Nitrate N loss by leaching and surface runoff in agricultural land: A global issue (a review)

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Nitrate N loss by leaching and surface runoff in agricultural land: A global issue (a review) Zhao-Hui Wanga,b,*, Sheng-Xiu Lib a

State Key Laboratory of Crop Stress Biology in Arid Areas, Northwest A&F University, Yangling, China College of Natural Resources and Environment, Northwest A&F University, Yangling, China *Corresponding author: e-mail address: [email protected] b

Contents 1. Introduction 2. Nitrate losses by leaching and surface runoff 3. Factors affecting nitrate N leaching 3.1 Excessive N input is the cause of nitrate leaching and surface runoff 3.2 Water: The driving force for nitrate leaching 3.3 Soil type 3.4 Land-use systems 3.5 Plant species richness 3.6 Tillage systems 3.7 Rotation systems 4. Methods for controlling N leaching 4.1 Application of N rate based on soil N supplying capacity 4.2 Control of water 4.3 Application of organic manure 4.4 Use of nitrification inhibitors 4.5 Split N fertilization 4.6 Cover crops 4.7 Choice of suitable crop types 4.8 Plastic mulched ridge cultivation 5. Research needs in the future 5.1 Simultaneous investigation of nitrate N leaching and N2O emission 5.2 Balanced fertilization 5.3 Relation of nitrate N leaching and water loss by evapotranspiration 5.4 Further understanding the process of N cycles 5.5 Integrated researches in major ecological regions of different countries Acknowledgments References Further reading

Advances in Agronomy ISSN 0065-2113 https://doi.org/10.1016/bs.agron.2019.01.007

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2019 Elsevier Inc. All rights reserved.

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Abstract Agricultural systems are nitrogen (N) deficient throughout the world and N is the most important nutrient in realizing the maximum potential of the crop and sustainability of the environment. The production and utilization of the chemical N fertilizer (CNF) have played a great role in increasing crop yield and meeting the demand of population growth. However, with excessive input of chemical N fertilizers, nitrate N leaching and surface runoff seriously threat people health and pollute biological environments. Based on a large number of publications, this paper has comprehensively reviewed the importance of application of CNF, the problems caused by the excessive chemical N input, with attention mainly to the seriousness of nitrate N loss by leaching and runoff, factors affecting nitrate N leaching and runoff, methods for controlling nitrate leaching and runoff, and research needs in the future globally in details.

1. Introduction Modern agriculture currently feeds almost 7.0 billion people in the world. This is possible because global cereal production has doubled in the past 40 years and yields have been increasing mainly due to intensified crop management involving improved germplasm (Cassman, 1999) and greater inputs of nutrients through application of chemical fertilizers (Erisman et al., 2008). Following the importance of water, nutrients are major constrains affecting plant growth and agricultural production. Data aggregated at a worldwide level and over several decades have shown a strong link between agriculture production and nutrient input (Tilman et al., 2002). Of the nutrients deficient in soil, N is the major one. Agricultural systems are N deficient throughout the world, and thus N has played the most important role in realizing the maximum potential of the crop yield and sustainability of the environment (Li et al., 2007). On an historic timescale, improving N availability has been the main driver in improvements of crop yields (Sinclair and Rufty, 2012). With increasing land use and intensive crop production, rainfed and irrigated agriculture cannot be sustainable without N input. In the current food production scenario across major cropping systems of the world, crop yield is still limited more by availability of N and water resources, rather than by crop genetics (Sinclair and Rufty, 2012). Wheat, maize and rice are major crops in the world that have consumed 50% of the total N. In China, maize-wheat rotation is the main cropping system, and wheat and maize occupy 22–26% of the total food sowing area and produce 21–30% of total grain production (National

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Bureau of Statistics of China, 2007). In many countries, of the estimated NPK nutrients, more than70% is N (Rao, 1999) and N is often considered essential to keep pace with food demand (Wang et al., 2008) and crucial to agriculture production systems (Sadras and Lawson, 2013). N inputs to farmland include biological N2 fixation of tightly bound N2 in the atmosphere (Galloway et al., 2003), livestock manures, landscape transformation from forestry to formland (Binkley et al., 1995) and fossil fuel combustion (Erisman et al., 2003). The biological N fixation was once the major one and played extremely important role in agriculture. However, since entering the 20th century, the natural N cycle mainly through the N fixation by microorganisms was interrupted by large amounts of CNF input. Such an input has increased crop yield by 50%, and cannot be replaced by other sources in crop production (Lin et al., 2003). By transported and transformed within the environment, this input has considerably provided major source of reactive N (Erisman et al., 2003). Through N cascades (Galloway and Cowling, 2002), a huge quantity of atmospheric N has been traced to land surface each year (Weinhold et al., 2000). Entering into 1980s, CNF rate increased much exceeded grain yield, and the difference between fertilizer and crop yield becomes greater and greater (Ye and Rozelle, 1994) At present, intensive CNF has been employed as a dominant agricultural practice to ensure high crop yields (Yang et al., 2013) and is applied very often at higher doses than crop requirement. The over N fertilization has caused a growing concern on the potential environmental impact (Fageria and Baligar, 2005). Reduction of N fertilizer efficiency is realized as first problem by formers. With regard to the world average fertilizer application of 132 kg N ha
1 in 2005/2007 (Alexandratos and Bruinsma, 2012), many countries such as China, India, South Korea utilize about 2.5 times more fertilizer and are therefore classified as the highest fertilizer application per hectare countries. In most situations, the efficiency of N fertilizer is fairly low ranging from 30% to 50% of the applied fertilizer, and 20–70% is lost from soil-crop systems (Dawson et al., 2008). In India, wheat and rice crops consume 70% of N fertilizers and N recoveries are about 50% and 25–30%, respectively, for the two crops (Tandon, 1992). In China, such low recoveries are even more serious. According to the data of FAO (2005), the total N fertilizer production and amounts used in China have reached 22–25% of the world total, respectively, being the first country for production and utilization of CNF in agriculture, and the largest country for energyconsumption in N fertilizer production. As early as in 1999, the expenses

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of CNF input in China had reached 50% of the total (including seeds, fertilizer, pesticides, machinery and irrigation) (Lin et al., 1999). Influence of environmental consequences becomes more and more serious. The excessive input of CNF has caused low N efficiency, but the low N efficiency is not entirely free from contaminative risk of nitrate. N input in natural and modified ecosystems over the world has several environmental consequences: the migration of N oxides (NOX) to water affects drinking water quality and results in eutrophication or more precisely hypertrophication (the enrichment of a water body with nutrients) of water bodies; the migration of ammonia and N oxides from land to the atmosphere could settle into land and sea and influence the function of forest ecological system and further intensify water bodies by eutrophication that damages the reverine environments (Vitousek et al., 1997); the increased concentration of N oxides leads to the formation of smog; the enhanced leaching of NO3
causes loss of other soil nutrients like Ca and K; the increased emissions of the greenhouse gas, nitrous oxide (N2O), from agricultural soils further strengthens the global warming effect (Vitousek et al., 2009); the acidification of soils and water bodies such as streams and lakes leads to the decline of biomass productivity and biodiversity; the increased rate of N input in natural terrestrial ecosystems results in the environmental N loading that has been reported to alter the structure and functioning of different ecosystems around the world (Phoenix et al., 2006), which in turn will affect the ecosystem services to the human beings in the coming decades. For such reasons, people all over the world have perturbed the changes. Decline of the biodiversity has been reported more and more increased. CNF additions have caused biodiversity decline and adverse effect on interactions among the closely associated animals and microbes in terrestrial ecosystems (Phoenix et al., 2006), Aggrandizement of N deposition has been noted gradually by researching workers. The increases of atmospheric reactive N will come down to the lands by wet or dry deposition, and this will in turn enhance environmental N loading as NHx (NH3 and NH4) and NOx (NO + NO2) in the earth ecosystems. Such N inputs from atmosphere vary widely from one ecosystem to another, and can balance or exceed nutrient losses at some sites while are unable to do so at other sites. The rate of N addition to a site is likely to be a crucial determinant of whether an ecosystem would be sustainable in the long-term or whether the soil or the vegetation be changed. However, as a whole, any additional N inputs through deposition will greatly impact all the various ecosystems and would accelerate N losses.

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This phenomenon is now considered as a new challenge for the environmentalists to quantify and understand the global threat of N deposition. Most of environmental N loading impacts were reported earlier in the European and North American ecosystems (Phoenix et al., 2006). Developing countries are emerging as the major emitters of such N because of increasing population and rapid industrialization. It is estimated that by 2020 these regions in the developing countries will account for more than half of the global anthropogenic N fixation (Singh and Tripathi, 2000). Nitrous oxide (N2O) emission is intensified. Large amounts of nitrate N accumulated in soil not only lead to nitrate leaching to groundwater, and drive surface water eutrophication from agricultural systems, but also intensify the gaseous evolution of N2O, one of the most important nonCO2 greenhouse gases with a global warming potential 298 times higher (100-year time horizon) than that of carbon dioxide (CO2) and a dominant driver or substance for the stratospheric ozone depleting (Ravishankara et al., 2009). Thus, nitrate N has been regarded as an indirect emission source for atmospheric N2O (Deng et al., 2011) and CNF input has been identified as the major source of soil N2O emissions (Bouwman et al., 2002). Of the detrimental effects from excessive CNF application, nitrate loss by leaching and runoff might be the most serious issues in the world, which have caused multiple problems. For this reason, we will discuss the two aspects in more detail in the following sections.

2. Nitrate losses by leaching and surface runoff Leaching, surface runoff and gaseous N emission (mainly NH3 and N2O) are three dominant pathways of agricultural soil N losses. The magnitude of the three N loss pathways differs greatly. On a global scale, CNFinduced N2O emission and nitrate leaching are estimated approximately 0.8–19% of N fertilizer input, respectively (Bouwman et al., 2002). Therefore, nitrate N loss by leaching and surface runoff is considered the dominant one with loss fluxes accounting for 6.7–19.0% of applied N in individual cropping years under normal agricultural practices in cultivated lands, followed by ammonia volatilization that certainly is an important N loss pathway in calcareous soils (Zhang et al., 2010). With the loss of natural forest and grassland ecosystems, all the agricultural practices, especially the long-term enhanced and accelerated N inputs can cause “N saturation” (Aber et al., 1989) and drive nonlinear changes of

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soil N retention, leading to dramatic increase of NO3 release (Perakis et al., 2005). Ammonium fertilizers and urea that are widely used in the world can be easily transformed into nitrate N, the final oxidized form of inorganic N, especially under aerobic soil conditions such as in dryland soils (Li et al., 2009). Therefore, agricultural fields are generally recognized as the main source for nitrate N production (Askegaard et al., 2005). Nitrate N is the main chemical form taken up by plants, and a large portion of it is absorbed by plants for metabolic processes. However, due to the low N fertilizer recovery (less than 50% in general and less than 33% in cereals) (Hardy and Havelka, 1975) and low N use efficiency of crops (Spiertz, 2010), the remaining portion becomes a potential source of contamination for both atmosphere and water resources, including surface and deep waters (Borin, 1997). In some soils, especially often occurring in light-textured ones (sandy soils), intensive use of CNF may lead to more nitrate leaching. If the longer drought periods interspersed with extreme rainfall events become more frequent due to global climate change, one has to expect even an increase of nitrate leaching. Due to its excessive production in field and rapid movement, nitrate is one of the most critical pollutants among the manifold agrochemicals in soil. Nitrate ions (NO3
) are soluble in water and can easily be removed away out of the soil mainly by two ways: one is leaching through the unsaturated zone to percolate into deeper horizons, and ultimately reach both aquifers (shallow and deep groundwater) and the other is the surface runoff from the top-soil through water flow to surface water resources. The two processes degrade water quality by increasing the risk of water contamination with nitrates (Powlson and Addiscott, 2005), resulting in eutrophication and non-potable water supplies (Barton and Colmer, 2006). The loss of nitrate N to aquifers and surface waters is an inevitable consequence of intensively managed agricultural systems with a large amount of CNF input (Nielsen et al., 2012). The intensively used agricultural areas are identified as hotspots of non-point pollution. Therefore, nitrate losses can occur either on-site or off-site, making its pollution a common problem in any areas. Surface waters facing a potential risk of nitrate contamination by runoff (Fenech et al., 2012) become a common way ( J
ego et al., 2012). It is estimated that of 532 rivers in China, 82% was polluted by nitrate N at different levels (Zhang et al., 2003); 92–88% of N entering to Yangtze and Yellow rivers annually originated from agriculture, and 50% of the river pollution was the contribution of CNF (Zhu et al., 2005). Via surface runoff,

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NO3
-N washed from the agricultural field could transfer and enter to streams, rivers, lakes, reservoirs, estuaries and coastal regions and finally to oceans (Matson et al., 1999). The accumulation of high levels of nitrate in such freshwater resources makes the resources deteriorated (Park et al., 2010), significantly raises N concentration in surface water, has a significant contribution to algal growth, and increases the mean annual algal abundance, declines water clarity, raise hypoxia, greatly impacts the riverine water quality (Boesch et al., 2001), and indirectly contributes to the atmosphere pollution with ammonia or N oxides (Bouwman et al., 2013). The adverse effect of the eutrophication for fish population has been found in the estuarine and coastal ecosystems (Bohlke et al., 2004). As a result, the high levels of nitrate N have made the water resources hostile for aquatic life as one of the major causes of eutrophication in freshwater bodies and hence lead to serious ecological problem process (Galloway and Cowling, 2002). In addition to surface water pollution by runoff, nitrate leaching making groundwater pollution becomes more serious. The nitrate leaching is not new to agricultural scientists, and the issue first came to the notice of scholars in European countries and was deemed to be an important approach to examine N loss from soils (Warrington, 1905). Since then, this problem has become more and more noted by workers. Undoubtedly, the increased amount of nitrate accumulation in cultivated lands and significant leaching to groundwater are associated with the excessive inputs of CNF (Ye et al., 2010) that in many agricultural areas has significantly enhanced nitrate accumulation in soil profile during dry periods. This accumulation would be a prerequisite ( Ju et al., 2009) for the downward movement of nitrate with water either from rainfall or irrigation through top soil to deeper layers below the rooting zone of crop plants during wet periods (RimskiKorsakov et al., 2004). Therefore, it has been inevitable to increase the possibilities of nitrate leaching, leading to decrease of N availability for crops (Delin and Stenberg, 2014). In intensive cropping production systems with both excessive CNF and irrigation in pursuit of high yields and great profits, the leaching problem is especially serious (Zhou and ButterbachBahl, 2014). The nitrate N loss by leaching and runoff has become a worldwide problem, either in developed or developing countries. In developed countries, this problem appeared much earlier. High nitrate concentrations in groundwater were previously found in Europe, the United States (Luo et al., 2008) and northeastern Australia (Keating et al., 1996). The European countries generally consider agriculture the main source of nitrate leaching to ground

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and surface waters (Askegaard et al., 2005), and the natural and modified ecosystems are world’s most saturated ecosystems. Continued high N deposition has resulted in increase of nitrate N loss by runoff in Europe and North America (Stoddard et al., 2001), the annual NO3
-N transported in subsurface flow from the watersheds ranged from 11.3 to 22.7 kg N ha
1 during the past 2–5 years at the North Appalachian Experimental Watershed (Owens et al., 2012), and ground water NO3
-N concentrations are commonly above the drinking water standard (10 mg L
1) in the Judith River Watershed of semiarid central Montana ( John et al., 2017). In Denmark, agriculture accounted for approximately 81% of the N load to Danish water courses and about 70% of the N inputs to the sea via water courses and atmospheric deposition during 1989–1996 (Hansen et al., 2000) although the nitrate leaching from agriculture in Denmark had been substantially reduced and N concentration reduction in surface water had been observed by then ( Jeppesen et al., 2011) through governmental regulations setting standards for limiting rates of manure and fertilizer application, use of catch crops (Kronvang et al., 2008) and encouraging additional measures such as organic farming (Askegaard et al., 2005). During 1985–2006, the average annual N leaching losses amounted to 15 kg ha
1 in Finland (Vuorenmaa et al., 2002), to 15–45 kg ha
1 in Sweden, to 20–43 kg ha
1 in Norway for cereal production (Vagstad, 2001), and to 36 kg ha
1 in Denmark (for arable farms averaged over soil types and fertility levels) according to a simulation model estimation (Knudsen et al., 2006). In developing countries, although such a problem occurred later, it gradually becomes a problem even more serious than the developed countries. For example, in some areas of Argentina, NO3
-N concentration in groundwater was reported above maximum levels accepted by Argentine food laws throughout the season for fertilized and irrigated cotton (Angella et al., 2002). In central China, a lysimeter study showed that nitrate leaching losses was as high as 63.3 kg N ha
1 when the N rate was higher than 370 kg ha
1, the NO3
-N concentration ranged from 3.7 to 43.1 mg N/L in the drainage water (Zhao et al., 2010), and the area-scaled N leaching losses amounted to 29.0–57.4 kg ha
1 in China for wheat and maize production, respectively (Zhou and Butterbach-Bahl, 2014). Due to huge demands for food, China has developed agriculture in both drylands and wet lands. The dryland including arid and semiarid areas mainly distributes in northern region, accounting for 52.5% of the acreage or the territory (Gao et al., 2003). With water deficit becoming more serious, dryland agriculture becomes the important agriculture area. Since 1980s, China

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has greatly increased application of CNF in agriculture (Yan et al., 2008), and nitrate leaching issues have been widely reported in southern China where a large amount of N fertilizers is used along with high precipitation and irrigation (Min et al., 2012). In view of the scarcity of irrigation and precipitation, the dryland areas have been regarded as a minor or no problem in nitrate leaching on cultivated lands. However, investigations have shown that the nitrate leaching in northern China is also serious, and becomes the main N loss pathway from agroecosystems (Wei et al., 2009). Although the rainfall is scarce in the areas, it concentrates in a few months of summer. During the rainy season, the torrential precipitation and professional flow have made the nitrate N moved down to accumulate in deep layer. Field trails have shown that the residual N left in soil increased after a crop harvest. In soil with a long-term application of CNF solely, a layer of 40–100 cm accumulates 73.5 kg N ha
1 (Chen et al., 2014). Application of N fertilizer of 180 kg N ha
1, the cumulative nitrate N in 0–300 cm soil layer, has reached as high as 1500 kg ha
1 after 23 years (Xue and Hao, 2009). Labeled 15N technique has shown the residual N proportion left in soil varied after wheat harvest from 38% to 45% of the annual N applied. Due to nitrate N leaching, contamination of groundwater becomes more and more significant. Based on 6-year observations, Zhou et al. (2012) showed that the estimated annual NO3
leaching fluxes varied significantly from a few kg N ha
1 to a relatively high amount with an average of 32.8 kg N ha
1. In an extreme rainfall event (211 mm within 36 h) following the drought period leached the accumulated NO3
equaling approx. 70% of annual cumulative NO3
in the soil profile (Zhou et al., 2012). Zhang et al. (1996) investigated nitrate contamination of groundwater and drinking water samples in 69 sites in northern China, and found that more than 50% of the samples exceeded the European maximum permissible nitrate concentration in groundwater (50 mg L
1), with the highest recorded being 300 mg L
1. A survey conducted in Beijing, Tianjin, Hebei, Shandong and Shaanxi showed that up to 45% of 600 groundwater samples exceeded the WHO drinking water standard (NO3
-N 10 mg L
1) ( Ju et al., 2006). Additionally, Zhao et al. (2007) examined nitrate contamination of 1139 groundwater samples in northern China and reported the average value of nitrate concentration being 11.9 mg L
1, with 34.1% exceeding the WHO standard. Currently, Asia, Europe and North America account for almost 90% of human generated reactive N. In the coming 50 years the developing countries are supposed to face increased N related problems due to dependence

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on CNF, growth of population densities and adoption of gasoline powered vehicles. So, if the N management measures in agricultural fields in coming decades are not taken into consideration, serious environmental consequence caused by the loss of NO3
to the surface and groundwater water bodies is expected to increase considerably. The leached nitrate reaching to surface and shallow groundwater tables (Keeney and Follett, 1991) has made the drinking water contaminated by nitrates exceeding the standard under agricultural fields in different parts of the world (Carneiro et al., 2012). Consumption of nitrate-rich water may be related to and increase the risk of methamoglobinamia (blue baby or child syndrome) that loses the oxygen carrying capacity of red blood corpuscles in babies and even leads to their death (Chambers et al., 2001). Other important healthy problems associated with NO3
toxicity in human beings are spontaneous abortions (Nolan, 1999), and different forms of cancer (Rivett et al., 2008) such as oral, colon, rectum and gastrointestinal cancer (Paul et al., 1999); Alzheimer’s disease (Tohgi et al., 1998); multiple sclerosis (Govannoni et al., 1997); non-Hodkins’s lymphoma (Michal, 1998). NO2
and NO3
ions under various clinical conditions cause diseases like hypertension, infection, renal and cardiac disease, and inflammatory diseases. Besides, several other diseases are caused in animals like pigs, rats, cattle, sheep, dogs, chickens and turkeys (Rao and Puutanna, 2000). Of the diseases as shown above, some have been refuted by more exact evidence (L’hirondel and L’hirondel, 2001), and some still need proving. However, in any case, too much nitrate in drinking water is believed to bring about dangers or damages to man and animals. For reduction and prevention of groundwater contamination by nitrate from agricultural sources, some European countries have established legislation to enforce N balance on a farm scale (Eichler and Schulz, 1998). The Nitrates Directive (91/676/EC) (Council of the European Communities, 1991) by the European Union (UE) is typical a law that designed to control N pollution by a maximum threshold of 50 mg NO3
L
1 or 10 mg NO3
-N L
1 in water. The European legislation limit for waters intended for human consumption and generally addresses concentration levels of contaminants in waters (European Union, 1998). Following this regulation, other organizations or countries also set up drinking water limits. For example, the U.S. Environmental Protection Agency (USEPA, 1990) set human drinking water at a critical value of 10 mg NO3
-N L
1. In Argentina, the national food law set the limit for human consumption at 11.3 mg NO3
-N L
1. A concentration maximum of 10 mg

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NO3
-N L
1 in drinking water was also recommended by the World Health Organization, while other countries such as South Korea, the official regulations are even less than 10 mg NO3
-N L
1 (Choi et al., 2007).

3. Factors affecting nitrate N leaching 3.1 Excessive N input is the cause of nitrate leaching and surface runoff Nitrate is the source for nitrate N leaching and runoff. Without nitrate in soil, there would be no such activities. However, soil is the media for crop production and nitrate is the major form of N nutrient for plant uptake. Without nitrate, the soil would lose its crop productivity, and thus there would be no foundation for agriculture. For this reason, nitrate is absolutely necessary in soil, but it should be in an adequate amount. Despite a number of interrelated factors controlling nitrate N accumulation and movements, such as mineralization of soil organic N (Thomsen et al., 2010) and large amount of animal excreta to arable lands in modern animal husbandry farms (Bryant et al., 2011), as a whole, the main cause for nitrate N leaching is the large inputs of CNF to soil by the farmers for seeking high crop yield. The high N loads have a low potential of utilization by cereal crops and have exceeded crop need for maximum yield (Porter et al., 1996). In addition, the substantial carryover effect of fertilizer N from one growing season to the next in cropped soils, especially under semiarid conditions (Corbeels et al., 1998) further intensifies the progressive accumulation of nitrates in the soil. As early as in 1998, Corbeels et al. (1998) had pointed out that the carryover effect of N fertilizer under rainfed Mediterranean conditions could be substantial. In the rainfed Mediterranean Vertisol, Herridge et al. (1998) observed that nitrate in the top 0.9 m of the soil at sowing varied between 19 and 158 kg N ha
1, the amounts reflecting site characteristics, fertilizer N rate and previous fertilizer N utilization. Lo´pez-Bellido et al. (2013) concluded that fluctuations in nitrate contents between years and a progressive increase in the soil profile was a function of CNF amounts. Any applied N unused by the crop progressively increases the N reserves in the soil, which can produce an over fertilization of the system and a reduced response of cereal to fertilizer. The continuous application of high rates of CNF can lead to a stable storage of residual nitrate N in the soil profile, and play an important role as N fertilization. The timing and placement of fertilizer N can impact N leaching by affecting the size of the soil nitrate N pool and the distribution of nitrate between rapid-flow and slow-flow

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drainage domains. Perego et al. (2012) measured the nitrate concentration in soil solution and leaching at a field scale at several sites in intensive agriculture lands and found that the whole set of measurement highlighted a high risk of leaching with concentration exceeding drinkability threshold of 50 mg L
1 of nitrate, and that the large N supply in some sites largely exceeded the crop demand, leading to high nitrate loss. Due mainly to both dramatic increases in CNF inputs and low increases in N uptake by crops, the residual soil N was doubled and a 42% increase in N loss through drainage residual soil N on a national basis of Canada (Yang et al., 2013), and an elevated nitrate concentrations in approximately half the wells in Australia (Thorburn et al., 2003). The serious contamination of groundwater has also been confirmed by broad surveys in the United States (Spalding and Exner, 1993), Europe (Chilton and Foster, 1991) and China (Zhang et al., 1996). In China, the traditional way for supplying nutrients to soil was to apply manure and night soil solely, and the amount applied to a unit of land was limited. Due to few N fertilizer plants setting up before 1980s, only a small amount of CNF was used in some places. As a result, there was almost no serious nitrate N accumulation in soil profile before that time. However, with application of a large amount of CNF since 1800s, nitrate N accumulation in soil profile was remarkably increased. Data from Yongshou County, Shaanxi Province, China showed that in 1989, over one-third fields in dryland areas had accumulated more than 100 kg ha
1 nitrate N in 0–100 cm depth; together with ammonium N, the total accumulative mineral N had outstripped 200 kg ha
1 in more than two-thirds of lands. According to IPCC (2006), the N fraction lost by leaching and runoff could be estimated using the factor 0.3 kg N per kg N applied. This default-leaching fraction is recommended for mineral fertilizer and animal waste applications, and is suggested for all countries, despite the differences between them with respect, for instance, to agricultural systems, soil types or climate. The methodological assumption that leaching losses are linearly related to N inputs is over-simplified for a complex N loss function that depends on the interactions between rainfall, soil type, cropping, rate, and timing of CNF or manure applications (Silgram et al., 2001). Carneiro et al. (2012) considered that from the mean values of N lost per kg N applied in treatments with more intensive use of CNF, the IPCC factor could reflect a realistic estimate of N losses in these fertilization systems. However, when organic residues were used in crop fertilization, except spilt applications of sewage sludge, the factor seems not true. Results showed that

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the proportion of the lost to applied N was very close to that measured with application of CNF only. With respect to results obtained in other treatments with organic residue incorporation, utilization of the IPCC factor (30%) has quite clearly overestimated the N leaching losses for application of cattle slurry, urban waste compost or for only spring soil incorporation of sewage sludge. Anyway, nitrate leaching and its pollution to groundwater has been attributed to excessive CNF application in many parts of the world under a broad spectrum of humid, semiarid and arid climatic conditions (Gustafson, 2012). The excessive N inputs have been also evaluated by the annual N surplus, defined as fertilization N supply exceeding the crop removal (kg N ha
1) or the difference between N fertilization and crop N removal (Grignani and Zavattaro, 2000) The N surplus to nitrate N leaching losses, year to year at each monitoring site, and found that the calculated surplus was 128–335 kg N ha
1 year
1 while the input ranged from 369 to 509 kg N ha
1 under similar cropping systems. There was a significant linear regression (P < 0.01) between N surplus and nitrate N leaching with a slope of 0.87, showing that the leaching appeared to be the main source of N loss for N exceeding the crop demands. Sacco et al. (2003) obtained the N surplus value of 40–320 kg N ha
1 for the main farm types; Mantovi et al. (2006) reported a mean annual distribution of 475 kg N ha
1 as pig slurry in cropping system based on silage maize and grain sorghum and winter wheat. The N surplus is different in different countries and regions of the same country. The relation of N supply to N loss by leaching is also proposed by several authors. Daud
en et al. (2004) revealed a relation between nitrate N leaching (NO3
-N kg ha
1 year
1) and total inorganic N applied, obtaining an R2 ¼ 0.99, and found a strict relation (R2 ¼ 0.99) between drainage nitrate N concentration (mg L
1) and total CNF applied (kg N ha
1), considering mineral fertilization and total ammonia N in manure (55% of total Kjeldahl N). Andraski et al. (2000) proposed a relationship between surplus of CNF applied and post-harvest soil NO3
-N concentration (1.3–1.5 m depth), obtaining a good relation (R2 ¼ 0.88). Sullivan and Cogger (2003) suggested assessment of N management by measuring the soil NO3
-N concentration in post-harvest in the upper 30 cm soil and categorized such NO3
-N concentration in three levels: less than 20 mg L
1; 20–45 mg L
1; higher than 45 mg L
1 with interpretation and suggestion for N management in each level. Perego et al. (2012) in evaluation of different relations from the literature showed that a significant linear correlation existed between leached

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NO3
-N (kg ha
1 year
1) and total inorganic N applied (N kg ha
1), between the post-harvest soil NO3
-N concentration (1.3–1.5 m depth) and N use efficiency, between NO3
-N drainage (soil NO3
-N concentration in mg L
1, scored after crop harvest at bottom layer) and total inorganic N applied (kg N ha
1) as well as N surplus.

3.2 Water: The driving force for nitrate leaching The nitrate leaching amount is the product of nitrate concentrations in soil solution and downward water fluxes, both of which together control nitrate leaching. Water is the most important factor for crop production. Agriculture is conducted either in wetlands including sub-humid and humid regions or drylands including arid and semiarid areas. In some countries, wet lands dominate agricultural production while in others drylands mainly performs agricultural activities. In the latter case ( Jalali, 2005), water shortage is the major stress constantly limiting crop production and excess irrigation, if possible, is usually used. Since occasional dry weather often occurs, excessive irrigation water is also used in the wetlands. Both the heavy rainfall and excessive irrigation have resulted in serious leaching of nitrate. Viewed from the global situation, water deficit is the most serious problem for crop production and due to constant or frequent water deficit occurrence, irrigation has been carried out worldwide even in Mediterranean countries such as Turkey and Syria. Control of NO3 leaching is difficult because its losses are often intermittent, and link with seasonal land management, irrigation practices and fertilizer applications and/or irregular rain events (Barton and Colmer, 2006). In most agricultural areas of the world, the yield potential of major cereal crops such as corn and wheat is commonly controlled and limited by water deficit and N availability. Soil water status is important in maintaining maximum corn yields, and maintaining optimal soil water is facilitated by high irrigation application in areas of light-textured soil (Derby et al., 2005). Therefore, irrigated agriculture is crucial for productivity of major crops grown in countries with extended and prolonged drought conditions that adversely impact productivity. For N fertilizer, water has two conflict functions: increase of fertilizer N recovery and its use efficiency in one hand and rise of N leaching risk on the other. Since water is the key factor limiting wheat production in northwest China, farmers have been taking summer fallow measure to store

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precipitation water in soil for wheat use in the next cycle (Shan, 2002). The summer fallow period (July to September) is the raining season and majority of rainfall occurs during the period. For instance, in the south of Loess Plateau, 60% of the average rainfall concentrates in the summer fallow period (Wang et al., 2009). Therefore, soil can store some water for the following wheat to use and this measure is very important for wheat production. However, due to significant increase of N fertilizer in dryland wheat ( Ju and Gu, 2014) and rainfall concentrating in the fallow period, leaching of the residual N left in soil after wheat harvest becomes more and more serious (Yuan et al., 2000). Results have shown that in Weibei dryland of Shaanxi Province of China, nitrate N accumulated in 0–200 cm layer has reached 59–284 kg ha
1 after continuous 2 years wheat plantation (Gao et al., 2005) while in south Shanxi reached 205 kg ha
1 in the same duration with 165 kgN ha
1 added each year (Li et al., 2011). Fan et al. (2010) claimed that when N rate increased from 90 kg to 180 kg ha
1, nitrate N accumulative in 0–400 cm varied from 460 kg to 1256kg ha
1 for several years application. The nitrate N leaching is closely related with precipitation or irrigation water amounts (Zhang et al., 2010). In dryland agricultural areas, precipitation is the only driven force for nitrate N leaching (Guo et al., 2003), and therefore affects nitrate N leaching depth. Peng et al. (1981) studied the seasonal changes of nitrate N in a loessial soil and found a significant accumulation and leaching of nitrate N during the summer fallow raining season; on average, each 10 mm precipitation could make nitrate N leached to 3–5 cm. With 364 mm precipitation during summer fallow period in south of the Loess Plateau, the downward movement of nitrate N could exceed 100 cm and each 10 mm precipitation had made nitrate N downward movement 2–4 cm (Dai et al., 2013). Zhang et al. (2010) found that in North China Plain, nitrate N in the fallow land gradually moved down with water, and when precipitation reached 556 mm, nitrate N would accumulate in 60–80 cm layer. Nitrate N leaching was closely correlated with N rate applied: when N rate was over 200 kg ha
1, the leached nitrate N from the residual N could reach more than 40 kg ha
1 (Wang et al., 2014). Xia et al. (2018) conducted nitrate N leaching experiments in three consecutive years from 2013 to 2015 during summer fallow period in Changwu and Yangling, Shaanxi Province of China, with different precipitation in fallow period. Soil samples in 0–200 cm layer were taken before and after summer fallow for determination of NO3
-N contents to evaluate the impact of different precipitation and N rates on NO3
-N leaching. Their results showed that in Yangling, addition of N significantly increased nitrate

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N accumulation in 0–200 cm layer after wheat harvest compared to control (without N application); when N rate was 240 kg ha
1, the accumulated nitrate N reached 366 kg ha
1, 5.6-fold higher than the control and also was significantly higher than application of 120 kg N. Determination of 26 farmers’ dryland sites in Changwu showed that after wheat harvest, nitrate N accumulated in 0–200 cm changed from 97 to 328 kg with an average of 193 kg ha
1. The peak depth of nitrate N down-movement was related with rainfall amount. The higher the amount, the deeper the nitrate N moved, and thus the velocity of nitrate N leaching accelerated (Chen et al., 2003). In contrast, when the precipitation is small, nitrate N leaching becomes negligible and in some case, even the nitrate N peak might move up due to evaporation of the top layer. Xia et al. (2018) further found a significant relationship between NO3
-N leaching and rainfall during the summer fallow. In Yangling, 2013 and 2015 were rainfall deficit years with precipitation of 220–288 mm, respectively, during summer fallow period, and thus there was no NO3
N leaching, and the accumulated peak of nitrate N in soil profile was slightly moved up while 2014 was a normal year with rainfall of 346 mm and even in such a year, a large amount of nitrate N leached to below 100 cm, and the peak of NO3
-N moved down to 140–160 cm depth, 60–80 cm deeper than that at the beginning of summer fallow regardless of application of N rate in 120 kg or in 240 kg ha
1. From the results we can conclude that in an abundant precipitation year, the risk of nitrate N leaching would be more serious. In Changwu, in 2013, rainfall was 296 mm during the fallow season, and NO3
-N peak depth was significantly moved to below 60 cm soil depth while in 2014, only 157 mm during the fallow period, the nitrate N peak in soil profile did not change after summer fallow. Therefore, rainfall is the key factor affecting NO3
-N leaching. The distance of nitrate N downward movement in soil by each unit of precipitation depended on locations or soil texture. In Yangling, each 10 mm precipitation made nitrate N leached from 1.4 in 2015 to 1.7 cm in 2014, while in Changwu with less clay particle in soil than that in Yangling, per 10 mm precipitation water made nitrate N leached to 1.8–3.7 cm in 2013. Although wheat root could penetrate to more than 200 cm, 92% root concentrates in 100 cm range, thus nitrate N leaching over 100 cm depth of soil is considered unavailable to plants (Dang et al., 2003). Therefore, reasonable reduction of N rate did not only decrease the risk of nitrate N leaching during summer fallow period but also affected crop yield.

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The nitrate N down-movement is also influenced by the accumulated peak of nitrate N in soil. When the peak locates in deep layer, water from precipitation is difficult to reach except that there is a large amount of continuous precipitation. Using 15N-tracer technique, Ju et al. (2009) showed that the peak depth of residual nitrate N in soil could affect its vertical movement. Under the small precipitation condition, water infiltrated depth is limited, and thus the deeper the peak depth of nitrate N, the less the nitrate N further move down. Liang et al. (2011) found that nitrate N in leachate was rapidly increased with N rate applied. However, the N rate did not influence the nitrate accumulate peak: with the same precipitation, the nitrate N downward movement peak was almost approaching the same depth, and there was no great difference in the leached depth due to N rate after summer fallow period. A similar result was also obtained later by Dai et al. (2013). Water also influences crop responses to N fertilizer. Experiments on dry Vertisols in southern Spain (Garrido-Lestache et al., 2004) showed that wheat yield increased with an N fertilizer application rate of up to 100 kg ha
1 during wet years but had little or no response to N fertilizer during dry years. Since water limitation substantially restricts total N uptake by crops (Haefele et al., 2008), supply of water through irrigation becomes the main factor affecting N uptake, translocation, distribution and accumulation in crops (Albina et al., 2013). Although various lines of evidence show that irrigation level is closely associated with plant N uptake, effects of water on crop production and N use efficiency is different under different conditions. Jin et al. (2009) found that N translocation amount from vegetative organs to and the N accumulation amount in wheat grains were changeable with irrigation times. Garabet et al. (1998) reported that wheat N uptake amount was lower by receiving 100% of full irrigation compared to that receiving 66% of full irrigation. In a Mediterranean environment, 100% of full irrigation resulted in N amount higher by 15 kg ha
1 than that 50% of full irrigation (Rossella et al., 2010). In Northern China, application of 192 kg N ha
1 was no significant difference in N use efficiency between that receiving irrigation once at jointing and twice at jointing and anthesis, but N fertilizer productivity was higher in the latter than in the former (Zang et al., 2012). In a semiarid environment of India, N use efficiency was higher in soil receiving 100% of full irrigation at 0–120 cm depth than that receiving only 30% of full irrigation (Pradhan et al., 2013). Cossani et al. (2012) found that water and N availabilities resulted in a wide range of N use efficiencies for bread wheat with 136 mm of irrigation water, whereas N-uptake efficiency increased by

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14% with 400 mm of irrigation water. It has been found that N concentration in dry matter was lower with decrease in irrigation (Gholamhoseini et al., 2013). Providing suitable amount of water (60–80 mm) at the jointing and flowering stages, enhancement of the uptake of available N of 76.6 kg kg
1 by wheat was found for optimum plant growth and development (Kibe et al., 2006). In a warm temperate semihumid monsoon environment, with an application of 240 kg N ha
1, the N uptake amount in grains under irrigation once (600–700 m3 ha
1) at jointing was higher than that under irrigation twice at jointing and anthesis. Likewise, N translocation amount under the single irrigation increased by 16.5 kg ha
1, compared to the double irrigations of wheat (Li, 2013). Liu et al. (2010) revealed that 15N-uptake in wheat grains with irrigation of 100 mm was higher than with 140 mm. A full soil irrigation regime, with the soil relative water content (SRWC) at the root layer of bread wheat raised to 100%, produced a higher grain yield (6.86 t ha
1) than one-third or two-thirds soil irrigation regimes in northern Syria (Karrou and Oweis, 2012). However, Karam et al. (2009) reported that a full soil irrigation regime with the soil relative water content to 100% in the 0–90 cm soil layers, decreased grain yield compared with a 50% irrigation regime in the central Bekaa Valley of Lebanon. Moreover, grain yield per unit N uptake was largely unchanged in Australian wheat varieties from 1958 to 2007, perhaps due to increase of N uptake that paralleled the increase in grain yield (Sadras and Lawson, 2013). The N use has been studied extensively in continuous-flood and limitedirrigation conditions in winter wheat (Krishna et al., 2013). In the North China Plain, the N use efficiency of wheat with 265 mm of applied water increased by 2.0 kg grain per kg N applied compared to that with 335 mm of applied water treatment. In the Mediterranean environment of Northern Syria, full irrigation was scheduled by monitoring soil water in the top 0.45–0.75 m soil layers. The root dry biomass under 66% of full irrigation increased more than that under 100% of full irrigation (Izzi et al., 2008). Several studies have reported on supplemental irrigation based on monitoring of water in some soil layers, i.e., 45–75 cm, 0–90 cm or the root zone (Karam et al., 2009; Karrou and Oweis, 2012) in different regions. Guo et al. (2014) found that when the soil relative water content in 0–40 cm layers was raised to 70% by supplemental irrigation at jointing and anthesis, the highest grain yields were recorded with a higher N use efficiency and lowest soil NO3
-N content and leaching. Therefore, optimizing irrigation regime is important and can improve N uptake and translocation in crops.

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In contrast, water is also the carrier and driving force for nitrate movement within soils. Nitrate distribution in soil is connected with the water movement through the entire soil profile and nitrate advanced as the water wetting front advanced through the soil profile (Nielsen et al., 1982). Water amount, doses of nitrate fertilizer applied, and the presence of residual nitrates in soil are three dominant factors to control nitrate N leaching. Optimizing N fertilization, irrigation scheduling and irrigation efficiency were shown to reduce N exports to drainage water (Quemada et al., 2013) and groundwater (Thorburn et al., 2003), and applying N at a rate less than optimal (Schroder et al., 1998) can reduce NO3 leaching. However, a high NO3 concentration in the root zone is one of the major concerns in extensively irrigated areas, and high rate of excess irrigation water is often used to control salinity of nitrate salt by deep percolation and to recharge groundwater aquifers beneath irrigated lands as one of water sources (Schaack et al., 1997). Water applied for frost prevention in some crops such as potato can cause nitrate N percolation or leaching. Nitrate dissolved in irrigation water can result in irrigation return flows (surface and subsurface water leaves the field following application of irrigation water) that have been recognized as the major diffuse or non-point pollution contributor to surface and ground water bodies (Arag€ u
es and Tanji, 2003), and high irrigation return flow can turn into aquifer recharge, and can predominantly affect the groundwater NO3 concentrations (Barros et al., 2012b). High NO3 concentrations have been measured in irrigation return flows during N fertilization to corn (Barros et al., 2012a). Isidoro et al. (2006) found that 75% of the total NO3
-N load was exported after the irrigation season in an irrigation district of Spain. The water-induced NO3 concentrations in seasonal patterns of irrigation return flows depend on N fertilization time (Ibrikci et al., 2015), irrigation scheduling and rainfall distribution. Different approaches such as experimental and tracer techniques and numerical modeling have been used to estimate irrigation return flow and nitrate leaching (Skhiri and Dechmi, 2012). Reduction of NO3
level in surface and ground water bodies is a prerequisite for sustainable management of natural and modified ecosystems. There is an ample potential to retain sufficient N in soil to meet the crop N demands and to lower the amount of NO3
-N percolating below the rooting zone of the crop by applying the best water and crop management practices. Well-established fertilizer and irrigation water management plans are critical to reduce NO3 pollution risks in any irrigated lands. For this purpose, agricultural management practices need to be better designed to

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synchronize the availability of NO3
and water with that of the crop N demand and optimization of the application of N fertilizer and water throughout the crop growing period is urgently needed (Galloway et al., 2008). Gained through irrigation or precipitation water, nitrate can mitigate the amount of N fertilizer required by crop, its accumulation in soil profile and its subsequent, potential leaching (Al-Jamal et al., 1997). Coles and Trudgill (1985) reported that NO3
-N can move rapidly downwards with the preferential flow of soil water through the structural pathways mainly observed in cracked clay soils and can reach deep to the drainage waters. It is believed that NO3
-N leaching is greater under wet climates and irrigated cropping systems where a considerable amount of water can percolate nitrate N below the root zone of the crop (Catchpoole, 1986). Crops generally could only recover 50% or less of the applied N (Bergstrom and Kirchmann, 2004). When leaching below 1.0 m soil depth, nitrate N is difficult for crop uptake and thus is the main pathway for N loss in a rotation system (Adrienn et al., 2012). Li and Li (2000) found that significant differences of nitrate leaching losses existed between years in semiarid districts due to different precipitation. Wang et al. (2010) studied nitrate leaching losses under three levels of irrigation water input and reported that water drainage and nitrate leaching losses dramatically increased as the irrigation rate increased. The positive correlation between nitrate leaching losses and water input was widely reported (Gheysari et al., 2009). Behera and Panda (2009) noted that although increased irrigation did not affect N uptake of winter wheat, it enhanced soil nitrate N leaching. The high risk of NO3
-N leaching was observed during the high summer rainfall period, the maize growth season, in the North China Plain even with low N fertilizer inputs (Liu et al., 2003). The increase in irrigation water can also lead to rise of ammonia volatilization and denitrifying N loss and thus decrease N use efficiency. The nitrate N leaching is related to the timing and amount of rainfall and irrigation scheduling (Klocke et al., 1996). Smika et al. (1977) measured NO3 leaching at 150 cm below the soil surface at a distance of 91, 182, and 273 m during the corn growing season of three fields in northeastern Colorado and found the average losses over 3 years ranging from 19 to 60 kg N ha
1 depending on the irrigation and N management. When the amount of water applied via irrigation does not meet the evapotranspiration demand of a crop, application of CNF based on fully irrigated conditions could induce an over application of N and, therefore, increase the potential of N losses to the groundwater (Tarkalson et al., 2006). Poch-Massegu´ et al.

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(2014) carried out experiments to assess the soil water and nitrate balance for a wide spectrum of crops and irrigation methods in the Western Mediterranean region with unevenly distributed intensive rainfall events in primarily irrigated drain-fed lands for agriculture. Results show that high percolation and nitrate leaching rates were observed in all crops. The intensive rainfall events plus high permanent soil water content had a great impact on the leaching of nitrate N resulting from soil organic matter mineralization to the aquifer along corn growth period after fallow. The nitrate leaching is also related with irrigation methods. Application of fertilizer through flood irrigation water has several shortcomings: uneven distribution of nutrients, high rate of nutrient loss and occurrence of crop injury (Playan and Faci, 1997). Under sprinkler irrigation, evaporation losses were higher due to a more frequent water application, and NO3 leaching was smaller compared to flood irrigation (Mack et al., 2005). Comparative studies report the advantages of drip irrigation compared to sprinkler or flood irrigation (Darwish et al., 2003; Mack et al., 2005). A variable deficit irrigation scheduling regime (Sexton et al., 1996) has been also proposed as an alternative to reduce nitrate leaching, but deficit irrigation decreased N uptake and increased final soil N compared to the full irrigation (Gheysari et al., 2009). For raising the N use efficiency, a method contracted as fertigation has been developed. This is to dissolve N fertilizer in water and apply the N-containing solute through sprinkler or mostly drip irrigation systems. Gheysari et al. (2009) determined the irrigation and N interaction on NO3 leaching for silage maize under fertigation via sprinkler irrigation and showed that NO3
-N leaching was dependent on both irrigation depth and N rate applied, and it increased in response to any additional N and/or water. Leaching of nitrate N during plant growing season could be controlled by implementing proper synchronized management of irrigation and fertigation, especially under sparse rainfall conditions. Asadi et al. (2002) studied the impact of fertigation via sprinkler irrigation on NO3 leaching and corn yield grown in field under four N levels (0, 100, 150, and 200 kg N ha
1) for an acid-sulfate soil in Thailand. They found that the NO3
-N leaching collected from ceramic suction cups installed at a depth of 60 cm, was 23 and 5.3 kg N ha
1 for the 200 kg N ha
1 treatment in the first and second years, respectively. In fertigation systems, drip irrigation is commonly used. Drip irrigation allows nutrients to be applied precisely and uniformly throughout the wetted root zone, and has obtained high fertilizer use efficiency (Hagin and Lowengart, 1996). However,

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fertigation does not always guarantee high recoveries of water and nutrients. Frequency of fertigation with NO3 can also induce NO3 leaching. Solute distribution in soils under drip irrigation systems has been extensively studied (Bar-Yosef, 1999), mainly concentrating on the distribution of inert salt ions, such as C1– and Br
in soils. Convection and diffusion are two main processes controlling the movement of salt ions in soil. Urea and ammonium fertilizers are commonly used in fertigation (Zhou et al., 2001). Unlike the inert ions, these fertilizers are active or potentially active solutes. When they are applied to the soil, several processes, such as enzymatic hydrolysis, oxidization, or fixation, may occur (Du et al., 2005). This makes tracing the movement and transformation of these fertilizers in soils during fertigation a formidable task. The transport and transformation of N fertilizers in soil are closely related to the physical, chemical, and biological properties of soils (Stevenson, 1982). Undoubtedly, the specific infiltration and distribution characteristics of soil water during drip irrigation have different effects on the movement and transformation of N fertilizers through the irrigation water. Clothier and Sauer (1988) modeled the spatial location of the transformation of urea to nitrate in soil under drip fertigation with a complex equation. With the gravimetric method, Li et al. (2003) found significant differences in the distribution of nitrate and ammonium in the soil after fertigation with ammonium nitrate. Ammonium ion is considered less mobile in soils since the cation exchange capacity of soil determines its movement in soils (Wang and Zhang, 2004). After drip fertigation, ammonium was primarily concentrated near the point source (about 2.5–7.5 cm from the emitter) (Li et al., 2003). Although soil colloids could adsorb NH4 + ions, high concentrations of NH4 + ions could overload the adsorption capacity of the soil, and thus could be easily leached into the deeper layers in the coarse-textured soil. The hydrogen-bonding may adsorb urea molecules to soil colloids, but high leaching rates of urea from the soil columns indicated that adsorption of urea by colloids was not strong. When urea was added to a soil, the urease enzyme hydrolyzes urea into NH4 + within several hours after application (Nkrumah et al., 1989) and this hydrolysis may occur during its transport through the soil profile (Zhou et al., 2006) by drip fertigation within 1–2 h after application (Clothier and Sauer, 1988). The produced ammonia might be lost by volatilization in the soil with pH above 7.0 (Zhang et al., 2004) or by ammonium fixation by 2:1 type minerals, such as vermiculite and illite that existed in the soil (Xiong and Li, 1987).

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3.3 Soil type Soil type, particularly soil texture, affects the movement and seepage of soil water and has significant influences on the loss of nitrate leaching (Cao et al., 2005). Highly permeable sandy soils with low water holding capacity and limited rooting depths could promote N leaching (Stenberg et al., 1999). Safadoust et al. (2014) reported that soil texture and cropping system affected the least limiting water range in relation with bulk density, clay and organic carbon contents. Maximum leaching losses of 65 kg N ha
1 were obtained in sandy textured soils planted with spring barley with application of 100 kg N ha
1 (Bergstr€ om and Johansson, 1991).

3.4 Land-use systems Leaching of nitrate from soil is mainly driven by land-use type, management (fertilization), land-use change, climate, and soil properties (Perego et al., 2012). Several studies have highlighted the relationship between land use and N leaching (Kvı´tek et al., 2009). Nitrate concentrations in drainage from grassland or forest are usually low (Addiscott, 2005). In the context of systems with rotating annual crops leaving the soil bare for part of the year means that soil N is vulnerable to leaching because no live plants are present to capture it. Grasses spread their roots in deep soil and allow the plants to capture nitrate before it can be leached (Franzluebbers et al., 2014). Moreover, the permanency of carbon flow through perennial vegetation to the soil under grassland enables a rapid N immobilization-mineralization turnover by soil microbes that prevents an accumulation of N as nitrate in soil and thus reduces its leaching risk (Premrov et al., 2012). As a result, even with higher N application rates than those used on maize, wheat and barley, grassland cover always produces both a lower level of nitrate leaching and a lower nitrate concentration in drainage water. This could be explained by the absence of soil disturbance and periods of bare soil during the grassland phase. In arable cropping systems, large quantities of N that remains in soil or is mineralized after harvest may be leached before the next crop is established (Sapkota et al., 2012). Therefore, increasing the proportion of grassland in cereal rotations would lead to a marked exhaustion of the nitrate in drainage water. Kunrath et al. (2015) studied the effect of sod-based rotations on reduction of nitrate leaching in a cereal cropping system in the central western France, and concluded that under such soil and climatic conditions, use of a pure cereal cropping system managed with sufficient N fertilization to

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achieve the target potential grain yield had caused an average nitrate concentration higher in drainage water than the 50 mg NO3
L
1 limit acceptable for drinking water; drainage periods were usually observed between October and April since the evaporative demand in summer was generally sufficient to restrict drainage to infrequent events of very intense rainfall at this time of year (Cuttle and Scholefield, 1995). The difference between land use systems was in that during the bare soil period, higher drainage occurred because of lower evapotranspiration than under permanent vegetation. As a result, introduction of mowed grassland sequences into the arable crop rotation significantly reduced the nitrate concentration of groundwater. Nevertheless, in regions where water availability is closely dependent on recharging aquifers in winter, there could be a trade-off between the benefits of introducing grassland into an arable cropping system in terms of groundwater quality and the possible disadvantages linked to reducing the volume of drainage water. Such a trade-off should be taken into particular account in terms of catchment management.

3.5 Plant species richness Plant species richness and functional composition of plant communities were shown to positively influence many ecosystem roles such as biomass production and the use of resources including N, the quantitatively most important plant nutrient (Allan et al., 2013). Plant-available nitrate concentrations in soil are an essential component for nitrate leaching, and nitrate leaching from soil was reported to decrease with increasing species richness (Leimer et al., 2014a,b). The species richness effect might originate from more exhaustive or complementary resource use by plant uptake of more diverse mixtures because of higher diversity in resource acquisition strategies, e.g., more variation in rooting depth or seasonal activity, irrespective of the functional composition of the plant communities (Niklaus et al., 2001). On the other hand, a decreasing percentage of legumes with increasing species richness might possibly be a reason for a significant species richness effect on NO3
-N leaching (Scherer-Lorenzen et al., 2003). “Complementarity effects” comprise processes like niche differentiation and facilitation that increase the performance of diverse mixtures. Plant functional composition and the presence of specific plant functional groups possibly affect nitrate concentrations even stronger than species richness (Scherer-Lorenzen et al., 2003). However, in the literature, results regarding plant diversity effects on the water cycle are inconsistent

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because the effects of species richness and functional group richness on the water cycle vary with the considered soil depth, temporal resolution, and meteorological conditions (Leimer et al., 2014a). In experimental grassland in Germany, Leimer et al. (2014a) found smaller downward water fluxes from the 0 to 0.3 m soil layer on plots containing grasses and larger downward water fluxes on plots containing legumes. Growing several species of crops together or sequentially may utilize nutrients more efficiently than monoculture if the different species exploit a larger soil volume or different parts of the profile (Francis, 1989). The efficient capture of nitrate from the topsoil requires a high rooting density. Given favorable soil conditions, winter wheat has one of the most rapidly growing and prolific root systems of all arable crops (Barraclough et al., 1991). By mowing and subsequent removal of the biomass, nutrients are removed from grassland ecosystems. In the transition phase after land-use change, the former land use can still affect ecosystem variables like nitrate leaching from the new system for several years (Oelmann et al., 2007b). Plant-available N from former fertilization decreases through plant uptake and leaching shortly after establishment of the new system (Leimer et al., 2014b). Later on, increasing organic matter concentrations, particularly in the more species-rich mixtures (Steinbeiss et al., 2008), could presumably result in enhanced ammonium release (Oelmann et al., 2011) that supplied additional substrate for nitrification, and subsequently, nitrate concentrations in soil solution increase again (Leimer et al., 2014b). Furthermore, plant diversity effects on nitrate leaching might change with time since last fertilizer application because of a progressive decline in fertilizer-derived plant available N. Dijkstra et al. (2007) reported that leaching of dissolved inorganic N from grassland monocultures and 16-species mixtures differed even more if inorganic N fertilizer was applied. Bingham and Biondini (2011) reported decreasing NO3
-N leaching with increasing functional group richness and explained it like the species richness effect, with complementary resource use because species richness and functional group richness were highly correlated. Leimer et al. (2015) also found a significant correlation (r ¼ 0.64, P < 0.001) between the species richness and functional group richness each other. In general, the presence of legumes tends to increase nitrate leaching because of their N2-fixing ability (Dijkstra et al., 2007) and also probably because of their positive effect on downward water fluxes (Leimer et al., 2014a). In contrast, the presence of grasses decreases downward water fluxes, particularly under wet soil conditions (Leimer et al., 2014a) and high NO3
-N concentrations in soil

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solution (Leimer et al., 2014b) and accordingly reduces NO3
-N leaching even more when the soil is usually moister likely because of a dense and extensive rooting system that allows for more exhaustive resource use (Scherer-Lorenzen et al., 2003). Numerous studies have found the reduction of components of nitrate leaching, nitrate concentrations in soil solution (Leimer et al., 2014b) and downward water fluxes (Leimer et al., 2014a) due to the presence of grasses. This is further supported by the finding that nitrate leaching decreased with increasing fine root biomass (Scherer-Lorenzen et al., 2003). Small herbs also have the decreasing effect on NO3
-N leaching due probably to the lower downward water fluxes (Leimer et al., 2014a). Increase of species richness during the vegetation period causes a stronger depletion of N pools (Oelmann et al., 2007b) and therefore a reduction of available N for leaching in species-rich mixtures. Shortly after land-use change from fertilized arable land to unfertilized grassland, the quantitative amount of NO3
-N leached from soil can probably be reduced if highly diverse plant mixtures without legumes are established. Christian and Riche (1998) reported a decrease of nitrate leaching by 95% (from 154 to 8 kg ha
1) from the first to the second year after conversion of previous arable land to unfertilized Miscanthus grassland. The pronounced decrease in NO3
-N leaching from the first to the second year after land-use change can be attributed to a reduction of available N in soil mainly originating from former fertilization, because of plant uptake and removal of N with mown biomass (Oelmann et al., 2007a). Then there was a subsequently slight increase in simulated NO3
-N leaching that coincides with a similar increase in soil organic matter concentrations (Steinbeiss et al., 2008) and with an enhanced ammonium release (Oelmann et al., 2011) allowing for a more pronounced nitrification. Marquard et al. (2009) reported that species richness effects on ecosystem functions strengthened with time because of increasing complementarity. However, there are different views on the function of the species and functional group richness on nitrate N leaching. Studies from Oelmann et al. (2007a) did not detect the effect of increasing species richness on nitrate leaching, the study of Hooper and Vitousek (1998) showed no such functional group richness effect, and Bingham and Biondini (2011) revealed that nitrate leaching increased with increasing biomass of C3 grasses in a fertilized C4-dominated temperate grassland ecosystem while decreasing nitrate leaching with increasing functional group richness. The controversial results might be related with the plant species themselves and climate variations. The climatically induced seasonal variations in the effects of specific

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functional groups could possibly explain the differing results for plant diversity effects on nitrate leaching (Scherer-Lorenzen et al., 2003). Arable land usually fertilized constitutes a nutrient input into the system and often causes increased nitrate leaching (Perego et al., 2012). Legume crop residue, particularly green manure, is an effective source of N. When released in synchrony with crop N demand, the crop residue N is a particularly desirable source of N because losses to the environment are minimized (Stute and Posner, 1995). The influence of crop residues on plant available N depends on how they affect the net mineralization of other soil N sources (Bremer and van Kessel, 1992).

3.6 Tillage systems Tillage and crop rotation with legume crops are two management practices that influence the N dynamics of soil–plant systems. Tillage affects on NO3
-N leaching (Strudley et al., 2008). Some studies have revealed the changes in the efficiency of available N under reduced tillage and no-tillage conditions (Campbell et al., 1993). Compared to plow-tillage, no-tillage generally increases hydraulic conductivities by preserving root or earthworm preferential-flow channels (Palmer et al., 2011), increases soil organic N due to reduced decomposition brought about by minimizing soil disturbance and protecting organic N within aggregates (Dolan et al., 2006; Zibilske and Bradford, 2007), slows the rate of residue decomposition by keeping residues on the surface where decomposition is slower than if mixed into the soil (Alvarez et al., 2008), and enhances soil water content (Blevins et al., 1971). Thus, tillage has multi-faceted effects on several soil processes that impact NO3
-N leaching, especially when tillage implements shortly before a season of high water-recharge. However, the effect of the tillage system on nitrate leaching remains controversial depending on the soil type, climate (Hansen and Djurhuus, 1997), and soil nitrate content (Randall and Iragavarapu, 1995). Different conclusions have been reached with regard to the effect of no tillage on nitrate leaching: higher (Tyler and Thomas, 1977), similar (Kitur et al., 1984), lower in no tillage than in conventional tillage with weed removal (Syswerda et al., 2012), or variable depending on climate conditions (Randall and Iragavarapu, 1995). This is because the effect of tillage on NO3
-N leaching will also interact with soil and weather conditions, and N management practices. The timing and amount of rainfall are two main factors affecting nitrate leaching, and tillage time can influence total

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and percentage of percolation and therefore can contribute to preferential flow. These factors and their interactions illustrate why reviews have concluded that tillage has dynamic and quite variable effects on NO3
-N leaching (Strudley et al., 2008). Taken into account the long-term effects of continuous treatment, nitrate contents were higher under conventional tillage than under no-tillage as reported by several studies (Lo´pez-Bellido et al., 2013; McConkey et al., 2002), perhaps due to a lower net N mineralization under no-tillage by slower decomposition and more N immobilization and nitrification differences. Lo´pez-Bellido et al. (1997) noted a more marked translocation in no till soil because it was favored by cracks formed in Vertisols during the drought period. Schlegel et al. (2005) showed that fallow systems can accumulate significant amounts of NO3
-N in the soil profile from mineralization during an extended fallow period.

3.7 Rotation systems Depending on year, tillage system and previous crops, the development of nitrate levels varies in soil over time. Nitrate amounts in soil are gradually differentiated with crop rotations (Lo´pez-Bellido et al., 2013). Burns (1980) noted that the effective maximum rooting depth for nitrate in faba beans may be less than half that of wheat, and therefore use less soil nitrate. The wheat–sunflower rotation has the lowest accumulation of soil nitrates, since sunflower is an excellent complement to wheat in terms of N utilization (Lo´pez-Bellido and Lo´pez-Bellido, 2001). Connor and Jones (1985) showed that the sunflower root density is high even at a depth of 120 cm under rainfed conditions, and can absorb nitrate N in the deeper soil layers. Thanks to this exhaustive exploitation of sunflower, the amount of nitrates in rotation with wheat remains virtually unchanged over time. Lo´pez-Bellido et al. (2013) observed that the content of nitrate N in Vertisols was not high in the surface horizons and the amount of soil nitrates accumulated at 0–30 cm was the smallest compared to other depth; with the rate of N fertilizer increased, major differences were observed between depths. In rotations with legumes, such differences between depths were observed for all rates of N (Lo´pez-Bellido et al., 2013).

4. Methods for controlling N leaching The nitrate leaching represents a resource loss and nitrate concentration presence in soil solution, groundwater and surface runoff can threaten

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drinking water quality (Oenema et al., 2005). Therefore, control of its leaching is of great significance. However, the control is not an easy job, because the seriousness of leaching depends on many factors such as mineral N excess, the hydrological regime, land use, soil type and climatic conditions, especially on the relation between uptake by plants and soil organisms, atmospheric N2 fixation, N mineralization (ammonification and nitrification), N deposition from the atmosphere, denitrification, and volatilization (Schimel and Bennett, 2004), Traditionally, agricultural practices to reduce NO3 losses are mainly designed to enhance and/or optimize soil C/N ratio for the proper micro floral development and complementary resource use in space and time by mixed cropping. These practices have been realized and are being carried out in different parts of the world. Improving N recovery is fundamental for lowering N losses to the environment and for improving economical returns to the producer. Fertilizer N costs have increased significantly since 2006, further underscoring the importance of efficient N management (Randall et al., 2008). The recovery of fertilizer N in global crop production is about 50% (Eickhout et al., 2006) and the rest may be lost to the ecosystem via various pathways or accumulated in soils. Despite the anticipated improvements in the N use efficiency, the total N loss is projected not to change much in the industrialized countries, but will increase strongly in the developing countries from 67 to 93 Tg N year
1, and thus it is expected to increase from 109 to 132 Tg N year
1 between 1995 and 2030 (Eickhout et al., 2006). For mitigating nitrate leaching risk in arable soils of higher risk regions, a range of management practices have been used, including proper soil management practices. The most effective management practices are those related to reducing interflow discharge or soil drainage while taking farm management response options to enhance crop N use efficiency in various ways, such as optimizing soil nutrient uptake by crop plants through increasing the time of soil N availability; retaining residues from previous crops and incorporating the residues into the soil with the help of varying tillage practices in agriculture systems (Kushwaha et al., 2000a,b); splitting N fertilizer applications (Sanchez-Martin et al., 2010); taking trails to synchronize the soil nutrient availability with that of crop N demand on field experiments by measuring the pools and fluxes of N in relation to varying soil resources using well established methodologies (Singh et al., 2007a,b); adopting different agricultural management practices proved to be most suitable to significantly reduce the high potential of mitigation options of nitrate leaching

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with better timing of N fertilizer applications in maintaining current crop yields (Kim et al., 2015); matching the N applied to crop demands, improving manure storage and land application; using catch crops; intercropping or rotations that can take up excess soil mineral N; using ensuring measures such as balanced application of nutrients, site-specific crop management, variable rate of N application, proper coordination of N and irrigation (Khind et al., 2005). All this can substantially control the leaching of nitrate N beyond crop rooting zone and enhance N fertilizer use efficiency in crop production. In addition, the nitrate discharge from agro-ecosystems can maximally be controlled by increasing crop complementarity in space and time for resource use (Rathke et al., 2006).

4.1 Application of N rate based on soil N supplying capacity The amount of the N applied to satisfy the N demand of the crop depends on factors such as the type of fertilizer, the timing of application and seasonal trends (Blankenau et al., 2002; Borghi, 2000; Huggins and Pan, 1993). To supply a correct amount of N at the time when the crop needs it and at the place where the crop can access it while minimizing losses to the environment (Schlegel et al., 2005) is the main goal for N fertilization. Consequently, the best strategy to establish the optimal rate of N fertilizer should be based specifically on the soil N supplying capacity (Li et al., 2009) to particular site, and enhanced synchronization of N availability with that of crop N demand. During a growing period, plants can use two sources of N from soil; the mineral N or residual N in soil that is as effective as N fertilizer and the mineralizable N that can be mineralized and absorbed during plant growing period (Li et al., 2009). However, recent studies show that the mineralizable N only had a small portion compared to the residual nitrate N. Using “balance method”(calculating the mineralizable N from crop uptake N plus mineral N (ammonium + nitrate N) left in soil profile after a crop harvest minus the initial mineral N presented before the crop seeding) (Cui et al., 2010), Miao et al. (2015) revealed that the accumulated nitrate N in the 0–100 cm layer had been about 165 kg ha
1 and the total amount of ammonium and nitrate N 181 kg ha
1 while the mineralizable N from organic N during wheat growth period was about 35 kg ha
1, much less than the mineral N, and had a minor role in contribution to wheat production in comparison with the mineral N. As a result, wheat was greatly responded to nitrate N cumulative in soil, but not to the mineralizable N.

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The residual soil nitrate buildup and consequent leaching to groundwater are the two key offshoots. Varvel and Peterson (1990) reported greater residual soil nitrate to 150 cm soil depth under continuous corn and grain sorghum systems with high N application rates of 180 kg ha
1. Cui et al. (2010) found that the continuous addition of high doses of inorganic N fertilizers had made a high soil nitrate N accumulation (172 kg N ha
1) in 90 cm soil depth in North China Plain and the extent of soil nitrate accumulation was so high that 55% farmers did not need to apply N fertilizers before sowing of wheat. Indiscriminate use of fertilizer N without considering contributions of residual soil nitrate N leads to low crop uptake and further growth of the residual pool. Therefore, the amount of residual N in a certain layer of the soil at sowing should be decided upon according to precipitation.

4.2 Control of water A proper coordination of N and irrigation can substantially control the leaching of nitrate N beyond crop rooting zone (Khind et al., 2005), and thus proper scheduling of irrigation, using sprinkler or drip irrigation with fertigation have become realized (Macdonald et al., 2005). Water table adjustment to crop rooting zone is also practiced to reduce NO3
discharge in water bodies by increased denitrification and to maximize plant uptake. Pattern of N use by crops and its consequent buildup in soil, apart from depending upon levels of N application, also depends on N source, water availability and their interaction. Lenka et al. (2013) undertook a study to evaluate the effect of different levels of water and N (organic and inorganic) on crop N recovery and soil nitrate accumulation in a maize–wheat cropping system grown for four consecutive cropping seasons in the Indo-Gangetic Plains of India. Results showed that crop N recovery and soil NO3
buildup were a function of soil water and N availability, and maximizing yield with increase in N dose from 120 to 180 kg ha
1 along with maximum number of recommended irrigation had reduced N recovery and contributed to a significant buildup of soil nitrate.

4.3 Application of organic manure The importance of organic sources of N is highlighted for sustaining crop productivity with a minimum environmental pollution. Soil organic matter has an extremely high water-holding capacity, about 10–1000 times higher

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than soil minerals (USDA, 2003), and thus can increase soil water-holding capacity, and reduce water loss by leakage and evaporation. In addition, soil organic matter can also promote the granulation of soil particles, and therefore reduce the water leakage in sand soil and the bonding power in clayey soil. Studies have shown soil nitrate levels to be unaffected by compost application but increased with CNF application (Hartl and Erhart, 2005). Nutrient-based application of N to corn resulted in greater residual soil nitrate to a depth of 120 cm under inorganic fertilizers than manure and compost treatments in dry years (Eghball, 2002). According to traditional viewpoints, use of organic manure may be beneficial for better crop yields and higher soil fertility, since organic fertilizer contains various nutrients that are slowly released, and thus can renew organic matter in soil, promote microorganism reproduction, alter the nematode community structure, improve soil physical and chemical properties, raise soil pH buffering capacity (Wen et al., 2015), and improve the quality of agricultural products. Lin et al. (2009) found that with application of organic fertilizer for 15 years, the yield became toward the same level with same mineral N input amount. Application of farmyard manure to agricultural soils, as an alternative to mineral fertilization, has made soil properties and fertility improved (Li et al., 2012). Both the quality and content of soil organic matter and microbial biomass are positively affected by organic N application in comparison with mineral N fertilization (Sradnick et al., 2013). Recently in China, use of organic fertilizer has been proposed to replace some CNF and considered it as a hot point for reduction of CNF and implementation of “green” yield increase. However, some research findings have indicated that unreasonable use of organic fertilizer can also increase the risk of inorganic N pollution (Macdonald et al., 2005). Water contamination from nitrate N can be directly related to the intensification of application of organic residues (Di and Cameron, 2002). Organic farming without sufficient soil cover causes nitrate leaching during summer when the accumulation of soil nitrate coincides with or is followed by higher precipitation. Maeda et al. (2003) reported the possibility of nitrate leaching from an organic farming field following excessive application of organic fertilizer. In many areas, efforts to develop improved management strategies have been made for the application of the organic residues to soil with attention particularly to better utilization of mineral N fertilizers. Results of studies on nitrate leaching losses from soils amended with organic residues or receiving applications of mineral fertilizers are often discordant (Diacono and

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Montemurro, 2010). Basso and Ritchie (2005) in a 6-year maize–alfalfa rotation conducted in southwest Michigan, United States, observed the highest amount of NO3
leaching in the manure treatment, followed by compost, inorganic N and control. In contrast, Mallory and Griffin (2007) in a 13-year experiment using soil collected from Presque Isle, Maine, United States, found that, despite similar ammonium (NH4 + ) inputs and rates of NH4 + uptake for manure and fertilizer N treatments, residual NO3
accumulation was lower in the manure treatment, because the N in manure became available more slowly than CNF. The incorporation in the soil of organic residues with high C/N ratios initiates mineral N immobilization, reducing the nutrient availability during plant development (Schomberg et al., 1994), and mainly removes the NO3
form (Sarrantonio, 2003). This process may contribute markedly to the reduction of nitrate leaching. Comparing NO3
and NH4 + losses by leaching with soil applications of pig slurry only or in combination with primary pulp mill sludge (C/N ¼ 80.3), Cabrita et al. (1996) found reductions of N losses with the presence of primary pulp mill sludge residue. The decrease in the C/N ratio could have initiated a reduction in the immobilization capacity and greater availability of N in soil to be leached (Cabral et al., 1998).

4.4 Use of nitrification inhibitors At present, for reducing nitrate N leaching, people often use nitrification inhibitors (NI) to control the activity of Nitrosomonas and thus to control the oxidation from NH4 + to NO3
, and nitrate N production (Ruser and Sehulz, 2015). A large number of laboratory and field experiments have shown that NI could decrease the nitrification velocity, N2O emission, and nitrate N concentration in runoff and seepage water. However, use of NI could make the ammonium N concentration in a higher level for a long time in soil and therefore increase the volatilization risk of ammonium N, particularly in high pH soil (Zhao et al., 2012). Qiao et al. (2015) reviewed 62 papers dealing with the use of NI in fields, and found NI could decrease 48% mineral N leaching, 44% N2O emission and 24% NO emission, but increased 20% volatilization of ammonium N. As a whole, NI shows positive effect on environments, and could reduce net N loss by 16.5%. In Mediterranean areas, the use of NI to control nitrate leaching has been successfully tested in irrigated crops (Dı´ez et al., 2010). In many countries, dicyandiamide (DCD) is the most widely used NI because it is economical, less volatile and relatively soluble in water (Zaman et al., 2008) compared to

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other commercially produced NI and has no lasting effect on the community composition of soil microbial populations but does reduce their abundance (Carneiro et al., 2010). In general, most of the research results agree on the beneficial effects of the use of NI for reducing N leaching, when incorporated with mineral fertilizers (Zaman et al., 2008) or applied with slurries (Vallejo et al., 2005).

4.5 Split N fertilization As already mentioned soil type, particularly soil texture, affects the movement and seepage of soil water and has significant influences on the loss of nitrate leaching (Cao et al., 2005). For such soils, split application of N fertilizer is useful. Lo´pez-Bellido et al. (2013) studied strategies of splitting and N timing in wheat crop to ensure an adequate amount of N in a rainfed Mediterranean system and suggested that maximizing N efficiency was an increasingly important objective in most crop management systems. Baethgen et al. (1995) studied the impacts on the yield of malting barley of differing levels of N fertilizer applied at sowing, mid-tillering and end of tillering and found that N fertilizer applied at the end of tillering gave the largest yields, but only if a sufficient amount of N was available at sowing to ensure crop establishment and initial tiller development. Similar empirical results have been obtained for wheat (McKenzie et al., 2006). These empirical results provide support for split N fertilization, but enable no more than tentative conclusions about its profitability. The economic desirability of split fertilization ultimately depends on the magnitude of additional labor and equipment costs in relation to the benefits from more efficient fertilizer use and the abatement of negative externalities.

4.6 Cover crops Previous studies have shown that nitrate leaching tended to be increased by irrigation, inorganic and organic N-fertilizer inputs in rainy season, and by the absence of a cover crop between two growing seasons. The amount of N leaching was found to be lower when a catch crop was grown during autumn and winter rainy season than when the soil was left bare. Therefore, establishment of a nutrient-scavenging cover or catch crop is an important strategy to reduce the risk of N leaching among the best management practices (Tosti et al., 2014). It is reported that winter cover crops decreased nitrate leaching in comparison with winter fallow (Constantin et al., 2010), since N was taken up when such a crop was grown during rainy seasons.

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The primary intention for using cover crops is to take up N from the soil, and thereby reduce the soil nitrate N content vulnerable to leaching during autumn and winter (Thorup-Kristensen et al., 2003), the raining season in European countries. Growing cover crops between two growing seasons with same or different crops during raining season can conserve nitrate N and hence reduce its leaching. The introduction of a mustard cover crop to replace a period of bare soil allows high ground cover during rainy periods, therefore maintaining nitrate N in the soil–plant system; afterward, when the cover crop residues are incorporated to the soil, the N absorbed by the cover crop becomes available for the following spring crop after mineralization (B
ecel et al., 2015). The potential of cover crops for reducing nitrate leaching have been demonstrated by numerous previous studies (Constantin et al., 2012). McCracken et al. (1994) reported that nitrate N loss was reduced from 20 kg N ha
1 year
1 under winter-fallow treatment to 1.2 kg N ha
1 year
1 under winter-cover crop treatment. The uptake of N by commonly used ryegrass species can reach up to 30–40 kg ha
1 (Thomsen and Hansen, 2014). It has indeed been demonstrated that including cover crops should reduce some of the risk of nitrate leaching under arable cropping systems ( Justes et al., 2012). Justes et al. (2012) and Ter Steege et al. (2001) showed that, based on experimental data and modeling simulations, introduction of cover crops into annual cropping systems could reduce nitrate leaching up to 50%. However, to be efficient, the design of cropping systems with cover crops needs to avoid the pre-emptive competition for water and nutrients in subsequent crops. In B
ecel et al. (2015) simulations, the effects of cover crops were particularly notable in reducing nitrate leaching. The combination of early sowing and cover crops in cropping system limited nitrate leaching more than for each scenario tested alone. Growing cover crops after harvesting the main crops has been identified as an even more useful strategy than reducing the application of fertilizer or no-till techniques (Constantin et al., 2010). Under sowing in the main crop in spring is particularly potential in Nordic countries, where climatic conditions after harvest of the main crops may reduce the development of a late-sown cover crop. The method for maximizing the time available for cover crop growth after harvesting the main crops can also avoid tillage in early autumn, which generally stimulates N mineralization (Aronsson, 2000). All the above results highlight the potential of cover crops in autumn/winter in reducing the annual N leaching in cereal cropping systems.

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Studies reported that use of under-sown cover crops considerably reduced nitrate loss (Lemola et al., 2000) and the content of soil nitrate N before winter (Lyngstad and Børresen, 1996). A disadvantage of undersown cover crops is competition with the main crops for nutrients, moisture and light, resulting in a few-percent loss of grain yield (K€ank€anen et al., 2001, 2003). The ability of cover crops to absorb mineral N from the soil profile is affected by their root growth rate and depth (Sapkota et al., 2012). Besides their effect on the reduction of N losses by leaching, cover crops can also provide many other ecosystem services such as an increase in the amount of soil organic matter, an improvement of soil structure, and suppression of diseases (Poeplau and Don, 2015). Once the cover crop is killed and incorporated into the soil, the N associated with the biomass gets mineralized and converted to NO3
compounds to be utilized by next cash crop. Cover crop not only retains the leachable NO3
-N but also maintains the soil in good physical health by promoting soil aggregation and increasing productivity due to providing additional residue and organic N (Sainju and Singh, 2001). However, the incorporation of straw, together with N-rich aboveground biomass of cover crops promoted immobilization of soil mineral N, restored soil organic matter and improved aeration (Talgre et al., 2011). The increased fertility, besides improving cash crop yields, might increase N losses through leaching or gaseous emissions, thereby stressing the importance of studying the effect of cover crop use and crop residue management on a long-term basis (Berntsen et al., 2006). In modeling the dynamics of soil water and mineral N, Plaza-Bonilla et al. (2015) found that the increase in the number of grain legumes in the rotations had led to higher N leaching given the higher amount of residual N found at soil depth after legumes harvest. This situation was reverted by the incorporation of cover crops during fallow periods for mitigating the leaching loss of N. The low N leaching losses due to dry conditions during the experimental period led to a risk of a pre-emptive competition for N when cover crops were used. However, the competition was avoided with the establishment of an early termination of cover crops and a slight increase in N fertilization in certain crops. In two lysimeter studies, an increase in N leaching was found when cropping peas or field bean as sole crops compared to barley, or to a pea-barley, or to field bean-barley intercropping (Mariotti et al., 2015). The decomposability of legume crop residues was higher than other crop residues given their low C:N ratio (Sa´nchez et al., 2004) that could

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accelerate the decomposition of native soil organic matter (Kuzyakov, 2010). The accelerative process could lead to an accumulation of mineral N in the soil susceptible to leaching. Grain legumes usually use less mineral N from deep soil layers given the superficial development of their root systems, and that can aggravate the loss of N as leaching in agricultural systems (Hauggaard-Nielsen et al., 2001). Due to these reasons, a higher amount of residual mineral N at soil depth (i.e., 60–120 cm) was found under spring and winter pea compared to the rest of cash crops. The adaptation of the grainlegume based rotations by the incorporation of cover crops helps to mitigate N loss and increase the recycling of N. Current cover crop management schemes (plant species, sowing date) might not be sufficient to avoid the potential future increase in N leaching.

4.7 Choice of suitable crop types Crop type and management typically affect N leaching losses. It has been observed that accumulation of NO3
-N in the soil profile and its leaching vary considerably with the type of crops and cropping system (Weed and Kanwar, 1996). Fields having deep rooted crops are generally found low in the amount of NO3
-N measured below the root zone because these crops can absorb NO3
-N from the deeper layers (Entz et al., 2000). N application efficiency was reported ranging from 30% to 50% for corn with average rooting depth of 120–150 cm (Martin et al., 1997), only 28% for onions with average rooting depth of 40–50 cm (Al-Jamal et al., 1997), and from 22% to 92% for Chile and cotton (Zhu et al., 2005). High NO3
-N leaching losses for onions were due to the large N fertilizer applications and a small rooting system (10–15 cm) during the early growing season (Halvorson et al., 2008), and therefore nitrate-N loadings to groundwater were greater for onion than alfalfa and Chile under a furrow irrigation system in arid New Mexico (Al-Jamal et al., 1997). Zavattaro et al. (2012) reported a strong reduction of N leaching in the case of maize-Italian ryegrass double cropping system. The choice of the most convenient strategy should be combined with an efficient schedule and application of fertilizers and manure handling according to the rotation requirements (Doltra et al., 2014).

4.8 Plastic mulched ridge cultivation Agricultural activities and their complex effects on nature conservation and the services that ecosystems deliver to humans remain controversial. To minimize the leaching risk of agrochemicals, precision agriculture was found

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to be a valuable tool (Di and Cameron, 2002). Zhang et al. (1996) stated that excessive fertilizer application should be prevented and more frequent, but smaller N applications during the rainy season with the additional use of slow-release fertilizer or controlled release fertilizer should help to maintain yield increase and minimize nitrate pollution of groundwater in northern China. Wallace (1994) proposed economic and environmental benefits by an adjusted fertilizer placement, an adapted timing of fertilizer application to the plant’s needs and an adapted leveling, draining and contouring of agricultural fields. The effect of ridge tillage has the potential to decrease NO3
leaching by isolating NO3
from the percolating water, especially if fertilizer is placed only in the upper part of the ridges (Waddell and Weil, 2006). Lament (1993) found an increased temperature in the ridge soil, which in turn induced earlier plant emergence for higher overall yield of several crop types. Plastic mulching was shown to be useful in terms of weed suppression and reducing evaporation loss. Locascio et al. (1985) observed that plastic mulch protected the fertilizer from infiltrating water and consequently enhanced the nutrient retention in the ridge soil and the nutrient use efficiency of crops. Accordingly, uncovered furrow positions are more prone to agrichemical leaching compared to ridge positions due to higher infiltration rates caused by the surface runoff from the ridges to the furrows (Leistra and Boesten, 2010). Vegetable production in plastic mulched ridge cultivation is practiced widespread over the region on dominating sandy soils with poor sorption characteristics (Kettering et al., 2012). Ruidisch et al. (2013) simulation showed that not only the plastic coverage but also the topography of the ridges potentially increased the nitrate availability in the root zone since surface runoff was channeled into the furrows and nitrate in the ridge soil was protected. Locascio et al. (1985) found that the plastic coverage had led to enhanced fertilizer retention underneath the ridges and protected the fertilizer from leaching. Fertilizer placement restricted to the ridges is a valuable tool to considerably reduce NO3
leaching losses compared to a broadcast fertilization. Waddell and Weil (2006) found that the fertilizer application in the upper portion of the ridge in corn cultivation had led to lower N leaching losses and higher yields. Clay et al. (1992) found that N placement in the ridge tops reduced N movement, while N movement in furrows increased due to the surface runoff from the ridges. Consequently, the local method of plastic mulched ridge cultivation in a flat terrain is a good step toward a sustainable management, which can be enhanced by additional fertilizer best management practices, when focusing solely on nitrate contamination of groundwater resources.

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Combining those management practices in a plastic mulched ridge cultivation system under monsoonal conditions would lead to economic benefits in terms of decreasing fertilizer inputs as well as ecological benefits by reducing substantially the risk of groundwater pollution. Results from Ruidisch et al. (2013) showed that ridge cultivation and plastic mulching of the ridges constituted a valuable tool to decrease nitrate leaching in a flat terrain where an excessive runoff from fields to the river network could not be expected. However, other studies show that plastic mulching in highland agriculture and vegetable production on slopes, especially during monsoon periods, have negative effects by substantially increasing surface runoff (Arnhold et al., 2013), which causes high soil erosion rates and supports transport of nutrients, particularly particle-bounded phosphorous, via surface runoff into water bodies (Park et al., 2010). Surface runoff in plastic mulched potato cultivation on slopes was found to increase up to 65%, whereas drainage water was reduced by 16% compared to ridge tillage without plastic coverage (Ruidisch et al., 2013). Moreover, water was observed to pond at the surface during monsoon events, when the infiltration capacity was exceeded, but percolated afterward through the soil matrix and contributed to groundwater. In hill side positions, ridge cultivation and plastic mulching can increase surface runoff, which supports the transport of agrochemicals via surface runoff directly into the rivers.

5. Research needs in the future Although large numbers of researches have been done on nitrate N leaching in croplands, and numerous discovers have been found for reduction of such a loss, there are still many problems that need studying in the future.

5.1 Simultaneous investigation of nitrate N leaching and N2O emission Management practices for reducing nitrate leaching may enhance N2O emissions and vice versa, leading to pollution swapping from NO3
to N2O. Irrigation management to reduce drainage and nitrate leaching may result in high topsoil NO3
contents and enhanced soil N2O emissions (Vallejo et al., 2005). Comparable adverse effects have also been reported for different tillage practices. While conventional tillage as compared to no-tillage may mitigate N2O emissions, this practice may on the other hand lead to increased N losses via nitrate leaching (Yao et al., 2010). All this needs

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a further studying comprehensively. However, so far, only a few studies are available for measuring N2O emissions and nitrate leaching simultaneously worldwide, and information about the link between nitrate leaching and N2O emissions from cropland is missing (Zhao et al., 2011). Zhu et al. (2009) reported that sites with high N leaching rates had low soil NO3
concentrations in the topsoil and vice versa. Therefore, simultaneously studying the relation of nitrate N to N2O emission is urgently needed.

5.2 Balanced fertilization Improvement of N recovery and its use efficiency is the key for reducing nitrate N leaching. The more N uptake by crop plants, the less N left in soil for leaching. For reaching such a purpose, balanced fertilization is one of the most important measures. Plants need various nutrients in a balanced status. If one nutrient is deficient, plants will not take other nutrients even if the other nutrients are sufficient enough. We have done an experiment in a P deficient field, and found that in such a field, the more the N fertilizer applied alone, the more the wheat yield reduced. In contrast, when phosphate fertilizer was added, wheat yield was almost doubled. Currently, most of workers have not paid attention to supplying some deficient nutrients for the balanced fertilization, but concentrated on N fertilizer itself. This idea and way should be changed, and field trials at different sites in different regions should be conducted to investigate the specific nutrient deficiency that previously limits crop plant growth and production and to supply such deficient nutrients for improvement of plant nutrition and crop production.

5.3 Relation of nitrate N leaching and water loss by evapotranspiration Nitrate N leaching is driven by water. When water enters into soil, there are different ways to lose, mainly by infiltration to deep layer, through which nitrate N is leaching, and by evapotranspiration (plant uptake and evaporation from soil surface). For reducing, or even eliminating the nitrate N leaching, the best management should be without water down-movement to deep layer. For achieving such a purpose, understanding the specific water holding capacity of soils and the evapotranspiration at different stages of different crops is extremely important. Based on these results, we can add water through irrigation to the root zone layer in such amount that can be consumed by evapotranspiration but not moved down to the deep layer. In such a manner of irrigation, nitrate N leaching could be totally controlled.

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5.4 Further understanding the process of N cycles Major task ahead among the environmental scientists is to understand the process of N cycling by using the principles of ecology and biogeochemistry at scales ranging from individual microbial processes to landscape level, which is greatly influenced by land-use change such as from forest to cropland and crop conservation management practices. The outcome of such studies will help us to minimize NO3
losses from agricultural ecosystems and understand how NO3
losses from agricultural land are attenuated by riparian buffers and affect the streams water chemistry and ultimately cause eutrophication of coastal regions (Tripathi, 2009).

5.5 Integrated researches in major ecological regions of different countries An integrated research in croplands of different ecological regions would be required worldwide to quantify the cascade effect and optimize N levels for different ecological systems, and to examine temporal variations in nitrification rate, NO3
availability and N associated with microbes so that we could further understand the origin and fate of NO3
generated by nitrification and its movement from cropland to streams using the modern techniques such as isotopic composition of N and O in NO3
. Using these data we can answer how the different ecological ecosystems affect the N cycling processes, how the agricultural conservation management practices affect the rate of nitrification and soil NO3
availability and microfloral developments, and whether the soil NO3
availability is synchronized with plant N demand, and to what extent these practices affect the stream water chemistry as well as how N retention and transformations affect the environmental at regional and global scales. As a result, we can find out the best crop management practices that had been proven the most productive by synchronizing soil N availability with that of crop N demand with least N discharge into the rivers.

Acknowledgments This work was part of the China Agricultural Research System (CARS–3), the National Key Research and Development Program of China (2018YFD0200400), and the projects supported by the National Natural Science Foundation of China (30971866, 30871596 and 30230230). The authors would like taking the opportunity to express their sincere thanks to the organization for its kindness in support of these programs and projects.

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Further reading Lal, R., 2004. Soil carbon sequestration to mitigate climate change. Geoderma 123, 1–22.