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Effects of water table and fertilization management on nitrogen loading to groundwater
Agricultural Water Management 82 (2006) 86–98 www.elsevier.com/locate/agwat
Effects of water table and fertilization management on nitrogen loading to groundwater Huaming Guo a,b,*, Guanghe Li a, Dayi Zhang a, Xu Zhang a, Chang’ai Lu c a b
Department of Environmental Science and Engineering, Tsinghua University, Beijing 100084, PR China School of Water Resource and Environment, China University of Geosciences, Beijing 100083, PR China c Chinese Academy of Agricultural Science, Beijing 100081, PR China Accepted 25 July 2005 Available online 12 September 2005
Abstract Groundwater contamination by nitrate associated with fertilization practices is a ubiquitous environmental issue, and consequently of world-wide concern. Controlling this contamination requires an ability to measure and predict nitrate loading from unsaturated zone to saturated zone. A field experiment was conducted in an intensively irrigated agricultural area in Dianchi catchment of Kunming, China. Two celery (Apium graveolens) crop sites with different water table depths (Site A is 2.0 m below the soil surface; Site B is 0.5 m below the soil surface) were selected for the experiment. Both of sites were applied fertilizers at two different rates, one the highest traditionally used by farmers in the region (about 4800 kg N/ha per year, HF) and the other three-eighth of the farmer (1800 kg N/ha per year, LF). The results showed that fertilization practices impacted few effects on the balance and dynamic of water in the plant-soil-aeration zone-saturated zone system. However, groundwater table controlled vertical infiltration recharge and evaporation–transpiration rate. The vertical infiltration recharge and the evaporation–transpiration rate were averagely 0.514 and 5.897 mm/d at Site B with a water table depth of 0.5 m below the soil surface, 0.335 and 6.420 mm/d at Site A with a water table depth of 2.0 m below the soil surface, respectively. Nitrate concentrations of soil water near groundwater table under HF subplot were much higher than that under LH subplot. High fertilization rate consequently resulted in great nitrogen (including nitrate,
* Corresponding author: Tel.: +86 10 82320552; fax: +86 10 82321081. E-mail addresses: [email protected], [email protected] (H. Guo). 0378-3774/$ – see front matter # 2005 Elsevier B.V. All rights reserved. doi:10.1016/j.agwat.2005.07.033
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nitrite and ammonium) loadings from aeration zone to groundwater. At Site B, nitrogen loadings were 316.03 and 223.89 kg/ha a under HF and LF, respectively. Nitrate was the dominant nitrogen component entering groundwater. Little ammonium and less nitrite transported into groundwater. Shallow water table made nitrate entering groundwater more easily and consequently determined the NO3 loading from vadose zone. For the same fertilization rate, nitrate loading to groundwater under Site B were much higher than those under Site A, with 47.11 kg NO3-N/ha a under Site A-HF and 311.73 kg NO3-N/ha a under Site B-HF. To avert or minimize the potential of groundwater nitrogen contamination in irrigated agricultural areas should determine and minimize the amounts of applied fertilizer by optimizing them to match crop requirements and environmental protection. # 2005 Elsevier B.V. All rights reserved. Keywords: Groundwater contamination; Nitrate loading; Water balance; Nitrogen fertilization; Water table
1. Introduction Groundwater contamination by nitrate associated with fertilization practices is a ubiquitous environmental issue, and consequently of world-wide concern (Holden et al., 1992; Hamilton and Helsel, 1995; Trauth and Xanthopoulos, 1997; Huang et al., 2000). Much knowledge has accumulated on the climate, soil, topography, land use and nitrogen management factors influencing the risk associated with nitrate leaching (Zebarth et al., 1998; Babiker et al., 2004). When utilization of animal manure and inorganic N fertilizers is incomplete or inefficient, or when water is applied in a pre-determined surplus, the resultant seepage ultimately deteriorates groundwater if shallow water table exists or in recharging the aquifers beneath the cropped land (Allison et al., 1983; Addiscott and Wagenet, 1985; Milburn et al., 1990; Randall and Irigavarapu, 1995). The amounts of water and leachates vary with irrigation efficiency, crop utilization of water and fertilizers, decomposition of the added organic materials, and absorption of the decomposed fractions. Great effort is directed to study the rate of recharge and the measures required to minimize the loads of contaminants seeping from agricultural areas toward the aquifers (Allison et al., 1983; Nielsen et al., 1986; Phillips, 1994). Little data is available on the nitrogen loads on groundwater under different groundwater table fields subjected to applying the different amounts of fertilizers. Groundwater table would dominate water budget in the agric-ecological system at the humid–semihumid area. There was relatively more irrigation water percolating into groundwater at irrigated agricultural area with a shallow groundwater table. While fertilizer application, which aimed to increase agricultural output, would affect the nitrogen balance in the system, and eventually result in remains of fertilizer nitrogen in the soils and soil water (Costa et al., 2002; Kellman and Hillaire-Marcel, 2003). Therefore, reducing loss of N-fertilizers and controlling groundwater pollution by agricultural activities could be carried out by effectively management of groundwater table and the amount of N-fertilizers applied for agriculture (Elmi et al., 2002). Our objective was to determine NO3, NO2, and NH4+ loadings to groundwater under irrigated vegetable fields with different groundwater table, at different nitrogen (N)-
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fertilizer rates, one the highest traditionally used by farmers in the region (about 4800 kg N/ha per year) and the other three-eighth of the farmer (1800 kg N/ha per year). We estimated nitrogen loadings from aeration zone under celery (Apium graveolens) greenhouses to groundwater using vertical infiltration recharges and nitrogen concentrations of waters at groundwater tables, based on intensive monitoring of the soil water content, soil water potential and extraction of the soil solution by a combination of time domain reflectometry (TDR), tensiometers and ceramic suction cups under the study fields.
2. Materials and methods 2.1. Experimental sites Experimental sites, located in approximately 25 km southeast of Kunming city in southwestern China, lie on the east shore of Dianchi Lake, one of the largest lakes in Yunnan province (Fig. 1). Mean temperature at this area was 14.7 8C with a maximum of
Fig. 1. Experimental site in Dianchi catchment, China.
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Table 1 Hydraulic characteristic of soils at different depths at Site A Depth (cm)
Soil type
Conductivity (m/d)
Porosity (%)
Density (103 kg/m3)
0–30 30–60 60–100 100–180 180–280
Arable layer, brown clay Brown clay, few gravels Brown clay, layered with gravels Yellow clay Yellow clay, silty clay
10.36 1.40 2.29 0.02 0.04
52.88 47.69 47.85 48.75 46.83
1.23 1.36 1.36 1.33 1.38
31 8C and minimum of 8.1 8C. Mean annual precipitation was 782.5 mm, more than ninety percent of which falls from May to October. The pore medium was generally heterogeneous in vertical and homogeneous in horizontal at experimental Site A (Fig. 1), with a groundwater table depth of 2.0 m below the soil surface. Due to cultivation practices, the arable layer had a high porosity with hydraulic conductivity of 10.36 m/d. The soils were compressed more tightly with increasing of depth due to long-term geological stress. Although the soils contained some gravels at depths between 30 and 100 cm, those soils had low hydraulic conductivity and less porosity. Hydraulic conductivity of the soils was 1.40 m/d at depths between 30 and 60 cm, and 2.29 m/d at depths between 60 and 100 cm. Hydraulic conductivities were between 0.02 and 0.04 m/d at depths between 100 and 280 cm (Table 1). At the experimental Site B (Fig. 1) with a groundwater table depth of 0.5 m below the soil surface, the pore medium was generally heterogeneous in vertical and homogeneous in horizontal as well. The soils’ characteristics at different depths were shown in Table 2. Site A is located near the foot of highland, while Site B is near the shore of Dianchi Lake (Fig. 1). Due to intensive agricultural activities in the study area with intensive irrigation, groundwater was recharged mainly by vertically infiltrating of irrigation water and partly by laterally penetrating water from the highland. It generally flowed from the highland to Dianchi Lake and eventually discharged into Dianchi Lake. 2.2. Crop management Field experiments were conducted in two greenhouses (Sites A and B, respectively). Each greenhouse was used to crop celeries, with an area of 46.5 m 4.0 m. The pilot occupied an area of 7.0 m 4.0 m within the greenhouse. In order to establish two Nfertilization treatments for each plot, each plot was divided into two subplots, including high fertilization subplot (HF) and low fertilization subplot (LF), with an area of Table 2 Hydraulic characteristic of soils at different depths at Site B Depth (cm)
Soil type
Conductivity (m/d)
Porosity (%)
Density (103 kg/m3)
0–10 10–30 30–80 80–100
Arable layer, brown clay Brown silty clay Brown clay Yellow silt
5.03 0.20 0.15 0.26
49.61 47.23 48.55 46.95
1.25 1.39 1.38 1.41
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Table 3 Fertilizations for the field experiment at Dianchi catchment, Kunming city N (kg/ha a) a
Sum Deep fertilization Top application 1 2 3 4 a b c
P2O5 (kg/ha a) b
K2O (kg/ha a) c
LF
HF
LF
HF
LF
HF
1800 180 1620 162 324 648 486
4800 480 4320 432 864 1728 1296
900 540 360 72 144 144 0
3600 2160 1440 288 576 576 0
1440 480 960 120 240 360 240
1440 480 960 120 240 360 240
N fertilizer was used as urea. P2O5 fertilizer was used as concentrated superphosphate. K2O fertilizer was used as potassium sulphate.
1.8 m 7.0 m for each subplot. Fertilization rates were presented in Table 3. Rates of N fertilizer as urea were 4800 and 1800 kg N/ha per year, representing the traditional N rate used in this area (HF) and the balanced fertilization N rate for celeries in the region (LF). Rates of P fertilizer as concentrated superphosphate were 3600 and 900 kg P2O5/ha per year in HF and LF, respectively. The same amount of K was applied to all subplots as potassium sulfate (1440 kg K2O/ha per year). Fertilization was applied at five times, one deep fertilization about 5 days before planting, and four top dressings at about 10, 30, 50, and 70 days after planting. The celeries at all subplots were irrigated at the rate traditionally used by farmers in the region. Irrigation intensity was similar for all subplots. The irrigation was generally carried out once every three days, with a fixed value of 28.6 mm. Guo et al. (2004) described the irrigation intensity in more detailed. The land surrounding the experimental plots was cropped with the same irrigation, which minimized the advection. 2.3. Measurements Prior to celery (A. graveolens) planting, TDR (TRIME-T3) and WM-1 tensiometers were installed. TDR was used to monitor soil moistures with measured ranges of 0 and 60% in volumetric moisture content and observed precision of 2% (Brandelik and Hubner, 1996; Musters and Bouten, 2000), and WM-1 was used to monitor matrix potential (Jing et al., 1994). TDR and WM-1 were performed once every day during celery cultivation. Simultaneously, ceramic suction cups were set up at locations corresponding to TDR and WM-1. In order to avoid disturbing soil structure, the soils were carefully dug out and replaced at their original locations after installation of experimental equipments. A well was installed at each experimental site, which was used to monitor groundwater table during the experiment. Soil waters were sampled generally once every week. Samples were collected in 650 mL glass bottles. These bottles had been rinsed with deionised water before sampling. NO3, NO2 and NH4+ were measured by colorimetry within 1 day after sampling.
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2.4. Determination of water balance and nitrogen loadings The zero flux-plane method was used to calculate vertical infiltration recharge and evaporation–transpiration, which was based on the soil water hydraulic theory that the water flux did not exist when a gradient of total potential of soil water was zero (Vachaud et al., 1978). On the zero flux plane, soil water complied with the following formula: @f ¼0 (1) @z where q represents water flux; u represents soil moisture content; K(u) represents hydraulic conductivity; f represents total potential of soil water; z represents the depth. Irrigation was uniform throughout the experimental site and the surface of the site was flat. Therefore, lateral flow of soil water could be considered to be negligible. The formula can be adapted to describe vertical water equilibrium for the quantities of our experiment. Z z @u 0 qðzÞ qðz Þ ¼ dz (2) 0 z @t Z z Z z QðzÞ Qðz0 Þ ¼ uðz; t1 Þdz uðz; t2 Þdz (3) q ¼ KðuÞ
z0
z0
0
where q(z) and q(z ) represent the water flux at depths z and z0 , respectively; Q(z) and Q(z0 ) represent the quantity of water through the plane at depth of z and z0 , respectively; u(z, t1) and u(z, t2) represent volumetric moisture content of soil at t1 and t2, respectively. When the zero flux plane was at a depth of z0 , q(z0 ) and Q(z0 ) were equal to zero. Eqs. (2) and (3) could be simplified as: Z z @u dz (4) qðzÞ ¼ 0 z @t Z z Z z QðzÞ ¼ uðz; t1 Þdz uðz; t2 Þdz: (5) z0
z0
The zero flux plane could be achieved by using total potential data in soil profiles obtained by WM-1 tensiometers. If the zero flux plane existed, flux and quantity of water through any plane at depth of z could be calculated by using Eqs. (4) and (5), respectively, based on moisture contents in soil profiles measured by TDR. When z was located near the groundwater table, Q(z) was considered to be the vertical infiltration recharge. When z was located near the land surface, Q(z) was considered to be the evaporation–transpiration rate. When the zero flux plane existed, the amounts of nitrogen (including nitrate, nitrite and ammonium) leached (LN) into groundwater at water table, were obtained from the equation: LN ¼ QðzÞCz
(6)
where Q(z) is the water flux calculated at water table (z) from Eq. (5) and Cz the nitrogen (including nitrate, nitrite and ammonium) concentration in the soil solution sampled by suction cups near the groundwater table.
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3. Results and discussions 3.1. Water balance at aeration zone For Site A with a deep groundwater table of 2.0 m below the soil surface (Guo et al., 2004), the zero flux plane was determined at depth of 0.6 m. Therefore, Z0 and Z were 0.6 and 2.0 m, respectively. Finally, the vertical infiltration recharge was calculated and evaluated as an average of 0.335 mm/d at two subplots with 0.330 mm/d at HF and 0.340 mm/d at LF, and the evaporation–transpiration rate as an average of 6.420 mm/d at two subplots with 6.463 mm/d at HF and 6.380 mm/d at LF (Table 4). The results demonstrated that fertilization practices impacted few effects on the balance and dynamic of water in the plant-soil-aeration zone-saturated zone system. Moreno et al. (1996) also concluded that, in terms of water balance, crop evapotranspiration was similar at both high N-fertilization rate (about 500 kg N/ha per year) and low N-fertilization rate (170 kg N/ha per year) when they conducted a field experiment on a furrow-irrigated maize crop. Irrigation intensity was averagely 8.276 mm/d during celery cultivation in Site A. The calculated output from the aeration zone was generally 6.755 mm/d, including vertical infiltration recharge and evaporation–transpiration. In other words, 4.05% of irrigation was averagely infiltrated into groundwater, and 77.57% of irrigation water was averagely evaporated and transpired through plants and surface soils. The derivation between the calculated value and the observed value was 18.38%, which probably resulted from experimental errors and observed precision, as well as water stored in soils. For Site B with a shallow groundwater table, the vertical infiltration recharge and the evaporation–transpiration rate were averagely 0.514 mm/d (0.493 mm/d at HF and 0.535 mm/d at LF) and 5.897 mm/d (6.010 mm/d at HF and 5.784 mm/d at LF), respectively. The derivation between the calculated value and the observed value was 22.54%, which was greater than that at Site A with a deep groundwater table. 3.2. Nitrate, nitrite and ammonium concentrations 3.2.1. NO3-N NO3 concentrations of soil waters near the water table at two subplots at Site A were shown in Fig. 2, which demonstrated that nitrate concentration under HF was much higher than that under LH. For Site A, NO3-N concentrations of soil waters near water table generally increased during celery cultivation. Especially under HF, NO3-N increased from 18.52 to 62.68 mg/L from March 4 to May 28. Moreno et al. (1996) also demonstrated that Table 4 Total water input (mainly irrigation), vertical infiltration recharge and evapotranspiration at Sites A and B Site A
Total water input (mm/d) Vertical infiltration recharge (mm/d) Evapotranspiration (mm/d)
Site B
HF
LF
HF
LF
8.276 0.330 6.463
8.276 0.340 6.380
8.276 0.493 6.010
8.276 0.535 5.784
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Fig. 2. NO3-N concentrations of soil waters near groundwater table under Site A. Solid arrows denote fertilizer N application.
the NO3-N contents in the profile for high N-fertilization site were much higher than those for the low N-fertilization site. High fertilization rate would load much more nitrate on groundwater than low fertilization rate. The NO3 concentrations of soil waters at Site B had the same trend as those at Site A (Fig. 3). Although there were the same fertilization, irrigation and cultivation, and similar soil textures at HF or LF between Sites A and B, nitrate concentrations near groundwater at Site B were much higher than those at Site A. The reason was that nitrate would enter groundwater more easily in the area with a shallow water table. Various natural physical processes, chemical reactions and microbial activities that occurred in the soil and unsaturated zone could cause contaminants to change their physical states and their chemical forms, which led to remediation of the groundwater pollution or change the characteristics of contaminants (Guo and Wang, 2004). The thicker the unsaturated zone, the more likely nitrate was degraded when it transported from surface to groundwater. In comparison with Figs. 2 and 3, we could find that the difference in NO3-N concentration of soil waters near water table between LF and HF at Site B was smaller than in Site A. This difference possibly resulted from limited natural attenuation of nitrate in the soil and unsaturated zone at both Sites A and B. It could be speculated that
Fig. 3. NO3-N concentrations of soil waters near groundwater table under Site B. Solid arrows denote fertilizer N application.
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relatively low concentration of nitrate of soil water under Site A-LF had been naturally degraded when it seeped from agricultural areas toward the aquifers, while relatively higher concentration of nitrate under Site A-HF exceeded the potential of natural attenuation and possibly restrained some microbial activities in the soil. This degradation consequently caused the relatively great difference in NO3-N concentration of soil waters between LF and HF at Site A. On the other hand, at Site B much higher concentration of NO3-N under both HF and LF far exceeded the low potential of natural attenuation of nitrate as the consequence of the thinnest unsaturated zone. Therefore, natural degradation had little effects on the decrease of nitrate during its transport in the soil at Site B, which resulted in the relatively small difference of NO3-N concentration of soil waters between LF and HF at Site B. 3.2.2. NO2-N and NH4-N NO2-N and NH4-N of soil waters were relatively low, with the maximum NO2-N of 2.47 mg/L and NH4-N of 4.00 mg/L for all subplots. Figs. 4 and 5 showed NO2-N and NH4N of soil waters near groundwater table at Sites A and B, respectively. As a whole, NO2 concentration under HF was greater than that under LF at both Sites A and B. The nitrite in soil water was mainly controlled by microbial activities. Denitrification required oxidized N (e.g. nitrate), available organic carbon and an absence of oxygen. The supply of nitrate or readily available organic C could limit denitrification rates in the surface soils (Groffman et al., 1991; Ellis et al., 1996; Sa´nchez et al., 2001). N-fertilizer application which increased nitrate content in the soil was one of the dominant factors promoting denitrification. Therefore, the nitrite concentration under HF was higher than that under LF for both Sites A and B. However, NH4+ concentration under HF of Site A was generally lower than that under LF of Site A, while NH4+ concentration under HF of Site B was generally greater than that under LF of Site B. In addition to microbial activities, adsorption of soils would dominate NH4+ content in soil waters. Heterogeneous adsorptive capacity of soils at Site A would contribute the difference of NH4+ concentration between HF and LF. Groundwater table impacted less effect on NO2 and NH4+ concentrations near water table. In contrast with NO3, the difference of NO2 and NH4+ concentrations in soil
Fig. 4. NO2-N and NH4-N of soil waters near groundwater table at Site A. Solid arrows denote fertilizer N application.
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Fig. 5. NO2-N and NH4-N of soil waters near groundwater table at Site B. Solid arrows denote fertilizer N application.
waters near water table between Sites A and B was much smaller (Figs. 4 and 5). For NH4+, the possible reason was that soil absorption of NH4+ accounted for the relatively uniform NH4+ concentration at Sites A and B. Although water table affected denitrification in the soils, intermediate product (such as NO2) of the reaction kept relatively constant in the study area. Possibly, specific denitrification process converting NO2 to NO (or N2O, N2) preferably existed in Site B and consequently contributed to the small difference in NO2 concentration between Sites A and B. 3.3. N loadings from aeration zone to groundwater Based on the vertical infiltration recharge from the unsaturated to the saturated zone and concentrations of NO3, NO2 and NH4+ near groundwater table, nitrogen loadings to groundwater under different experimental subplots were calculated, according to Eq. (6). The results were shown in Table 5. Nitrate loading was 32–404 times more than nitrite loading and 9–162 times more than ammonium loading, which demonstrated that nitrate was the dominant nitrogen component entering groundwater from vadose zone in agricultural areas. Power and Schepers (1989) and Spalding and Exner (1993) also reported that groundwater nitrate contamination was a major environmental problem in irrigated agriculture areas. Table 5 N loadings to groundwater under different subplots (kg/ha a) NO3-N loading
NO2-N loading
NH4-N loading
Total N loading
N-fertilizer applied
Percentage of N loadings to fertilizer N (%)
Site A
HF LF
47.11 13.46
1.24 0.41
0.29 1.35
48.64 15.22
4800 1800
1.01 0.85
Site B
HF LF
311.73 221.30
0.77 0.63
3.53 1.96
316.03 223.89
4800 1800
6.58 12.44
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With regard to the same groundwater table, the NO3-N loading under HF was higher than that under LF, which indicated that application of abundant fertilizers would result in severe loss of NO3-N with an increasing contamination of groundwater. Granlund et al. (2000) also indicated that nitrate leaching increased if fertilization rates increased. Therefore, to reduce the potential of groundwater contamination in agricultural areas, the amount of fertilizers applied must effectively be minimized in view of systematic combination of agricultural production and environmental protection. Hadas et al. (1999) showed that intensification of agricultural activities led to increased hazards to surface and groundwater pollution and this could be diminished by balanced irrigation–fertilization programs. Although the amount of applied fertilizers at HF of Site A was the same as that at HF of Site B, NO3-N loading at Site B was 5.62 times more than that at Site A. Because groundwater table significantly controlled the vertical infiltration recharge and definitely affected degradation of nitrate at unsaturated zone, it eventually determined the NO3 input from vadose zone to groundwater. In analogical hydrogeological settings, shallow water table favoured the contaminant loadings to groundwater. Delin and Landon (2002) also observed that the estimated mass flux of chloride and nitrate to the water table were five to two times greater, respectively, at the lowland site compared to the upland site. Kraft and Stites (2003) found that nitrate-N loading to shallow groundwater from the sweet corn ranged from 126 to 169 kg/ha a; loading from potato was 228 kg/ha a, which amounted to 61% of total available N and 77% of fertilizer N in the Wisconsin Central Sand Plain possessing a humid climate, coarse soil, and shallow water table. Consequently, lowering groundwater table would mitigate contamination of groundwater below the soils contaminated by nitrate. It could be applied to enhance drastically microbial degradation of nitrate and further reduce the potential of groundwater nitrate contamination at nitratecontaminated sites.
4. Conclusions The results demonstrated that fertilization practices impacted few effects on the balance and dynamic of water in the plant-soil-aeration zone-saturated zone system. However, groundwater table controlled vertical infiltration recharge and evaporation–transpiration rate. The vertical infiltration recharges were 0.330 and 0.340 mm/d at HF and LF of Site A, 0.493 and 0.535 mm/d at HF and LF of Site B, respectively. The evaporation–transpiration rates were 6.463, 6.380, 6.010 and 5.784 mm/d at Site A-HF, Site A-LF, Site B-HF and Site B-LF, respectively. Nitrate was the dominant nitrogen component entering groundwater from vadose zone at agricultural areas. Little ammonium and less nitrite transported into groundwater. Nitrate concentrations of soil water near groundwater table under HF subplot were much higher than that under LH subplot. The averages of NO3-N were 49.17 and 14.50 mg/L under HF and LF of Site A, respectively. Application of abundant fertilizers would result in severe loss of NO3-N with increasing contamination of groundwater. At Site B, for example, nitrogen loading was 316.03 kg/ha a under HF. Groundwater table determined the NO3 loading to groundwater. For the same N-fertilization rate, nitrate loading at Site B with a
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shallow groundwater table was much higher than that at Site A with a deep groundwater table. For HF, nitrate loading was 47.11 kg NO3-N/ha a under Site A and 311.73 kg NO3N/ha a under Site B. The nitrate loading to groundwater in the Dianchi catchment had resulted in nitrate contamination of groundwater, NO3-N concentration of which was generally 5.3–15.8 mg/ L. Because of recharging the Dianchi Lake, contaminated groundwater contributed partly to the lake eutrophication.
Acknowledgements We thank two anonymous referees for their valuable comments on the manuscript. Funding by China Postdoctoral Science Foundation (2003034021), China’s National Key Technologies R&D Program in the 10th Five-Year Plan (No. 2001BA610A-04) is also gratefully acknowledged.
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Holden, I.R., Graham, J.A., Alexander, W.J., Pratt, R., Liddle, S.K., Piper, L.L., 1992. Results of the national alachlor well water survey. Environ. Sci. Technol. 26, 935–943. Huang, S.M., Zhang, H.C., Bao, D.J., Huangpu, X.R., 2000. Effcet of applying fertilizer on content and distribution of NO3-N in chao soil and reasonable fertilzer application. Soil Environ. Sci. 9 (3), 201–203 (in Chinese with English abstract). Jing, E., Fei, J., Zhang, X., Han, S., Xu, W., 1994. Experimental Study on Soilwater Flux. Seismal Press, Beijing, pp. 138–148 (in Chinese). Kellman, L.M., Hillaire-Marcel, C., 2003. Evaluation of nitrogen isotopes as indicators of nitrate contamination sources in an agricultural watershed. Agric. Ecosyst. Environ. 95, 87–102. Kraft, G.J., Stites, W., 2003. Nitrate impacts on groundwater from irrigated-vegetable systems in a humid northcentral US sand plain. Agric. Ecosyst. Environ. 100 (1), 63–74. Milburn, P., Richards, J.E., Gartley, C., Pollock, T., O’Neil, H., Bailcy, H., 1990. Nitarte leaching from systematically tiled potato fields in New Brunswick. Can. J. Environ. Qual. 19, 448–454. Moreno, F., Cayuela, J.A., Ferna´ndez, J.E., Ferna´ndez-Boy, E., Murillo, J.M., Cabrera, F., 1996. Water balance and nitrate leaching in an irrigated maize crop in SW Spain. Agric. Water Manag. 32, 71–83. Musters, P.A.D., Bouten, W., 2000. A method for identifying optimum strategies of measuring soil water contents for calibrating a root water uptake model. J. Hydrol. 227, 273–286. Nielsen, D.R., Van Genuchten, M.Th., Biggar, J.W., 1986. Water flow and solute transport processes in the unsaturated zone. Water Resour. Res. 22, 89–108. Phillips, F.M., 1994. Environmental tracers for water movement in desert soils in the American Southwest. Soil Sci. Soc. Am. J. 58, 15–24. Power, J.F., Schepers, J.S., 1989. Nitrate contamination of groundwater in North America. Agric. Ecosyst. Environ. 26, 165–187. Randall, G.W., Irigavarapu, T.K., 1995. Impact of long-term tillage systems for continuous corn on nitrate leaching to tile drainage. J. Environ. Qual. 24, 360–366. Sa´nchez, L., Dı´ez, J.A., Vallejo, A., Cartagena, M.C., 2001. Denitrification losses from irrigated crops in central Spain. Soil Biol. Biochem. 33, 1201–1209. Spalding, R.F., Exner, M.E., 1993. Occurrence of nitrate in groundwater: A review. J. Environ. Qual. 22, 392–402. Trauth, R., Xanthopoulos, C., 1997. Non-point pollution of groundwater in urban areas. Water Res. 31 (11), 2711– 2718. Vachaud, G., Dancette, C., Sonko, M., Thony, J.L., 1978. Me´thodes de caracte´risation hydrodynamique in-situ d’un sol non sature´. Application a´ deux types de sol du se´ne´gal en vue de la de´termination des termes du bilan hydrique. Ann. Agron. 29, 1–36. Zebarth, B.J., Hii, B., Liebscher, H., Chipperfield, K., Paul, J.W., Grove, G., Szeto, S.Y., 1998. Agricultural land use practices and nitrate contamination in the Abbotsford Aquifer, British Columbia. Can. Agric. Ecosyst. Environ. 69, 99–112.
Effects of water table and fertilization management on nitrogen loading to groundwater Huaming Guo a,b,*, Guanghe Li a, Dayi Zhang a, Xu Zhang a, Chang’ai Lu c a b
Department of Environmental Science and Engineering, Tsinghua University, Beijing 100084, PR China School of Water Resource and Environment, China University of Geosciences, Beijing 100083, PR China c Chinese Academy of Agricultural Science, Beijing 100081, PR China Accepted 25 July 2005 Available online 12 September 2005
Abstract Groundwater contamination by nitrate associated with fertilization practices is a ubiquitous environmental issue, and consequently of world-wide concern. Controlling this contamination requires an ability to measure and predict nitrate loading from unsaturated zone to saturated zone. A field experiment was conducted in an intensively irrigated agricultural area in Dianchi catchment of Kunming, China. Two celery (Apium graveolens) crop sites with different water table depths (Site A is 2.0 m below the soil surface; Site B is 0.5 m below the soil surface) were selected for the experiment. Both of sites were applied fertilizers at two different rates, one the highest traditionally used by farmers in the region (about 4800 kg N/ha per year, HF) and the other three-eighth of the farmer (1800 kg N/ha per year, LF). The results showed that fertilization practices impacted few effects on the balance and dynamic of water in the plant-soil-aeration zone-saturated zone system. However, groundwater table controlled vertical infiltration recharge and evaporation–transpiration rate. The vertical infiltration recharge and the evaporation–transpiration rate were averagely 0.514 and 5.897 mm/d at Site B with a water table depth of 0.5 m below the soil surface, 0.335 and 6.420 mm/d at Site A with a water table depth of 2.0 m below the soil surface, respectively. Nitrate concentrations of soil water near groundwater table under HF subplot were much higher than that under LH subplot. High fertilization rate consequently resulted in great nitrogen (including nitrate,
* Corresponding author: Tel.: +86 10 82320552; fax: +86 10 82321081. E-mail addresses: [email protected], [email protected] (H. Guo). 0378-3774/$ – see front matter # 2005 Elsevier B.V. All rights reserved. doi:10.1016/j.agwat.2005.07.033
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nitrite and ammonium) loadings from aeration zone to groundwater. At Site B, nitrogen loadings were 316.03 and 223.89 kg/ha a under HF and LF, respectively. Nitrate was the dominant nitrogen component entering groundwater. Little ammonium and less nitrite transported into groundwater. Shallow water table made nitrate entering groundwater more easily and consequently determined the NO3 loading from vadose zone. For the same fertilization rate, nitrate loading to groundwater under Site B were much higher than those under Site A, with 47.11 kg NO3-N/ha a under Site A-HF and 311.73 kg NO3-N/ha a under Site B-HF. To avert or minimize the potential of groundwater nitrogen contamination in irrigated agricultural areas should determine and minimize the amounts of applied fertilizer by optimizing them to match crop requirements and environmental protection. # 2005 Elsevier B.V. All rights reserved. Keywords: Groundwater contamination; Nitrate loading; Water balance; Nitrogen fertilization; Water table
1. Introduction Groundwater contamination by nitrate associated with fertilization practices is a ubiquitous environmental issue, and consequently of world-wide concern (Holden et al., 1992; Hamilton and Helsel, 1995; Trauth and Xanthopoulos, 1997; Huang et al., 2000). Much knowledge has accumulated on the climate, soil, topography, land use and nitrogen management factors influencing the risk associated with nitrate leaching (Zebarth et al., 1998; Babiker et al., 2004). When utilization of animal manure and inorganic N fertilizers is incomplete or inefficient, or when water is applied in a pre-determined surplus, the resultant seepage ultimately deteriorates groundwater if shallow water table exists or in recharging the aquifers beneath the cropped land (Allison et al., 1983; Addiscott and Wagenet, 1985; Milburn et al., 1990; Randall and Irigavarapu, 1995). The amounts of water and leachates vary with irrigation efficiency, crop utilization of water and fertilizers, decomposition of the added organic materials, and absorption of the decomposed fractions. Great effort is directed to study the rate of recharge and the measures required to minimize the loads of contaminants seeping from agricultural areas toward the aquifers (Allison et al., 1983; Nielsen et al., 1986; Phillips, 1994). Little data is available on the nitrogen loads on groundwater under different groundwater table fields subjected to applying the different amounts of fertilizers. Groundwater table would dominate water budget in the agric-ecological system at the humid–semihumid area. There was relatively more irrigation water percolating into groundwater at irrigated agricultural area with a shallow groundwater table. While fertilizer application, which aimed to increase agricultural output, would affect the nitrogen balance in the system, and eventually result in remains of fertilizer nitrogen in the soils and soil water (Costa et al., 2002; Kellman and Hillaire-Marcel, 2003). Therefore, reducing loss of N-fertilizers and controlling groundwater pollution by agricultural activities could be carried out by effectively management of groundwater table and the amount of N-fertilizers applied for agriculture (Elmi et al., 2002). Our objective was to determine NO3, NO2, and NH4+ loadings to groundwater under irrigated vegetable fields with different groundwater table, at different nitrogen (N)-
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fertilizer rates, one the highest traditionally used by farmers in the region (about 4800 kg N/ha per year) and the other three-eighth of the farmer (1800 kg N/ha per year). We estimated nitrogen loadings from aeration zone under celery (Apium graveolens) greenhouses to groundwater using vertical infiltration recharges and nitrogen concentrations of waters at groundwater tables, based on intensive monitoring of the soil water content, soil water potential and extraction of the soil solution by a combination of time domain reflectometry (TDR), tensiometers and ceramic suction cups under the study fields.
2. Materials and methods 2.1. Experimental sites Experimental sites, located in approximately 25 km southeast of Kunming city in southwestern China, lie on the east shore of Dianchi Lake, one of the largest lakes in Yunnan province (Fig. 1). Mean temperature at this area was 14.7 8C with a maximum of
Fig. 1. Experimental site in Dianchi catchment, China.
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Table 1 Hydraulic characteristic of soils at different depths at Site A Depth (cm)
Soil type
Conductivity (m/d)
Porosity (%)
Density (103 kg/m3)
0–30 30–60 60–100 100–180 180–280
Arable layer, brown clay Brown clay, few gravels Brown clay, layered with gravels Yellow clay Yellow clay, silty clay
10.36 1.40 2.29 0.02 0.04
52.88 47.69 47.85 48.75 46.83
1.23 1.36 1.36 1.33 1.38
31 8C and minimum of 8.1 8C. Mean annual precipitation was 782.5 mm, more than ninety percent of which falls from May to October. The pore medium was generally heterogeneous in vertical and homogeneous in horizontal at experimental Site A (Fig. 1), with a groundwater table depth of 2.0 m below the soil surface. Due to cultivation practices, the arable layer had a high porosity with hydraulic conductivity of 10.36 m/d. The soils were compressed more tightly with increasing of depth due to long-term geological stress. Although the soils contained some gravels at depths between 30 and 100 cm, those soils had low hydraulic conductivity and less porosity. Hydraulic conductivity of the soils was 1.40 m/d at depths between 30 and 60 cm, and 2.29 m/d at depths between 60 and 100 cm. Hydraulic conductivities were between 0.02 and 0.04 m/d at depths between 100 and 280 cm (Table 1). At the experimental Site B (Fig. 1) with a groundwater table depth of 0.5 m below the soil surface, the pore medium was generally heterogeneous in vertical and homogeneous in horizontal as well. The soils’ characteristics at different depths were shown in Table 2. Site A is located near the foot of highland, while Site B is near the shore of Dianchi Lake (Fig. 1). Due to intensive agricultural activities in the study area with intensive irrigation, groundwater was recharged mainly by vertically infiltrating of irrigation water and partly by laterally penetrating water from the highland. It generally flowed from the highland to Dianchi Lake and eventually discharged into Dianchi Lake. 2.2. Crop management Field experiments were conducted in two greenhouses (Sites A and B, respectively). Each greenhouse was used to crop celeries, with an area of 46.5 m 4.0 m. The pilot occupied an area of 7.0 m 4.0 m within the greenhouse. In order to establish two Nfertilization treatments for each plot, each plot was divided into two subplots, including high fertilization subplot (HF) and low fertilization subplot (LF), with an area of Table 2 Hydraulic characteristic of soils at different depths at Site B Depth (cm)
Soil type
Conductivity (m/d)
Porosity (%)
Density (103 kg/m3)
0–10 10–30 30–80 80–100
Arable layer, brown clay Brown silty clay Brown clay Yellow silt
5.03 0.20 0.15 0.26
49.61 47.23 48.55 46.95
1.25 1.39 1.38 1.41
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Table 3 Fertilizations for the field experiment at Dianchi catchment, Kunming city N (kg/ha a) a
Sum Deep fertilization Top application 1 2 3 4 a b c
P2O5 (kg/ha a) b
K2O (kg/ha a) c
LF
HF
LF
HF
LF
HF
1800 180 1620 162 324 648 486
4800 480 4320 432 864 1728 1296
900 540 360 72 144 144 0
3600 2160 1440 288 576 576 0
1440 480 960 120 240 360 240
1440 480 960 120 240 360 240
N fertilizer was used as urea. P2O5 fertilizer was used as concentrated superphosphate. K2O fertilizer was used as potassium sulphate.
1.8 m 7.0 m for each subplot. Fertilization rates were presented in Table 3. Rates of N fertilizer as urea were 4800 and 1800 kg N/ha per year, representing the traditional N rate used in this area (HF) and the balanced fertilization N rate for celeries in the region (LF). Rates of P fertilizer as concentrated superphosphate were 3600 and 900 kg P2O5/ha per year in HF and LF, respectively. The same amount of K was applied to all subplots as potassium sulfate (1440 kg K2O/ha per year). Fertilization was applied at five times, one deep fertilization about 5 days before planting, and four top dressings at about 10, 30, 50, and 70 days after planting. The celeries at all subplots were irrigated at the rate traditionally used by farmers in the region. Irrigation intensity was similar for all subplots. The irrigation was generally carried out once every three days, with a fixed value of 28.6 mm. Guo et al. (2004) described the irrigation intensity in more detailed. The land surrounding the experimental plots was cropped with the same irrigation, which minimized the advection. 2.3. Measurements Prior to celery (A. graveolens) planting, TDR (TRIME-T3) and WM-1 tensiometers were installed. TDR was used to monitor soil moistures with measured ranges of 0 and 60% in volumetric moisture content and observed precision of 2% (Brandelik and Hubner, 1996; Musters and Bouten, 2000), and WM-1 was used to monitor matrix potential (Jing et al., 1994). TDR and WM-1 were performed once every day during celery cultivation. Simultaneously, ceramic suction cups were set up at locations corresponding to TDR and WM-1. In order to avoid disturbing soil structure, the soils were carefully dug out and replaced at their original locations after installation of experimental equipments. A well was installed at each experimental site, which was used to monitor groundwater table during the experiment. Soil waters were sampled generally once every week. Samples were collected in 650 mL glass bottles. These bottles had been rinsed with deionised water before sampling. NO3, NO2 and NH4+ were measured by colorimetry within 1 day after sampling.
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2.4. Determination of water balance and nitrogen loadings The zero flux-plane method was used to calculate vertical infiltration recharge and evaporation–transpiration, which was based on the soil water hydraulic theory that the water flux did not exist when a gradient of total potential of soil water was zero (Vachaud et al., 1978). On the zero flux plane, soil water complied with the following formula: @f ¼0 (1) @z where q represents water flux; u represents soil moisture content; K(u) represents hydraulic conductivity; f represents total potential of soil water; z represents the depth. Irrigation was uniform throughout the experimental site and the surface of the site was flat. Therefore, lateral flow of soil water could be considered to be negligible. The formula can be adapted to describe vertical water equilibrium for the quantities of our experiment. Z z @u 0 qðzÞ qðz Þ ¼ dz (2) 0 z @t Z z Z z QðzÞ Qðz0 Þ ¼ uðz; t1 Þdz uðz; t2 Þdz (3) q ¼ KðuÞ
z0
z0
0
where q(z) and q(z ) represent the water flux at depths z and z0 , respectively; Q(z) and Q(z0 ) represent the quantity of water through the plane at depth of z and z0 , respectively; u(z, t1) and u(z, t2) represent volumetric moisture content of soil at t1 and t2, respectively. When the zero flux plane was at a depth of z0 , q(z0 ) and Q(z0 ) were equal to zero. Eqs. (2) and (3) could be simplified as: Z z @u dz (4) qðzÞ ¼ 0 z @t Z z Z z QðzÞ ¼ uðz; t1 Þdz uðz; t2 Þdz: (5) z0
z0
The zero flux plane could be achieved by using total potential data in soil profiles obtained by WM-1 tensiometers. If the zero flux plane existed, flux and quantity of water through any plane at depth of z could be calculated by using Eqs. (4) and (5), respectively, based on moisture contents in soil profiles measured by TDR. When z was located near the groundwater table, Q(z) was considered to be the vertical infiltration recharge. When z was located near the land surface, Q(z) was considered to be the evaporation–transpiration rate. When the zero flux plane existed, the amounts of nitrogen (including nitrate, nitrite and ammonium) leached (LN) into groundwater at water table, were obtained from the equation: LN ¼ QðzÞCz
(6)
where Q(z) is the water flux calculated at water table (z) from Eq. (5) and Cz the nitrogen (including nitrate, nitrite and ammonium) concentration in the soil solution sampled by suction cups near the groundwater table.
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3. Results and discussions 3.1. Water balance at aeration zone For Site A with a deep groundwater table of 2.0 m below the soil surface (Guo et al., 2004), the zero flux plane was determined at depth of 0.6 m. Therefore, Z0 and Z were 0.6 and 2.0 m, respectively. Finally, the vertical infiltration recharge was calculated and evaluated as an average of 0.335 mm/d at two subplots with 0.330 mm/d at HF and 0.340 mm/d at LF, and the evaporation–transpiration rate as an average of 6.420 mm/d at two subplots with 6.463 mm/d at HF and 6.380 mm/d at LF (Table 4). The results demonstrated that fertilization practices impacted few effects on the balance and dynamic of water in the plant-soil-aeration zone-saturated zone system. Moreno et al. (1996) also concluded that, in terms of water balance, crop evapotranspiration was similar at both high N-fertilization rate (about 500 kg N/ha per year) and low N-fertilization rate (170 kg N/ha per year) when they conducted a field experiment on a furrow-irrigated maize crop. Irrigation intensity was averagely 8.276 mm/d during celery cultivation in Site A. The calculated output from the aeration zone was generally 6.755 mm/d, including vertical infiltration recharge and evaporation–transpiration. In other words, 4.05% of irrigation was averagely infiltrated into groundwater, and 77.57% of irrigation water was averagely evaporated and transpired through plants and surface soils. The derivation between the calculated value and the observed value was 18.38%, which probably resulted from experimental errors and observed precision, as well as water stored in soils. For Site B with a shallow groundwater table, the vertical infiltration recharge and the evaporation–transpiration rate were averagely 0.514 mm/d (0.493 mm/d at HF and 0.535 mm/d at LF) and 5.897 mm/d (6.010 mm/d at HF and 5.784 mm/d at LF), respectively. The derivation between the calculated value and the observed value was 22.54%, which was greater than that at Site A with a deep groundwater table. 3.2. Nitrate, nitrite and ammonium concentrations 3.2.1. NO3-N NO3 concentrations of soil waters near the water table at two subplots at Site A were shown in Fig. 2, which demonstrated that nitrate concentration under HF was much higher than that under LH. For Site A, NO3-N concentrations of soil waters near water table generally increased during celery cultivation. Especially under HF, NO3-N increased from 18.52 to 62.68 mg/L from March 4 to May 28. Moreno et al. (1996) also demonstrated that Table 4 Total water input (mainly irrigation), vertical infiltration recharge and evapotranspiration at Sites A and B Site A
Total water input (mm/d) Vertical infiltration recharge (mm/d) Evapotranspiration (mm/d)
Site B
HF
LF
HF
LF
8.276 0.330 6.463
8.276 0.340 6.380
8.276 0.493 6.010
8.276 0.535 5.784
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Fig. 2. NO3-N concentrations of soil waters near groundwater table under Site A. Solid arrows denote fertilizer N application.
the NO3-N contents in the profile for high N-fertilization site were much higher than those for the low N-fertilization site. High fertilization rate would load much more nitrate on groundwater than low fertilization rate. The NO3 concentrations of soil waters at Site B had the same trend as those at Site A (Fig. 3). Although there were the same fertilization, irrigation and cultivation, and similar soil textures at HF or LF between Sites A and B, nitrate concentrations near groundwater at Site B were much higher than those at Site A. The reason was that nitrate would enter groundwater more easily in the area with a shallow water table. Various natural physical processes, chemical reactions and microbial activities that occurred in the soil and unsaturated zone could cause contaminants to change their physical states and their chemical forms, which led to remediation of the groundwater pollution or change the characteristics of contaminants (Guo and Wang, 2004). The thicker the unsaturated zone, the more likely nitrate was degraded when it transported from surface to groundwater. In comparison with Figs. 2 and 3, we could find that the difference in NO3-N concentration of soil waters near water table between LF and HF at Site B was smaller than in Site A. This difference possibly resulted from limited natural attenuation of nitrate in the soil and unsaturated zone at both Sites A and B. It could be speculated that
Fig. 3. NO3-N concentrations of soil waters near groundwater table under Site B. Solid arrows denote fertilizer N application.
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relatively low concentration of nitrate of soil water under Site A-LF had been naturally degraded when it seeped from agricultural areas toward the aquifers, while relatively higher concentration of nitrate under Site A-HF exceeded the potential of natural attenuation and possibly restrained some microbial activities in the soil. This degradation consequently caused the relatively great difference in NO3-N concentration of soil waters between LF and HF at Site A. On the other hand, at Site B much higher concentration of NO3-N under both HF and LF far exceeded the low potential of natural attenuation of nitrate as the consequence of the thinnest unsaturated zone. Therefore, natural degradation had little effects on the decrease of nitrate during its transport in the soil at Site B, which resulted in the relatively small difference of NO3-N concentration of soil waters between LF and HF at Site B. 3.2.2. NO2-N and NH4-N NO2-N and NH4-N of soil waters were relatively low, with the maximum NO2-N of 2.47 mg/L and NH4-N of 4.00 mg/L for all subplots. Figs. 4 and 5 showed NO2-N and NH4N of soil waters near groundwater table at Sites A and B, respectively. As a whole, NO2 concentration under HF was greater than that under LF at both Sites A and B. The nitrite in soil water was mainly controlled by microbial activities. Denitrification required oxidized N (e.g. nitrate), available organic carbon and an absence of oxygen. The supply of nitrate or readily available organic C could limit denitrification rates in the surface soils (Groffman et al., 1991; Ellis et al., 1996; Sa´nchez et al., 2001). N-fertilizer application which increased nitrate content in the soil was one of the dominant factors promoting denitrification. Therefore, the nitrite concentration under HF was higher than that under LF for both Sites A and B. However, NH4+ concentration under HF of Site A was generally lower than that under LF of Site A, while NH4+ concentration under HF of Site B was generally greater than that under LF of Site B. In addition to microbial activities, adsorption of soils would dominate NH4+ content in soil waters. Heterogeneous adsorptive capacity of soils at Site A would contribute the difference of NH4+ concentration between HF and LF. Groundwater table impacted less effect on NO2 and NH4+ concentrations near water table. In contrast with NO3, the difference of NO2 and NH4+ concentrations in soil
Fig. 4. NO2-N and NH4-N of soil waters near groundwater table at Site A. Solid arrows denote fertilizer N application.
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Fig. 5. NO2-N and NH4-N of soil waters near groundwater table at Site B. Solid arrows denote fertilizer N application.
waters near water table between Sites A and B was much smaller (Figs. 4 and 5). For NH4+, the possible reason was that soil absorption of NH4+ accounted for the relatively uniform NH4+ concentration at Sites A and B. Although water table affected denitrification in the soils, intermediate product (such as NO2) of the reaction kept relatively constant in the study area. Possibly, specific denitrification process converting NO2 to NO (or N2O, N2) preferably existed in Site B and consequently contributed to the small difference in NO2 concentration between Sites A and B. 3.3. N loadings from aeration zone to groundwater Based on the vertical infiltration recharge from the unsaturated to the saturated zone and concentrations of NO3, NO2 and NH4+ near groundwater table, nitrogen loadings to groundwater under different experimental subplots were calculated, according to Eq. (6). The results were shown in Table 5. Nitrate loading was 32–404 times more than nitrite loading and 9–162 times more than ammonium loading, which demonstrated that nitrate was the dominant nitrogen component entering groundwater from vadose zone in agricultural areas. Power and Schepers (1989) and Spalding and Exner (1993) also reported that groundwater nitrate contamination was a major environmental problem in irrigated agriculture areas. Table 5 N loadings to groundwater under different subplots (kg/ha a) NO3-N loading
NO2-N loading
NH4-N loading
Total N loading
N-fertilizer applied
Percentage of N loadings to fertilizer N (%)
Site A
HF LF
47.11 13.46
1.24 0.41
0.29 1.35
48.64 15.22
4800 1800
1.01 0.85
Site B
HF LF
311.73 221.30
0.77 0.63
3.53 1.96
316.03 223.89
4800 1800
6.58 12.44
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With regard to the same groundwater table, the NO3-N loading under HF was higher than that under LF, which indicated that application of abundant fertilizers would result in severe loss of NO3-N with an increasing contamination of groundwater. Granlund et al. (2000) also indicated that nitrate leaching increased if fertilization rates increased. Therefore, to reduce the potential of groundwater contamination in agricultural areas, the amount of fertilizers applied must effectively be minimized in view of systematic combination of agricultural production and environmental protection. Hadas et al. (1999) showed that intensification of agricultural activities led to increased hazards to surface and groundwater pollution and this could be diminished by balanced irrigation–fertilization programs. Although the amount of applied fertilizers at HF of Site A was the same as that at HF of Site B, NO3-N loading at Site B was 5.62 times more than that at Site A. Because groundwater table significantly controlled the vertical infiltration recharge and definitely affected degradation of nitrate at unsaturated zone, it eventually determined the NO3 input from vadose zone to groundwater. In analogical hydrogeological settings, shallow water table favoured the contaminant loadings to groundwater. Delin and Landon (2002) also observed that the estimated mass flux of chloride and nitrate to the water table were five to two times greater, respectively, at the lowland site compared to the upland site. Kraft and Stites (2003) found that nitrate-N loading to shallow groundwater from the sweet corn ranged from 126 to 169 kg/ha a; loading from potato was 228 kg/ha a, which amounted to 61% of total available N and 77% of fertilizer N in the Wisconsin Central Sand Plain possessing a humid climate, coarse soil, and shallow water table. Consequently, lowering groundwater table would mitigate contamination of groundwater below the soils contaminated by nitrate. It could be applied to enhance drastically microbial degradation of nitrate and further reduce the potential of groundwater nitrate contamination at nitratecontaminated sites.
4. Conclusions The results demonstrated that fertilization practices impacted few effects on the balance and dynamic of water in the plant-soil-aeration zone-saturated zone system. However, groundwater table controlled vertical infiltration recharge and evaporation–transpiration rate. The vertical infiltration recharges were 0.330 and 0.340 mm/d at HF and LF of Site A, 0.493 and 0.535 mm/d at HF and LF of Site B, respectively. The evaporation–transpiration rates were 6.463, 6.380, 6.010 and 5.784 mm/d at Site A-HF, Site A-LF, Site B-HF and Site B-LF, respectively. Nitrate was the dominant nitrogen component entering groundwater from vadose zone at agricultural areas. Little ammonium and less nitrite transported into groundwater. Nitrate concentrations of soil water near groundwater table under HF subplot were much higher than that under LH subplot. The averages of NO3-N were 49.17 and 14.50 mg/L under HF and LF of Site A, respectively. Application of abundant fertilizers would result in severe loss of NO3-N with increasing contamination of groundwater. At Site B, for example, nitrogen loading was 316.03 kg/ha a under HF. Groundwater table determined the NO3 loading to groundwater. For the same N-fertilization rate, nitrate loading at Site B with a
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shallow groundwater table was much higher than that at Site A with a deep groundwater table. For HF, nitrate loading was 47.11 kg NO3-N/ha a under Site A and 311.73 kg NO3N/ha a under Site B. The nitrate loading to groundwater in the Dianchi catchment had resulted in nitrate contamination of groundwater, NO3-N concentration of which was generally 5.3–15.8 mg/ L. Because of recharging the Dianchi Lake, contaminated groundwater contributed partly to the lake eutrophication.
Acknowledgements We thank two anonymous referees for their valuable comments on the manuscript. Funding by China Postdoctoral Science Foundation (2003034021), China’s National Key Technologies R&D Program in the 10th Five-Year Plan (No. 2001BA610A-04) is also gratefully acknowledged.
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