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Regional differentiation of non-point source pollution of agriculture-derived nitrate nitrogen in groundwater in northern China
Agriculture, Ecosystems and Environment 107 (2005) 211–220 www.elsevier.com/locate/agee
Regional differentiation of non-point source pollution of agriculture-derived nitrate nitrogen in groundwater in northern China G.D. Liu a, W.L. Wu a,*, J. Zhang b a
College of Resources and Environments, China Agricultural University, 2 Yuanmingyuan West Road, Beijing, 100094 Haidian District, PR China b Department of Agronomy, Iowa State University, Ames, IA 50011, USA
Received 23 January 2004; received in revised form 12 October 2004; accepted 16 November 2004
Abstract Nitrate is one of the major agriculture-derived pollutants leaching to groundwater. This paper discusses the contamination process and spatial distribution of nitrate-N concentrations in groundwater in a typical high-yielding area of northern China. Onsite monitoring of nitrate-N concentrations in groundwater from two separated regions that had been managed similarly indicated different trends of nitrate-N elevation during the past 4 years: one remained relatively stable while the other one significantly increased. Statistical analyses show that nitrate-N concentrations in groundwater were significantly correlated with the sampling depth of 60 m and deeper and nitrate had diffused to the depth of 150–200 m. Results of surface interpolation analyses based on a large number of groundwater samples indicate that the distribution and pollution patterns of nitrate-N were mainly influenced by groundwater flow in horizontal direction. It was found that some regions were more obviously contaminated than others and a few regions were less polluted or even absent of pollution. The spatial variation in pollution patterns of nitrate-N could be mainly attributed to the diffusion and translocation of nitrate-N in three dimensions after being leached into groundwater. The differentiation of nitrate-N concentration in groundwater resulted in poor correlation between land use types on the ground and nitrate-N concentrations in the underlying water. Nitrate-N pollution of groundwater might be directly related to management practices in situations where relatively stable groundwater tables and negligible water flow existed. Therefore, the observed nitrate-N concentration in groundwater at a specific site could be a static appearance of dynamic distribution of the agriculture-derived nitrate-N within a watershed. The importance of this study is in recognition of the guidelines on managing and controlling agriculture-derived pollution within a watershed—the basic ecologic unit. # 2005 Published by Elsevier B.V. Keywords: Nitrate-N; Non-point source pollution; Groundwater; China
* Corresponding author. Tel.: +86 10 62892387; fax: +86 10 62892498. E-mail address: [email protected] (W.L. Wu). 0167-8809/$ – see front matter # 2005 Published by Elsevier B.V. doi:10.1016/j.agee.2004.11.010
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1. Introduction Since the 1960s, nitrate loading in both surface water and groundwater has drawn great attention worldwide. Nitrate pollution cases that exceed the threshold as recommended by the World Health Organization (50 mg NO3 l 1) and by the USA (10 mg N l 1) for drinking water have been reported in many countries such as UK, Denmark, Belgium, France (Strebel et al., 1989; Fried, 1991), USA (Nolan et al., 1997; Hudak, 2000) and India (Agrawal et al., 1999). The agriculture-derived nitrate pollution of groundwater has become an environmental issue (Spalding and Exner, 1993) which may cause algal blooms and eutrophication in aquifers, and even produce potential hazards to human health (Knobeloch et al., 1992; Fan and Steinberg, 1996; Gelberg et al., 1999; Gulis et al., 2002) though the link is still disputable (Forman et al., 1985; Van loon et al., 1998). The heavy use of nitrogen (N) fertilizer for intensive farming and cropping systems with low N use efficiency are often responsible for nitrate overloading into groundwater (Strebel et al., 1989). The averaged N use efficiency in Chinese grain production was 30–41% (Zhu, 2000). About 30–70% of fertilizer N applied to paddy fields and 20–50% of fertilizer N applied to uplands were lost, and about 5–10% of N was directly lost through leaching (Yuan et al., 1996). Since 1978, fertilizer consumption in China has been growing rapidly and China recently became the biggest producer and consumer of N fertilizer in the world. The average annual application rate of N in China was gradually increased from 38 kg N ha 1 in 1975 to 130 kg N ha 1 in 1985, and rapidly increased to 236 kg N ha 1 in 1995 and 262 kg N ha 1 in 2001 (Zhang et al., 1996; Anonymous, 2002). Nitrate concentration in groundwater usually exceeds the standard for drinking water in the regions where N fertilizer application rates are above 500 kg N ha 1 and N use efficiency is less than 40% (Zhang et al., 1996). In recent years, many studies have reported growing incidences of nitrate pollution and dramatic increases of nitrate concentration of groundwater in intensive farming regions (Ma and Qian, 1987; Lu et al., 1998; Liu et al., 2001). It is predicted that agriculture-derived nitrate pollution would become worse in the future because China will have to rely
more on N fertilizer for crop production in order to feed the increasing population (Zhang et al., 1996). Also, improved irrigation may further enhance nitrate leaching to aquifers from the farmlands (Singh et al., 1995; Hudak, 2000). Other factors such as land use types, vadose zone thickness, soil texture, timing of fertilizer application, and land use types have significant impacts on N loading in shallow groundwater (Mclay et al., 2001). This might be mainly due to non-linear effects of spatial variation of heterogeneity on hydraulic and geological processes within a watershed (Nikolaidis et al., 1998; Townsend et al., 1996). Observed pollution may be the consequence of farming practices many years ago rather than current practices (Singh and Skelon, 1978; Mutch, 1998). The pollutant source may not spatially and temporally coincide with the polluted site. For example, in a vegetable production region in Shunyi, Beijing, nitrate concentration in groundwater was unexpectedly high though land use type and N application rate were similar to other adjacent regions. This inconsistency was attributed to the hydrographical conditions (Liu et al., 2001). Non-point source pollutants such as nutrients, pesticides, heavy metals, and sediments are transported from the land by surface water and groundwater pathways since the pollutant sources cannot be attributed to one particular discharge location rather to a large area (Nikolaidis et al., 1998). Mutch (1998) suggested considering the sample location within a watershed and the scale of water flow when trying to interpret the nitrate data and to assess the success of Best Management Practices (BMPs). However, hydraulic and geographic conditions were seldom taken into account for the assessment. Nevertheless, random sampling with limited numbers of samples usually provides no more information than investigating site-specific pollution within an isolated region. It is difficult to reveal the essential characteristics of regional distribution of agriculture-derived pollution due to limited sampling density or lack of necessary hydraulic data. In one early study, we found that nitrate-N concentration in the groundwater varied widely from place to place though relative uniform field managements were practiced in the study area (Liu and Wu, 2003). The spatial variation could not be explained by
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the site-specific data, which suggested an unidentified mechanism plays an important role. The present study is intended to find the spatial characteristics of groundwater nitrate diffusion in three dimensions and to investigate the possible mechanism that affects spatial distribution of agriculture-derived non-point source pollution. This information is expected to be important for land use planning and risk assessment. It would encourage policy makers to make decisions on controlling agriculture-derived non-point source pollution at a watershed scale based on limited data.
2. Methods 2.1. Study area This study covered an area of 509.5 km2, the entire Huantai county in Shandong province, China. It is located south of the northern Shandong Plain, representing the transient zone from the mountain alluvial plain in central Shandong to the Yellow River plain in the north (Fig. 1). The soil parent material mainly consists of mountain diluvium and Yellow River alluvial deposits, which formed the clay loam soils classified as Calcaric Fluvisols. The crust about 300 m thick contains abundant pore water and was originally formed during the Tertiary and Quaternary periods. The landscape of the study area is relatively
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flat to gently rolling with relative elevations of 6.5– 29.5 m and gradient ratios of 1/800–1/3500. Both groundwater and surface water run roughly in the same direction as the land slopes from southwest down to northeast (Anonymous, 1993). Terrestrial waters in the study area are subject to the Xiaoqing River watershed of the Yellow River basin. The groundwater tables have been declining due to extreme pump drainage of water for irrigation in recent years. Rainfall becomes a major supplemental source for groundwater. In the past 13 years, about 993– 1123 mm of water (565 mm as precipitation and 429– 558 mm as flood irrigation) were annually provided. The average groundwater table was 13.6 8.2 m across the entire county in 2002. On average, groundwater accounts for 76% of the total fresh water consumed by industrial and agricultural productions in the county, of which agriculture consumed 87.5% (Zhang and Yu, 2002). The study area has been dominated by a rotational double cropping practice of winter wheat (Triticum aestivum L.) and summer corn (Zea mays L.) that account for 82% of all crops planted for decades. More than 15 t ha 1 of grain was produced across the entire area in 1990 (Wang and Li, 1991). Sufficient to excess amounts of fertilizer and water have been applied annually. According to an unpublished survey, N fertilizer was applied at 680 and 634 kg N ha 1 for the double cropping practice in 1996 and 2002, respec-
Fig. 1. Geographic location of the study area, Huantai county.
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tively. Furthermore, the farming practices have been typically characterized by prevalent crop-straw-return for 15 years in the study area. This caused an increase of organic matter content in soil to a relatively stable level of 14.9 0.2 g kg 1. In recent years, nearly 100% of wheat and corn straw was returned to fields. Under such an intensive management practice, nitrateN leaching from plant rooting zones and seasonal nitrate-N contamination of groundwater were usually observed (Liu and Wu, 2002, 2003). 2.2. Experimental design and data collection Two farmlands, Guojia and Lijia, under the rotational wheat–corn cropping system were monitored. They covered areas of 66 and 53 ha, respectively and mainly differed in the depth of groundwater tables, fertilization rates, and soil texture. There were 10 and 8 shallow irrigation wells evenly scattered in the two monitored regions. Water samples taken from these wells were analyzed for nitrate-N concentration. The sampling was repeated seven times from September 1998 to October 1999 to assess seasonal variation of nitrate-N concentrations in groundwater. All wells were sampled again in November 2002 and March 2003 in order to compare the nitrate-N contamination process in both regions. A second survey with intensive sampling was carried out in November 2002 to study the spatial distribution of nitrate-N pollution of groundwater beneath the entire county, except for wetlands around the Mata Lake. In total, 1186 water samples were collected from drinking and irrigation wells, each representing about 43 ha of area. On average, three or four samples including one from the drinking well were colleted from each village within the study area. The selected drinking well was usually located in the center of the village. Other sampled wells were reasonably separated from each other in different directions from the village. All the water samples were pumped out unless the wells were shallow or there were no pumps installed. The water samples were kept below 18 8C in a freezer soon after being filtered. A continuous flowing analyzer (TRAACS 2000) was used to analyze nitrate-N concentration. At the time of sampling, basic information such as well location and depth, contact information of the well owner, cropping system and land use type surrounding the well was collected. The geographic coordinates of
the wells were precisely recorded by using a global position system (GPS) receiver. The land use map on 1:50,000 scales were used as a base map to create point themes for wells. ArcView GIS 3.2 (ESRI, California) was used for surface analysis, interpolation, and isoclines creation of groundwater depths based on 637 samples. The ArcView extension ODF XTool was used to classify pollution levels in the study area. Groundwater table data were collected by local monitoring stations. The monthly and annual water table data was averaged across observations at 5-day intervals from 32 monitored wells in the study area. A survey on N input for various cropping systems was conducted with 76 randomly chosen households. 2.3. Statistical methods According to our experience on previous studies, only those samples taken in early winter and late spring were compared because nitrate-N concentration in groundwater was relatively stable during this period (Liu and Wu, 2003). A pairwised t-test was performed to test the significance of mean differences between 1988/1989 and 2002/2003. The average N rates applied to five major cropping systems, i.e., wheat–corn, vegetables, cereal–vegetable, cotton (Gossypium hirsutum L.) and fruit trees were calculated from the household survey data. The measurements of well depth and nitrate-N concentration for each cropping system were averaged as a separated category. Mean value of the N applied to the entire county was weighted by the N rates applied to each category. However, the means of well depth and nitrate-N concentration for the entire county was simply an arithmetic average because all groundwater samples were almost evenly scattered. The standard deviations of N rates, well depth and nitrate-N concentration were calculated as well.
3. Results and discussion 3.1. Differential pollution process and influential land use Monitoring data collected in 1998/1999 showed that nitrate-N concentrations in groundwater had been fluctuating seasonally in the study area. It reached a
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Table 1 Differential process of groundwater pollution by nitrate-N underlying two typical farmlands Mean of NO3 -Na (mg l 1)
Site
Year
No. of samples
Guojia
1998–99 2002–03
18 22
5.3 (3.8) 6.4 (3.9)
Lijia
1998–99 2002–03
16 15
9.7 (3.6) 36.6 (10.9)
a
P for t-test
Annual increase of NO3 -N (mg l 1 per year)
0.384
+0.3
<0.001
+6.7
Number in parentheses indicates standard deviation.
peak in late fall or early winter and dropped to the lowest level in early spring (data not shown). Table 1 shows little difference of the mean nitrate-N concentrations between two sites at the initial date (1998/1999) and no significant changes was found for the Guojia Site later (P = 0.384). However, the mean nitrate-N concentrations at the Lijia Site had tripled between 1988/1989 and 2002/2003 (P < 0.001) with an annual increase of 6.7 mg l 1. The observed trend in Table 1 suggests that nitrateN concentration in local groundwater had increased and that dramatic differentiation had taken place between the two monitored regions. Thus, there are reasons to believe that agriculture-derived nitrate-N pollution might have developed non-uniformly even in a relatively uniform farming area. Statistical data of the survey given in Table 2 show the association of vegetation types and nitrate pollution in shallow groundwater (<60 m). The average N rate of 639.8 N kg ha 1 indicated great intensity of the farming systems in this area. However, a clear correlation was not found between N application rates for different cropping systems and nitrate-N concentrations in the underlying groundwater. This result apparently is not consistent with earlier reports (Zhang et al., 1996; Liu et al., 2001).
It is generally believed that intensive farming with high application rates of N would lead to more severe groundwater pollution. Nitrate pollution of groundwater for vegetable cropping systems was found to be worse than for cereal cropping systems (Zhang et al., 1996; Hudak and Blanchard, 1997; Liu et al., 2001). In our study, however, an application rate of N fertilizer as high as 884.0 kg N ha 1 in the vegetable cropping system was acting similarly as a rate of 633.7 kg N ha 1 in the rotational wheat–corn cropping system in terms of nitrate-N concentration in groundwater. Moreover, the nitrate-N concentration in groundwater beneath cotton fields had reached the highest level among all cropping systems though the lowest rate of N (182.6 kg N ha 1) was applied for this crop. More careful investigations revealed that more groundwater samples were taken from areas with higher nitrate-N concentration, where cotton was commonly planted. This finding supports the assumption that nitrate-N levels in groundwater should be closely related to the geographical location of the system instead of the specific system management including N rates applied. Therefore, the regional differentiation of non-point source pollution of agriculture-derived nitrate-N may have played a more important role in groundwater pollution.
Table 2 Nitrate-N concentration in shallow groundwater in relation to aboveground cropping systems in 2002a System
N rateb (kg ha 1)
No. of samples
Well depth (m)
NO3 -N (mg l 1)
Wheat–corn Vegetables Cereal–vegetable Cotton Fruit trees
633.7 884.0 507.0 182.6 492.0
(46.0) (87.2) (62.0) (12.1) (35.4)
546 32 21 28 10
30.4 32.9 25.9 33.6 37.0
8.1 8.4 6.5 12.4 3.0
Mean
639.8 (48.9)
637
30.5 (15.3)
a b
Number in parentheses indicates standard deviation. Mean of N input is weighted by the planting areas for each cropping system.
(15.1) (17.7) (15.8) (14.7) (14.6)
(8.8) (9.5) (6.3) (15.5) (2.7)
8.0 (8.9)
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High application rates of N fertilizer combined with flood irrigation on high fertility soils seem to have greater potential for nitrate leaching (Agrawal et al., 1999). However, the groundwater flow may gradually retard the nitrate-N pollution in some regions and enhance the accumulation of nitrate-N in other regions. For example, a large portion of clean groundwater was found in the southwest part, around ZJ Town, though this region had been intensively farmed with a relatively large proportion of vegetable productions, accounting for 9.8% of the arable land. It is very likely that nitrate-N pollution would even happen under virgin lands due to translocation of nitrate-N in groundwater. On the other hand, clean water could be found beneath an intensive farming region if the groundwater does not remain static. Therefore, nitrate-N pollution of groundwater is more associated with the intensity of fertilizer N input only in the regions with very static groundwater. For a specific region, farming practices and land use types may interfere with nitrate-N pollution via root zones, but not necessarily be the sole factor guiding spatial patterns of nitrate-N pollution of groundwater. 3.2. Horizontal distribution of nitrate-N pollution Spatial distribution of nitrate-N was estimated with interpolation of 12 neighboring values by using the inverse distance weighted method (ArcView 3.2, ESRI). As shown in Fig. 2a, the isolines of groundwater tables marked with elevation are the monthly averaged levels at the sampling time in November 2002. The background nitrate-N values of 3, 6 and 10 mg l 1 shown in this figure are in accordance with the threshold and sub-critical levels as defined early in our study and with the world-wide accepted critical level, respectively. The levels of pollution are: clean (0–3 mg l 1), lightly polluted (3–6 mg l 1), polluted (6–10 mg l 1) and severely polluted (10 mg l 1). The nitrate-N pollution map overlaid with the isograms in Fig. 2a and b clearly identifies the spatial relationships between nitrate-N concentration categories and isolines of groundwater depths. This figure also illustrates the characteristics of regional differentiation of nitrate-N pollution in shallow groundwater in the entire study area. For example, most of the severely polluted regions were located in the central part from east to west, except for regions along the
mid-west county boundary. Obviously, the nitrate-N pollution was not evenly distributed over the county. The groundwater flowed from surrounding regions towards two funnels in TZ and SZ Towns as indicated by the arrows in Fig. 2a. The decline of water tables was hardly recovered all year round because of groundwater exploitation. The groundwater carrying concentrated nitrate-N flowed from a pollution source to recharge regions. This can be shown by the groundwater flow from the mid-west boundary area to the central funnel in TZ Town through CZ Town and nitrate-N pollution following a similar trend. In another case the internally polluted groundwater near GQ Town spread towards the southeast funnel (see Fig. 2a and b). Therefore, the recharge regions are usually vulnerable to nitrate pollution in a watershed. The phenomenon of spreading nitrate pollution in recharge regions is observed in some European countries (Strebel et al., 1989; Fried, 1991). This finding also implies that more efficient control of nonpoint source pollution of groundwater should be directed at a larger scale such as catchments instead of being limited to an administrative regional scale. Some regions benefited from the dilution process. In Fig. 2, there was a clean pathway of groundwater from southwest ZJ Town via central TS Town to TZ Town. On the other hand, there may be less nitrate-N transportation in groundwater if the water table is relatively static, for instance, from XJ Town to GQ Town in the east part of the central polluted region in Fig. 2, both intensity and scope of nitrate-N pollution supposedly depended on internal cycles of nitrate-N in groundwater through local farmland ecosystem. It was unlikely that these regions were influenced by external groundwater out of the system. 3.3. Vertical distribution of nitrate-N pollution The vertical distribution of nitrate-N concentration illustrated in Fig. 3 is a function of sampling depth for all groundwater samples. The inverse association of nitrate-N concentrations and well depths can be explained by the concept of land-surface origin for contamination (Hudak, 1999; Gelberg et al., 1999). The relationship can be approximately fitted with an exponential equation. With a focus on the shallow groundwater where the vertical change can be reasonably neglected, a horizontal distribution of
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Fig. 2. Horizontal distribution of nitrate-N concentration in groundwater and spatial relation to groundwater flow (a) [arrows roughly indicating direction of groundwater flow] within the study area (b).
nitrate-N concentration can be assumed to reflect the spreading dynamics of pollution. If an arbitrarily stratified depth of 60 m was assumed as reasonable boundary and all groundwater
samples from wells were divided into the shallow and deep groups, then significant contrasts for the two groups are as follows: average depth at 30.7 m versus 150.6 m, number of samples, accounting for 637
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Fig. 3. Vertical distribution of nitrate-N concentration in groundwater.
versus 415 samples, nitrate-N concentration ranged from 0.02 to 79.7 mg l 1 with a mean of 7.0 8.2 mg l 1 versus 0–41.2 mg l 1 with a mean of 3.8 4.8 mg l 1. Large variation of nitrate-N concentration in the shallow group might be due to strong disturbance such as frequent water pumping or mixing with percolating water from surface via crop root zones, while small variation in the deep groundwater might reflect a relatively static status of free diffusion of nitrate. As shown in Fig. 4, nitrate-N concentrations in groundwater at the depth of 60 m and above were not significantly correlated with the depth (P = 0.286). However, a significantly negative correlation between nitrate-N concentrations and sampling depths (P < 0.001) was found for the depths below 60 m. At the depth of 200 m, nitrate-N concentration approached to 2 mg l 1. The results shown in Fig. 4 imply that human activities had exerted an impact on groundwater as
deep as 150–200 m. Nitrate-N concentration in groundwater below 200 m was 1.4 2.1 mg l 1 based on 142 samples. At depths from 100 to 350 m, six samples were suspicious for unusually high nitrate-N concentration (about 10 mg l 1). This could be a problem caused either by direct leakage of nitrate-N from the polluted shallow water through unclosed well wall (Anonymous, 1993) or by unexpected sampling errors. Excluding the abnormal samples, nitrate-N concentration in the remaining 136 samples gave an average of 0.97 0.90 mg l 1, which fell below the critical level (3 mg l 1) for undisturbed groundwater as defined by the Environmental Protection Agency of the United States (Anonymous, 1987). Therefore, we use 2 mg l 1 of nitrate-N as a background value for the study area. Any concentration above the threshold will be a suspect for nitrate-N pollution due to human activities. Surface water may affect distribution patterns of nitrate-N in shallow groundwater in some regions. However, the surface water system in the study region seems not abundant enough to exert influence on the pollution distribution. Except for the Dongzhulong and Xiaoqing Rivers, all other rivers in the county temporarily reserved water during wet season and kept dry in other seasons (Anonymous, 1993). Although Dongzhulong River was mainly polluted by effluents of the waste water from the upstream city, Zibo, there was no obvious pattern for nitrate-N pollution along the river. It should be noticed that there was an area low in nitrate-N concentration under a hydraulic head with relatively shallow water table in the southwest part of Mata Lake. A clean water channel was constructed for discharging fresh water to the lake in fall. That could be an evidence of the effect of surface water on shallow water pollution. Nevertheless, it is supposed that the seasonal river could affect the pollution distribution in the shallow groundwater though it was not observed directly from the pollution map. Further investigation is needed.
4. Conclusions
Fig. 4. Averaged nitrate-N concentration of each arbitrary stratum in relation to sampling depth.
This study clearly revealed the association of spatial patterns of nitrate-N pollution and groundwater flow directions. A new outlook on the relationship
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between land use and N pollution should be enhanced by recognizing this mechanism. It is reasonable to conclude that land use types or cropping systems might be a necessary condition but not the only condition to be taken into account for groundwater pollution at a larger scale. Our findings support the conclusion that elevation of nitrate-N concentration in groundwater is a synchronizing result of intensive farming though it has not been concluded by some other studies. Agriculture-derived nitrate-N pollution dynamically moves or diffuses in three dimensions within a watershed. Any regional water cycles may divide a watershed into recharge and discharge sub-regions. Hydraulic conditions may play an important role on the region’s vulnerability to agriculture-derived nonpoint source pollution. To some extent, nitrate-N distribution patterns and hydraulic properties in a region can spatially indicate with each other. Our study suggests that the vulnerability of agriculture-derived nitrate-N loading in different regions should vary with differences in differential mechanism of groundwater pollution. The importance of understanding this process is to help guide land use planning, alleviate environmental impacts, and minimize ecologic risk of intensive farming by taking watershed as an ecologic unit instead of administrative division.
Acknowledgements This study was part of the project. ‘‘Modeling Approach of Estimate Environmental Cost Based on Budgets and Cycles of Carbon and Nitrogen in AgroEcosystem’’ (30270220), funded by the National Natural Science Foundation of China (NSFC). The authors wish to acknowledge endeavours of the government of Huantai County for organizing groundwater sampling. We express our gratitude to Huantai Water Conservation Bureau for providing the hydraulic and geologic data. The constructive comments and suggestions on the manuscript given by Peter Kyveryga, Brad van de Woestyne and other anonymous reviewers are greatly appreciated. We would like to thank participating colleagues and farmers for their assistance and cooperation during the field experiment setup and data collection.
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References Agrawal, G.D., Lunkad, S.K., Malkhed, T., 1999. Diffuse agricultural nitrate pollution of groundwaters in India. Water Sci. Technol. 29 (3), 67–75. Anonymous, 1987. Nitrate: a question of time? Water Quality International, No.1, IAWPRC, London, pp. 24–28. Anonymous, 1993. Report of exploiting groundwater of Huantai county, Shandong Province, Geologic Environment Monitor Station of Ministry of Geologic Mineral Resources, pp. 20– 28 (in Chinese). Anonymous, 2002. Report of Agriculture Development of China. Ministry of Agriculture, The People’s Republic of China, China Agriculture Press, Beijing, pp. 129–132 (in Chinese). Fan, A.M., Steinberg, V.E., 1996. Health implications of nitrate and nitrite in drinking water. An update on methemoglobinemia occurrence and reproductive and developmental toxicity. Regul. Toxicol. Pharmacol. 23, 35–43. Forman, D., Al-Dabbagh, S., Doll, R., 1985. Nitrates, nitrites and gastric cancer in Great Britain. Nature 313 (21), 620–625. Fried, J.J., 1991. Sources of nitrates in fissure groundwater in the humid tropical zone—the example of Ivory Coast. J. Hydrol. 113, 231–264. Gelberg, K.H., Church, L., Casey, G., London, M., Roerig, D.S., Boyd, J., Hill, M., 1999. Nitrate levels in drinking water in rural New York State. Environ. Res. Sect. A 80, 30–40. Gulis, G., Czompolyova, M., Cerhan, J.R., 2002. An ecologic study of nitrate in municipal drinking water and cancer incidence in Trnava District, Slovakia. Environ. Res. Sect. A 88, 182–187. Hudak, P.F., 1999. Chloride and nitrate distributions in the Hickory aquifer, central Texas. U.S.A. Environ. Int. 25 (4), 393–401. Hudak, P.F., 2000. Regional trends in nitrate content of Texas groundwater. J. Hydrol. 228, 37–47. Hudak, P.F., Blanchard, S., 1997. Land use and groundwater quality in the Trinity Group outcrop of north-central Texas. U.S.A. Environ. Int. 23 (4), 507–517. Knobeloch, L., Krenz, K., Anderson, H., Hovel, C., 1992. Methemoglobinemia in an infant—Wisconsin. Morbidity Mortality Weekly Report 42 (12), 217–219. Liu, G.D., Wu, W.L., 2002. The dynamics of nitrate nitrogen leaching through soil in high-yield farmland ecosystem. Chin. J. Eco-Agric. 10 (4), 71–74 (in Chinese). Liu, G.D., Wu, W.L., 2003. The dynamics of soil nitrate leaching and contamination of groundwater in high-yielding-farmlands. Chin. J. Eco-Agric. 11 (1), 91–93 (in Chinese). Liu, H.B., Lei, B.K., Zhang, Y.G., Zhang, W.L., Lin, B., 2001. Investigation and evaluation on nitrate pollution of groundwater of Shunyi District. Plant Nutr. Fert. Sci. 4, 385–390 (in Chinese). Lu, D.Q., Tong, Y.A., Sun, B.H., 1998. Study on effect of nitrogen fertilizer use on environment pollution. Plant. Nutr. Fert. Sci. 4 (1), 8–15 (in Chinese). Ma, L.S., Qian, M.R., 1987. Contamination of nitrate and nitrite of water environment in Taihu Lake watershed. Environ. Sci. 8 (2), 60–65 (in Chinese). Mclay, C.D.A., Dragten, R., Sparling, G., Selvarajah, N., 2001. Predicting groundwater nitrate concentrations in a region of
220
G.D. Liu et al. / Agriculture, Ecosystems and Environment 107 (2005) 211–220
mixed agricultural land use: a comparison of three approaches. Environ. Pollut. 115, 191–204. Mutch, J., 1998. The hydrologic cycle and water movement. Nitrate – agricultural sources and fate in the environment – perspectives and direction. In: Proceedings of the workshop. Eastern Canada Soil and Water Conservation Centre, pp. 3–7. Nikolaidis, N.P., Heng, H., Semagin, R., Clausen, J.C., 1998. Nonlinear response of a mixed land use watershed to nitrogen loading. Agric. Ecosyst. Environ. 67, 251–265. Nolan, B.T., Ruddy, B.C., Hitt, K.J., Helsel, D.R., 1997. Risk of nitrate in groundwaters of the United States—a national perspective. Environ. Sci. Technol. 31, 2229–2236. Singh, B., Sekhon, G.S., 1978-9. Nitrate pollution of groundwater from farm use of nitrogen fertilizers—a review. Agric. Environ. 4, 207–225. Singh, B., Singh, Y., Sekhon, G.S., 1995. Fertilizer-N use efficiency and nitrate pollution of groundwater in developing countries. J. Contam. Hydrol. 20, 167–184. Spalding, R.F., Exner, M.E., 1993. Occurrence of nitrate in groundwater—a review. J. Environ. Qual. 22, 392–402. Strebel, O., Duynisveld, W.H.M., Bottcher, J., 1989. Nitrate pollution of groundwater in Western Europe. Agric. Ecosyst. Environ. 26, 189–214.
Townsend, M.A., Sleezer, R.O., Macko, S.A., 1996. Effects of agricultural practices and vadose zone stratigraphy on nitrate concentration in ground water in Kansas. U.S.A. Water Sci. Technol. 33 (4–5), 219–226. Van Loon, A.J., Botterweck, A.A., Goldbohm, R.A., Brants, H.A., van Klaveren, J.D., van den Brandt, P.A., 1998. Intake of nitrate and nitrite and the risk of gastric cancer: a prospective cohort study. Br. J. Cancer 78 (1), 129–135. Wang, Y.M., Li, K.F., 1991. Integrated techniques of cultivation for producing grain yield of 15 t ha 1. J. Shandong Agric. Sci. 3, 9– 11 (in Chinese). Yuan, F.M., Chen, Z.M., Yao, Z.H., 1996. Nitrogen leaching from soils. In: Chen, Z.M. (Ed.), Nitrogen, Yield and Environment. Chinese Agri-Tech Press, Beijing, p. 191. Zhang, W.L., Tian, Z.X., Zhang, N., Li, X.Q., 1996. Nitrate pollution of groundwater in northern China. Agric. Ecosyst. Environ. 59, 223–231. Zhang, Z.X., Yu, G.R., 2002. Counter measures on sustainable utilization of agricultural water resources in high-yield grain area in north China—a case study in Huantai county, Shandong Province. Resour. Sci. 24 (1), 68–71 (in Chinese). Zhu, Z.L., 2000. Loss of fertilizer N from plant–soil system and the strategies and techniques for its reduction. Soil Environ. Sci. 9 (1), 1–6 (in Chinese).
Regional differentiation of non-point source pollution of agriculture-derived nitrate nitrogen in groundwater in northern China G.D. Liu a, W.L. Wu a,*, J. Zhang b a
College of Resources and Environments, China Agricultural University, 2 Yuanmingyuan West Road, Beijing, 100094 Haidian District, PR China b Department of Agronomy, Iowa State University, Ames, IA 50011, USA
Received 23 January 2004; received in revised form 12 October 2004; accepted 16 November 2004
Abstract Nitrate is one of the major agriculture-derived pollutants leaching to groundwater. This paper discusses the contamination process and spatial distribution of nitrate-N concentrations in groundwater in a typical high-yielding area of northern China. Onsite monitoring of nitrate-N concentrations in groundwater from two separated regions that had been managed similarly indicated different trends of nitrate-N elevation during the past 4 years: one remained relatively stable while the other one significantly increased. Statistical analyses show that nitrate-N concentrations in groundwater were significantly correlated with the sampling depth of 60 m and deeper and nitrate had diffused to the depth of 150–200 m. Results of surface interpolation analyses based on a large number of groundwater samples indicate that the distribution and pollution patterns of nitrate-N were mainly influenced by groundwater flow in horizontal direction. It was found that some regions were more obviously contaminated than others and a few regions were less polluted or even absent of pollution. The spatial variation in pollution patterns of nitrate-N could be mainly attributed to the diffusion and translocation of nitrate-N in three dimensions after being leached into groundwater. The differentiation of nitrate-N concentration in groundwater resulted in poor correlation between land use types on the ground and nitrate-N concentrations in the underlying water. Nitrate-N pollution of groundwater might be directly related to management practices in situations where relatively stable groundwater tables and negligible water flow existed. Therefore, the observed nitrate-N concentration in groundwater at a specific site could be a static appearance of dynamic distribution of the agriculture-derived nitrate-N within a watershed. The importance of this study is in recognition of the guidelines on managing and controlling agriculture-derived pollution within a watershed—the basic ecologic unit. # 2005 Published by Elsevier B.V. Keywords: Nitrate-N; Non-point source pollution; Groundwater; China
* Corresponding author. Tel.: +86 10 62892387; fax: +86 10 62892498. E-mail address: [email protected] (W.L. Wu). 0167-8809/$ – see front matter # 2005 Published by Elsevier B.V. doi:10.1016/j.agee.2004.11.010
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1. Introduction Since the 1960s, nitrate loading in both surface water and groundwater has drawn great attention worldwide. Nitrate pollution cases that exceed the threshold as recommended by the World Health Organization (50 mg NO3 l 1) and by the USA (10 mg N l 1) for drinking water have been reported in many countries such as UK, Denmark, Belgium, France (Strebel et al., 1989; Fried, 1991), USA (Nolan et al., 1997; Hudak, 2000) and India (Agrawal et al., 1999). The agriculture-derived nitrate pollution of groundwater has become an environmental issue (Spalding and Exner, 1993) which may cause algal blooms and eutrophication in aquifers, and even produce potential hazards to human health (Knobeloch et al., 1992; Fan and Steinberg, 1996; Gelberg et al., 1999; Gulis et al., 2002) though the link is still disputable (Forman et al., 1985; Van loon et al., 1998). The heavy use of nitrogen (N) fertilizer for intensive farming and cropping systems with low N use efficiency are often responsible for nitrate overloading into groundwater (Strebel et al., 1989). The averaged N use efficiency in Chinese grain production was 30–41% (Zhu, 2000). About 30–70% of fertilizer N applied to paddy fields and 20–50% of fertilizer N applied to uplands were lost, and about 5–10% of N was directly lost through leaching (Yuan et al., 1996). Since 1978, fertilizer consumption in China has been growing rapidly and China recently became the biggest producer and consumer of N fertilizer in the world. The average annual application rate of N in China was gradually increased from 38 kg N ha 1 in 1975 to 130 kg N ha 1 in 1985, and rapidly increased to 236 kg N ha 1 in 1995 and 262 kg N ha 1 in 2001 (Zhang et al., 1996; Anonymous, 2002). Nitrate concentration in groundwater usually exceeds the standard for drinking water in the regions where N fertilizer application rates are above 500 kg N ha 1 and N use efficiency is less than 40% (Zhang et al., 1996). In recent years, many studies have reported growing incidences of nitrate pollution and dramatic increases of nitrate concentration of groundwater in intensive farming regions (Ma and Qian, 1987; Lu et al., 1998; Liu et al., 2001). It is predicted that agriculture-derived nitrate pollution would become worse in the future because China will have to rely
more on N fertilizer for crop production in order to feed the increasing population (Zhang et al., 1996). Also, improved irrigation may further enhance nitrate leaching to aquifers from the farmlands (Singh et al., 1995; Hudak, 2000). Other factors such as land use types, vadose zone thickness, soil texture, timing of fertilizer application, and land use types have significant impacts on N loading in shallow groundwater (Mclay et al., 2001). This might be mainly due to non-linear effects of spatial variation of heterogeneity on hydraulic and geological processes within a watershed (Nikolaidis et al., 1998; Townsend et al., 1996). Observed pollution may be the consequence of farming practices many years ago rather than current practices (Singh and Skelon, 1978; Mutch, 1998). The pollutant source may not spatially and temporally coincide with the polluted site. For example, in a vegetable production region in Shunyi, Beijing, nitrate concentration in groundwater was unexpectedly high though land use type and N application rate were similar to other adjacent regions. This inconsistency was attributed to the hydrographical conditions (Liu et al., 2001). Non-point source pollutants such as nutrients, pesticides, heavy metals, and sediments are transported from the land by surface water and groundwater pathways since the pollutant sources cannot be attributed to one particular discharge location rather to a large area (Nikolaidis et al., 1998). Mutch (1998) suggested considering the sample location within a watershed and the scale of water flow when trying to interpret the nitrate data and to assess the success of Best Management Practices (BMPs). However, hydraulic and geographic conditions were seldom taken into account for the assessment. Nevertheless, random sampling with limited numbers of samples usually provides no more information than investigating site-specific pollution within an isolated region. It is difficult to reveal the essential characteristics of regional distribution of agriculture-derived pollution due to limited sampling density or lack of necessary hydraulic data. In one early study, we found that nitrate-N concentration in the groundwater varied widely from place to place though relative uniform field managements were practiced in the study area (Liu and Wu, 2003). The spatial variation could not be explained by
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the site-specific data, which suggested an unidentified mechanism plays an important role. The present study is intended to find the spatial characteristics of groundwater nitrate diffusion in three dimensions and to investigate the possible mechanism that affects spatial distribution of agriculture-derived non-point source pollution. This information is expected to be important for land use planning and risk assessment. It would encourage policy makers to make decisions on controlling agriculture-derived non-point source pollution at a watershed scale based on limited data.
2. Methods 2.1. Study area This study covered an area of 509.5 km2, the entire Huantai county in Shandong province, China. It is located south of the northern Shandong Plain, representing the transient zone from the mountain alluvial plain in central Shandong to the Yellow River plain in the north (Fig. 1). The soil parent material mainly consists of mountain diluvium and Yellow River alluvial deposits, which formed the clay loam soils classified as Calcaric Fluvisols. The crust about 300 m thick contains abundant pore water and was originally formed during the Tertiary and Quaternary periods. The landscape of the study area is relatively
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flat to gently rolling with relative elevations of 6.5– 29.5 m and gradient ratios of 1/800–1/3500. Both groundwater and surface water run roughly in the same direction as the land slopes from southwest down to northeast (Anonymous, 1993). Terrestrial waters in the study area are subject to the Xiaoqing River watershed of the Yellow River basin. The groundwater tables have been declining due to extreme pump drainage of water for irrigation in recent years. Rainfall becomes a major supplemental source for groundwater. In the past 13 years, about 993– 1123 mm of water (565 mm as precipitation and 429– 558 mm as flood irrigation) were annually provided. The average groundwater table was 13.6 8.2 m across the entire county in 2002. On average, groundwater accounts for 76% of the total fresh water consumed by industrial and agricultural productions in the county, of which agriculture consumed 87.5% (Zhang and Yu, 2002). The study area has been dominated by a rotational double cropping practice of winter wheat (Triticum aestivum L.) and summer corn (Zea mays L.) that account for 82% of all crops planted for decades. More than 15 t ha 1 of grain was produced across the entire area in 1990 (Wang and Li, 1991). Sufficient to excess amounts of fertilizer and water have been applied annually. According to an unpublished survey, N fertilizer was applied at 680 and 634 kg N ha 1 for the double cropping practice in 1996 and 2002, respec-
Fig. 1. Geographic location of the study area, Huantai county.
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tively. Furthermore, the farming practices have been typically characterized by prevalent crop-straw-return for 15 years in the study area. This caused an increase of organic matter content in soil to a relatively stable level of 14.9 0.2 g kg 1. In recent years, nearly 100% of wheat and corn straw was returned to fields. Under such an intensive management practice, nitrateN leaching from plant rooting zones and seasonal nitrate-N contamination of groundwater were usually observed (Liu and Wu, 2002, 2003). 2.2. Experimental design and data collection Two farmlands, Guojia and Lijia, under the rotational wheat–corn cropping system were monitored. They covered areas of 66 and 53 ha, respectively and mainly differed in the depth of groundwater tables, fertilization rates, and soil texture. There were 10 and 8 shallow irrigation wells evenly scattered in the two monitored regions. Water samples taken from these wells were analyzed for nitrate-N concentration. The sampling was repeated seven times from September 1998 to October 1999 to assess seasonal variation of nitrate-N concentrations in groundwater. All wells were sampled again in November 2002 and March 2003 in order to compare the nitrate-N contamination process in both regions. A second survey with intensive sampling was carried out in November 2002 to study the spatial distribution of nitrate-N pollution of groundwater beneath the entire county, except for wetlands around the Mata Lake. In total, 1186 water samples were collected from drinking and irrigation wells, each representing about 43 ha of area. On average, three or four samples including one from the drinking well were colleted from each village within the study area. The selected drinking well was usually located in the center of the village. Other sampled wells were reasonably separated from each other in different directions from the village. All the water samples were pumped out unless the wells were shallow or there were no pumps installed. The water samples were kept below 18 8C in a freezer soon after being filtered. A continuous flowing analyzer (TRAACS 2000) was used to analyze nitrate-N concentration. At the time of sampling, basic information such as well location and depth, contact information of the well owner, cropping system and land use type surrounding the well was collected. The geographic coordinates of
the wells were precisely recorded by using a global position system (GPS) receiver. The land use map on 1:50,000 scales were used as a base map to create point themes for wells. ArcView GIS 3.2 (ESRI, California) was used for surface analysis, interpolation, and isoclines creation of groundwater depths based on 637 samples. The ArcView extension ODF XTool was used to classify pollution levels in the study area. Groundwater table data were collected by local monitoring stations. The monthly and annual water table data was averaged across observations at 5-day intervals from 32 monitored wells in the study area. A survey on N input for various cropping systems was conducted with 76 randomly chosen households. 2.3. Statistical methods According to our experience on previous studies, only those samples taken in early winter and late spring were compared because nitrate-N concentration in groundwater was relatively stable during this period (Liu and Wu, 2003). A pairwised t-test was performed to test the significance of mean differences between 1988/1989 and 2002/2003. The average N rates applied to five major cropping systems, i.e., wheat–corn, vegetables, cereal–vegetable, cotton (Gossypium hirsutum L.) and fruit trees were calculated from the household survey data. The measurements of well depth and nitrate-N concentration for each cropping system were averaged as a separated category. Mean value of the N applied to the entire county was weighted by the N rates applied to each category. However, the means of well depth and nitrate-N concentration for the entire county was simply an arithmetic average because all groundwater samples were almost evenly scattered. The standard deviations of N rates, well depth and nitrate-N concentration were calculated as well.
3. Results and discussion 3.1. Differential pollution process and influential land use Monitoring data collected in 1998/1999 showed that nitrate-N concentrations in groundwater had been fluctuating seasonally in the study area. It reached a
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Table 1 Differential process of groundwater pollution by nitrate-N underlying two typical farmlands Mean of NO3 -Na (mg l 1)
Site
Year
No. of samples
Guojia
1998–99 2002–03
18 22
5.3 (3.8) 6.4 (3.9)
Lijia
1998–99 2002–03
16 15
9.7 (3.6) 36.6 (10.9)
a
P for t-test
Annual increase of NO3 -N (mg l 1 per year)
0.384
+0.3
<0.001
+6.7
Number in parentheses indicates standard deviation.
peak in late fall or early winter and dropped to the lowest level in early spring (data not shown). Table 1 shows little difference of the mean nitrate-N concentrations between two sites at the initial date (1998/1999) and no significant changes was found for the Guojia Site later (P = 0.384). However, the mean nitrate-N concentrations at the Lijia Site had tripled between 1988/1989 and 2002/2003 (P < 0.001) with an annual increase of 6.7 mg l 1. The observed trend in Table 1 suggests that nitrateN concentration in local groundwater had increased and that dramatic differentiation had taken place between the two monitored regions. Thus, there are reasons to believe that agriculture-derived nitrate-N pollution might have developed non-uniformly even in a relatively uniform farming area. Statistical data of the survey given in Table 2 show the association of vegetation types and nitrate pollution in shallow groundwater (<60 m). The average N rate of 639.8 N kg ha 1 indicated great intensity of the farming systems in this area. However, a clear correlation was not found between N application rates for different cropping systems and nitrate-N concentrations in the underlying groundwater. This result apparently is not consistent with earlier reports (Zhang et al., 1996; Liu et al., 2001).
It is generally believed that intensive farming with high application rates of N would lead to more severe groundwater pollution. Nitrate pollution of groundwater for vegetable cropping systems was found to be worse than for cereal cropping systems (Zhang et al., 1996; Hudak and Blanchard, 1997; Liu et al., 2001). In our study, however, an application rate of N fertilizer as high as 884.0 kg N ha 1 in the vegetable cropping system was acting similarly as a rate of 633.7 kg N ha 1 in the rotational wheat–corn cropping system in terms of nitrate-N concentration in groundwater. Moreover, the nitrate-N concentration in groundwater beneath cotton fields had reached the highest level among all cropping systems though the lowest rate of N (182.6 kg N ha 1) was applied for this crop. More careful investigations revealed that more groundwater samples were taken from areas with higher nitrate-N concentration, where cotton was commonly planted. This finding supports the assumption that nitrate-N levels in groundwater should be closely related to the geographical location of the system instead of the specific system management including N rates applied. Therefore, the regional differentiation of non-point source pollution of agriculture-derived nitrate-N may have played a more important role in groundwater pollution.
Table 2 Nitrate-N concentration in shallow groundwater in relation to aboveground cropping systems in 2002a System
N rateb (kg ha 1)
No. of samples
Well depth (m)
NO3 -N (mg l 1)
Wheat–corn Vegetables Cereal–vegetable Cotton Fruit trees
633.7 884.0 507.0 182.6 492.0
(46.0) (87.2) (62.0) (12.1) (35.4)
546 32 21 28 10
30.4 32.9 25.9 33.6 37.0
8.1 8.4 6.5 12.4 3.0
Mean
639.8 (48.9)
637
30.5 (15.3)
a b
Number in parentheses indicates standard deviation. Mean of N input is weighted by the planting areas for each cropping system.
(15.1) (17.7) (15.8) (14.7) (14.6)
(8.8) (9.5) (6.3) (15.5) (2.7)
8.0 (8.9)
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High application rates of N fertilizer combined with flood irrigation on high fertility soils seem to have greater potential for nitrate leaching (Agrawal et al., 1999). However, the groundwater flow may gradually retard the nitrate-N pollution in some regions and enhance the accumulation of nitrate-N in other regions. For example, a large portion of clean groundwater was found in the southwest part, around ZJ Town, though this region had been intensively farmed with a relatively large proportion of vegetable productions, accounting for 9.8% of the arable land. It is very likely that nitrate-N pollution would even happen under virgin lands due to translocation of nitrate-N in groundwater. On the other hand, clean water could be found beneath an intensive farming region if the groundwater does not remain static. Therefore, nitrate-N pollution of groundwater is more associated with the intensity of fertilizer N input only in the regions with very static groundwater. For a specific region, farming practices and land use types may interfere with nitrate-N pollution via root zones, but not necessarily be the sole factor guiding spatial patterns of nitrate-N pollution of groundwater. 3.2. Horizontal distribution of nitrate-N pollution Spatial distribution of nitrate-N was estimated with interpolation of 12 neighboring values by using the inverse distance weighted method (ArcView 3.2, ESRI). As shown in Fig. 2a, the isolines of groundwater tables marked with elevation are the monthly averaged levels at the sampling time in November 2002. The background nitrate-N values of 3, 6 and 10 mg l 1 shown in this figure are in accordance with the threshold and sub-critical levels as defined early in our study and with the world-wide accepted critical level, respectively. The levels of pollution are: clean (0–3 mg l 1), lightly polluted (3–6 mg l 1), polluted (6–10 mg l 1) and severely polluted (10 mg l 1). The nitrate-N pollution map overlaid with the isograms in Fig. 2a and b clearly identifies the spatial relationships between nitrate-N concentration categories and isolines of groundwater depths. This figure also illustrates the characteristics of regional differentiation of nitrate-N pollution in shallow groundwater in the entire study area. For example, most of the severely polluted regions were located in the central part from east to west, except for regions along the
mid-west county boundary. Obviously, the nitrate-N pollution was not evenly distributed over the county. The groundwater flowed from surrounding regions towards two funnels in TZ and SZ Towns as indicated by the arrows in Fig. 2a. The decline of water tables was hardly recovered all year round because of groundwater exploitation. The groundwater carrying concentrated nitrate-N flowed from a pollution source to recharge regions. This can be shown by the groundwater flow from the mid-west boundary area to the central funnel in TZ Town through CZ Town and nitrate-N pollution following a similar trend. In another case the internally polluted groundwater near GQ Town spread towards the southeast funnel (see Fig. 2a and b). Therefore, the recharge regions are usually vulnerable to nitrate pollution in a watershed. The phenomenon of spreading nitrate pollution in recharge regions is observed in some European countries (Strebel et al., 1989; Fried, 1991). This finding also implies that more efficient control of nonpoint source pollution of groundwater should be directed at a larger scale such as catchments instead of being limited to an administrative regional scale. Some regions benefited from the dilution process. In Fig. 2, there was a clean pathway of groundwater from southwest ZJ Town via central TS Town to TZ Town. On the other hand, there may be less nitrate-N transportation in groundwater if the water table is relatively static, for instance, from XJ Town to GQ Town in the east part of the central polluted region in Fig. 2, both intensity and scope of nitrate-N pollution supposedly depended on internal cycles of nitrate-N in groundwater through local farmland ecosystem. It was unlikely that these regions were influenced by external groundwater out of the system. 3.3. Vertical distribution of nitrate-N pollution The vertical distribution of nitrate-N concentration illustrated in Fig. 3 is a function of sampling depth for all groundwater samples. The inverse association of nitrate-N concentrations and well depths can be explained by the concept of land-surface origin for contamination (Hudak, 1999; Gelberg et al., 1999). The relationship can be approximately fitted with an exponential equation. With a focus on the shallow groundwater where the vertical change can be reasonably neglected, a horizontal distribution of
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Fig. 2. Horizontal distribution of nitrate-N concentration in groundwater and spatial relation to groundwater flow (a) [arrows roughly indicating direction of groundwater flow] within the study area (b).
nitrate-N concentration can be assumed to reflect the spreading dynamics of pollution. If an arbitrarily stratified depth of 60 m was assumed as reasonable boundary and all groundwater
samples from wells were divided into the shallow and deep groups, then significant contrasts for the two groups are as follows: average depth at 30.7 m versus 150.6 m, number of samples, accounting for 637
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Fig. 3. Vertical distribution of nitrate-N concentration in groundwater.
versus 415 samples, nitrate-N concentration ranged from 0.02 to 79.7 mg l 1 with a mean of 7.0 8.2 mg l 1 versus 0–41.2 mg l 1 with a mean of 3.8 4.8 mg l 1. Large variation of nitrate-N concentration in the shallow group might be due to strong disturbance such as frequent water pumping or mixing with percolating water from surface via crop root zones, while small variation in the deep groundwater might reflect a relatively static status of free diffusion of nitrate. As shown in Fig. 4, nitrate-N concentrations in groundwater at the depth of 60 m and above were not significantly correlated with the depth (P = 0.286). However, a significantly negative correlation between nitrate-N concentrations and sampling depths (P < 0.001) was found for the depths below 60 m. At the depth of 200 m, nitrate-N concentration approached to 2 mg l 1. The results shown in Fig. 4 imply that human activities had exerted an impact on groundwater as
deep as 150–200 m. Nitrate-N concentration in groundwater below 200 m was 1.4 2.1 mg l 1 based on 142 samples. At depths from 100 to 350 m, six samples were suspicious for unusually high nitrate-N concentration (about 10 mg l 1). This could be a problem caused either by direct leakage of nitrate-N from the polluted shallow water through unclosed well wall (Anonymous, 1993) or by unexpected sampling errors. Excluding the abnormal samples, nitrate-N concentration in the remaining 136 samples gave an average of 0.97 0.90 mg l 1, which fell below the critical level (3 mg l 1) for undisturbed groundwater as defined by the Environmental Protection Agency of the United States (Anonymous, 1987). Therefore, we use 2 mg l 1 of nitrate-N as a background value for the study area. Any concentration above the threshold will be a suspect for nitrate-N pollution due to human activities. Surface water may affect distribution patterns of nitrate-N in shallow groundwater in some regions. However, the surface water system in the study region seems not abundant enough to exert influence on the pollution distribution. Except for the Dongzhulong and Xiaoqing Rivers, all other rivers in the county temporarily reserved water during wet season and kept dry in other seasons (Anonymous, 1993). Although Dongzhulong River was mainly polluted by effluents of the waste water from the upstream city, Zibo, there was no obvious pattern for nitrate-N pollution along the river. It should be noticed that there was an area low in nitrate-N concentration under a hydraulic head with relatively shallow water table in the southwest part of Mata Lake. A clean water channel was constructed for discharging fresh water to the lake in fall. That could be an evidence of the effect of surface water on shallow water pollution. Nevertheless, it is supposed that the seasonal river could affect the pollution distribution in the shallow groundwater though it was not observed directly from the pollution map. Further investigation is needed.
4. Conclusions
Fig. 4. Averaged nitrate-N concentration of each arbitrary stratum in relation to sampling depth.
This study clearly revealed the association of spatial patterns of nitrate-N pollution and groundwater flow directions. A new outlook on the relationship
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between land use and N pollution should be enhanced by recognizing this mechanism. It is reasonable to conclude that land use types or cropping systems might be a necessary condition but not the only condition to be taken into account for groundwater pollution at a larger scale. Our findings support the conclusion that elevation of nitrate-N concentration in groundwater is a synchronizing result of intensive farming though it has not been concluded by some other studies. Agriculture-derived nitrate-N pollution dynamically moves or diffuses in three dimensions within a watershed. Any regional water cycles may divide a watershed into recharge and discharge sub-regions. Hydraulic conditions may play an important role on the region’s vulnerability to agriculture-derived nonpoint source pollution. To some extent, nitrate-N distribution patterns and hydraulic properties in a region can spatially indicate with each other. Our study suggests that the vulnerability of agriculture-derived nitrate-N loading in different regions should vary with differences in differential mechanism of groundwater pollution. The importance of understanding this process is to help guide land use planning, alleviate environmental impacts, and minimize ecologic risk of intensive farming by taking watershed as an ecologic unit instead of administrative division.
Acknowledgements This study was part of the project. ‘‘Modeling Approach of Estimate Environmental Cost Based on Budgets and Cycles of Carbon and Nitrogen in AgroEcosystem’’ (30270220), funded by the National Natural Science Foundation of China (NSFC). The authors wish to acknowledge endeavours of the government of Huantai County for organizing groundwater sampling. We express our gratitude to Huantai Water Conservation Bureau for providing the hydraulic and geologic data. The constructive comments and suggestions on the manuscript given by Peter Kyveryga, Brad van de Woestyne and other anonymous reviewers are greatly appreciated. We would like to thank participating colleagues and farmers for their assistance and cooperation during the field experiment setup and data collection.
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References Agrawal, G.D., Lunkad, S.K., Malkhed, T., 1999. Diffuse agricultural nitrate pollution of groundwaters in India. Water Sci. Technol. 29 (3), 67–75. Anonymous, 1987. Nitrate: a question of time? Water Quality International, No.1, IAWPRC, London, pp. 24–28. Anonymous, 1993. Report of exploiting groundwater of Huantai county, Shandong Province, Geologic Environment Monitor Station of Ministry of Geologic Mineral Resources, pp. 20– 28 (in Chinese). Anonymous, 2002. Report of Agriculture Development of China. Ministry of Agriculture, The People’s Republic of China, China Agriculture Press, Beijing, pp. 129–132 (in Chinese). Fan, A.M., Steinberg, V.E., 1996. Health implications of nitrate and nitrite in drinking water. An update on methemoglobinemia occurrence and reproductive and developmental toxicity. Regul. Toxicol. Pharmacol. 23, 35–43. Forman, D., Al-Dabbagh, S., Doll, R., 1985. Nitrates, nitrites and gastric cancer in Great Britain. Nature 313 (21), 620–625. Fried, J.J., 1991. Sources of nitrates in fissure groundwater in the humid tropical zone—the example of Ivory Coast. J. Hydrol. 113, 231–264. Gelberg, K.H., Church, L., Casey, G., London, M., Roerig, D.S., Boyd, J., Hill, M., 1999. Nitrate levels in drinking water in rural New York State. Environ. Res. Sect. A 80, 30–40. Gulis, G., Czompolyova, M., Cerhan, J.R., 2002. An ecologic study of nitrate in municipal drinking water and cancer incidence in Trnava District, Slovakia. Environ. Res. Sect. A 88, 182–187. Hudak, P.F., 1999. Chloride and nitrate distributions in the Hickory aquifer, central Texas. U.S.A. Environ. Int. 25 (4), 393–401. Hudak, P.F., 2000. Regional trends in nitrate content of Texas groundwater. J. Hydrol. 228, 37–47. Hudak, P.F., Blanchard, S., 1997. Land use and groundwater quality in the Trinity Group outcrop of north-central Texas. U.S.A. Environ. Int. 23 (4), 507–517. Knobeloch, L., Krenz, K., Anderson, H., Hovel, C., 1992. Methemoglobinemia in an infant—Wisconsin. Morbidity Mortality Weekly Report 42 (12), 217–219. Liu, G.D., Wu, W.L., 2002. The dynamics of nitrate nitrogen leaching through soil in high-yield farmland ecosystem. Chin. J. Eco-Agric. 10 (4), 71–74 (in Chinese). Liu, G.D., Wu, W.L., 2003. The dynamics of soil nitrate leaching and contamination of groundwater in high-yielding-farmlands. Chin. J. Eco-Agric. 11 (1), 91–93 (in Chinese). Liu, H.B., Lei, B.K., Zhang, Y.G., Zhang, W.L., Lin, B., 2001. Investigation and evaluation on nitrate pollution of groundwater of Shunyi District. Plant Nutr. Fert. Sci. 4, 385–390 (in Chinese). Lu, D.Q., Tong, Y.A., Sun, B.H., 1998. Study on effect of nitrogen fertilizer use on environment pollution. Plant. Nutr. Fert. Sci. 4 (1), 8–15 (in Chinese). Ma, L.S., Qian, M.R., 1987. Contamination of nitrate and nitrite of water environment in Taihu Lake watershed. Environ. Sci. 8 (2), 60–65 (in Chinese). Mclay, C.D.A., Dragten, R., Sparling, G., Selvarajah, N., 2001. Predicting groundwater nitrate concentrations in a region of
220
G.D. Liu et al. / Agriculture, Ecosystems and Environment 107 (2005) 211–220
mixed agricultural land use: a comparison of three approaches. Environ. Pollut. 115, 191–204. Mutch, J., 1998. The hydrologic cycle and water movement. Nitrate – agricultural sources and fate in the environment – perspectives and direction. In: Proceedings of the workshop. Eastern Canada Soil and Water Conservation Centre, pp. 3–7. Nikolaidis, N.P., Heng, H., Semagin, R., Clausen, J.C., 1998. Nonlinear response of a mixed land use watershed to nitrogen loading. Agric. Ecosyst. Environ. 67, 251–265. Nolan, B.T., Ruddy, B.C., Hitt, K.J., Helsel, D.R., 1997. Risk of nitrate in groundwaters of the United States—a national perspective. Environ. Sci. Technol. 31, 2229–2236. Singh, B., Sekhon, G.S., 1978-9. Nitrate pollution of groundwater from farm use of nitrogen fertilizers—a review. Agric. Environ. 4, 207–225. Singh, B., Singh, Y., Sekhon, G.S., 1995. Fertilizer-N use efficiency and nitrate pollution of groundwater in developing countries. J. Contam. Hydrol. 20, 167–184. Spalding, R.F., Exner, M.E., 1993. Occurrence of nitrate in groundwater—a review. J. Environ. Qual. 22, 392–402. Strebel, O., Duynisveld, W.H.M., Bottcher, J., 1989. Nitrate pollution of groundwater in Western Europe. Agric. Ecosyst. Environ. 26, 189–214.
Townsend, M.A., Sleezer, R.O., Macko, S.A., 1996. Effects of agricultural practices and vadose zone stratigraphy on nitrate concentration in ground water in Kansas. U.S.A. Water Sci. Technol. 33 (4–5), 219–226. Van Loon, A.J., Botterweck, A.A., Goldbohm, R.A., Brants, H.A., van Klaveren, J.D., van den Brandt, P.A., 1998. Intake of nitrate and nitrite and the risk of gastric cancer: a prospective cohort study. Br. J. Cancer 78 (1), 129–135. Wang, Y.M., Li, K.F., 1991. Integrated techniques of cultivation for producing grain yield of 15 t ha 1. J. Shandong Agric. Sci. 3, 9– 11 (in Chinese). Yuan, F.M., Chen, Z.M., Yao, Z.H., 1996. Nitrogen leaching from soils. In: Chen, Z.M. (Ed.), Nitrogen, Yield and Environment. Chinese Agri-Tech Press, Beijing, p. 191. Zhang, W.L., Tian, Z.X., Zhang, N., Li, X.Q., 1996. Nitrate pollution of groundwater in northern China. Agric. Ecosyst. Environ. 59, 223–231. Zhang, Z.X., Yu, G.R., 2002. Counter measures on sustainable utilization of agricultural water resources in high-yield grain area in north China—a case study in Huantai county, Shandong Province. Resour. Sci. 24 (1), 68–71 (in Chinese). Zhu, Z.L., 2000. Loss of fertilizer N from plant–soil system and the strategies and techniques for its reduction. Soil Environ. Sci. 9 (1), 1–6 (in Chinese).