in DIET, NUE, and REC scenarios. The estimated mitigation potential of the ALL scenario is 7.2 Tg NH3–N (61% reduction relative to BAU), of which an

80% reduction is achieved by improved agricultural management and technical solutions, while the remaining 20% reduction is due to decreasing the consumption of animal products. The ALL

scenario suggests that after achieving the technical mitigation potential, reducing consumption of animal products could offer a further 12% of NH3 mitigation potential by 2050. However, the

abatement costs vary among the four mitigation scenarios: The DIET scenario has the smallest mitigation cost (<US$1.0 billion) given that shifting to sustainable diets mainly depends on

cost-free adjustment of consumers’ preference and behaviors. The ALL scenario has the greatest abatement cost (up to US$11 billion in 2050) because it requires comprehensive and coordinated

actions. The benefits from improved human and ecosystem health increase with the extent of NH3 mitigation for all mitigation scenarios. The NUE, DIET, and REC scenarios have similar

mitigation potentials and human and ecosystem health benefits of US$22–26 billion while the ALL scenario generates the highest health benefit of US$72 billion in 2050 (Fig. 3). The

agricultural GHG emissions vary between the NH3 mitigation scenarios. The DIET, NUE, and ALL scenarios can significantly reduce agricultural GHG emissions and generate positive climate

benefits, while REC slightly decreases GHG emissions because manure addition to soil increases GHG (N2O) emission over time36, but that could be partially offset by indirect N2O mitigation.

The reduction of agricultural NH3 emissions may aggravate acid rain over China and may result in economic loss under all mitigation scenarios (Fig. 3). Nevertheless, the overall societal

benefits under all four mitigation scenarios still exceed the corresponding abatement costs and result in net national social welfare. INTERNATIONAL COMPARISON The EU27 (27 countries of the

EU) has a history of NH3 mitigation over about two decades6, while Canada has just revealed its national mitigation potential24. In the USA, several studies have indicated that agricultural

NH3 emissions are an important driver of PM2.5 pollution causing huge health costs5,37. The current average NH3 mitigation potential of China (53%) is around twice that of the EU27 (24%)30

and Canada (29%)24 (Table 1). This is not surprising because China has the highest quantities of NH3 emissions in the world but has not yet implemented mitigation policies. There are two

main reasons for the high NH3 emissions currently in China. First, the total amount of fertilizer applied to Chinese croplands has increased more than threefold since 1980, accounting for

one-third of global synthetic N fertilizer consumption10. The excessive N input, poor farming practices and small farm size have led to a low NUE and high NH3 losses in Chinese cropping

systems25. Second, the livestock population increased from 142 to 441 million LUs, almost tripling between 1980 and 201027. The rapid growth in intensive industrial livestock production and

unsustainable management practices arising from the unbalanced spatial distribution of livestock farms to poor animal house cleansing and manure handling have resulted in large waste of

nutrients in manure11,27, which further increases NH3 emissions from animal husbandry. The weighted unit abatement cost for China is estimated to be US$0.8–2.1 per kg NH3–N, which is lower

than that of the EU27 at US$1.2–3.5 per kg NH3–N38. This discrepancy is attributed to the large differences in farm sizes, labor costs, and agricultural mechanization levels. Recent studies

indicate that larger-scale farms are more efficient in fertilizer use, labor, and professional management than smaller ones25. However, smallholder farming dominates the agricultural

landscape in China39. Land fragmentation and small farm size in China reduce the efficiency of machinery and services25,40. However, the unit labor cost in China is around one-sixth of the

EU, which can partly explain the lower implementation cost for the early stage of NH3 mitigation (defined as the implementation for the first 60% of mitigation potential) in China as

compared to the EU27. The low-hanging fruit for achieving reductions in NH3 reductions is the direct reduction of urea-based fertilizer use and protein-rich feeding in animal production. The

remaining mitigation will become increasingly more expensive due to the greater requirements of technologies and equipment, with a marginal abatement cost in the range of US$2–15 per kg

NH3–N mitigated in China (Fig. 1). The transition to large-scale and mechanized agriculture in China is restricted by inherent social barriers and weak technical foundation26, which takes

time and effort to overcome. POLICY IMPLICATIONS To clean up the air, Chinese governments have already made major efforts to reduce anthropogenic SO2 and NO_x_ emissions, which have declined

by 41% and 34%, respectively, from their peaks to 201941. Although continuing the stringent policies to reduce SO2 and NO_x_ emissions could further improve air quality, and may partially

offset the effects of NH3 mitigation, studies have suggested that current policies are not sufficient or cost-effective in achieving the targets of clean air in China16,23,42. A recent study

has found that reducing livestock NH3 emissions would be highly effective in reducing PM2.5 during severe winter haze events43. Our quantitative assessments of the implementation cost and

societal effects of NH3 mitigation in China further demonstrate that NH3 mitigation could generate net societal benefits, even though it may worsen regional acid rain. Therefore, coordinated

mitigation of multiple air pollutants (SO2, NO_x_, and NH3) should be implemented to more rationally and effectively achieve the dual benefits of protecting human and ecosystem health in

China at both national and regional scales21. For farm holders, strategic designs of cost-effective mitigation pathways are needed. The aforementioned cheap and easy mitigation options

(direct reduction of N fertilizer use and improved animal feeding practices) should be introduced first to pick the low-hanging fruit of NH3 mitigation in China. The remaining mitigation

measures (e.g., HA and manure handling systems) are expensive due to the higher requirements of the investments in technologies and infrastructures. It is necessary to increase government

support (e.g., technical guidance and training) and subsidies (e.g., enhanced efficiency fertilizers and agricultural machinery) to encourage farmers to adopt these mitigation measures10.

Perhaps even more challenging, the government should also promote the reform of China’s land tenure system to facilitate large-scale farming44. Large-scale farms will be a better platform

for the implementation of advanced management practices and mechanization (e.g., deep application of fertilizers) and can reduce the abatement cost per unit cropland area10,25. Livestock

production is shifting from small-scale outdoor systems to large-scale intensive indoor systems27, which causes decoupling between croplands for feed production and industrial feedlots10. As

a consequence, manure is increasingly dumped or discharged instead of being recycled or reused owing to high transportation costs, resulting in huge NH3 emissions in China11. In 2015 only

30% of livestock manure N was recycled to croplands in China11 while in the EU the proportion was more than 65%45. Relocating feedlots to feed croplands can greatly improve manure recycling

and reduce the associated implementation costs where livestock densities being kept within the cropland carrying capacity for manure application11. Financial incentives (e.g., subsidies,

discounted interest, technical guidance, taxation exemption, etc.) are required to help farmers develop a region-specific farming structure that facilitates manure recycling, optimizes N

management and promotes large-scale operation27. Furthermore, it should be noted that NH3 mitigation through human dietary changes, also benefits human health46,47. Dietary change is a

nontechnical measure with little implementation cost but requires other interventions to change consumers’ preferences. The government can play an essential role in setting up campaigns to

promote low-protein diets. LIMITATIONS AND UNCERTAINTY Agricultural NH3 mitigation strategies are linked to the overall N cycle and could affect agricultural production and farmers’

incomes7, which may further influence food security and rural economies. This study did not incorporate the effects of NH3 mitigation on crop yield or animal productivity in the cost-benefit

assessment of scenarios owing to the lack of comprehensive Chinese-specific data. In fact, fertilizer N application in China far exceeds the crop demand; NH3 mitigation by improved farming

practices would unlikely create N limitation or reduce crop yields19,31. If the yield benefits attributed to NH3 mitigation could be quantified rationally and accurately, it would greatly

improve the cost-effectiveness of NH3 mitigation and therefore engage farmers to adopt these measures. Besides, this study does not address the regional difference in China due to the lack

of detailed regional data. Given the large differences in regional agricultural structures and environmental conditions, mitigation strategies and targets may vary considerably, which

affects the accuracy of current national estimates. In this paper we limited the climate benefits to non-CO2 GHG (CH4 and N2O) emissions resulting from NH3 mitigation. This is mainly because

CO2 emission from agriculture is more related to fossil fuel consumption, such as fertilizer production and transportation48, which is beyond the scope of this study. Furthermore, we did

not consider the effects of NH3 mitigation on climate change, including changes in aerosols and carbon sinks in terrestrial ecosystems, owing to limited research and models that target

China49,50. There are complex chemical interactions among SO2, NO_x_, and NH3 in the atmosphere51. Thus, the future policies to control SO2 and NO_x_ emissions may partially offset the

effects of NH3 mitigation, which also bring uncertainties to our estimations. While the projections of NH3 mitigation potential and costs toward 2050 are based on current technologies,

future technological advancement, and policy optimization may further reduce the implementation cost to yield a higher NEB. Nevertheless, as the first attempt, this study provides a basis

and reference for on-going improvement in NH3 mitigation potential and cost-benefit assessment. METHODS SELECTION OF AVAILABLE MITIGATION OPTIONS To identify feasible and cost-effective NH3

abatement measures for Chinese agriculture, we reviewed all currently available mitigation options in China and other countries. Our criteria for the selection of abatement measures focus on

five aspects: * (1) Mitigation efficiency: measures that could significantly reduce NH3 emissions are included, for example, deep manure placement has a very high mitigation potential at

93–99%52. * (2) Implementation cost: measures with lower cost or labor inputs are more acceptable to farmers, for example, reduced use of urea-based fertilizer and lower crude protein diet.

* (3) Practical applicability: measures with current limited applicability due to technical, political or obvious social barriers in China, were excluded, for example, soil testing has been

ruled out in this study due to high costs for the small farm size and high spatial and temporal variability, although it is an effective measure to optimize fertilizer use in the US and

Europe where farm sizes are much larger. * (4) Limitations: measures that likely and significantly reduce agricultural productivity (crop yield or animal productivity) were adopted with

caution, for example, the full substitution of synthetic fertilizers by manure decreases the yield of upland crops and lowland rice by 9.6% and 4.1%, respectively53; and low crude protein

(LCP) feeding should only be adopted to an applicable level to avoid undermining animal productivity and welfare. Besides, LCP is mainly applicable to indoor animals (pig, poultry, and

dairy). * (5) Presence of co-benefit: measures that could reduce both NH3 emission and total GHG (CH4 and N2O) emissions are included, for example, biochar additives could reduce NH3, N2O

and CH4 emissions during manure composting54. Based on the selection criteria and literature review, a total of 27 technical mitigation options for specific crops and animals were included

in this study, with a coded version provided in Supplementary Tables 1 and 2. Detailed descriptions of these options are listed in Supplementary Tables 3–6 and Supplementary Notes 1–3. Most

of these mitigation measures have been validated and adopted in the EU, while some of the measures (e.g., optimum N fertilization techniques) have been validated in China. For the measures

that have been validated in China we directly adopted their parameters, whereas for measures that have not been validated in China, we calculated their potential implementation costs based

on China-specific parameters such as labor cost, fertilizer prices, machinery cost. Only cost-effective measures and their combinations were selected for the analysis. Most agricultural NH3

and GHG emissions originate from the same activities (Supplementary Fig. 2) and their emission rates depend on common factors, such as management practice, weather conditions and soil type7.

NH3 abatement options can increase or decrease GHG emission20. This study aims to explore the maximum NH3 mitigation potential while achieving the co-benefits of GHG reduction. In this

context, optimal combinations of NH3 mitigation options for different crops and animals are proposed in Supplementary Table 7 with their abatement efficiencies. MITIGATION POTENTIAL OF NH3

EMISSIONS NH3 emissions from agriculture generally are assessed by multiplying the activity level with specific emission factors for each sector. The NH3 mitigation potential is calculated

as Eq. (1): $${\mathrm{\Delta }}E_{NH_3} = \mathop {\sum}\limits_i {{\mathrm{A}}_{i,k}\, * \,{\mathrm{EF}}_{i,k} \times \eta _{i,k} \times X_{i,k}},$$ (1) where \({\mathrm{\Delta

}}E_{NH_3}\) is the reduction of agricultural NH3 emissions in mainland China; _i_ represents the source type; _k_ means a specific abatement option or the combination of multiple options;

\({\mathrm{A}}_{i,k}\) is the activity data of the source type; \({\mathrm{EF}}_{i,k}\) is the original emission factor; \(\eta _{i,k}\) is the NH3 abatement efficacy; \(X_{i,k}\) is the

implementation rate of the abatement technique or options. CALCULATION OF NH3 ABATEMENT COST Abatement cost of NH3 emissions in this study is defined as direct expenditures (the sum of

investment costs and operation costs) for implementation of measures to reduce NH3 emissions from agriculture, while the possible public costs (e.g., subsidy to promote the control policy)

are not considered. Here, we mainly refer to the methodology of cost assessment from the GAINS model55 to calculate the abatement costs of implementing various NH3 mitigation measures.

China-specific commodity prices were collected mainly from the _China Agricultural Products Cost-Benefit Yearbook_ (2000–2018)26, European cost data were adopted by conversion at market

exchange rates where data supply is insufficient. All cost data from the literature were adjusted by the purchasing power parity (PPP) index and measured in constant 2015 US$ (e.g., 100 EUR 

= US$113.49, 100 CNY = US$14.89) by assuming a 2% annual inflation and setting 2015 as the base year for future projection. The calculation of abatement costs is formulated in Eq. (2):

$${\mathrm{AC}}_{i,k} = {\mathrm{I}}_{i,k}\, * \,\left[ {\frac{{\left( {1 + r} \right)^{lt_{i,k}} \times r}}{{\left( {1 + r} \right)^{lt_{i,k}}\, - \, 1}}} \right] + {\mathrm{FO}}_{i,k} +

{\mathrm{VO}}_{i,k} - {\mathrm{FS}}_{i,k},$$ (2) where AC_i_,_k_ represents the annual implementation cost; _I__i,k_ refers to the investment cost; _r_ is the discount rate; _lt__i,k_

represents the lifetime of abatement technique (10–15 years); \({\mathrm{FO}}_{i,k}\) is the fixed operating cost; \({\mathrm{VO}}_{i,k}\) is the variable operating costs (e.g., feed, gas,

electricity, labor, and water); FS_i,k_ means saving costs from reduced use of N fertilizer. Investment cost _I__i_,_k_ is calculated as a function of the average farm size

(\({\mathrm{AFS}}_i\)) by Eq. (3): $${I}_{i,k} = ci_{i,k}^f \cdot st_i \cdot mp_i \cdot pc_i + \frac{{ci_{i,k}^v}}{{{\mathrm{AFS}}_i}},$$ (3) where \(ci^f,ci^v\) represents the fixed and

variable coefficients derived from Klimont and Winiwarter (Annex: Table A1)56; _st__i_ represents manure storage time (in year); _mp__i_ represents manure production of a single animal per

year; _pc__i_ represents production cycles per year; parameters used in the function are available in an online GAINS report. Annual fixed operating costs \({\mathrm{FO}}_{i,k}\) are

estimated as the 0.05% of the total investments by Eq. (4) according to GAINS cost calculation21. $${\mathrm{FO}}_{i,k} = {I}_{i,k} \cdot 0.05\%.$$ (4) Variable operating costs

\({\mathrm{VO}}_{i,k}\) are calculated by cost summation of the quantity (_Q_) of a certain extra supply (e.g. feed, gas, electricity, water, and labor) for a specific abatement option

(_k_), as shown in Eq. (5): $${\mathrm{VO}}_{i,k} = \mathop {\sum}\limits_p {\left( {{Q}_{i,k} \cdot c_{i,k}} \right)},$$ (5) where _p_ represents parameter type (additional feed, gas,

electricity, water and labor input); \(c_{i,k}\) means the unit price of these extra supply, which is mainly derived from the _China Agricultural Products Cost-Benefit Yearbook_26 and market

survey or adjusted by a coefficient if no direct data source could be accessed. The unit labor cost of farmworkers in 2015 is 15.7 Chinese yuan (CNY) per hour according to the national

averaged salary for individual persons26,57. Other relevant parameters used in the calculation of FO and VO are obtained from GAINS. The cost-effectiveness of various NH3 mitigation options

was calculated following Eq. (6)55,58 to yield MACC curve according to increasing cost-effectiveness, as shown in Fig. 1. $${\mathrm{CE}}_{i,k} = \frac{{{\mathrm{CE}}_{i,k}\, * \,\eta _{i,k}

- {\mathrm{CE}}_{i,k - 1}\, * \,\eta _{i,k - 1}}}{{\eta _{i,k} - \eta _{i,k - 1}}},$$ (6) where \(CE_{i,k}\) is the cost-effectiveness for mitigation option k; \(\eta _{i,k}\) is the NH3

mitigation efficiency. SCENARIO ANALYSIS OF FUTURE NH3 EMISSIONS To explore the feasibility of NH3 mitigation in the future, the CHANS model was employed in this study to make systematic and

comprehensive analyses of NH3 sources, emissions, and environmental fates8. A detailed introduction of the model can be found in Zhang et al.8 and Gu et al.59. Taking into consideration the

impacts of policy, and other external factors on Chinese agricultural production and consumption, the baseline NH3 emission budgets during 2020–2050 were built in the first place, then four

abatement scenarios (DIET, NUE, REC, and ALL) with corresponding packages of mitigation measures (detailed description in Table 3) were integrated into the CHANS model to quantify the new

NH3 emission budgets and identify the feasible NH3 reduction potential in China. Human population and the per capita gross domestic product are two important parameters that affect future

NH3 emission budgets. These two parameters are assumed to remain the same as the BAU for the four mitigation scenarios while other input drivers and parameters, such as diet structure, NUE,

cropping area, animal numbers, will vary with scenarios (Supplementary Fig. 6). Details about the data sources, prediction methods and parameters can be found in Supplementary Tables 8–18

and Supplementary Note 4. It should be noted that change in human diet structure as a nontechnical measure was also included in the scenario analysis to obtain a more comprehensive

assessment of the mitigation potential and pathways. SOCIETAL BENEFIT ASSESSMENT OF NH3 MITIGATION Societal benefits (\({\mathrm{SOC}}_{{\mathrm{benefit}}}\)) of NH3 mitigation in this study

are defined as the sum of benefits for human health (\({\mathrm{HH}}_{{\mathrm{benefit}}}\)), ecosystem health (\({\mathrm{EH}}_{{\mathrm{benefit}}}\)), GHG mitigation benefit

(\({\mathrm{GHG}}_{{\mathrm{benefit}}}\)) minus the cost of damage by increased acidity of precipitation (\({\mathrm{AR}}_{{\mathrm{damage}}}\), as shown in Eq. (7)

$${\mathrm{SOC}}_{{\mathrm{benefit}}} = {\mathrm{HH}}_{{\mathrm{benefit}}} + {\mathrm{EH}}_{{\mathrm{benefit}}} + {\mathrm{GHG}}_{{\mathrm{benefit}}} - {\mathrm{AR}}_{{\mathrm{damage}}}.$$

(7) The human health benefit assessment was performed based on the exposure–response function and the Value of Statistical Life (VSL) as applied in earlier studies both at the global and

national scales1. Five categories of diseases causing premature mortality via PM2.5 pollution are considered in this study, namely lower respiratory tract illness, chronic obstructive

pulmonary disease (COPD), ischemic heart disease (IHD), lung cancer (LC) and cerebrovascular disease (CEV). The impacts of NH3 mitigation on annual PM2.5 concentration were assessed based on

the model simulation of Weather WRF-CMAQ performed by Xu et al.60. A deduced nonlinear function between PM2.5 concentration and NH3 reduction was built in Eq. (8); detailed description of

WRF-CMAQ simulation can be found in Xu et al.60 and Supplementary Note 5. Then, an exposure–response function (Eq. (9)) was combined with the health effect function (Eq. (10)) based on

Global Burden of Disease61 to estimate the 1-year premature mortality attributable to PM2.5 exposure. Afterward, we used the updated Chinese-specific VSL following the method in Giannadaki

et al.1 to derive corresponding economic benefits of NH3 abatement by Eq. (11) in China. $$C_j = C_{2015}\, \times \,(1 - 0.0173\, \times \,e^{2.9532 \times \eta _j}),$$ (8)

$${\mathrm{HE}}_{j,q} = \mathop {\sum}\limits_j {e^{\beta _q\, \times \,(C_j - C_o)}\, \times \,{\mathrm{HE}}_{0,q}},$$ (9) $${\mathrm{\Delta }}M_j = \mathop {\sum}\limits_q {\left(

{{\mathrm{HE}}_{j,q} - {\mathrm{HE}}_{0,q}} \right)\, \times \,{\mathrm{Pop}}_j},$$ (10) $${\mathrm{HH}}_{{\mathrm{benefit}},j} = VSL_j\, \times \,{\mathrm{\Delta }}M_j,$$ (11) where _C__j_

is the annual average PM2.5 concentration in year _j_; _C_2015 is the annual average PM2.5 concentration in year 2015 (50 µg m−3); \(\eta _j\) is the reduction rate (%) of NH3 emission; _q_

represents the category of diseases (IRL, COPD, IHD, LC, CEV); \(\beta _q\) is the coefficient in the exposure–response function which refers to the proportion of change in the endpoint of

each health effect of the population for unit change in PM2.5 concentration; _C_0 is the background concentration below which no health impact is assumed (10 µg m−3 as suggested by the

WHO62); HE0,_q_ is the baseline health effect (the mortality risk) due to a particular disease category for China estimated by the WHO61; HE_j,q_ is the actual health effect under

significant PM2.5 pollution levels; Pop_j_ is the population exposed to air population in China; \({\mathrm{\Delta }}M_j\) is the avoided death toll; _VSL__j_ is the Chinese-specific value

of a statistical life derived from Giannadaki et al.1; \({\mathrm{HH}}_{{\mathrm{benefit}},j}\) means the human health benefits by NH3 mitigation. Ecosystem benefits in this study are

regarded as the avoided damage cost of decreased acidification and eutrophication of ecosystems due to NH3 mitigation. We assume the unit NH3 damage cost to the ecosystem in the European

Nitrogen Assessment21 is also applicable to China after correction for differences in the willingness to pay (WTP) for ecosystem service and PPP in China and EU, as shown in Eq. (12).

$${\mathrm{EH}}_{{\mathrm{benefit}},j} = {\mathrm{\Delta }}E_{{\mathrm{NH}}_3,j}\, \times \,\partial _{{\mathrm{EU}}}\, \times

\,\frac{{{\mathrm{WTP}}_{{\mathrm{China}}}}}{{{\mathrm{WTP}}_{{\mathrm{EU}}}}}\, \times \,\frac{{{\mathrm{PPP}}_{{\mathrm{China}},j}}}{{{\mathrm{PPP}}_{{\mathrm{EU}},j}}},$$ (12) where

\(\partial _{{\mathrm{EU}}}\) is the estimated unit ecosystem damage cost of NH3 emission in relation to soil acidification and water eutrophication in Europe based on the European Nitrogen

Assessment63; WTPChina and WTPEU are the values of the WTP for ecosystem service in China and Europe; PPPChina,_j_ and PPPEU,_j_ stand for the PPP of China and the EU. GHG benefit from NH3

mitigation is regarded as the avoided abatement cost of GHG emissions, as shown in Eqs. (13)–(14). $${\mathrm{GHG}}_{{\mathrm{benefit}},j} = {\mathrm{\Delta }}E_{{\mathrm{GHG}},j}\, *

\,{\mathrm{MCost}}_{{\mathrm{GHG}},j},$$ (13) $${\mathrm{\Delta }}E_{{\mathrm{GHG}},j} = \left( {{\mathrm{\Delta }}E_{{\mathrm{directN}}_2{\mathrm{O}},j} + {\mathrm{\Delta

}}E_{{\mathrm{indirectN}}_2{\mathrm{O}},j}} \right)\, * \,298 + {\mathrm{\Delta }}E_{{\mathrm{CH}}_4,j}\, * \,34,$$ (14) where \({\mathrm{\Delta }}E_{{\mathrm{GHG}},j}\) is the estimated

reduction in agricultural GHG emissions, presented as kg CO2-eq, using the default values of 298 kg CO2-eq for N2O emissions, and 34 kg CO2-eq for CH4 emissions at a 100-year time horizon64;

both reduction of direct and indirect N2O emissions are included, the indirect N2O reduction is calculated as 1% of reduced NH3 deposition according to the IPCC guideline65.

\({\mathrm{MCost}}_{{\mathrm{GHG}},j}\) represents the marginal abatement cost (the carbon price) to reduce one tonne of GHG emissions in $ per tonne CO2-eq, the Chinese-specific (East Asia)

value is derived from West et al.66. Acid rain damage (\({\mathrm{AR}}_{{\mathrm{damage}},j}\)) induced by NH3 mitigation refers to the economic loss of reduced crop yields

(\({\mathrm{Crop}}_{{\mathrm{damage}},j}\)) and forestry (\({\mathrm{Forest}}_{{\mathrm{damage}},j}\)) in Eq. (15). Based on the experimental results reported in Feng et al.22 and model

simulation of precipitation acidity in Liu et al.21, we estimated the economic cost of increased acid rain under different mitigation scenarios. $${\mathrm{AR}}_{{\mathrm{damage}},j} =

{\mathrm{Crop}}_{{\mathrm{damage}},j} + {\mathrm{Forest}}_{{\mathrm{damage}},j}.$$ (15) REPORTING SUMMARY Further information on research design is available in the Nature Research Reporting

Summary linked to this article. DATA AVAILABILITY Data supporting the findings of this study are available within the article and its supplementary information files, or are available from

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Effects and Policy Perspectives_ (Cambridge University Press, 2011). Download references ACKNOWLEDGEMENTS This study was supported by the National Key Research and Development Project of

China (2018YFC0213300 and 2016YFC0207906), National Natural Science Foundation of China (41822701 and 41773068), Discovery Early Career Researcher Award of the Australian Research Council

(DE170100423), the Melbourne Research Scholarship, Australia-China Joint Research Centre of Healthy Soils for Sustainable Food Production and Environmental Quality (ACSRF48165) and Linkage

Projects of the Australian Research Council (LP160101417). We would like to thank Professor Wilfried Winiwarter, Arvin Mosier and Ian Willett in particular for their helpful comments on an

earlier draft. AUTHOR INFORMATION AUTHORS AND AFFILIATIONS * School of Agriculture and Food, The University of Melbourne, Melbourne, VIC, 3010, Australia Xiuming Zhang, Shu Kee Lam, Xia

Liang, Mei Bai & Deli Chen * College of Environmental and Resource Sciences, Zhejiang University, Hangzhou, 310058, PR China Baojing Gu * PBL Netherlands Environmental Assessment Agency,

PO BOX 30314, 2500 GH, The Hague, The Netherlands Hans van Grinsven Authors * Xiuming Zhang View author publications You can also search for this author inPubMed Google Scholar * Baojing Gu

View author publications You can also search for this author inPubMed Google Scholar * Hans van Grinsven View author publications You can also search for this author inPubMed Google Scholar

* Shu Kee Lam View author publications You can also search for this author inPubMed Google Scholar * Xia Liang View author publications You can also search for this author inPubMed Google

Scholar * Mei Bai View author publications You can also search for this author inPubMed Google Scholar * Deli Chen View author publications You can also search for this author inPubMed 

Google Scholar CONTRIBUTIONS B.G. and X.Z. designed the study. X.Z. performed the research. X.Z. and B.G. analyzed the data and interpreted the results. H.G. provided cost and benefit

analysis supports. X.Z. and B.G. wrote the paper, D.C., S.L., X.L., and M.B. contributed to the discussion and revision of the paper. CORRESPONDING AUTHORS Correspondence to Baojing Gu or

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_et al._ Societal benefits of halving agricultural ammonia emissions in China far exceed the abatement costs. _Nat Commun_ 11, 4357 (2020). https://doi.org/10.1038/s41467-020-18196-z

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