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Global Change Biology Global Change Biology (2015) 21, 1249–1257, doi: 10.1111/gcb.12802 How inhibiting nitrification affects nitrogen cycle and reduces environmental impacts of anthropogenic nitrogen input CHUNLIAN QIAO1,2, LINGLI LIU1, SHUIJIN HU3, JANA E. COMPTON4, T A R A L . G R E A V E R 5 and Q U A N L I N L I 6 1 State Key Laboratory of Vegetation and Environmental Change, Institute of Botany, Chinese Academy of Sciences, No. 20 Nanxincun, Xiangshan, Beijing 100093, China, 2University of Chinese Academy of Sciences, No. 19 Yuquan Road, Beijing 100049, China, 3Department of Plant Pathology, North Carolina State University, Raleigh, NC 27695, USA, 4Western Ecology Division, US Environmental Protection Agency, Corvallis, OR 97333, USA, 5National Center for Environmental Assessment, US Environmental Protection Agency, Research Triangle Park, NC 27695, USA, 6Biostatistics and Bioinformatics, Department of Medicine, Cedars-Sinai Medical Center, Los Angeles, CA 90048, USA Abstract Anthropogenic activities, and in particular the use of synthetic nitrogen (N) fertilizer, have doubled global annual reactive N inputs in the past 50–100 years, causing deleterious effects on the environment through increased N leaching and nitrous oxide (N2O) and ammonia (NH3) emissions. Leaching and gaseous losses of N are greatly controlled by the net rate of microbial nitrification. Extensive experiments have been conducted to develop ways to inhibit this process through use of nitrification inhibitors (NI) in combination with fertilizers. Yet, no study has comprehensively assessed how inhibiting nitrification affects both hydrologic and gaseous losses of N and plant nitrogen use efficiency. We synthesized the results of 62 NI field studies and evaluated how NI application altered N cycle and ecosystem services in N-enriched systems. Our results showed that inhibiting nitrification by NI application increased NH3 emission (mean: 20%, 95% confidential interval: 33–67%), but reduced dissolved inorganic N leaching (48%, 56% to 38%), N2O emission (44%, 48% to 39%) and NO emission (24%, 38% to 8%). This amounted to a net reduction of 16.5% in the total N release to the environment. Inhibiting nitrification also increased plant N recovery (58%, 34–93%) and productivity of grain (9%, 6–13%), straw (15%, 12–18%), vegetable (5%, 0–10%) and pasture hay (14%, 8–20%). The cost and benefit analysis showed that the economic benefit of reducing N’s environmental impacts offsets the cost of NI application. Applying NI along with N fertilizer could bring additional revenues of $163 ha1 yr1 for a maize farm, equivalent to 8.95% increase in revenues. Our findings showed that NIs could create a win-win scenario that reduces the negative impact of N leaching and greenhouse gas production, while increases the agricultural output. However, NI’s potential negative impacts, such as increase in NH3 emission and the risk of NI contamination, should be fully considered before large-scale application. Keywords: cost–benefit analysis, ecosystem services, N2O emission, NH3 emission, nitrogen fertilizer, nitrogen leaching, nitrogen management, NO emission Received 5 May 2014; revised version received 21 October 2014 and accepted 21 October 2014 Introduction Synthetic nitrogen fertilizers were developed via the Haber–Bosch process over a century ago and have been applied intensively since then to increase plant growth (Galloway et al., 2008); they now exceed terrestrial biological N fixation as the main source of new reactive N (Nr) to the global N cycle (Erisman et al., 2011; Fowler et al., 2013; Vitousek et al., 2013). N fertilizers have made a remarkable contribution in the alleviation of global food shortage, increasing food production by almost 50% (Sutton et al., 2011). However, N fertilizers Correspondence: Lingli Liu, tel. +86 10 62836160, fax +86 10 82596146, e-mail: [email protected] © 2014 John Wiley & Sons Ltd not taken up by the target system tend to mobilize, causing serious environmental consequences as the N cascades into the atmosphere, terrestrial and aquatic systems (Sutton et al., 2011; De Vries et al., 2013; Fowler et al., 2013). In most regions, about 20–70% of the fertilizer is lost to the environment as dissolved N, greenhouse gases (GHG), for example, N2O, or other air pollutants, for example, NH3 and NO (Smil, 1999; Bouwman et al., 2009; Sutton et al., 2011). The massive release of the excess Nr greatly disturbs the natural biogeochemical cycle of N, resulting in severe environmental problems in water, air and soil (Davidson et al., 2012; De Vries et al., 2013; Fowler et al., 2013). Taking China as an 1249 1250 C . Q I A O et al. example, N fertilizer contributed 66% of the country’s N2O emission, caused severe eutrophication of 62% of total freshwater areas (Ju et al., 2004; Le et al., 2010) and led to atmospheric N deposition as high as 89– 104 kg ha1 yr1 in intensive agricultural regions (Ju et al., 2009). Similarly, in Europe, anthropogenic nitrogen contributed to 10% of the global N2O emission and 80% of European fresh waters exceed the level of 1.5 mg N L1, which is the threshold for high risk of biodiversity loss (Sutton et al., 2011). Taken together, the environmental problems caused by anthropogenic N threaten global sustainability (Rockstr€ om et al., 2009; De Vries et al., 2013; Fowler et al., 2013). Maintaining food production while reducing the detrimental effects of anthropogenic N application is an urgent priority for global food security and environmental sustainability (Giller et al., 2004; Erisman et al., 2011; Davidson et al., 2012). When N fertilizers are applied, usually as urea or anhydrous ammonia, a microbial process called nitrification converted most N fertilizers into the highly mobile NO 3 within 2–3 weeks (Huber et al., 1977), which causes low retention in the target system. Most NO 3 losses occur through leaching or denitrification before plants can utilize it; thus, the system has low nitrogen use efficiency (NUE). Nitrification inhibitors (NIs) are a group of chemical compounds that suppress the first step of nitrification by inhibiting Nitrosomonas spp. bacteria that oxidize NHþ 4 to nitrite (NO2 ), and therefore delay the nitrification process (Huber et al., 1977; Zerulla et al., 2001). Inhibiting nitrification through use of NI can encourage the retention of soil N as the less mobile NHþ 4 form, which could significantly reduce leaching loss of NO 3 from soil. Lower soil NO3 concentration resulting from the inhibition of nitrification also reduces substrate availability for denitrification and decreases N2O emissions (Akiyama et al., 2010). NIs are recommended by IPCC as a potential mitigation option for cutting agricultural N2O emissions (IPCC, 2014). However, inappropriate use of NI can cause adverse impacts. NIs are often applied to pastures to reduce GHG emission and N leaching. NIs degrade over time, but can be a concern if livestock are fed with hay containing NI residues, as happened in New Zealand in 2012. Trace amounts of dicyandiamide (DCD) were found in some dairy products, raising great concerns about the safe use of NI in feedlot areas (MPI, 2013). Hundreds of field and laboratory studies have investigated how inhibiting nitrification through use of NIs affected N cycling and plant productivity since the 1960s. Although the effects varied depending on factors such as soil texture, physicochemical characters of NIs and plant species, appropriate applications of NI with N fertilizer often increased plant NUE and alleviate environmental damages caused by N fertilizers (Huber et al., 1977; Giller et al., 2004). To better understand how inhibiting nitrification by NI application affects the environmental impacts caused by N fertilizers, we conducted a comprehensive meta-analysis including soil acidification, N leaching, air pollutant emission, GHG emission and plant productivity. We also assessed whether the responses of those variables will be altered by NI forms, ecosystem types, fertilizer types and soil texture. In addition, we performed a cost–benefit analysis (CBA) based on ecosystem services following studies by Compton et al. (2011) and Van Grinsven et al. (2013) and quantified the overall impacts of NI application on ecosystem services in agricultural systems by monetary values. Materials and methods Data selection Peer-reviewed NI studies were searched using Web of Science (1984–2013). The impacts of NI were assessed by the following six categories with 19 variables, including (i) soil chemistry: soil pH, and the concentrations of extractable NHþ 4 , extractable NO3 and dissolved inorganic N (DIN); (ii) N leaching: the leaching of NHþ 4 , NO3 and DIN; (iii) GHG emission: the emissions of N2O and CH4, and the uptake of CH4; (iv) air pollutant emission: the emissions of NO and NH3; (v) crop productivity: the productivity of grain, straw, vegetable and pasture hay; and (vi) NUE of plants: N concentration in plant tissue, plant N uptake, which was defined as the total N content in the harvested aboveground biomass per unit area, and N recovery (RN), which was defined as RN = (PN-P N0)/N, where PN was plant N uptake (kg N/ha) in N fertilized plots , PN0 was plant N uptake (kg N/ha) in control plots, N was the N fertilization rate (kg N/ha). The control and treatment means of the 19 assessed variables, including NI forms [DCD, 3, 4-Dimethylpyrazole phosphate (DMPP), nitrapyrin, Ca-carbide and organic NI], ecosystem types (agriculture and pasture), fertilizer types (synthetic N fertilizer, manure and the mixture of the both) and soil texture classes (sand, loam, clay and silt), were extracted from the available studies. Where data were presented graphically, figures were digitized using software ENGAUGE DIGITIZER (Free Software Foundation, Inc., Boston, MA, USA). In addition, experimental location, latitude, plant species, NI and N manipulation level were also included in the dataset. Details of the dataset structure were described in Table S1 (See Supporting information). Meta-analysis In total, 62 peer-reviewed publications with 859 datasets across the world were selected for our analysis (Fig. S1 and © 2014 John Wiley & Sons Ltd, Global Change Biology, 21, 1249–1257 N I T R I F I C A T I O N I N H I B I T O R S F O R N I T R O G E N M A N A G E M E N T 1251 Table S1). For parameters of soil chemistry, including soil pH and concentrations of NHþ 4 , NO3 and DIN, we used the average value across a whole crop rotation. For variables showing great seasonal variability, such as the leaching loss of NHþ 4, NO 3 and DIN, or the emissions of N2O, CH4, NO and NH3, we only included studies which reported the cumulative values for a whole crop rotation, or the cumulative values for a whole crop rotation can be estimated from the reported data. If data of multiple crop rotations with same vegetation and NI treatments were reported in a study, data of the latest crop rotation were chosen because the meta-analysis model required independence between observations. Measurements were considered as independent observations if different NI forms, fertilizer types or addition levels were manipulated in the same experiment. The response ratio (RR) was used to measure the effect of NI application on the assessed variables (Hedges et al., 1999). For each variable, the natural log of RR is calculated by: ln RR = ln NþNI =X N ), where X N is the mean of the variable when N fer(X NþNI is the mean of the variable tilizer was applied alone, and X when N fertilizer was applied with NI. Among 859 field data entries in our dataset, only 266 reported standard errors/deviations of the experimental and control group responses. Adams et al. (1997) developed resampling methods for metaanalysis of ecological data. Different from conventional parametric methods, resampling methods do not require computation of standard errors of the individual effect size estimations. Studies that did not report standard errors/deviations therefore can be included in the analysis (Adams et al., 1997). We adopted the resampling methods and estimated the mean RR with bias-corrected 95% confidential intervals (CI) using bootstrap resampling procedure (Adams et al., 1997). For each variable, total heterogeneity among the categorical group (QT) was partitioned into within-group heterogeneity (QW) and between-group heterogeneity (QB), where QT, QW and QB have chi-square distributions. A significance of QB indicates that the RRs are significantly different between the levels of the categorical group. Publication bias was assessed with studies that reported standard errors/deviations. We used Egger’s regression test to quantify publication bias, which has higher power to detect the bias than the rank correlation method (Sterne et al., 2001). Meta-analysis was performed by the software METAWIN 2.1 (Sinauer Associates, Inc. Sunderland, MA, USA). The response of each variable was considered significant if the 95% CI of RR did not overlap 1. of FN among experimental conditions such as climate, soil type, cropping time, seasonality fertilization and fertilization method. To evaluate whether NI application is a feasible approach to reduce N’s environmental impacts and increase agricultural outputs, a CBA was conducted using a US maize farm as a case study. The net economic impact of NI was assessed by summing the cost/benefit associated with (i) NI’s impacts on ecosystem services, including DIN leaching, GHG emission and air pollutant emission; (ii) NI’s impacts on crop productivity; and (iii) the expense of purchasing NI. For each individual ecosystem service, NI’s impact on the economic value (M) of this service is estimated by the formula: M = N 9 (FN+NI FN) 9 P. Here, N is a fertilizer input rate that is 125 kg N ha1 yr1, which is the mean annual N fertilizer input rate in US cropland (FAO/IFA, 2001). FN+NI and FN are defined above. P is the monetary value of the related environmental impacts given by previous CBA studies (Jaynes et al., 2010; Kusiima & Powers, 2010; Compton et al., 2011), and the details were listed in Table 2. Results Soil chemistry Averaged across all studies, NI application raised soil pH by 0.23 pH units (95% CI: 0.17–0.34, Table S2). NI application significantly increased soil-extractable NHþ 4 concentration (41%, 27–60%) and decreased extractable NO 3 concentration (41%, 48% to 33%, Fig. 1, Table S2), although publication bias was detected for NO 3 concentration (Egger’s P-value = 0.05, Table S9). When data were examined based on the chemical nature of NIs, all NIs significantly increased NHþ 4 concentration, except for nitrapyrin. All NIs significantly decreased NO 3 concentration (Table S2). Averaged across all studies, NI application did not alter the concentration of DIN (Fig. 1, Table S2). However, when data were subdivided by NI types, DIN decreased in soil applied with organic NI, but was not affected by DCD, DMPP and other NIs (Table S2). DIN increased (20%, 9–32%) in pasture soils, but decreased (12%, 19% to 3%) in agricultural soils. Fertilizer types and soil texture also altered DIN’s response to NI addition (Table S2). Nitrogen loss factor and cost–benefit analysis N loss factor under conventional fertilizer practice (FN) is defined as the ratio of the amount of N fertilizer lost to environment to the amount of N fertilizer applied to soil. FN values were derived from the literatures (FAO/IFA, 2001; Bouwman et al., 2009; Liu et al., 2010). For the variable whose RR was significantly altered by NI application, its N loss factor under NI application (FN+NI) was estimated by multiplying FN with the corresponding RR calculated from this meta-analysis. Because of data limitation, we did not assess the heterogeneity © 2014 John Wiley & Sons Ltd, Global Change Biology, 21, 1249–1257 N leaching On average, NI reduced NO 3 leaching (47%, 59% to 32%) and total DIN leaching (48%, 56% to 38%) (Fig. 1, Table S3). The reduction in NO 3 and DIN leaching was significant for different types of ecosystem, fertilizer and soil texture (Table S3). For NHþ 4 leaching, the overall effect of NIs was not significant (Table S3). However, when data were subdivided in different 1252 C . Q I A O et al. responses among N fertilizer types were similar (Fig. 1, Table S5). Averaged across all studies, NI increased NH3 emission (20%, 7–33%, Fig. 1, Table S5), although publication bias was detected (Egger’s P-value < 0.01, Table S9). The responses of NH3 emission to NI addition varied for different NI forms, ecosystem types, fertilizer types and soil types. DCD increased NH3 emission (34%, 20–50%), but DMPP and Ca-carbide had no significant impact on it (Table S5). NH3 emission was significantly increased in agricultural soils (21%, 4–44%), but not in pasture soils. Fertilizer type and soil texture also affected the responses of NH3, emission to NI addition (Table S5). Crop productivity Fig. 1 Effects of nitrification inhibitors on soil chemistry, environment impacts, plant productivity and plant nitrogen use efficiency. ecosystem type, ammonium leaching increased in agricultural lands (25%, 17–34%), but decreased in pastures (33%, 54% to 4%). NI’s impacts on NHþ 4 leaching also differed among fertilizer types and soil texture (Table S3). Greenhouse gas emission from soil When averaged across all studies, NI application significantly decreased N2O emission (44%, 48% to 39%). The reduction in N2O emission induced by NI was significant for all types of ecosystem, fertilizer and soil texture (Table S4). NI had no significant impact on the emission and uptake of CH4 (Fig. 1, Table S4), and the responses of CH4 emission and uptake were similar under different experimental conditions, such as NI form, ecosystem type, fertilizer type and soil texture (Table S4). Air pollutant emission from soil Nitrification inhibitors application significantly decreased NO emission (24%, 38% to 8%), and the Across all studies, NI applications significantly increased productivity of grain (9%, 6–13%), straw (15%, 12–18%), vegetables (5%, 0–10%) and pasture hay (14%, 8–20%, Fig. 1, Table S6). Among the four most commonly studied crops, grain productivity of barley showed the greatest increase (17%, 7–29%), as compared to wheat (2%, 0–5%), maize (9%, 3–17%) and rice (7%, 3–12%, Fig. 1, Table S6). For all crop systems, DCD increased plant productivity except for vegetable. DMPP increases plant productivity as well, except for pasture hay (Table S6). In addition, the effects of NI on crop production varied among different soil texture classes (Table S6). Nitrogen use efficiency On average, NI increased plant N concentration (7%, 1–16%), N uptake (15%, 11–20%) and N recovery (58%, 34–93%, Fig. 1, Table S7), although publication bias was detected for N concentration (Egger’s P-value < 0.01, Table S9). For different NI forms, DMPP showed no significant effect on plant N concentration, and nitrapyrin showed no significant impact on N recovery. When data were divided into different ecosystem types, NI did not affect plant N concentration in pastures. In general, the responses of N concentration, N uptake and N recovery to NI application were not affected by fertilizer types and soil texture classes (Table S7). Nitrogen loss factor and cost–benefit analysis Compared to conventional N fertilizer practice, application of NI with N fertilizer increased N loss factor for gaseous emission from 0.156 to 0.178, but decreased N loss factor for DIN leaching from 0.154 to 0.081 (Table 1). Overall, NI application resulted in a decrease in total N loss factor from 0.310 to 0.259 (Table 1), © 2014 John Wiley & Sons Ltd, Global Change Biology, 21, 1249–1257 N I T R I F I C A T I O N I N H I B I T O R S F O R N I T R O G E N M A N A G E M E N T 1253 equivalent to a net reduction of the total N loss by 16.5% (Table 1). The CBA case study indicated that NI application brought a total environmental benefit of $25.16 ha1, mostly by reducing the impacts associated with DIN leaching and NO emission (Table 2). The increases in maize yield increased the revenue by $163.83 ha1 (Table 2). After deducting the cost of purchasing NI, NI application resulted in a net monetary benefit of $162.70 ha1 for a maize farm (Table 2). Based on the average maize yield (9.24 ton ha1) and its market price in 2012 ($197 ton1), the mean revenue for a maize farm in US is $1820 ha1. NI application will raise the revenue to $162.70 ha1, which is about 8.95% increase if the environmental benefit was accounted. Discussion Previously, it has been unclear whether the benefits of inhibiting nitrification through use of NI could offset or even outweigh the cost because there had not been a systematic analysis considering environmental effects, forms of NI and types of crop systems. Using meta-analysis and CBA, we conducted a comprehensive assessment on how NI alters N cycling and ecosystem services and evaluated whether NI application is an effective approach for improving N management. Soil acidification Soil acidification is one of the most common consequences of soil degradation caused by N overuse (Cui et al., 2013; Zhang et al., 2013). Our results indicated NI application increased soil pH by 0.23 units, suggesting that NI application could alleviate soil acidifi- cation. The increase in soil pH under NI application could be caused by the combination of the decelerated rate of nitrification and increased plant N use efficiency, as indicated by higher N uptake (Fig. 1, Table S7). Instead of loss through cation-depleting N leaching, most NO 3 released by nitrifiers was absorbed by roots, which released OH and balanced soil pH. Considering that many farm fields add lime to optimize pH for nutrient availability, the benefit of NI application could include a small reduction in lime additions to the fields. Nitrogen losses to environment The greatest impacts of anthropogenic N addition beyond the intended use are impairment of ecosystem services in adjacent and downstream environments related to soil, air and water pollution (De Vries et al., 2013; Van Grinsven et al., 2013). Although NI application did not change soil DIN concentration, it altered DIN composition by increas ing NHþ 4 concentration and reducing NO3 concentration in soil (Fig. 1). The change in inorganic N composition altered the pathways of N losses to environment and resulted in different impacts on water and air. Higher soil NHþ 4 concentration led to more NH3 volatilization, which may exaggerate NH3 pollution and result in more damage on human health (Compton et al., 2011). On the other hand, NI reduced the emission of NO and N2O by depressing nitrification. However, the decrease in the amount of N emitted as NO and N2O is less than the increase in N emitted as NH3 (Table 1). Overall, NI application increased gaseous N loss factor by 14.3% (Table 1). Nitrate is highly mobile in soil and comprises the major fraction of N leaching loss from agricultural sys- Table 1 N loss factors with and without nitrification inhibitors (NI) application in croplands. RR: response ratio; FN: N loss factor without NI; FN+NI: N loss factor with NI; Change in N loss: the percentage difference between N loss factor with and without NI. Positive and negative values indicated the increase and decrease, respectively, in N loss when N fertilizer was applied with NI Nitrogen loss RR FN FN+NI Change in N loss (%) NH3 emission N2O emission NO emission Dissolved inorganic N (DIN) leaching Total gaseous N loss‡ Total N loss§ 1.201 0.560 0.758 0.524 0.140* 0.010* 0.006* 0.154† 0.156 0.310 0.168 0.006 0.005 0.081 0.178 0.259 20.1 44.0 24.2 47.6 14.3 16.5 *The source of the data was FAO/IFA (2001). †Calculated by using N budget data from three global N synthesis studies: Smil (1999), Bouwman et al. (2009) and Liu et al. (2010). ‡The sum of N loss through NH3, N2O and NO emission. §The sum of N loss through NH3, N2O and NO emission and DIN leaching. © 2014 John Wiley & Sons Ltd, Global Change Biology, 21, 1249–1257 1254 C . Q I A O et al. Table 2 The cost/benefit analysis for a maize farm applying nitrification inhibitors (NI) with fertilizer rate of 125 kg N ha1 yr1. For change in N loss under NI, positive values indicate that NI increases N losses, and negative ones indicate N reduces N loss. For the monetary response, the positive numbers indicate the amount of the economic benefit, whereas the negative ones indicate the amount of the economic cost Variables Assessed impacts Cost (data source) NH3 emission The cost of human health damage The cost of climate change $1.30 kg1 N (Compton et al., 2011) $1.24 kg1 N (Kusiima & Powers, 2010) $23.00 kg1 N (Compton et al., 2011) $2.71 kg1 N (Jaynes et al., 2010) N2O emission NO emission The cost human health damage Dissolved inorganic The abatement cost of N leaching reducing N from agricultural drainage water Sum of the environmental impacts Change in N loss under NI (kg N1 ha1)* Monetary response ($ha1) 3.52 4.58 0.55 0.69 0.18 4.18 9.17 24.84 25.16 Changes in yield (ton ha1)# Monetary response ($ ha1) Variables Assessed impacts Unit price (data source) Maize production The benefit of increase in yield $197.00 ton1 (USDA, 2013) Variables Assessed impacts Unit price Application rate (kg ha1)† Monetary response ($ ha1) $1.75 kg1¶ 15.00 26.25 Dicyandiamide The cost of (DCD) purchasing DCD Sum of the monetary responses 0.83 163.83 162.70 *Changes in N loss under NI = 125 kg N ha1 9 (FN+NI FN). The positive value indicated an increase in N loss under NI, and the negative values indicated a decrease in N loss under NI. #The change in maize production was calculated by multiplying the mean maize production in US (9.24 ton ha1, USDA, 2012) with mean response ratio of maize yield (1.09) estimated by the current study (Table S6). †The recommended DCD application rate (15 kg ha1 yr1) was from Di & Cameron (2003). ¶The price of DCD was the mean of the market price of DCD from the website of Alibaba. tems (Smil, 1999). The decrease in soil NO 3 concentration under NI addition resulted in a significant decrease þ in NO 3 leaching. Although soil NH4 concentration increased under NI application, ammonium leaching was not changed (Fig. 1), probably because NHþ 4 is not as mobile as NO 3 . Compared to conventional N fertilizer practice, NI application reduced DIN loss factor by 47.6% (Table 1). The environmental impacts of NI application depend on whether NI could reduce the net amount of N release to the environment. Our estimation suggested that when NI was applied along with N fertilizers, the decrease in N leaching outweighed the increase in N gaseous loss. Overall, NI application decreased total N loss by 16.5% (Table 1), suggesting that NI application is an effective practice to reduce N lost to environment. However, it should be aware that many factors, such as climate, soil types, cropping time, seasonality fertilization and fertilization method, would affect the value of N loss factor (FN) and eventually alter the estimation for the amount of N loss. Because of data limitation, we could not assess how those factors will affect N loss factor under NI application. More studies are needed to address the uncertainties. Greenhouse gas emissions Agriculture accounts for 60% of N2O and 50% of CH4 global anthropogenic emission (IPCC, 2007). Synthetic N fertilizer is the largest contributor to N2O emission in agricultural systems and may alter emissions of CO2 and CH4 as well (Liu & Greaver, 2009). We found that NI application reduced N2O emission by 44%, which is slightly more than the 38% reported in a previous meta-analysis (Akiyama © 2014 John Wiley & Sons Ltd, Global Change Biology, 21, 1249–1257 N I T R I F I C A T I O N I N H I B I T O R S F O R N I T R O G E N M A N A G E M E N T 1255 et al., 2010). N2O is released during both nitrification and denitrification processes. NI application not only reduces N2O emission by inhibiting nitrification, it also reduces emission via denitrification by reducing soil NO 3 concentration. Our results also suggested that NIs had no significant effect on soil CH4 emission and uptake. However, the small number of field studies on CH4 emission and uptake limited the statistical power of the analysis. Plant productivity and NUE Previous studies have not identified a consistent effect of NI application on plant productivity. A meta-evaluation on those studies is needed to assess whether NI is an effective approach to improve crop yield. Until now, only Wolt (2004) used meta-analysis to assess NI research trials in Midwest USA and found that nitrapyrin increased maize production by 7%. In the current study, we assessed how NI application affected productivity in different plant systems. Although the responses varied among crops, NI application significantly increased the productivity of grain, straw, vegetable and hay (Fig. 1, Table S6). The increase in plant productivity was attributed to the higher NUE under NI application, as indicated by greater plant N uptake and N recovery rate (Fig. 1). Higher NUE not only enhanced plant productivity, but also partially reduced N’s environmental impact (Table 1). With more N being recovered by plant, less of the applied fertilizer was subjected to be lost to the environment. Nitrification inhibitors application caused a shift in the relative proportion of NHþ 4 and NO3 in soil, which may influence the plant response. Although plants can take up both NHþ 4 and NO3 , species may differ in their preference for the two chemical forms of N (Li et al., 2013). Such preference could affect species’ response to NI application. In the current study, we found that NI application induced less increase in the productivity of vegetable, compared to that of grain, straw and hay. This is probably due to the fact that vegetables strongly prefer þ NO 3 to NH4 (Huber et al., 1977; Li et al., 2013). In addition to enhancing NUE, NI application also increased N concentration in plant tissue by 8% (Fig. 1, Table S7). N concentration is an important indicator for food and forage quality as N concentration is highly correlated with the concentration of Fig. 2 Effects of nitrification inhibitors (NI) on plant productivity and other ecosystem services in N-amended ecosystems. Mean effects of NI on each variable were shown in parentheses. NI, nitrification inhibitor; DIN, dissolved inorganic nitrogen. © 2014 John Wiley & Sons Ltd, Global Change Biology, 21, 1249–1257 1256 C . Q I A O et al. proteins. Higher N concentration in plant tissues suggested that NI might further enhance the nutrition quality of crops. As discussed above, NI could increase plant productivity and reduce N’s environmental impacts. Still, the safety in use of NI should be assessed for different plant systems before large-scale application. While most studies focused on the application of NI in row crops, there are concerns about application to pastures and animal feed. The residence time of DCD on plant canopy is between 6 and 16 days depending on climate condition (Kim et al., 2012). Because dairy cows ate grass before the sprayed DCD was completely degraded, trace amounts of DCD were detected in some milk products produced in New Zealand in 2012. Although DCD is of very low toxicity and the concentration was lower than any level known to cause human health concern, it raised great public concerns and New Zealand Government suspended the use of DCD (MPI, 2013). Appropriate application procedures to avoid direct consumption of NI by farm animals could greatly reduce the risk of contamination. local trials are needed to examine the effectiveness and also the potential negative impacts of NI for particular crops and in feedlots. Appropriate use of NI could help reduce the cost of mitigating pollution and climate change. Implication for N policy development Cui S, Shi Y, Groffman PM, Schlesinger WH, Zhu YG (2013) Centennial-scale analysis of the creation and fate of reactive nitrogen in China (1910–2010). Proceedings of the National Academy of Sciences of the United States of America, 110, 2052–2057. Davidson E, David M, Galloway J et al. (2012) Excess nitrogen in the US environment: trends, risks, and solutions. Issues in Ecology, 15, 1–16. De Vries W, Kros J, Kroeze C, Seitzinger SP (2013) Assessing planetary and regional Our results indicated that inhibiting nitrification by NI application altered the pathways of the excess N entering the environment, decreased overall loss and increased the productivity of crops (Fig. 2). Although the changes in the environmental pathways of N released from the system induced both positive and negative environmental impacts, the overall budget estimation suggests that NI application alleviates water pollution and lowers the global warming potential induced by N fertilizer. Our CBA suggests that the monetary benefit associated with NI application outweighs the cost, resulting in an increase in revenue of $162.70 ha1, which is equivalent to an increase of 8.95% in financial gain for a maize farm. Abatements of air pollution, water eutrophication and global warming require significant investment of public dollars. Our analysis suggests that inhibiting nitrification by NI application is an effective approach to increase agricultural output while also alleviating the environmental damage caused by reactive N leaching and GHG production. NI did increase the amount of N released to the environment as NH3, meaning that NI application may exacerbate environmental costs in areas with concerns about NH3 emissions. To encourage the use of NI, governments could design effective policies, such as providing subsidies for farmers using NIs. On the other hand, before large-scale application, Acknowledgements This study was supported financially by Chinese National Key Development Program for Basic Research (2013CB956304, 2014CB954003) and National 1000 Young Talents Program. This work has been subjected to EPA review and approved for publication. Mention of trade names or commercial products does not constitute endorsement or recommendation for use. References Adams DC, Gurevitch J, Rosenberg MS (1997) Resampling tests for meta-analysis of ecological data. Ecology, 78, 1277–1283. Akiyama H, Yan X, Yagi K (2010) Evaluation of effectiveness of enhanced-efficiency fertilizers as mitigation options for N2O and NO emissions from agricultural soils: meta-analysis. Global Change Biology, 16, 1837–1846. Bouwman A, Beusen A, Billen G (2009) Human alteration of the global nitrogen and phosphorus soil balances for the period 1970–2050. Global Biogeochemical Cycles, 23, GB0A04. Compton JE, Harrison JA, Dennis RL et al. (2011) Ecosystem services altered by human changes in the nitrogen cycle: a new perspective for US decision making. Ecology Letters, 14, 804–815. nitrogen boundaries related to food security and adverse environmental impacts. Current Opinion in Environmental Sustainability, 5, 392–402. Di HJ, Cameron KC (2003) Mitigation of nitrous oxide emissions in spray-irrigated grazed grassland by treating the soil with dicyandiamide, a nitrification inhibitor. Soil Use and Management, 19, 284–290. Erisman JW, Galloway J, Seitzinger S, Bleeker A, Butterbach-Bahl K (2011) Reactive nitrogen in the environment and its effect on climate change. Current Opinion in Environmental Sustainability, 3, 281–290. FAO/IFA (2001) Global Estimates of Gaseous Emissions of NH3, NO and N2O from Agricultural Land. Food and Agriculture Organization/International Fertilizer Industry Association, Rome, Italy. Fowler D, Coyle M, Skiba U et al. (2013) The global nitrogen cycle in the twenty-first century. Philosophical Transactions of the Royal Society of London. Series B, Biological Sciences, 368, 20130164. Galloway JN, Townsend AR, Erisman JW et al. (2008) Transformation of the nitrogen cycle: recent trends, questions, and potential solutions. Science, 320, 889–892. Giller KE, Chalk P, Dobermann A et al. (2004) Emerging technologies to increase the efficiency of use of fertilizer nitrogen. In: Agriculture and the Nitrogen Cycle: Assessing the Impacts of Fertilizer Use on Food Production and the ment (eds Mosier A, Delgado J, Cochran V, Valentine D, Parton 35–52. Island Press, Washington, DC, USA. Hedges LV, Gurevitch J, Curtis PS (1999) The meta-analysis of response experimental ecology. Ecology, 80, 1150–1156. Huber D, Warren H, Nelson D, Tsai C (1977) Nitrification inhibitors: new food production. BioScience, 27, 523–529. EnvironW), pp. ratios in tools for IPCC (2007) Climate Change 2007: Mitigation of Climate Change: Working Group III Contribution to the Fourth Assessment Report of the Intergovernmental Panel on Climate Change. Cambridge University Press: Cambridge, UK and New York, NY, USA. IPCC (2014) Climate Change 2014: Mitigation of Climate Change. Contribution of Working Group III to the Fifth Assessment Report of the Intergovernmental Panel on Climate Change. Cambridge University Press, Cambridge, UK and New York, NY, USA. © 2014 John Wiley & Sons Ltd, Global Change Biology, 21, 1249–1257 N I T R I F I C A T I O N I N H I B I T O R S F O R N I T R O G E N M A N A G E M E N T 1257 Jaynes D, Thorp K, James D (2010) Potential water quality impact of drainage water management in the Midwest USA. American Society of Agricultural and Biological Wolt JD (2004) A meta-evaluation of nitrapyrin agronomic and environmental effectiveness with emphasis on corn production in the Midwestern USA. Nutrient Engineers Annual International Meeting. Ju XT, Liu XJ, Zhang FS, Roelcke M (2004) Nitrogen fertilization, soil nitrate accumulation, and policy recommendations in several agricultural regions of China. Ambio: A Journal of the Human Environment, 33, 300–305. Ju XT, Xing GX, Chen XP et al. (2009) Reducing environmental risk by improving N management in intensive Chinese agricultural systems. Proceedings of the National Academy of Sciences of the United States of America, 106, 3041–3046. Cycling In Agroecosystems, 69, 23–41. Zerulla W, Barth T, Dressel J et al. (2001) 3, 4-Dimethylpyrazole phosphate (DMPP) – a new nitrification inhibitor for agriculture and horticulture – an introduction. Biology and Fertility of Soils, 34, 79–84. Zhang W, Dou Z, He P et al. (2013) New technologies reduce greenhouse gas emissions from nitrogenous fertilizer in China. Proceedings of the National Academy of Sciences of the United States of America, 110, 8375–8380. Kim DG, Giltrap D, Saggar S, Palmada T, Berben P, Drysdale D (2012) Factors controlling disappearance of nitrification inhibitor, dicyandiamide (DCD) in a grazed pasture soil in Manawatu. Proceedings of the Workshop “Advanced Nutrient Management: Gains from the Past–Goals for the Future”. pp. 7–9. Kusiima JM, Powers SE (2010) Monetary value of the environmental and health externalities associated with production of ethanol from biomass feedstocks. Energy Policy, 38, 2785–2796. Le C, Zha Y, Li Y, Sun D, Lu H, Yin B (2010) Eutrophication of lake waters in China: cost, causes, and control. Environmental Management, 45, 662–668. Li SX, Wang ZH, Stewart BA (2013) Responses of crop plants to ammonium and nitrate N. In: Advances in Agronomy (ed Donald L), pp. 205–397. Academic Press, New York. Liu LL, Greaver TL (2009) A review of nitrogen enrichment effects on three biogenic GHGs: the CO2 sink may be largely offset by stimulated N2O and CH4 emission. Ecology Letters, 12, 1103–1117. Liu J, You L, Amini M, Obersteiner M, Herrero M, Zehnder AJ, Yang H (2010) A high-resolution assessment on global nitrogen flows in cropland. Proceedings of the National Academy of Sciences of the United States of America, 107, 8035–8040. MPI (2013) News: DCD Suspension Supported. Ministry for Primary Industries, Wellington, New Zealand. Available at: http://www.mpi.govt.nz/newsresources/news/dcd-suspension-supported (accessed 24 January 2013). Rockstr€ om J, Steffen W, Noone K et al. (2009) A safe operating space for humanity. Nature, 461, 472–475. Smil V (1999) Nitrogen in crop production: an account of global flows. Global Biogeochemical Cycles, 13, 647–662. Sterne JAC, Egger M, Smith GD (2001) Systematic reviews in health care – investigating and dealing with publication and other biases in meta-analysis. British Medical Journal, 323, 101–105. Sutton MA, Howard CM, Erisman JW et al. (2011) The European Nitrogen Assessment: Sources, Effects and Policy Perspectives. Cambridge University Press, Cambridge, UK. USDA (2012) Crop Production. The National Agricultural Statistics Service (NASS), Agricultural Statistics Board. United States Department of Agriculture, Washington, DC. USDA (2013) Feed Grains Data Yearbook Tables: Table 12-Corn: Cash Prices at Principal Markets. Economic Research Service, United States Department of Agriculture, Washington, DC. Available at: http://www.ers.usda.gov/datafiles/Feed_Grains_ Yearbook_Tables/Domestic_and_International_Prices/FGYearbookTable 12.xls (accessed 15 January 2013). Van Grinsven HJ, Holland M, Jacobsen BH, Klimont Z, Sutton MA, Jaap Willems W (2013) Costs and benefits of nitrogen for Europe and implications for mitigation. Environmental Science & Technology, 47, 3571–3579. Vitousek PM, Menge DNL, Reed SC, Cleveland CC (2013) Biological nitrogen fixation: rates, patterns and ecological controls in terrestrial ecosystems. Philosophical Transactions of the Royal Society of London. Series B, Biological Sciences, 368, 20130119. © 2014 John Wiley & Sons Ltd, Global Change Biology, 21, 1249–1257 Supporting Information Additional Supporting Information may be found in the online version of this article: Figure S1. Global distribution of study sites included in this synthesis. Red triangles represent agriculture sites and purple triangles represent pasture sites. Table S1. Publications and observations numbers of each variable for this meta-analysis. Table S2. Response ratios of variables for soil chemistry with 95% bootstrap CI. Table S3. Response ratios of variables for soil N leaching with 95% bootstrap CI. Table S4. Response ratios of variables for GHG emission with 95% bootstrap CI. Table S5. Response ratios of variables for air pollutant emission with 95% bootstrap CI. Table S6. Response ratios of variables for plant productivity with 95% bootstrap CI. Table S7. Response ratios of variables for plant nitrogen use efficiency with 95% bootstrap CI. In Table S2–S7, number of studies for each category contained is shown in following parentheses. Numbers highlighted by bolded font indicate that the 95% bootstrap CI does not overlap 1. Table S8. Effects of NI on between-group heterogeneity (Qb) for each assessed variable. Numbers highlighted with bolded font indicate that the response ratios are different among the subgroups. Table S9. Results of publication bias tested by Egger’s regression test. P-values < 0.05 indicated publication bias existed. Data S1. Source references of all datasets in current syntheses.