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Optimizing Nitrogen Fertilizer Application for Rice Production in the Taihu Lake Region, China
Pedosphere 22(1): 48–57, 2012 ISSN 1002-0160/CN 32-1315/P c 2012 Soil Science Society of China Published by Elsevier B.V. and Science Press
Optimizing Nitrogen Fertilizer Application for Rice Production in the Taihu Lake Region, China∗1 DENG Mei-Hua1,3 , SHI Xiao-Jun2 , TIAN Yu-Hua1 , YIN Bin1,∗2 , ZHANG Shao-Lin1 , ZHU Zhao-Liang1 and S. D. KIMURA3 1
State Key Laboratory of Soil and Sustainable Agriculture, Institute of Soil Science, Chinese Academy of Sciences, Nanjing 210008 (China) 2 College of Resources and Environments, Southwest University, Chongqing 310047 (China) 3 United Graduate School of Agriculture Science, Tokyo University of Agriculture and Technology, Tokyo 1838509 (Japan) (Received June 27, 2011; revised October 31, 2011)
ABSTRACT To determine the optimal amount of nitrogen (N) fertilizer for achieving a sustainable rice production at the Taihu Lake region of China, two-year on-farm field experiments were performed at four sites using various N application rates. The results showed that 22%–30% of the applied N was recovered in crop and 7%–31% in soils at the rates of 100–350 kg N ha−1 . Nitrogen losses increased with N application rates, from 44% of the applied fertilizer N at the rate of 100 kg N ha−1 to 69% of the N applied at 350 kg N ha−1 . Ammonia volatilization and apparent denitrification were the main pathways of N losses. The N application rate of 300 kg N ha−1 , which is commonly used by local farmers in the study region, was found to lead to a significant reduction in economic and environmental efficiency. Considering the cost for mitigating environmental pollution and the maximum net economic income, an application rate of 100–150 kg N ha−1 would be recommended. This recommended N application rate could greatly reduce N loss from 199 kg N ha−1 occurring at the N application rate of 300 kg N ha−1 to 80–110 kg N ha−1 , with the rice grain yield still reaching 7 300–8 300 kg DW ha−1 in the meantime. Key Words:
economic efficiency, environmental efficiency, N application rate, N loss, rice grain yield
Citation: Deng, M. H., Shi, X. J., Tian, Y. H., Yin, B., Zhang, S. L., Zhu, Z. L. and Kimura, S. D. 2012. Optimizing nitrogen fertilizer application for rice production in the Taihu Lake region, China. Pedosphere. 22(1): 48–57.
INTRODUCTION The use of nitrogen (N) fertilizer has changed global N cycle markedly and has been causing various negative environmental consequences such as eutrophication of surface water, global warming, and ozone layer depletion (Gruber and Galloway, 2008; Seitzinger, 2008). However, the total amount of N fertilizer in the world will undoubtedly continue to increase to meet the increasing demand for food (Mosier et al., 2001; Dobermann and Cassman, 2005). The mitigation of negative environmental influence of N fertilizer has become an important issue for sustainable agriculture development (Galloway et al., 2008). ∗1
In the Taihu Lake region, one of main food production regions in China, the intensive agricultural activities have contributed to a marked eutrophication of the surface water and deterioration of drinking water quality (Qin, 2009), and increased atmospheric deposition of N (Luo et al., 2007; Hayashi and Yan, 2010). The annual chemical fertilizer N input in the whole Taihu Lake region reached 450 kg N ha−1 in 2002, 2 times the Chinese national average in that year (Deng et al., 2007). In general, the rotations of irrigated summer rice and winter upland crops (wheat or rapeseed) are the typical crop system in this region, which accounts for about 80% of the arable land area (National Bureau of Statistics of China, 2003). To maximize rice yields,
Supported by the Knowledge Innovation Program of the Chinese Academy of Sciences (No. KZCX2-YW-440-1), the National Natural Science Foundation of China (Nos. 30390080 and 41071197), and the National High Technology Research and Development Program (863 Program) of China (No. 2006AA10Z418). ∗2 Corresponding author. E-mail: [email protected].
N FERTILIZER APPLICATION FOR RICE PRODUCTION
farmers often use excessive N fertilizer (Ju et al., 2009; Zhao et al., 2009). Many studies reported that rice production in the Taihu Lake region was a dominant nonpoint pollution source (Mosier et al., 2001; Roelcke et al., 2004; Ju et al., 2009). To ensure food security and an environmentally friendly rice fertilizer management system, many researchers have investigated the optimal N application rate for rice production. An optimum N fertilize application rate of 225–270 kg N ha−1 was suggested to achieve good rice productivity (Wang et al., 2003) and Zhu et al. (2010) recommended a rate of 200 kg N ha−1 for rice cultivation for the Taihu Lake region. However, we suspect that these recommended N application rates may be still too high, when considering already high ambient N deposition and severe N losses to the environment. According to the investigation at Hunan, Jiangsu, Zhejiang, and Guangdong in China, Peng et al. (2006) suggested that a good rice yield (ranging from 5 300 to 9 900 kg ha−1 ) can be achieved at 60–120 kg N ha−1 for irrigated rice production. However, the N application rate employed by farmers in the Taihu Lake region is still around 300 kg N ha−1 in a single rice season, corresponding to 500–600 kg N ha−1 year−1 under the double-cropping system (Ju et al., 2009; Zhao et al., 2009). This high N application rate was mainly attributed to the pursuit of maximum grain yields by policy-makers and farmers. For policy-makers, national food security is the main concern due to China’s huge population. As for farmers, the unintentionally excessive use of N fertilizer is aimed at achieving a high yield and consequently a high economic income. Hence, an optimized N application rate is required to balance the concern of farmers and policy-makers about high and stable grain yields and the need to protect the environment in the Taihu Lake region from severe N pollution, also taking into account the cost of N fertilizers. In this study, two-year on-farm field experiments were carried out at four sites using seven different N application rates and eight varieties of rice to investigate the impact of N fertilizer use on rice yield and economic and environmental efficiency of N in the rice system. The objectives of this study were to 1) determine the economic and environmental losses incurred on farmers’ rice fields in the Taihu Lake region and to 2) seek a new optimal N application rate and potential solutions to mitigate N pollution in rice systems in the Taihu Lake region.
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MATERIALS AND METHODS Study sites The experiments were conducted in 2003 and 2004 at Changshu City (31◦ 30 –31◦ 50 N, 120◦ 33 –121◦ 3 E) of Jiangsu Province in southeastern China, which is located in the heartland of the Taihu Lake region. This region has a typical subtropical humid climate. The annual average temperature is 15.5 ◦ C and the annual precipitation is 1 038 mm. The soil is classified as a Gleyi-stagnic Anthrosol according to the FAO soil taxonomy system. An irrigated summer rice and winter upland crop (wheat or rapeseed) double-crop rotation is the main cropping system in this region. Four farmers’ fields at Xinzhuang (31◦ 32 45 N, 120◦ 41 57 E), Baimao (31◦ 35 30 N, 120◦ 54 9 E), Dayi (31◦ 42 10 N, 120◦ 41 3 E), and Meili (31◦ 42 50 N, 120◦ 52 55 E) were chosen. Experiment 1: yield responses to N application rate All 4 farmers’ fields were employed to investigate the impact of N rates on rice productivity. The N application rates included 0, 100, 200, 300 (local farmers’ practice), and 350 kg N ha−1 in 2003 and 0, 100, 150, 200, 250, 300, and 350 kg N ha−1 in 2004, which are subsequently marked as control, N100, N150, N200, N250, N300F and N350, respectively. The plots (7 m × 6 m) were arranged randomly with 4 replicates for each site. To decrease the bias of the effects of nitrogen application rates on yield, economic benefit and environmental pollution, rice varieties were Oryza sativa L. cv. Wuyuejing, Suying 201, 9915, and Sanyou 418 for Xinzhuang, Baimao, Dayi, and Meili, respectively, in 2003, and Oryza sativa L. cv. Suxiangjing, 97-46, 93-1, and Sanyou 422, respectively, in 2004. Transplanting of the seedlings at age around 20 d was done in late June with a density of about 16 hills m−2 and harvest was carried out in late October in both years. Urea was applied 3 times: prior to transplanting, in late July, and in mid-August at a ratio of 5:2:3 and 4:2:4 for 2003 and 2004, respectively. Phosphorus (P) and potassium (K) were applied prior to rice transplanting at the same amounts for every plot according to farmers’ practice. Phosphorus fertilizer was applied in the form of calcium superphosphate at 60 kg P2 O5 ha−1 , and K fertilizer was applied in the form of potassium chloride at 120 kg K2 O ha−1 . Water and pesticide managements were carried out according to the usual farmers’ practice; the soil was flooded with 5 cm of po-
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nded water for the entire rice growth period, except for one week in mid-August and the period from midOctober to rice harvest. The grain yield was determined by harvesting three 1-m2 areas from each plot in mid-October. 0.5 kg grains were sampled and air-dried. After air-drying, the water content was measured at 70 ◦ C and the dry weight of the grains were determine after 8 h. Experiment 2: N fate at different N application rates A detailed investigation of the fate of applied N was conducted at Xinzhuang. A micro-plot was established in each plot, with a plastic tube 40 cm in diameter and 80 cm in height. The plastic tube was inserted into the soil to a depth of 60 cm. The tube was located at least 1.5 m from the boundary ridge of the field. 15 Nurea (enrichment: 5.31 atom% 15 N excess) was applied to these micro-plots at the same N rate as for the field conditions. Phosphorus and K fertilizer and water management practices were the same as the macro-plot experiment. Irrigation water was poured manually into the micro-plots every day to maintain the same level as for the surrounding macro-plot areas. The microplots were established at different locations inside the macro-plots during the two-year investigation to avoid contamination. The fate of N in soil and plant system was investigated in mid-October during the rice harvesting period. The soil N content in the 0–20, 20– 40, and 40–60 cm soil depth, and the N content of the plant materials which were separated into grain, straw and root, were measured using the Kjeldahl method. 15 N abundance of the soil and plant samples was determined using isotope mass spectrometry (MAT 251, USA). Nitrogen recovery in plants (Nrec , %), retention in soil (Nret , %), and loss to the environment (N unaccounted for, %) were calculated as follows: Nrec =
plant N uptake × plant 15 N atom% excess × fertilizer 15 N atom% excess × N applied 100 (1)
Nret =
N remained in soil × soil 15 N atom% excess × fertilizer 15 N atom% excess × N applied 100 (2)
N unaccounted for = 100 − Nrec − Nret
(3)
To understand the dynamics of N fertilizer transformation after application to the field, the surface water was sampled every day within two weeks following each fertilization. Ammonium N (NH+ 4 -N), ni-
trate N (NO− 3 -N), and total dissolved N (TDN) in the surface water were measured using a Bran Luebbe Auto Analyzer III, via ultraviolet spectrophotometry and K2 S2 O8 oxidation ultraviolet spectrophotometry. Ammonia (NH3 ) volatilization was measured in situ using a continuous airflow dynamic chamber method (Kissel et al., 1977). The flow rate was approximately 8 L min−1 . Volatilized NH3 from the soil or water surface was gathered using a PVC chamber of 20 cm in diameter and 20 cm in height and was trapped using 20 mg mL−1 H3 BO3 as an absorbent. Ammonia volatilization was measured twice daily at 7:30–10:30 a.m. and 15:00–18:00 p.m. following fertilization. The sampling was carried out until the NH3 volatilization of the field treated with fertilizer became equivalent to that of the control field (around 2 weeks after fertilizer application). The absorbent H3 BO3 was titrated with 0.01 mol L−1 H2 SO4 , and hourly NH3 volatilization fluxes were calculated. Economic budgets The farmers’ net returns from grain sales minus fertilizer costs were calculated by subtracting the cost of N fertilizer from the income of rice grain. The grain yield was multiplied by the current price of latematuring unhusked japonica rice in this region for both years, which was 2.4–2.8 RMB Yuan kg−1 , and the cost of urea, which is the most common fertilizer, was 2.2–2.3 RMB Yuan kg−1 (Chinagrain, 2011). In addition, in this study we also took into account the direct CO2 emissions during the N chemical fertilizer production (0.88–1.74 kg CO2 -C kg−1 N) (West and Marland, 2002; Lu et al., 2008) and N2 O emissions from rice fields with 0.12% of applied N fertilizer (Zhao et al., 2009). The greenhouse gas emissions were converted to CO2 equivalents (global warming potential of 1 mol N2 O: 296 mol CO2 equivalents) (IPCC, 2007) and then calculated with the carbon price. The carbon price was obtained from Blyth et al. (2009). However, this study did not calculate the cost for mitigation of NH3 volatilization and NO− 3 leaching losses due to the lack of data for the relative price. Statistical analysis Statistical analysis was conducted using SigmaStat 3.11 (Systat Software, Inc., USA). The difference between the fertilization treatments was tested by oneway analysis of variance (ANOVA). The Pearson coefficient was used to test for significant correlations. Statistical significance was indicated at P < 0.05.
N FERTILIZER APPLICATION FOR RICE PRODUCTION
RESULTS Relationships between rice yield and N application rate The rice grain yield had a significant polynomial correlation with the amount of N applied at all sites and during different years (Table I; Fig. 1). The grain yield ranged from 3 876 to 6 325 kg DW ha−1 when no fertilizer was applied. When N was applied at 195– 302 kg N ha−1 , the highest yield of 6 888–8 883 kg DW ha−1 could be obtained. Considering the whole region for both years, the potential highest rice grain yield could be up to 8 164 kg DW ha−1 with N application rate at around 247 kg N ha−1 . These results suggested that a nitrogen fertilization rate around 250 kg N ha−1 was sufficient to return the highest yields in this region.
Fig. 1 Relationships between rice grain yield and N application rate at all sites in 2003 and 2004.
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Since the increased yield results from N application in comparison with no-N application treatment, the agronomic N use efficiency (kg grain yield increase per kg N applied) can be calculated. In this study, it ranged from 3 to 23 kg kg−1 . On average, the agronomic N use efficiency was 13, 14, 15, 11, 8, and 6 kg kg−1 for the N application rates of N100, N150, N200, N250, N300F (farmers’ practice), and N350, respectively. Dynamics of N components in surface water The concentrations of all N components in surface water of the rice fields displayed similar dynamics with time following N applications in 2003 and 2004, and thus only the results from 2003 are shown in Fig. 2. The TDN concentration in surface water was highest one day after fertilization, and then rapidly decreased to the level of the control within 3–7 days after the three applications of urea (Fig. 2a). The significant differences in treatments were maintained (P < 0.05) until 3–7 days after fertilization. The highest TDN concentration was found for the treatments N300F and N350 for all applications. With respect to the NH+ 4N concentration in surface water, peaks appeared 2–3 days after fertilization (Fig. 2b). Significant differences among treatments were found for surface water 4–5 days after fertilization. The NO− 3 -N concentration in surface water increased slowly and showed no peak until 4 days after fertilization (Fig. 2c). Significant differences among treatments began to appear 1–2 days af-
TABLE I Correlations between the amount of N applied and rice grain yield in 2003 and 2004 Year
2003
Site
Xinzhuang Baimao Dayi Meili Average 2004 Xinzhuang Baimao Dayi Meili Average Whole region and years
Regression equation of rice yield (y, kg DW ha−1 ) and N application rates (x, kg N ha−1 ) y y y y y y y y y y y
= −0.0618x + 24.143x + 5650 = −0.039x2 + 19.694x + 5819 = −0.0388x2 + 18.061x + 6048 = −0.0446x2 + 20.188x + 6325 = −0.0448x2 + 20.178x + 5966 = −0.0432x2 + 19.645x + 4655 = −0.0494x2 + 29.81x + 3876 = −0.0232x2 + 13.576x + 6182 = −0.0501x2 + 23.349x + 6163 = −0.0396x2 + 20.968x + 5263 = −0.0405x2 + 20.004x + 5609
**Significant at P < 0.01 level. a) Coefficient of determination.
2
R2 a)
0.877** 0.889** 0.814** 0.747** 0.716** 0.818** 0.811** 0.867** 0.821** 0.529** 0.542**
n
17 18 20 17 72 24 23 28 23 98 170
Crop N requirement for potential highest yield
Potential highest yield
kg N ha−1 195 252 233 226 225 227 302 293 233 265 247
kg DW ha−1 8 008 8 305 8 150 8 610 8 238 6 888 8 373 8 168 8 883 8 039 8 164
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− Fig. 2 Time course of total dissolved N (a), NH+ 4 -N (b), and NO3 -N (c) concentrations in the surface water of the rice fields after application of urea at Xinzhuang in 2003. Arrows indicate fertilization dates. Vertical bars indicate Fisher’s LSD values (P = 0.05). Control, N100, N200, N300F , and N350 are the N application rates of 0, 100, 200, 300 (local farmers’ practice), and 350 kg N ha−1 , respectively.
ter fertilization. The NO− 3 -N concentration was much + lower than the NH4 -N concentration. Similar to TDN, the N300 and N350 treatments were associated with + the highest NO− 3 -N and NH4 -N concentrations. It was demonstrated that the increased N concentration in the surface water was correlated with the amount of N fertilizer used. NH 3 emission from soil and water surface Ammonia emission from the field increased with increased amount of N applied, and ranged from 1.1 kg N ha−1 for the control to 35.4 kg N ha−1 at the N350 rate (Fig. 3). As a result, 8.5%–12.3% of applied fertilizer N was lost by NH3 volatilization, with an average of 10% of applied fertilizer N. No difference was found between the two years. The NH3 flux showed a significant and positive relationship with the NH+ 4 -N concentration in the surface water (Fig. 4).
Fig. 3 Cumulative NH3 volatilization losses during the whole rice season in 2003 and 2004. Vertical bars indicate standard deviation.
Nitrogen fate under different N treatments Plant N recovery rates ranged from 22% to 30%
N FERTILIZER APPLICATION FOR RICE PRODUCTION
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44 kg N ha−1 at N100 to 240 kg N ha−1 at N350 (Table II). Ammonia volatilization represented 9–34 kg N ha−1 . Assuming that the other N was lost through denitrification in the forms of N2 O and N2 , N2 O emission accounts for 0.12% of fertilizer N applied (Zhao et al., 2009), and then N loss in the form of N2 ranged from 35 to 206 kg N ha−1 . The ratio of NH3 -N to N2 -N was from 1:3 to 1:6 (Table II). In addition, the CO2 emission from fossil fuel combustion during N fertilizer production was also markedly influenced by the N application rate. It ranged from 88 to 609 kg C ha−1 (Table II). Fig. 4 Relationships between NH3 emission and NH+ 4 -N concentration in surface water at Xinzhuang in 2003 and 2004 (P < 0.01).
with an average of 26% (Fig. 5). No significant correlation was found with the amount of N applied. The highest N recovery was found for N200, which may have resulted from the higher rice yield compared with the other N treatments. The lowest recovery was found for N300 due to the slightly lower rice yield. The ratio of soil residual N varied between 7% and 31%, decreasing as the amount of N applied increased (Fig. 5). Soil residual N was mainly found in the 0–20 cm top soil layer, while no fertilizer 15 N was found in the 20–60 cm soil layer. The N portion unaccounted for varied between 44% and 69% and showed a significant positive correlation with the amount of N applied (Fig. 5). Environmental responses to different N rates in rice fields Nitrogen application rates significantly influenced fertilizer N loss to the environment, which varied from
Economic budgets Economic income was the top priority considered by farmers. Based on the current price of rice grain and chemical N fertilizer, the returns from sales minus fertilizer costs could be calculated from the benefit of rice grain and the cost of fertilizer N (Table III). At farmers’ current N practice (N300F ), these returns were 1.76 × 104 –2.09 × 104 RMB Yuan ha−1 . The highest returns were found at N200, which were 1.89 × 104 –2.23 × 104 RMB Yuan ha−1 , higher than those at N300F . The lowest returns were at N100, 1.64 × 104 –1.92 × 104 RMB Yuan ha−1 , lower than those at N300F due to the lower grain yield (90% of the rice yield at N300F ). The cost to solve the pollution of CO2 emission due to fossil fuel combustion during N fertilizer production and N2 O emission from the rice field ranged from 2.1 × 102 to 13.6 × 102 RMB Yuan ha−1 (Table III) and increased with fertilizer N application rate. At farmers’ practice (N300F ), it was up to 6.7 × 102 –11.7 × 102 RMB Yuan ha−1 .
Fig. 5 Relationships between fate of fertilizer N and N application rate of the at Xinzhuang. Vertical bars indicate minimum and maximum values.
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N micro-plot experiment in the rice season
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TABLE II Environmental effect of fertilizer N of the N rate
15
N micro-plot experiment in the rice season at Xinzhuang
N lossa) NH3 emission
Total
N2 O emission
N2 emission (apparent)
NH3 :N2
35 64 91 133 169 206
1:4 1:3 1:5 1:5 1:6 1:6
−1
100 150 200 250 300 350
44±2c) fd) 82±2 e 111±13 d 158±12 c 199±9 b 240±7 a
9±1 18±4 20±5 25±4 29±5 34±5
kg N ha (8.5%)e) d 0.12 (12.3%) c 0.18 (9.5%) bc 0.24 (10.1%) abc 0.30 (10.6%) ab 0.36 (9.7%) a 0.42
CO2 emission from N fertilizer productionb) kg C ha−1 88–174 132–261 176–348 220–435 264–522 308–609
a)
N2 O emission is calculated with 0.12% of fertilizer N application (Zhao et al., 2009), assuming no N loss from leaching and runoff, and N2 emission is the total N loss minus NH3 and N2 O emissions. b) C emission from N fertilizer production was calculated with fossil fuel combustion of 0.88 t C t−1 N (West and Marland, 2002) and 1.74 t C t−1 N (Lu et al., 2008). c) Mean±standard deviation. d) Values followed by the same letter(s) in a column indicate no significant difference (P < 0.05). e) Values in the parentheses indicate the percentage of NH3 emission from fertilizer N. TABLE III Comparison of grain yield, net income, and cost for greenhouse gas emission between the farmers’ N application rate (300F ) and alternative N strategies (ANS) N rate
Grain yielda)
kg N ha−1 100 150 200 250 300 350
kg DW ha−1 7 033±667 7 258±863 8 302±646 8 223±800 7 976±492 7 570±830
Grain-yield ratio of 300F :ANS 88 91 104 103 95
Net incomeb)
Cost for greenhouse gas emissionc)
× 104 RMB Yuan ha−1 1.64–1.92 1.67–1.96 1.89–2.23 1.85–2.18 1.76–2.09 1.64–1.95
× 102 RMB Yuan ha−1 2.12–3.89 3.19–5.84 4.25–7.79 5.31–9.73 6.73–11.68 7.43–13.63
a)
Averages derived from the 4 experimental sites in two years and standard deviations. Returns from the grain sales minus the fertilizer costs. The prices of fertilizer and rice grain were obtained from http://www.chinagrain.cn in 2011. The retail price of urea was 2.2–2.3 RMB Yuan kg−1 and the nitrogen content of the urea was 460 g kg−1 . The returns derived from the grain sales were calculated using a price of 2.4–2.8 RMB Yuan kg−1 for the late-maturing unhusked japonica rice in this region. c) Here only including the cost for N2 O (global warming potential: 296 CO2 equivalents) and CO2 emissions from N fertilizer production. The carbon price ranges from 0 to 196 EUR t−1 CO2 equivalents with an average of 55 EUR t−1 CO2 equivalents (Blyth et al., 2009). 100 RMB Yuan = 9.80 EUR. b)
DISCUSSION Environmental impacts of fertilizer N in rice systems The volatilized NH3 exerts a negative impact on the quality of atmospheric environment, and often returns to soil and surface water through dry and wet deposition, aggravating water eutrophication and soil acidification (Bull and Sutton, 1998; Camargo and Alonso, 2006). In the present study, loss of N via NH3 volatilization was one of the dominant N losses and
represented 8.5%–12.4% of the applied N. This is comparable to the other published results for rice fields (Tian et al., 2001; Watanabe et al., 2009). Most of the NH3 volatilization occurred within one week after the N application. Those NH3 volatilization losses resulted from the high concentration of NH+ 4 -N in flooding water (Fig. 4) and high temperature in the summer time (Zhou et al., 2009). At farmers’ common N application rate of 300 kg N ha−1 , the NH3 volatilization loss was up to 29 kg N ha−1 (Table II) and over 3 times that of
N FERTILIZER APPLICATION FOR RICE PRODUCTION
the treatment of N100. Indeed, NH3 volatilization from agriculture has become an important atmospheric pollutant in the Taihu Lake area (Xie et al., 2008), which could be dramatically reduced if the N application rate was reduced to 100–150 kg N ha−1 . Little N loss through runoff was found in rice system in some previous studies (Tian et al., 2007; Zhao et al., 2009). This maybe resulted from the fact that the research fields were chosen at flat areas, where surface water level was carefully controlled. For mountainous areas and some areas with difficulties in managing the irrigation water, however, the runoff of rice field is likely to happen during heavy rainfall events. For the present study, the dissolved fertilizer N (calculated as the difference in N concentration between the control and N treatments in the 5 cm surface water layer) was up to 41%–79% and 24%–78% of applied fertilizer N one day after application, with an average of 56% and 45% for 2003 and 2004, respectively (data not shown). The proportion decreased to 1% of the total N applied at 6–7 days after application. If runoff of the surface water occurred within one week after fertilization, up to 79% of the applied urea fertilizer N would likely be lost. Woli et al. (2004) also reported that the total N concentration in river increased during the rice planting period in central Hokkaido of Japan. Those N losses such as NH3 volatilization (Hayashi et al., 2006) and runoff (Cho, 2003) mainly depended on the concentration of N in surface water. All N components in the surface water of the treatments with N application in the present study dropped to the levels of the control within one week after fertilization (Fig. 2). Therefore, the key period for managing surface water to prevent N loss in rice fields was the week following urea fertilizer application. The irrigation water should be carefully managed during this period so that there is no surplus water which may easily be lost by runoff with heavy rain. Runoff occurred in the plain areas close to the Taihu Lake every few years, as in the summer of 2010, when very heavy rainfalls happened for several days in the week after the tillering fertilization. Large amounts of fertilizer were lost with the runoff and/or drainage water and entered the irrigation and drainage ditches and small canals connected to the Taihu Lake. Also, with the shift to direct seeding instead of transplanting in recent years, though not in the experiments of this study, the field is re-drained again for a few days after germination, which is after the basal fertilization has taken place. This is also a strong cause for loss of
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N to aquifers (surface and subsurface waters) during the summer rice growing season. Several studies reported that there was very little N loss by leaching in rice system in the Taihu Lake region. Zhao et al. (2009) reported that only 1–2 kg N ha−1 was lost by leaching in the rice season when the amount of N applied was 100 and 300 kg N ha−1 . Tian et al. (2007) found that the N concentration in the water that had been leached at 20 cm soil depth during the rice season was less than 1 mg N L−1 and that there was no relationship with the amount of N applied. In this study, it was found that soil residual N was mainly maintained in the 0–20 cm top soil layer (data not shown). No fertilizer 15 N was found in the 20–60 cm top soil layer. This suggested that there was little N fertilizer loss by leaching during the period of summer rice growth at that region. This was probably due to the hard and compact plough pan which prevents water from penetrating into deeper soil layers, a feature of rice fields which can decrease leaching rate compared with upland fields. Denitrification is the main N loss during the rice growing season (Aulakh et al., 2001; Ju et al., 2009) although in-situ studies have yet to confirm this. Zhou et al. (2009) reported that approximately 30%–40% of N fertilizer was lost via denitrification in the rice field. In this study, the apparent N loss with denitrification process ranged from 35% to 57% (Table II). Usually, anaerobic conditions prevail in the rice field during the flooding period and induce denitrification of any NO− 3N present to N2 and N2 O (Davidson et al., 2000). But in irrigated rice systems, many studies reported that N loss via N2 O was only around 0.12% of applied fertilizer N (Zhao et al., 2009) and almost all N loss by denitrification process was in the form of N2 . In the same way, the N loss via denitrification in this study was dominated by N2 and the amount significantly increased as fertilizer N applied increased. Therefore, the rice system has a big potential to remove N load. All of the results suggested that NH3 volatilization and N2 emission by denitrification were the main N loss ways in the rice system although the N2 O emissions were not measured and the total denitrification loss (N2 + N2 O) was only calculated using the difference method in this study. Optimal N application rate in rice systems Many studies suggested that over-application of N fertilizer in farmers’ fields causes very serious environ-
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mental problems in the rice fields in China and the reduction of the N application rates is required (Mosier et al., 2001; Roelcke et al., 2004; Ju et al., 2009). A recommendation for an optimized N application rate should also consider the economic efficiency, so that farmers would accept the recommendation. In this study, the rice grain yields of eight varieties showed that the highest grain yields could be obtained at the N rate of about 250 kg N ha−1 in the Taihu Lake region (Table I; Fig. 1). This application rate was significantly lower than the farmers’ common application rate of 300 kg N ha−1 . Moreover, the agronomic N use efficiency at farmers’ practice (N300F ) was only 8 kg grain kg−1 N, much lower than that at the N rates of 150–200 kg N ha−1 , where the agronomic N use efficiency was nearly 14 kg grain kg−1 N. A low N use efficiency also means a low economic efficiency. Here, the highest returns were 1.89 × 104 –2.23 × 104 RMB Yuan ha−1 at the rate of 200 kg N ha−1 with the highest grain yield (Table III). However, when the N application rate decreased to 150 kg N ha−1 , 91% of the net income at the farmers’ current practice of N300F and 89% of the potential highest grain yield could still be achieved. On the other hand, a high risk of the N environmental pollution is also related to the low N use efficiency. In the present study, the risk of environmental pollution was mainly from N loss after fertilizer application and the CO2 emission from N fertilizer production and transportation. At the farmers’ common practice (N300F ), the N portion unaccounted for reached 199 ± 9 kg N ha−1 , which accounted for 66% of the applied fertilizer N. Due to the scarcity of references for the cost price for NH3 volatilization and NO− 3 leaching, only the cost for CO2 emission was budgeted. The carbon cost for greenhouse gas emission ranged from 2.12 × 102 to 13.63 × 102 RMB Yuan ha−1 , and tended to increase with increasing amounts of applied N (Table III). The farmers’ N application rate resulted in at least 6.73 × 102 –11.68 × 102 RMB Yuan ha−1 for the greenhouse gas emission. If considering the cost to compensate for the N loss due to NH3 emission, runoff, and leaching, the total cost would sharply increase. If the amount of N applied was reduced to 150 kg N ha−1 , the N loss from the field would decrease to 82 kg N ha−1 and the cost for mitigating environmental pollution would also significantly decrease (Tables II and III). CONCLUSIONS The farmers’ common N application rate of 300 kg
N ha−1 was far higher than that required for rice production because the N loss was up to 66% of the applied fertilizer N (199 kg N ha−1 ) and the N use efficiency decreased to 8 kg grain kg−1 N. Based on the results, we thus recommended an application rate of N of 150–200 kg N ha−1 , which can maintain a grain yield of 7 600– 8 300 kg DW ha−1 and decrease N loss to 80–110 kg N ha−1 , and would also significantly decrease the cost for mitigating the environmental pollution. The results of this study showed that NH3 volatilization and denitrification to N2 , which attributed to N losses of 9.6% and 35%–57% of applied fertilizer N, respectively, were the dominant N loss pathways in the rice fields of the Taihu Lake region in China. REFERENCES Aulakh, M. S., Khera, T. S., Doran, J. W. and Bronson, K. F. 2001. Denitrification, N2 O and CO2 fluxes in ricewheat cropping system as affected by crop residues, fertilizer N and legume green manure. Biol. Fert. Soils. 34: 375–389. Blyth, W., Bunn, D., Kettunen, J. and Wilson, T. 2009. Policy interactions, risk and price formation in carbon markets. Energy Policy. 37: 5192–5207. Bull, K. R. and Sutton, M. A. 1998. Critical loads and the relevance of ammonia to an effects-based nitrogen protocol. Atmos. Environ. 32: 565–572. Camargo, J. A. and Alonso, A. 2006. Ecological and toxicological effects of inorganic nitrogen pollution in aquatic ecosystems: A global assessment. Environ. Int. 32: 831–849. Chinagrain. 2011. The database of Chinagrain. Available online at http://www.chinagrain.cn (verified on October 31, 2011). Cho, J. Y. 2003. Seasonal runoff estimation of N and P in a paddy field of central Korea. Earth Environ. Sci. 65: 43–52. Davidson, E. A., Keller, M., Erickson, H. E., Verchot, L. V. and Veldkamp, E. 2000. Testing a conceptual model of soil emissions of nitrous and nitric oxides. Bioscience. 50: 667–680. Deng, M. H., Xie, Y. X., Xiong, Z. Q., Xing, G. X. and Yan, X. Y. 2007. Nitrogen budgets of Yangtse delta region and their effects on environment. Acta Sci. Circumst. (in Chinese). 27: 1709–1716. Dobermann, A. and Cassman, K. G. 2005. Cereal area and nitrogen use efficiency are drivers of future nitrogen fertilizer consumption. Sci. China Ser. C. 48: 745–758. Galloway, J. N., Townsend, A. R., Erisman, J. W., Bekunda, M., Cai, Z. C., Freney, J. R., Martinelli, L. A., Seitzinger, S. P. and Sutton, M. A. 2008. Transformation of the nitrogen cycle: recent trends, questions, and potential solutions. Science. 320: 889–892.
N FERTILIZER APPLICATION FOR RICE PRODUCTION Gruber, N. and Galloway, J. N. 2008. An Earth-system perspective of the global nitrogen cycle. Nature. 451: 293–296. Hayashi, K., Nishimura, S. and Yagi, K. 2006. Ammonia volatilization from the surface of a Japanese paddy field during rice cultivation. Soil Sci. Plant Nutr. 52: 545–555. Hayashi, K. and Yan, X. Y. 2010. Airborne nitrogen load in Japanese and Chinese agroecosystems. Soil Sci. Plant Nutr. 56: 2–18. IPCC. 2007. Working Group III Report “Mitigation of Climate Change”: Chapter 1. Introduction. Available online at http://www.ipcc.ch/pdf/assessment-report/ ar4/wg3/ar4-wg3-chapter1.pdf (verified on October 31, 2011). Ju, X. T., Xing, G. X., Chen, X. P., Zhang, S. L., Zhang, L. J., Liu, X. J., Cui, Z. L., Yin, B., Christie, P., Zhu, Z. L. and Zhang, F. S. 2009. Reducing environmental risk by improving N management in intensive Chinese agricultural systems. P. Natl. Acad. Sci. USA. 106: 3041–3046. Kissel, D. E., Brewer, H. L. and Arkin, G. F. 1977. Design and test of a field sampler for ammonia volatilization. Soil Sci. Soc. Am. J. 41: 1133–1138. Lu, F., Wang, X. K., Han, B., Ouyang, Z. Y., Duan, X. N. and Zheng, H. 2008. Assessment on the availability of nitrogen fertilization in improving carbon sequestration potential of China’s cropland soil. Chin. J. Appl. Ecol. (in Chinese). 19: 2239–2250. Luo, L. C., Qin, B. Q., Yang, L. Y. and Song, Y. Z. 2007. Total inputs of phosphorus and nitrogen by wet deposition into Lake Taihu, China. Hydrobiologia. 581: 63–70. Mosier, A. R., Bleken, M. A., Chaiwanakupt, P., Ellis, E. C., Freney, J. R., Howarth, R. B., Matson, P. A., Minami, K., Naylor, R., Weeks, K. N. and Zhu, Z. L. 2001. Policy implications of human-accelerated nitrogen cycling. Biogeochemistry. 52: 281–320. National Bureau of Statistics of China. 2003. Statistical Yearbook of Jiangsu (in Chinese). Chinese Stastical Publishing House, Beijing. Peng, S. B., Buresh, R. J., Huang, J. L., Yang, J. C., Zou, Y. B., Zhong, X. H., Wang, G. H. and Zhang, F. S. 2006. Strategies for overcoming low agronomic nitrogen use efficiency in irrigated rice systems in China. Field Crops Res. 96: 37–47. Qin, B. Q. 2009. Lake eutrophication: control countermeasures and recycling exploitation. Ecol. Eng. 35: 1569–
57 1573. Roelcke, M., Han, Y., Schleef, K. H., Zhu, J. G., Liu, G., Cai, Z. C. and Richter, J. 2004. Recent trends and recommendations for nitrogen fertilization in intensive agriculture in eastern China. Pedosphere. 14: 449–460. Seitzinger, S. 2008. Nitrogen cycle: out of reach. Nature. 452: 162–163. Tian, G. M., Cai, Z. C., Cao, J. L. and Li, X. P. 2001. Factors affecting ammonia volatilisation from a rice-wheat rotation system. Chemosphere. 42: 123–129. Tian, Y. H., Yin, B., Yang, L. Z., Yin, S. X. and Zhu, Z. L. 2007. Nitrogen runoff and leaching losses during rice-wheat rotations in Taihu Lake region, China. Pedosphere. 17: 445–456. Wang, D. J., Lin, J. H., Sun, R. J., Xia, L. Z. and Lian, G. 2003. Optimum nitrogen rate for a high productive rice-wheat system and its impact on the groundwater in the Taihu Lake area. Acta Pedol. Sin. (in Chinese). 40: 426–432. Watanabe, T., Son, T. T., Hung, N. N., Van Truong, N., Giau, T. Q., Hayashi, K. and Ito, O. 2009. Measurement of ammonia volatilization from flooded paddy fields in Vietnam. Soil Sci. Plant Nutr. 55: 793–799. West, T. O. and Marland, G. 2002. A synthesis of carbon sequestration, carbon emissions, and net carbon flux in agriculture: comparing tillage practices in the United States. Agr. Ecosyst. Environ. 91: 217–232. Woli, K. P., Nagumo, T., Lie, L. and Hatano, R. 2004. Evaluation of the impact of paddy fields on stream water nitrogen concentration in central Hokkaido. Soil Sci. Plant Nutr. 50: 45–55. Xie, Y. X., Xiong, Z. Q., Xing, G. X., Yan, X. Y., Shi, S. L., Sun, G. Q. and Zhu, Z. L. 2008. Source of nitrogen in wet deposition to a rice agroecosystem at Tai Lake region. Atmos. Environ. 42: 5182–5192. Zhao, X., Xie, Y. X., Xiong, Z. Q., Yan, X. Y., Xing, G. X. and Zhu, Z. L. 2009. Nitrogen fate and environmental consequence in paddy soil under rice-wheat rotation in the Taihu Lake region, China. Plant Soil. 319: 225– 234. Zhou, S., Nishiyama, K., Watanabe, Y. and Hosomi, M. 2009. Nitrogen budget and ammonia volatilization in paddy fields fertilized with liquid cattle waste. Water Air Soil Poll. 201: 135–147. Zhu, Z. L., Zhang, S. L., Yin, B. and Yan, X. Y. 2010. Historical comparison on the response curves of rice yield-nitrogen application rate in Tai Lake region. Plant Nutr. Fert. Sci. (in Chinese). 16: 1–5.
Optimizing Nitrogen Fertilizer Application for Rice Production in the Taihu Lake Region, China∗1 DENG Mei-Hua1,3 , SHI Xiao-Jun2 , TIAN Yu-Hua1 , YIN Bin1,∗2 , ZHANG Shao-Lin1 , ZHU Zhao-Liang1 and S. D. KIMURA3 1
State Key Laboratory of Soil and Sustainable Agriculture, Institute of Soil Science, Chinese Academy of Sciences, Nanjing 210008 (China) 2 College of Resources and Environments, Southwest University, Chongqing 310047 (China) 3 United Graduate School of Agriculture Science, Tokyo University of Agriculture and Technology, Tokyo 1838509 (Japan) (Received June 27, 2011; revised October 31, 2011)
ABSTRACT To determine the optimal amount of nitrogen (N) fertilizer for achieving a sustainable rice production at the Taihu Lake region of China, two-year on-farm field experiments were performed at four sites using various N application rates. The results showed that 22%–30% of the applied N was recovered in crop and 7%–31% in soils at the rates of 100–350 kg N ha−1 . Nitrogen losses increased with N application rates, from 44% of the applied fertilizer N at the rate of 100 kg N ha−1 to 69% of the N applied at 350 kg N ha−1 . Ammonia volatilization and apparent denitrification were the main pathways of N losses. The N application rate of 300 kg N ha−1 , which is commonly used by local farmers in the study region, was found to lead to a significant reduction in economic and environmental efficiency. Considering the cost for mitigating environmental pollution and the maximum net economic income, an application rate of 100–150 kg N ha−1 would be recommended. This recommended N application rate could greatly reduce N loss from 199 kg N ha−1 occurring at the N application rate of 300 kg N ha−1 to 80–110 kg N ha−1 , with the rice grain yield still reaching 7 300–8 300 kg DW ha−1 in the meantime. Key Words:
economic efficiency, environmental efficiency, N application rate, N loss, rice grain yield
Citation: Deng, M. H., Shi, X. J., Tian, Y. H., Yin, B., Zhang, S. L., Zhu, Z. L. and Kimura, S. D. 2012. Optimizing nitrogen fertilizer application for rice production in the Taihu Lake region, China. Pedosphere. 22(1): 48–57.
INTRODUCTION The use of nitrogen (N) fertilizer has changed global N cycle markedly and has been causing various negative environmental consequences such as eutrophication of surface water, global warming, and ozone layer depletion (Gruber and Galloway, 2008; Seitzinger, 2008). However, the total amount of N fertilizer in the world will undoubtedly continue to increase to meet the increasing demand for food (Mosier et al., 2001; Dobermann and Cassman, 2005). The mitigation of negative environmental influence of N fertilizer has become an important issue for sustainable agriculture development (Galloway et al., 2008). ∗1
In the Taihu Lake region, one of main food production regions in China, the intensive agricultural activities have contributed to a marked eutrophication of the surface water and deterioration of drinking water quality (Qin, 2009), and increased atmospheric deposition of N (Luo et al., 2007; Hayashi and Yan, 2010). The annual chemical fertilizer N input in the whole Taihu Lake region reached 450 kg N ha−1 in 2002, 2 times the Chinese national average in that year (Deng et al., 2007). In general, the rotations of irrigated summer rice and winter upland crops (wheat or rapeseed) are the typical crop system in this region, which accounts for about 80% of the arable land area (National Bureau of Statistics of China, 2003). To maximize rice yields,
Supported by the Knowledge Innovation Program of the Chinese Academy of Sciences (No. KZCX2-YW-440-1), the National Natural Science Foundation of China (Nos. 30390080 and 41071197), and the National High Technology Research and Development Program (863 Program) of China (No. 2006AA10Z418). ∗2 Corresponding author. E-mail: [email protected].
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farmers often use excessive N fertilizer (Ju et al., 2009; Zhao et al., 2009). Many studies reported that rice production in the Taihu Lake region was a dominant nonpoint pollution source (Mosier et al., 2001; Roelcke et al., 2004; Ju et al., 2009). To ensure food security and an environmentally friendly rice fertilizer management system, many researchers have investigated the optimal N application rate for rice production. An optimum N fertilize application rate of 225–270 kg N ha−1 was suggested to achieve good rice productivity (Wang et al., 2003) and Zhu et al. (2010) recommended a rate of 200 kg N ha−1 for rice cultivation for the Taihu Lake region. However, we suspect that these recommended N application rates may be still too high, when considering already high ambient N deposition and severe N losses to the environment. According to the investigation at Hunan, Jiangsu, Zhejiang, and Guangdong in China, Peng et al. (2006) suggested that a good rice yield (ranging from 5 300 to 9 900 kg ha−1 ) can be achieved at 60–120 kg N ha−1 for irrigated rice production. However, the N application rate employed by farmers in the Taihu Lake region is still around 300 kg N ha−1 in a single rice season, corresponding to 500–600 kg N ha−1 year−1 under the double-cropping system (Ju et al., 2009; Zhao et al., 2009). This high N application rate was mainly attributed to the pursuit of maximum grain yields by policy-makers and farmers. For policy-makers, national food security is the main concern due to China’s huge population. As for farmers, the unintentionally excessive use of N fertilizer is aimed at achieving a high yield and consequently a high economic income. Hence, an optimized N application rate is required to balance the concern of farmers and policy-makers about high and stable grain yields and the need to protect the environment in the Taihu Lake region from severe N pollution, also taking into account the cost of N fertilizers. In this study, two-year on-farm field experiments were carried out at four sites using seven different N application rates and eight varieties of rice to investigate the impact of N fertilizer use on rice yield and economic and environmental efficiency of N in the rice system. The objectives of this study were to 1) determine the economic and environmental losses incurred on farmers’ rice fields in the Taihu Lake region and to 2) seek a new optimal N application rate and potential solutions to mitigate N pollution in rice systems in the Taihu Lake region.
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MATERIALS AND METHODS Study sites The experiments were conducted in 2003 and 2004 at Changshu City (31◦ 30 –31◦ 50 N, 120◦ 33 –121◦ 3 E) of Jiangsu Province in southeastern China, which is located in the heartland of the Taihu Lake region. This region has a typical subtropical humid climate. The annual average temperature is 15.5 ◦ C and the annual precipitation is 1 038 mm. The soil is classified as a Gleyi-stagnic Anthrosol according to the FAO soil taxonomy system. An irrigated summer rice and winter upland crop (wheat or rapeseed) double-crop rotation is the main cropping system in this region. Four farmers’ fields at Xinzhuang (31◦ 32 45 N, 120◦ 41 57 E), Baimao (31◦ 35 30 N, 120◦ 54 9 E), Dayi (31◦ 42 10 N, 120◦ 41 3 E), and Meili (31◦ 42 50 N, 120◦ 52 55 E) were chosen. Experiment 1: yield responses to N application rate All 4 farmers’ fields were employed to investigate the impact of N rates on rice productivity. The N application rates included 0, 100, 200, 300 (local farmers’ practice), and 350 kg N ha−1 in 2003 and 0, 100, 150, 200, 250, 300, and 350 kg N ha−1 in 2004, which are subsequently marked as control, N100, N150, N200, N250, N300F and N350, respectively. The plots (7 m × 6 m) were arranged randomly with 4 replicates for each site. To decrease the bias of the effects of nitrogen application rates on yield, economic benefit and environmental pollution, rice varieties were Oryza sativa L. cv. Wuyuejing, Suying 201, 9915, and Sanyou 418 for Xinzhuang, Baimao, Dayi, and Meili, respectively, in 2003, and Oryza sativa L. cv. Suxiangjing, 97-46, 93-1, and Sanyou 422, respectively, in 2004. Transplanting of the seedlings at age around 20 d was done in late June with a density of about 16 hills m−2 and harvest was carried out in late October in both years. Urea was applied 3 times: prior to transplanting, in late July, and in mid-August at a ratio of 5:2:3 and 4:2:4 for 2003 and 2004, respectively. Phosphorus (P) and potassium (K) were applied prior to rice transplanting at the same amounts for every plot according to farmers’ practice. Phosphorus fertilizer was applied in the form of calcium superphosphate at 60 kg P2 O5 ha−1 , and K fertilizer was applied in the form of potassium chloride at 120 kg K2 O ha−1 . Water and pesticide managements were carried out according to the usual farmers’ practice; the soil was flooded with 5 cm of po-
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nded water for the entire rice growth period, except for one week in mid-August and the period from midOctober to rice harvest. The grain yield was determined by harvesting three 1-m2 areas from each plot in mid-October. 0.5 kg grains were sampled and air-dried. After air-drying, the water content was measured at 70 ◦ C and the dry weight of the grains were determine after 8 h. Experiment 2: N fate at different N application rates A detailed investigation of the fate of applied N was conducted at Xinzhuang. A micro-plot was established in each plot, with a plastic tube 40 cm in diameter and 80 cm in height. The plastic tube was inserted into the soil to a depth of 60 cm. The tube was located at least 1.5 m from the boundary ridge of the field. 15 Nurea (enrichment: 5.31 atom% 15 N excess) was applied to these micro-plots at the same N rate as for the field conditions. Phosphorus and K fertilizer and water management practices were the same as the macro-plot experiment. Irrigation water was poured manually into the micro-plots every day to maintain the same level as for the surrounding macro-plot areas. The microplots were established at different locations inside the macro-plots during the two-year investigation to avoid contamination. The fate of N in soil and plant system was investigated in mid-October during the rice harvesting period. The soil N content in the 0–20, 20– 40, and 40–60 cm soil depth, and the N content of the plant materials which were separated into grain, straw and root, were measured using the Kjeldahl method. 15 N abundance of the soil and plant samples was determined using isotope mass spectrometry (MAT 251, USA). Nitrogen recovery in plants (Nrec , %), retention in soil (Nret , %), and loss to the environment (N unaccounted for, %) were calculated as follows: Nrec =
plant N uptake × plant 15 N atom% excess × fertilizer 15 N atom% excess × N applied 100 (1)
Nret =
N remained in soil × soil 15 N atom% excess × fertilizer 15 N atom% excess × N applied 100 (2)
N unaccounted for = 100 − Nrec − Nret
(3)
To understand the dynamics of N fertilizer transformation after application to the field, the surface water was sampled every day within two weeks following each fertilization. Ammonium N (NH+ 4 -N), ni-
trate N (NO− 3 -N), and total dissolved N (TDN) in the surface water were measured using a Bran Luebbe Auto Analyzer III, via ultraviolet spectrophotometry and K2 S2 O8 oxidation ultraviolet spectrophotometry. Ammonia (NH3 ) volatilization was measured in situ using a continuous airflow dynamic chamber method (Kissel et al., 1977). The flow rate was approximately 8 L min−1 . Volatilized NH3 from the soil or water surface was gathered using a PVC chamber of 20 cm in diameter and 20 cm in height and was trapped using 20 mg mL−1 H3 BO3 as an absorbent. Ammonia volatilization was measured twice daily at 7:30–10:30 a.m. and 15:00–18:00 p.m. following fertilization. The sampling was carried out until the NH3 volatilization of the field treated with fertilizer became equivalent to that of the control field (around 2 weeks after fertilizer application). The absorbent H3 BO3 was titrated with 0.01 mol L−1 H2 SO4 , and hourly NH3 volatilization fluxes were calculated. Economic budgets The farmers’ net returns from grain sales minus fertilizer costs were calculated by subtracting the cost of N fertilizer from the income of rice grain. The grain yield was multiplied by the current price of latematuring unhusked japonica rice in this region for both years, which was 2.4–2.8 RMB Yuan kg−1 , and the cost of urea, which is the most common fertilizer, was 2.2–2.3 RMB Yuan kg−1 (Chinagrain, 2011). In addition, in this study we also took into account the direct CO2 emissions during the N chemical fertilizer production (0.88–1.74 kg CO2 -C kg−1 N) (West and Marland, 2002; Lu et al., 2008) and N2 O emissions from rice fields with 0.12% of applied N fertilizer (Zhao et al., 2009). The greenhouse gas emissions were converted to CO2 equivalents (global warming potential of 1 mol N2 O: 296 mol CO2 equivalents) (IPCC, 2007) and then calculated with the carbon price. The carbon price was obtained from Blyth et al. (2009). However, this study did not calculate the cost for mitigation of NH3 volatilization and NO− 3 leaching losses due to the lack of data for the relative price. Statistical analysis Statistical analysis was conducted using SigmaStat 3.11 (Systat Software, Inc., USA). The difference between the fertilization treatments was tested by oneway analysis of variance (ANOVA). The Pearson coefficient was used to test for significant correlations. Statistical significance was indicated at P < 0.05.
N FERTILIZER APPLICATION FOR RICE PRODUCTION
RESULTS Relationships between rice yield and N application rate The rice grain yield had a significant polynomial correlation with the amount of N applied at all sites and during different years (Table I; Fig. 1). The grain yield ranged from 3 876 to 6 325 kg DW ha−1 when no fertilizer was applied. When N was applied at 195– 302 kg N ha−1 , the highest yield of 6 888–8 883 kg DW ha−1 could be obtained. Considering the whole region for both years, the potential highest rice grain yield could be up to 8 164 kg DW ha−1 with N application rate at around 247 kg N ha−1 . These results suggested that a nitrogen fertilization rate around 250 kg N ha−1 was sufficient to return the highest yields in this region.
Fig. 1 Relationships between rice grain yield and N application rate at all sites in 2003 and 2004.
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Since the increased yield results from N application in comparison with no-N application treatment, the agronomic N use efficiency (kg grain yield increase per kg N applied) can be calculated. In this study, it ranged from 3 to 23 kg kg−1 . On average, the agronomic N use efficiency was 13, 14, 15, 11, 8, and 6 kg kg−1 for the N application rates of N100, N150, N200, N250, N300F (farmers’ practice), and N350, respectively. Dynamics of N components in surface water The concentrations of all N components in surface water of the rice fields displayed similar dynamics with time following N applications in 2003 and 2004, and thus only the results from 2003 are shown in Fig. 2. The TDN concentration in surface water was highest one day after fertilization, and then rapidly decreased to the level of the control within 3–7 days after the three applications of urea (Fig. 2a). The significant differences in treatments were maintained (P < 0.05) until 3–7 days after fertilization. The highest TDN concentration was found for the treatments N300F and N350 for all applications. With respect to the NH+ 4N concentration in surface water, peaks appeared 2–3 days after fertilization (Fig. 2b). Significant differences among treatments were found for surface water 4–5 days after fertilization. The NO− 3 -N concentration in surface water increased slowly and showed no peak until 4 days after fertilization (Fig. 2c). Significant differences among treatments began to appear 1–2 days af-
TABLE I Correlations between the amount of N applied and rice grain yield in 2003 and 2004 Year
2003
Site
Xinzhuang Baimao Dayi Meili Average 2004 Xinzhuang Baimao Dayi Meili Average Whole region and years
Regression equation of rice yield (y, kg DW ha−1 ) and N application rates (x, kg N ha−1 ) y y y y y y y y y y y
= −0.0618x + 24.143x + 5650 = −0.039x2 + 19.694x + 5819 = −0.0388x2 + 18.061x + 6048 = −0.0446x2 + 20.188x + 6325 = −0.0448x2 + 20.178x + 5966 = −0.0432x2 + 19.645x + 4655 = −0.0494x2 + 29.81x + 3876 = −0.0232x2 + 13.576x + 6182 = −0.0501x2 + 23.349x + 6163 = −0.0396x2 + 20.968x + 5263 = −0.0405x2 + 20.004x + 5609
**Significant at P < 0.01 level. a) Coefficient of determination.
2
R2 a)
0.877** 0.889** 0.814** 0.747** 0.716** 0.818** 0.811** 0.867** 0.821** 0.529** 0.542**
n
17 18 20 17 72 24 23 28 23 98 170
Crop N requirement for potential highest yield
Potential highest yield
kg N ha−1 195 252 233 226 225 227 302 293 233 265 247
kg DW ha−1 8 008 8 305 8 150 8 610 8 238 6 888 8 373 8 168 8 883 8 039 8 164
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− Fig. 2 Time course of total dissolved N (a), NH+ 4 -N (b), and NO3 -N (c) concentrations in the surface water of the rice fields after application of urea at Xinzhuang in 2003. Arrows indicate fertilization dates. Vertical bars indicate Fisher’s LSD values (P = 0.05). Control, N100, N200, N300F , and N350 are the N application rates of 0, 100, 200, 300 (local farmers’ practice), and 350 kg N ha−1 , respectively.
ter fertilization. The NO− 3 -N concentration was much + lower than the NH4 -N concentration. Similar to TDN, the N300 and N350 treatments were associated with + the highest NO− 3 -N and NH4 -N concentrations. It was demonstrated that the increased N concentration in the surface water was correlated with the amount of N fertilizer used. NH 3 emission from soil and water surface Ammonia emission from the field increased with increased amount of N applied, and ranged from 1.1 kg N ha−1 for the control to 35.4 kg N ha−1 at the N350 rate (Fig. 3). As a result, 8.5%–12.3% of applied fertilizer N was lost by NH3 volatilization, with an average of 10% of applied fertilizer N. No difference was found between the two years. The NH3 flux showed a significant and positive relationship with the NH+ 4 -N concentration in the surface water (Fig. 4).
Fig. 3 Cumulative NH3 volatilization losses during the whole rice season in 2003 and 2004. Vertical bars indicate standard deviation.
Nitrogen fate under different N treatments Plant N recovery rates ranged from 22% to 30%
N FERTILIZER APPLICATION FOR RICE PRODUCTION
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44 kg N ha−1 at N100 to 240 kg N ha−1 at N350 (Table II). Ammonia volatilization represented 9–34 kg N ha−1 . Assuming that the other N was lost through denitrification in the forms of N2 O and N2 , N2 O emission accounts for 0.12% of fertilizer N applied (Zhao et al., 2009), and then N loss in the form of N2 ranged from 35 to 206 kg N ha−1 . The ratio of NH3 -N to N2 -N was from 1:3 to 1:6 (Table II). In addition, the CO2 emission from fossil fuel combustion during N fertilizer production was also markedly influenced by the N application rate. It ranged from 88 to 609 kg C ha−1 (Table II). Fig. 4 Relationships between NH3 emission and NH+ 4 -N concentration in surface water at Xinzhuang in 2003 and 2004 (P < 0.01).
with an average of 26% (Fig. 5). No significant correlation was found with the amount of N applied. The highest N recovery was found for N200, which may have resulted from the higher rice yield compared with the other N treatments. The lowest recovery was found for N300 due to the slightly lower rice yield. The ratio of soil residual N varied between 7% and 31%, decreasing as the amount of N applied increased (Fig. 5). Soil residual N was mainly found in the 0–20 cm top soil layer, while no fertilizer 15 N was found in the 20–60 cm soil layer. The N portion unaccounted for varied between 44% and 69% and showed a significant positive correlation with the amount of N applied (Fig. 5). Environmental responses to different N rates in rice fields Nitrogen application rates significantly influenced fertilizer N loss to the environment, which varied from
Economic budgets Economic income was the top priority considered by farmers. Based on the current price of rice grain and chemical N fertilizer, the returns from sales minus fertilizer costs could be calculated from the benefit of rice grain and the cost of fertilizer N (Table III). At farmers’ current N practice (N300F ), these returns were 1.76 × 104 –2.09 × 104 RMB Yuan ha−1 . The highest returns were found at N200, which were 1.89 × 104 –2.23 × 104 RMB Yuan ha−1 , higher than those at N300F . The lowest returns were at N100, 1.64 × 104 –1.92 × 104 RMB Yuan ha−1 , lower than those at N300F due to the lower grain yield (90% of the rice yield at N300F ). The cost to solve the pollution of CO2 emission due to fossil fuel combustion during N fertilizer production and N2 O emission from the rice field ranged from 2.1 × 102 to 13.6 × 102 RMB Yuan ha−1 (Table III) and increased with fertilizer N application rate. At farmers’ practice (N300F ), it was up to 6.7 × 102 –11.7 × 102 RMB Yuan ha−1 .
Fig. 5 Relationships between fate of fertilizer N and N application rate of the at Xinzhuang. Vertical bars indicate minimum and maximum values.
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N micro-plot experiment in the rice season
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TABLE II Environmental effect of fertilizer N of the N rate
15
N micro-plot experiment in the rice season at Xinzhuang
N lossa) NH3 emission
Total
N2 O emission
N2 emission (apparent)
NH3 :N2
35 64 91 133 169 206
1:4 1:3 1:5 1:5 1:6 1:6
−1
100 150 200 250 300 350
44±2c) fd) 82±2 e 111±13 d 158±12 c 199±9 b 240±7 a
9±1 18±4 20±5 25±4 29±5 34±5
kg N ha (8.5%)e) d 0.12 (12.3%) c 0.18 (9.5%) bc 0.24 (10.1%) abc 0.30 (10.6%) ab 0.36 (9.7%) a 0.42
CO2 emission from N fertilizer productionb) kg C ha−1 88–174 132–261 176–348 220–435 264–522 308–609
a)
N2 O emission is calculated with 0.12% of fertilizer N application (Zhao et al., 2009), assuming no N loss from leaching and runoff, and N2 emission is the total N loss minus NH3 and N2 O emissions. b) C emission from N fertilizer production was calculated with fossil fuel combustion of 0.88 t C t−1 N (West and Marland, 2002) and 1.74 t C t−1 N (Lu et al., 2008). c) Mean±standard deviation. d) Values followed by the same letter(s) in a column indicate no significant difference (P < 0.05). e) Values in the parentheses indicate the percentage of NH3 emission from fertilizer N. TABLE III Comparison of grain yield, net income, and cost for greenhouse gas emission between the farmers’ N application rate (300F ) and alternative N strategies (ANS) N rate
Grain yielda)
kg N ha−1 100 150 200 250 300 350
kg DW ha−1 7 033±667 7 258±863 8 302±646 8 223±800 7 976±492 7 570±830
Grain-yield ratio of 300F :ANS 88 91 104 103 95
Net incomeb)
Cost for greenhouse gas emissionc)
× 104 RMB Yuan ha−1 1.64–1.92 1.67–1.96 1.89–2.23 1.85–2.18 1.76–2.09 1.64–1.95
× 102 RMB Yuan ha−1 2.12–3.89 3.19–5.84 4.25–7.79 5.31–9.73 6.73–11.68 7.43–13.63
a)
Averages derived from the 4 experimental sites in two years and standard deviations. Returns from the grain sales minus the fertilizer costs. The prices of fertilizer and rice grain were obtained from http://www.chinagrain.cn in 2011. The retail price of urea was 2.2–2.3 RMB Yuan kg−1 and the nitrogen content of the urea was 460 g kg−1 . The returns derived from the grain sales were calculated using a price of 2.4–2.8 RMB Yuan kg−1 for the late-maturing unhusked japonica rice in this region. c) Here only including the cost for N2 O (global warming potential: 296 CO2 equivalents) and CO2 emissions from N fertilizer production. The carbon price ranges from 0 to 196 EUR t−1 CO2 equivalents with an average of 55 EUR t−1 CO2 equivalents (Blyth et al., 2009). 100 RMB Yuan = 9.80 EUR. b)
DISCUSSION Environmental impacts of fertilizer N in rice systems The volatilized NH3 exerts a negative impact on the quality of atmospheric environment, and often returns to soil and surface water through dry and wet deposition, aggravating water eutrophication and soil acidification (Bull and Sutton, 1998; Camargo and Alonso, 2006). In the present study, loss of N via NH3 volatilization was one of the dominant N losses and
represented 8.5%–12.4% of the applied N. This is comparable to the other published results for rice fields (Tian et al., 2001; Watanabe et al., 2009). Most of the NH3 volatilization occurred within one week after the N application. Those NH3 volatilization losses resulted from the high concentration of NH+ 4 -N in flooding water (Fig. 4) and high temperature in the summer time (Zhou et al., 2009). At farmers’ common N application rate of 300 kg N ha−1 , the NH3 volatilization loss was up to 29 kg N ha−1 (Table II) and over 3 times that of
N FERTILIZER APPLICATION FOR RICE PRODUCTION
the treatment of N100. Indeed, NH3 volatilization from agriculture has become an important atmospheric pollutant in the Taihu Lake area (Xie et al., 2008), which could be dramatically reduced if the N application rate was reduced to 100–150 kg N ha−1 . Little N loss through runoff was found in rice system in some previous studies (Tian et al., 2007; Zhao et al., 2009). This maybe resulted from the fact that the research fields were chosen at flat areas, where surface water level was carefully controlled. For mountainous areas and some areas with difficulties in managing the irrigation water, however, the runoff of rice field is likely to happen during heavy rainfall events. For the present study, the dissolved fertilizer N (calculated as the difference in N concentration between the control and N treatments in the 5 cm surface water layer) was up to 41%–79% and 24%–78% of applied fertilizer N one day after application, with an average of 56% and 45% for 2003 and 2004, respectively (data not shown). The proportion decreased to 1% of the total N applied at 6–7 days after application. If runoff of the surface water occurred within one week after fertilization, up to 79% of the applied urea fertilizer N would likely be lost. Woli et al. (2004) also reported that the total N concentration in river increased during the rice planting period in central Hokkaido of Japan. Those N losses such as NH3 volatilization (Hayashi et al., 2006) and runoff (Cho, 2003) mainly depended on the concentration of N in surface water. All N components in the surface water of the treatments with N application in the present study dropped to the levels of the control within one week after fertilization (Fig. 2). Therefore, the key period for managing surface water to prevent N loss in rice fields was the week following urea fertilizer application. The irrigation water should be carefully managed during this period so that there is no surplus water which may easily be lost by runoff with heavy rain. Runoff occurred in the plain areas close to the Taihu Lake every few years, as in the summer of 2010, when very heavy rainfalls happened for several days in the week after the tillering fertilization. Large amounts of fertilizer were lost with the runoff and/or drainage water and entered the irrigation and drainage ditches and small canals connected to the Taihu Lake. Also, with the shift to direct seeding instead of transplanting in recent years, though not in the experiments of this study, the field is re-drained again for a few days after germination, which is after the basal fertilization has taken place. This is also a strong cause for loss of
55
N to aquifers (surface and subsurface waters) during the summer rice growing season. Several studies reported that there was very little N loss by leaching in rice system in the Taihu Lake region. Zhao et al. (2009) reported that only 1–2 kg N ha−1 was lost by leaching in the rice season when the amount of N applied was 100 and 300 kg N ha−1 . Tian et al. (2007) found that the N concentration in the water that had been leached at 20 cm soil depth during the rice season was less than 1 mg N L−1 and that there was no relationship with the amount of N applied. In this study, it was found that soil residual N was mainly maintained in the 0–20 cm top soil layer (data not shown). No fertilizer 15 N was found in the 20–60 cm top soil layer. This suggested that there was little N fertilizer loss by leaching during the period of summer rice growth at that region. This was probably due to the hard and compact plough pan which prevents water from penetrating into deeper soil layers, a feature of rice fields which can decrease leaching rate compared with upland fields. Denitrification is the main N loss during the rice growing season (Aulakh et al., 2001; Ju et al., 2009) although in-situ studies have yet to confirm this. Zhou et al. (2009) reported that approximately 30%–40% of N fertilizer was lost via denitrification in the rice field. In this study, the apparent N loss with denitrification process ranged from 35% to 57% (Table II). Usually, anaerobic conditions prevail in the rice field during the flooding period and induce denitrification of any NO− 3N present to N2 and N2 O (Davidson et al., 2000). But in irrigated rice systems, many studies reported that N loss via N2 O was only around 0.12% of applied fertilizer N (Zhao et al., 2009) and almost all N loss by denitrification process was in the form of N2 . In the same way, the N loss via denitrification in this study was dominated by N2 and the amount significantly increased as fertilizer N applied increased. Therefore, the rice system has a big potential to remove N load. All of the results suggested that NH3 volatilization and N2 emission by denitrification were the main N loss ways in the rice system although the N2 O emissions were not measured and the total denitrification loss (N2 + N2 O) was only calculated using the difference method in this study. Optimal N application rate in rice systems Many studies suggested that over-application of N fertilizer in farmers’ fields causes very serious environ-
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mental problems in the rice fields in China and the reduction of the N application rates is required (Mosier et al., 2001; Roelcke et al., 2004; Ju et al., 2009). A recommendation for an optimized N application rate should also consider the economic efficiency, so that farmers would accept the recommendation. In this study, the rice grain yields of eight varieties showed that the highest grain yields could be obtained at the N rate of about 250 kg N ha−1 in the Taihu Lake region (Table I; Fig. 1). This application rate was significantly lower than the farmers’ common application rate of 300 kg N ha−1 . Moreover, the agronomic N use efficiency at farmers’ practice (N300F ) was only 8 kg grain kg−1 N, much lower than that at the N rates of 150–200 kg N ha−1 , where the agronomic N use efficiency was nearly 14 kg grain kg−1 N. A low N use efficiency also means a low economic efficiency. Here, the highest returns were 1.89 × 104 –2.23 × 104 RMB Yuan ha−1 at the rate of 200 kg N ha−1 with the highest grain yield (Table III). However, when the N application rate decreased to 150 kg N ha−1 , 91% of the net income at the farmers’ current practice of N300F and 89% of the potential highest grain yield could still be achieved. On the other hand, a high risk of the N environmental pollution is also related to the low N use efficiency. In the present study, the risk of environmental pollution was mainly from N loss after fertilizer application and the CO2 emission from N fertilizer production and transportation. At the farmers’ common practice (N300F ), the N portion unaccounted for reached 199 ± 9 kg N ha−1 , which accounted for 66% of the applied fertilizer N. Due to the scarcity of references for the cost price for NH3 volatilization and NO− 3 leaching, only the cost for CO2 emission was budgeted. The carbon cost for greenhouse gas emission ranged from 2.12 × 102 to 13.63 × 102 RMB Yuan ha−1 , and tended to increase with increasing amounts of applied N (Table III). The farmers’ N application rate resulted in at least 6.73 × 102 –11.68 × 102 RMB Yuan ha−1 for the greenhouse gas emission. If considering the cost to compensate for the N loss due to NH3 emission, runoff, and leaching, the total cost would sharply increase. If the amount of N applied was reduced to 150 kg N ha−1 , the N loss from the field would decrease to 82 kg N ha−1 and the cost for mitigating environmental pollution would also significantly decrease (Tables II and III). CONCLUSIONS The farmers’ common N application rate of 300 kg
N ha−1 was far higher than that required for rice production because the N loss was up to 66% of the applied fertilizer N (199 kg N ha−1 ) and the N use efficiency decreased to 8 kg grain kg−1 N. Based on the results, we thus recommended an application rate of N of 150–200 kg N ha−1 , which can maintain a grain yield of 7 600– 8 300 kg DW ha−1 and decrease N loss to 80–110 kg N ha−1 , and would also significantly decrease the cost for mitigating the environmental pollution. The results of this study showed that NH3 volatilization and denitrification to N2 , which attributed to N losses of 9.6% and 35%–57% of applied fertilizer N, respectively, were the dominant N loss pathways in the rice fields of the Taihu Lake region in China. REFERENCES Aulakh, M. S., Khera, T. S., Doran, J. W. and Bronson, K. F. 2001. Denitrification, N2 O and CO2 fluxes in ricewheat cropping system as affected by crop residues, fertilizer N and legume green manure. Biol. Fert. Soils. 34: 375–389. Blyth, W., Bunn, D., Kettunen, J. and Wilson, T. 2009. Policy interactions, risk and price formation in carbon markets. Energy Policy. 37: 5192–5207. Bull, K. R. and Sutton, M. A. 1998. Critical loads and the relevance of ammonia to an effects-based nitrogen protocol. Atmos. Environ. 32: 565–572. Camargo, J. A. and Alonso, A. 2006. Ecological and toxicological effects of inorganic nitrogen pollution in aquatic ecosystems: A global assessment. Environ. Int. 32: 831–849. Chinagrain. 2011. The database of Chinagrain. Available online at http://www.chinagrain.cn (verified on October 31, 2011). Cho, J. Y. 2003. Seasonal runoff estimation of N and P in a paddy field of central Korea. Earth Environ. Sci. 65: 43–52. Davidson, E. A., Keller, M., Erickson, H. E., Verchot, L. V. and Veldkamp, E. 2000. Testing a conceptual model of soil emissions of nitrous and nitric oxides. Bioscience. 50: 667–680. Deng, M. H., Xie, Y. X., Xiong, Z. Q., Xing, G. X. and Yan, X. Y. 2007. Nitrogen budgets of Yangtse delta region and their effects on environment. Acta Sci. Circumst. (in Chinese). 27: 1709–1716. Dobermann, A. and Cassman, K. G. 2005. Cereal area and nitrogen use efficiency are drivers of future nitrogen fertilizer consumption. Sci. China Ser. C. 48: 745–758. Galloway, J. N., Townsend, A. R., Erisman, J. W., Bekunda, M., Cai, Z. C., Freney, J. R., Martinelli, L. A., Seitzinger, S. P. and Sutton, M. A. 2008. Transformation of the nitrogen cycle: recent trends, questions, and potential solutions. Science. 320: 889–892.
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