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Article

Nitrate Source and Transformation in Groundwater under Urban and Agricultural Arid Environment in the Southeastern Nile Delta, Egypt

by
Alaa M. Kasem
1,2,3,
Zhifang Xu
1,2,4,*,
Hao Jiang
5,
Wenjing Liu
1,2,4,
Jiangyi Zhang
1,2,4 and
Ahmed M. Nosair
6
1
Key Laboratory of Cenozoic Geology and Environment, Institute of Geology and Geophysics, Chinese Academy of Sciences, Beijing 100029, China
2
University of Chinese Academy of Sciences, Beijing 100049, China
3
Department of Natural Resources, Faculty of African Postgraduate Studies, Cairo University, Giza 12613, Egypt
4
CAS Center for Excellence in Life and Paleoenvironment, Beijing 100044, China
5
Key Laboratory of Aquatic Botany and Watershed Ecology, Wuhan Botanical Garden, Chinese Academy of Sciences, Wuhan 430074, China
6
Environmental Geophysics Lab (ZEGL), Geology Department, Faculty of Science, Zagazig University, Zagazig 44519, Egypt
*
Author to whom correspondence should be addressed.
Water 2024, 16(1), 22; https://doi.org/10.3390/w16010022
Submission received: 21 November 2023 / Revised: 15 December 2023 / Accepted: 18 December 2023 / Published: 20 December 2023
(This article belongs to the Special Issue Impacts of Environmental Change and Human Activities on Aquatic Ecosystems)
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Abstract

:
With the intensification of human activities, nitrate pollutants in groundwater are receiving increasing attention worldwide. Especially in the arid Nile Delta of Egypt, groundwater is one of the most valuable water resources in the region. Identifying the source of nitrate in groundwater with strong human disturbances is important to effective water resource management. This paper examined the stable isotopes (δ15N/δ18O-NO3 and δ2H/δ18O-H2O) and the hydrogeochemical parameters of the shallow groundwaters in the arid southeast of the Nile Delta to assess the potential sources and transformation processes of nitrate under severe urban and agricultural activities. The results revealed that the groundwaters were recharged by the Nile River. Meanwhile, the infiltration of irrigation water occurred in the west, while the mixing with the deep groundwater occurred in the east regions of the study area. The TDS, SO42−, NO3−, and Mn2+ concentrations of groundwaters (n = 55) exceeded the WHO permissible limit with 34.6%, 23.6%, 23.6%, and 65.5%, respectively. The NO3− concentrations in the shallow groundwaters ranged from 0.42 mg/L to 652 mg/L, and the higher levels were observed in the middle region of the study area where the unconfined condition prevailed. It extended to the deep groundwater and eastward of the study area in the groundwater flow direction. The δ15N-NO3 and δ18O-NO3 values suggested that the groundwater NO3− in the west and east regions of semi-confined condition were largely from the nitrification of soil organic nitrogen (SON) and chemical fertilizer (CF). In contrast, wastewater input (e.g., domestic sewage and unlined drains) and prevalent denitrification were identified in the middle region. The denitrification might be tightly coupled with the biogeochemical cycling of manganese. This study provides the first report on the groundwater NO3− dynamics in the Nile Delta, which generated valuable clues for effective water resource management in the arid region.

Graphical Abstract

1. Introduction

Groundwater nitrate (NO3−) pollution resulting from human activities is a pressing global concern [1,2]. Particularly, regions with high population density and intensive land use face increased vulnerability to this contamination [3,4,5,6,7,8]. This issue is prevalent in arid and semi-arid areas where water scarcity is already a significant challenge [9,10,11]. High NO3− levels in groundwater cause ecological issues such as eutrophication of nearby surface water and hypoxia [1,12,13,14,15,16,17,18]. Furthermore, NO3− in drinking water has been associated with health risks such as methemoglobinemia and cancer [19,20,21,22]. In response to the prevailing health concerns associated with NO3− contamination, the World Health Organization (WHO) has set a maximum threshold of 50 mg/L for drinking water [23].
To effectively manage NO3− pollution in groundwaters, a comprehensive understanding of its sources and transformation processes is necessary. While NO3− stems from various sources, including agricultural practices, wastewater, organic matter in the soil, atmospheric nitrogen deposition [24,25,26], and complex transformation processes such as denitrification and nitrification [27,28,29] make it more difficult to determine the exact source of nitrate. The use of isotopic tracer techniques, particularly dual NO3− stable isotopes, has been widely employed in various areas to investigate NO3− sources and transformations based on the distinct isotopic compositions of NO3− sources and the predictable isotopic fractionations [22,30,31,32,33,34,35,36,37,38]. For instance, studies have utilized the dual NO3− isotopes to explore the relationship between NO3− contamination in groundwaters and agricultural practices [39] and urban activities [30].
Despite the widespread use of this approach, uncertainties can arise when attempting to identify NO3− sources [40,41,42]. First, isotopic compositions of NO3− arising from various sources can exhibit some degree of overlap [38]. Second, the isotope fractionation process may obscure the initial isotopic compositions [38]. Therefore, other isotopes (e.g., δ2H/δ18O-H2O and δ11B), water chemistry (e.g., Cl− and SO42−), and hydrogeochemical parameters were jointly applied to enhance the accuracy of NO3− source identification [22,43,44,45,46,47,48]. The integration of these additional data sets is expected to provide valuable and definitive insights into both water and NO3− cycling dynamics.
In the arid Nile Delta, the groundwater is a valuable resource because people rely on it for different purposes. The growing population, rapid urbanization, and increased usage of chemical fertilizers have led to severe NO3− pollution in the groundwaters. The authors of [49] highlighted significant urban expansion in the Nile Delta region from 1987 to 2015. It is presumed that this expansion has led to an increase in the discharge of untreated wastewater into water bodies. Furthermore, the two main wastewater unlined drains (Belbies and Kalyobiya) cross the study areas. These pollutants will inevitably infiltrate into the groundwater. High NO3− levels in the groundwaters have been reported (210 mg/L); however, the sources of pollution have not yet been identified, and the driving mechanisms still remain to be studied. There has been a lack of isotopic studies on NO3− pollution in the groundwaters in this region [50,51]. This seriously hinders the implementation of groundwater nitrate pollution prevention and control work in the region.
Located at the eastern fringe of the Nile Delta, the studied area is characterized with unlined drainage channels crisscrossing, a wide distribution of illegal shallow wells, and the expansion of urban areas, which further deteriorate the problem of groundwater pollution. In this study, for the first time, we use an integrating approach of a multiple isotope tracer (δ15N/δ18O-NO3 and δ2H/δ18O-H2O) combined with hydrogeochemical parameters to identify the sources of NO3− contamination and its transportation and transformation processes in the shallow groundwater of this area. The results are expected to provide insights to the effective management of the water resources in the arid region with intense human disturbance.

2. Materials and Methods

2.1. Study Area

The study area is located in the southeastern Nile Delta, a natural extension of the Nile Delta floodplain stretching toward the east. It extends between longitudes 31°3′22.42″–31°40′25.4″ E and latitudes 31°14′4.87″–30°37′16.87″ N and covers an area of approximately 2050 km2 (Figure 1a). The study area had a population of roughly 12 million people, and the population density was calculated to be 5800 individuals/km2 [52]. The land use is dominated by the interrelation of rural–urban and agricultural lands (Figure S2). The cultivation in the area primarily relies on Nile River water and groundwater for irrigation as a water source. The climate is arid and rainless in summer (May to September), with temperatures ranging from 30 to 40 °C, while the winter season (November to February) is relatively mild, with precipitation ranging from 10 to 20 mm/year and temperatures ranging from 10 to 20 °C. The study area is typically characterized by its flat plains forming the Delta landscape, gently sloping toward the north–northeast and east. It exhibits low elevation and is surrounded by a moderately elevated plateau to the south and southeast (Figure 1b).
The Quaternary aquifer is semi-confined and consists of two main units, the semi-pervious upper Holocene silt, sand, and silty clay unit (upper Holocene layer) (Figures S1 and S3), with thickness ranging from 0 to 20 m (Figure 1c) [53]. It is underlain by unconsolidated Pleistocene sand and gravel with clay lenses, with thickness ranging from a few meters in the south to about 400 m north (Figure 1b and Figure S3). Moreover, the basalt and clay layer underlain the Quaternary aquifer acts as an aquiclude south of the study area, where the aquifer is thin (Figure S4). In contrast, in other areas, the Quaternary is uncomfortably underlain by the Tertiary aquifer, which is hydraulically connected particularly with the pumping increase from the Quaternary aquifer [54,55]. The consequences of the high extraction due to the irrigation of the newly reclaimed areas at the southeast corner of the study area create a cone of depression. At the same time, the increase in wastewater and surface water infiltration to the groundwater south of Shibean Elqanater City has increased the water table (Figure 1c).
The groundwater depth was measured in January 2022; it ranges from 2 m to 47 m, with the shallowest levels found near Khanka city and the deepest levels located in newly reclaimed areas southeast of the study area, at well SH35 (Figure 1a). The elevation of the water table fluctuates between 14 and −1 masl, with the highest values observed in the south and southwestern part of the study area, in the Khanka city and its vicinity. The groundwater flow direction exhibits a southwest-to-northeast pattern in the central, southern, and northern regions and a northwest-to-southeast direction in the southeastern part of the study area (Figure 1c). The groundwater recharge primarily occurs through two main sources: Nile River and irrigation canal seepage and infiltration of excess irrigation water, industrial usage, and surface water bodies (e.g., cesspools and drains) [51,56,57]. On the other hand, groundwater discharge is mainly attributed to pumping withdrawals for various purposes, such as agricultural irrigation, drinking water supply, and industrial usage.

2.2. Sample Collection

A total of 71 samples (16 surface water and 55 shallow groundwater) were collected between December 2022 and January 2023. The surface water samples were collocated as an endmember to represent the potential pollution sources for the shallow groundwater, not for water quality purposes. There are 4 samples from two different wastewater drains (Bilbies and Kalyobiya drains), 2 from Brackish lakes, 2 from agricultural drains, and 8 from surface freshwater. The majority of wells were established for private domestic purposes, with depths ranging from 9 to 43 m below the ground surface, and the average depth is 29.7 m. Each well was pumped 5 to 10 min before sampling to eliminate any stagnant water in the pipes. Of the 55 groundwater samples, 3 were from relatively deep agricultural wells (well depths of more than 50 m), and 52 were from shallow domestic wells. All samples were passed through 0.45 μm cellulose membranes and then collected into pre-washed high-density polyethylene (HDPE) bottles without headspace to avoid evaporation for the analysis of dissolved ions, heavy metals, δ2H/δ18O-H2O, and δ15N/δ18O-NO3. The samples were acidified for major cation and heavy metal analysis by adding a few drops of ultrapure nitric acid to lower the pH below 2 and stored in 60 mL HDPE bottles. Another 125 mL was stored in HDPE bottles for anion, δ15N-NO3, and δ18O-NO3 analyses. In addition, two 15 mL bottles were prepared for other anions and δ2H-H2O and δ18O-H2O, respectively. The collected samples were carefully stored in a cold and dark environment until they can be transported back to the laboratory for analysis.

2.3. Analytical Methods

In situ measurements of water temperature (T), electrical conductivity (EC), and pH were conducted using portable pH (JENCO-6010N, USA) and EC (JENCO-EC3175, USA) meters. In addition, the HCO3− concentration was determined by titration within 24 h after sampling using HCl and Methyl Orange as the indicator [58]. The concentrations of Cl−, SO42−, F−, SiO2, NO3−, and NO2− were analyzed through ion chromatography (IC, DIONEX, ICS-1500, USA), with analytical precisions of ±5%. The concentration of cations (K+, Na+, Mg2+, and Ca2+) and trace elements (As3+, B3+, Li+, Sr3+, Al3+, Mn2+, Cr3+, Fe2+, Co2+, Cu2+, Zn2+, Ba2+, and Ni2+) were measured by an inductively coupled plasma optical emission spectrometer (ICP-OES, IRIS Intrepid II XSP, USA) with analytical precisions of ±5%. All the analyses were conducted at the Key Laboratory of Cenozoic Geology and Environment, Institute of Geology and Geophysics, Chinese Academy of Sciences, Beijing. δ2H-H2O and δ18O-H2O were analyzed using an IRMS at the Institute of Resources and Environment, Henan Polytechnic University, and the isotope data were reported in per mil (‰) relative to the ratios of the Vienna Standard Mean Ocean Water (VSMOW). The analytical precisions were ±0.5‰ for δ2H and ±0.1‰ for δ18O. The δ15N-NO3 and δ18O-NO3 analyses were conducted with the bacterial denitrification method [59,60]. This method examines the isotopic analysis of nitrous oxide (N2O) reduced from NO3− through the activity of denitrifying bacteria that lack N2O reductase activity, which is analyzed using an IRMS at the Environmental Stable Isotope Lab of the Chinese Academy of Agricultural Sciences. The isotope data were reported in per mil (‰) relative to atmospheric N2 (AIR) and VSMOW. The analytical uncertainties were 0.3‰ for the δ15N-NO3− and 0.1‰ for the δ18O-NO3−.

3. Results and Discussion

3.1. Spatial Distribution of NO3− in the Shallow Groundwaters

Significant variability in NO3− concentrations across the studied area was observed (Figure 2a). The NO3− concentrations mainly ranged from 0.42 mg/L to approximately 350 mg/L, except in well SH26 (9 m depth, with a value of 651.8 mg/L). The presence of basalt at the bottom of this well acts as a NO3− trapping mechanism, forming a concentrated NO3− pool. Overall, 23.6% of the groundwater samples exceeded the permissible limit of 50 mg/L established by WHO [23]. The high NO3− levels were governed by unconfined conditions and the thickness of the aquifer, resulting in higher concentrations in the Abu Zaabal area (i.e., vicinity of Khanka city) at the central of the study area, extending toward the northwest and east regions following the groundwater flow path (Figure 2a). The study area was characterized into three regions based on the NO3− concentrations: the west, middle, and east regions. The west region exhibited low NO3− concentrations, which can be attributed to the dilution process from recharging by the Nile River (Damietta branch) or the presence of high thickness of the upper Holocene unit that regulates the infiltration of NO3− into the shallow groundwater. The middle region was generally characterized by unconfined conditions (Figure 1c), along with a small thickness of the aquifer, making it highly vulnerable to the infiltration of surface pollutants. Moreover, due to the point pollution sources (e.g., cesspools and drains), NO3− exhibited a wide concentration range [57,61]. Finally, the east region showed little impact from surface contaminants due to the increasing thickness of the upper Holocene layer.

3.2. Spatial Distribution of Hydrogeochemical Parameters

The hydrogeochemical parameters in the shallow groundwater aquifer exhibited a distinct spatial distribution pattern (Figure 2 and Figure S5). The pH values ranged from 6.2 to 7.85, with a mean value of 7.08, indicating the groundwater’s slightly acidic to neutral nature. The EC varied from 700 µS/cm to 3650 µS/cm, with an average of 1476 µS/cm (Table 1). The total dissolved solids (TDS), a critical indicator of dissolved chemical concentrations, exhibited a range of 455 mg/L and 2373 mg/L, with a mean value of 960 mg/L. The waters displayed elevated salinity levels, primarily attributed to the seepage of surface pollutants, particularly in the middle region. Furthermore, the increased extraction from well SH35 in the southeast part contributed to more significant mixing with the deep saline aquifer (Miocene aquifer) (Figure 2b). These observations aligned with previous studies by [55,57,62]. The SO42− and Cl− concentrations ranged from 2.5 mg/L to 827.2 mg/L (mean: 200.7 mg/L) and from 33.3 mg/L to 604 mg/L (mean: 170.3 mg/L), respectively. Compared with the previously reported values [61], the current study showed higher SO42− and Cl− concentrations, which can be attributed to the enhanced seepage of surface pollutants, particularly notable in the middle region (Figure 2c,d).
The high SO42− and Cl− concentrations in the middle region of the study area were consistent with the high NO3− concentrations (R2 = 0.2; p = 0.3 and R2 = 0.33; p < 0.01, respectively). As SO42− and Cl− are typically enriched in wastewater, the correlations strengthened the scenario of increased recharge from the surface contaminant source. However, the insignificant Cl− and NO3− correlation can be attributed to biochemical mediated NO3− concentration attenuation. In contrast, well SH35 in the southeast part records high SO42− and Cl− values, which can be mainly due to the mixing with the saline water from the underlain Miocene aquifer. In the west and east regions, the levels of these anions were not high, likely due to the thick upper Holocene layer in these regions and the dilution process in the west (Figure 2c,d). The HCO3− concentration showed high values in the west and southwest areas and decreased from the Damietta branch toward the east, with values ranging from 165.2 mg/L to 753.4 mg/L (mean: 390.9 mg/L).

3.3. Groundwater Recharge

The stable isotopes of water (δ2H/δ18O-H2O) provide valuable insights into the origin and movement of groundwater, rendering them indispensable tracers in hydrological investigations. The Global Meteoric Water Line (GMWL) is a crucial tool to interpret the isotopic tracers as it is a reference for understanding fractionation and mixing processes in natural water circulation [63,64]. The Nile River, irrigation canals, and drains were the potentially important shallow groundwater sources. However, before they seeped into the groundwater, the water can undergo isotopic fractionation associated with evapotranspiration during the infiltration from the ground surface, leading to enriched values. The Quaternary aquifer can be isotopically depleted because of the hydraulic connection with the depleted underlain Tertiary aquifer.
The stable isotope composition of water samples gathered from the study area is depicted on the conventional δ18O-H2O and δ2H-H2O diagram (Figure 3). The groundwater values for δ18O-H2O and δ2H-H2O exhibit considerable diversity, spanning from −1.86‰ to 3.32‰ and from −4.46‰ to 26.61‰, respectively (Table 1). The water samples deviate below the GMWL (Figure 3), suggesting that they had been influenced by evaporation, in contrast with the old Nile water sample before Aswan High Dam (AHD) compiled by [65], which was distributed close to the GMWL. Most of the samples were distributed near the Nile water, the irrigation return flow, and wastewater endmembers (Figure 3), suggesting that these sources primarily recharged the shallow groundwater.
In particular, the groundwater in the middle and western regions of the study area has undergone modern recharge, as indicated by its isotopic composition closely resembling that of the Nile water. In the west region, the isotopic values of groundwater experienced noticeable enrichment, likely due to an increase in evaporation (Figure 4). Several factors can contribute to this phenomenon. Firstly, the availability of Nile water and the prevalent traditional irrigation method, such as flood irrigation, enhance evaporation due to prolonged time on the surface before seeping downward. Secondly, the shallow depth of the water table contributed to the enrichment of the isotopic signature. Lastly, the presence of a thick upper Holocene unit exacerbated the evaporation process. Conversely, in the middle region, where this layer was absent, and there was a shortage of Nile water coupled with increased groundwater abstraction to compensate for this shortage, the isotopic composition of the groundwater remained closer to that of the Nile water and drainage systems. In the east region, where the upper Holocene layer is thick and there is shortage of the Nile water, the abstraction of the groundwater increased, which resulted in well nos. SH43, SH17, SH35, SH50, SH46, and SH38.2 being affected by mixing with the Tertiary aquifer. These findings are consistent with previous studies conducted by [51,66]. These results show that, although the NO3− groundwater pollution in the west/east regions was low, it can result from irrigation water enriched with CF and SON. This low concentration was due to the dilution of excess irrigation water and the Nile water recharging in the west.
In contrast, in the east, the dilution that resulted from the mixing with the deep groundwater contributed to the low NO3− concentrations. However, the relatively high NO3− (28.8 mg/L in well SH35) southeast of the study area could have originated from the lateral groundwater flow with high NO3− concentration due to relatively intense abstraction (250 m3/day). In contrast, in the middle region, more wastewater is loaded into the groundwater due to unconfined conditions.

3.4. Sources and Transformations of NO3−

Cl− serves as a reliable indicator of wastewater and fertilizer impacts, which can contribute to elevated Cl− levels and is unaffected by chemical, physical, and biological processes [45,46,67]. The correlation between Cl− and NO3−/Cl− is widely utilized for source identification of NO3− sources in water, specifically for distinguishing agricultural or/and wastewater sources [45,47] and for distinguishing between the influences of dilution and denitrification on NO3− levels [68]. A high NO3−/Cl− ratio coupled with low Cl− levels indicates that agricultural sources (e.g., chemical fertilizer) and soil organic nitrogen (SON) are the main contributors of NO3− in the groundwater. Conversely, a low NO3−/Cl− ratio and high Cl− content suggest wastewater sources as the primary origin of NO3− contamination [27,69].
Most of the groundwater samples had NO3−/Cl− ratios < 1 and relatively high Cl− concentrations (Figure 5a), suggesting a probable dominance of wastewater as the NO3− source [68,70]. However, a few samples in the middle region (SH26 and SH36; Figure 5a) exhibited high NO3−/Cl− ratios > 1, which can be linked to a combination of SON and CF with wastewater inputs [70] (Figure 5a). The significant correlation between SO42− and NO3− of the groundwaters in the middle region supported the dominant role of wastewater (Figure 5b and Figure 6b) [71,72,73]. A few points were scattered from the correlation, which can be ascribed to the evaporate (e.g., gypsum) dissolution in the arid region [74,75] or as a result of using CF enriched with SO42− [13,76].
In addition, most of the groundwater samples in the study area exhibited TDS concentrations below 1000 mg/L associated with a low (NO3− + Cl−)/HCO3− ratio. The correlation is strongly positive (R2 = 0.71; p < 0.01) in the middle region, which can be attributed to the substantial reduction condition and then denitrification, which attenuates the NO3− and/or produce more HCO3− [70,77] (Figure 5c). This speculation can then be consistent with the theoretical denitrification defined in Equation (1) [78,79]. In the west and east regions, this relationship is a weak positive correlation as a result of the impacts of SON and CF [73,80] associated with biodegradation of organic matter, which is rich in the upper Holocene layer in these regions and consequently increase HCO3− concentration [81].
NO3− + 1.08 CH3OH + 0.24 H2CO3 → 0.56 C5H7O2N + 0.47 N2 + HCO3− + 1.68 H2O
The dual NO3− isotope approach (δ15N-NO3 and δ18O-NO3) is widely recognized as a powerful technique for constraining NO3− sources and behaviors [9,38,45,50,82]. Different NO3− sources exhibit characteristic isotopic signatures of δ15N-NO3 and δ18O-NO3, which can be used to trace their origins [38,83,84]. Figure 7a shows the δ15N-NO3 and δ18O-NO3 values of the groundwaters, Nile water, and potential endmembers in the study area. The isotopic compositions of δ15N-NO3 and δ18O-NO3 in the groundwaters showed wide ranges, ranging from −23.4‰ to 75.4‰ (mean: 8.92‰) and from −14.3‰ to 39.8‰ (mean: 10.53‰), respectively (Table 1). The groundwaters in the middle region have δ15N-NO3 and δ18O-NO3 ranging from 4‰ to 24.4‰ and from 10‰ to 23.9‰, respectively (Figure 4 and Figure 7). As expected, the data are distributed near the endmembers for SON and CF and wastewater, suggesting that these anthropogenic sources should be the major sources of the groundwaters (Figure 7a). Due to their analogous isotopic compositions and transport routes, we merged SON and CF as one source. The direct input of atmospheric precipitation should be minor because the study area experiences minimal precipitation (Figure 7a).
Notably, a number of the isotopic values, especially those in the middle region, are higher than all the endmembers, and there is a significant positive correlation between the dual isotopes with a slope of 0.48 (R2 = 0.39, p < 0.0001) (Figure 4, Figure 6b and Figure 7), which is the typical isotopic signal for biological NO3− removal [38]. Denitrification in groundwaters is well documented, leading to an exponential co-increase in δ15N-NO3 and δ18O-NO3, with values exceeding 100‰ recorded [38]. The significant denitrification is also supported by the simultaneous decreasing NO3− levels and increasing dual isotopic values (Figure 6b and Figure 7a). The observation gives us the confidence that the prevalent denitrification process counterbalanced the excess NO3− loading to the groundwater to some extent (Figure 7a). The presence of clay lenses and clayey sediments in the aquifer can enhance the denitrification [38,85] by providing soluble organic carbon and anaerobic conditions [77]. However, more excess NO3− loading into the groundwater can inhibit this process and result in high NO3− levels (e.g., SH26 and SH32).
By contrast, no relationships existed between the dual isotopes in the samples collected in the west/east regions (Figure 6a and Figure 7b), indicating minor denitrification. The groundwater samples in the west and east regions exhibited relatively depleted and constrained δ15N-NO3 values (Figure 4 and Figure 7a). Their δ15N-NO3 and δ18O-NO3 compositions ranged from −23.3‰ to 4‰ and from −13.6‰ to 10‰, respectively. Most of these samples had δ15N-NO3 values lower than most 15N-depleted SON and CF endmember, which can be ascribed to nitrification processes. Nitrification has a substantial fractionation effect, with 14N being preferentially nitrified in NH4+ plentiful environments. The anthropogenic activities should discharge a large amount of NH4+ containing wastewater and irrigation discharge to the groundwaters, resulting in NH4+ plentiful conditions. Yet, the overall anaerobic environments in the groundwaters would inhibit nitrification. The speculation is supported by the low NO3− concentrations (Figure 4 and Figure 7a). Collectively, the low isotopic values and concentrations of NO3− in the west/east regions should be attributed to the nitrification process, which also can be enhanced in the west due to the long residence time of water on the surface before it seeps into the groundwater [86]. Notably, although sample SH35 is collected from a deep well (depth = 156 m), it showed a NO3− value of 28.8 mg/L; this observation can result from NO3− recharging to the deep zone through the lateral groundwater flow.
Overall, according to the water chemistry and isotopic tracers, we found that wastewater and SON and CF were the major sources of shallow groundwaters. Denitrification is the primary process that occurs in the middle region accompanied by excess NO3− loadings, whereas the west/east regions can be affected by zonal nitrification.

3.5. Identification of NO3− Transformation Process Coupled with Mn Oxides Reduction

The presence of trace elements in groundwater provides valuable insights into NO3− dynamics [12,87]. Among these trace elements, Mn2+, Fe2+, and Cu2+ play crucial roles in denitrification by enhancing the rate of this process [44]. Notably, the relationship between oxidation and reduction of Mn and Fe in natural waters extends beyond the typical redox conditions necessary for their mobilization [88]. Mn and Fe in minerals serve as electron acceptors for denitrifying bacteria [77], thereby facilitating denitrification and ultimately enhancing denitrification rates [43,44,89].
The concentrations of Mn2+ in the groundwaters ranged from 0.002 mg/L to 3.38 mg/L, with a mean value of 0.83 mg/L. Furthermore, it is noteworthy that 65.5% of the groundwater samples examined exceeded the permissible limit of 0.4 mg/L set by the WHO (Figure S6). There were significant correlations between the Mn2+ concentrations and the δ15N-NO3 and δ18O-NO3 values in the middle region (p < 0.01; Figure 6b and Figure 8a,b), implying the role of Mn in regulating the prevalent denitrification. The speculation was also supported by the negative correlation between Mn2+ and NO3− concentrations (Figure 6b and Figure 8c). In addition, in the west/east region, the above correlations were absent (Figure 6a and Figure 8a,c), further giving us the confidence that the biogeochemistry of Mn and denitrification were tightly coupled. This observed relationship was in line with the results of an experimental study that showed that MnO2 addition enhanced the denitrifying bacteria’s metabolism and removed about 99% of the NO3− and released Mn2+ in weak acidic and neutral conditions [43].
The potential source of Mn in the study area is the upper Holocene layer [90], which is also enriched in organic matter and Mn minerals, leading to reduced environments [91]. Mn is commonly present in insoluble Mn(IV) and Mn(III), and their reduction will generate soluble Mn2+ [88,90]. In addition, Mn can result from CF utilized in the fruit farms in the western region [92]. We observed co-increasing of Mn2+ with HCO3−, Fe2+, and Ba2+ (Figure 6 and Figure 8e–g). As HCO3−, Fe2+, and Ba2+ concentrations are closely related to chemical weathering, high Mn2+ in the groundwaters should be associated with chemical weathering processes. The mafic rocks are enriched with Mn [88] and the origin of the Nile silts and the Quaternary deposits as a result of weathering of these rocks [93], which lead to sediment enriched with Mn oxides. In contrast, there was no significant relationship between Mn2+ and SO42− in the middle region (Figure 6b and Figure 8h), indicating that the anthropogenic contribution of Mn2+ was less significant. However, Mn2+ and SO42− showed a significant positive correlation in the west/east regions (Figure 6a and Figure 8h), indicating the direct input of Mn2+ from anthropogenic sources [81,91]. That is, Mn was largely in an oxidized form in this layer before entering the water body, and it cannot participate in the reduction of NO3−. As the predominant factors influencing the concentrations of Mn2+ in the shallow groundwaters in the west/east and middle regions of the study area are quite different, the increase in Mn2+ concentrations and the associated redox process in the middle region enhanced denitrification, while in the west/east regions, denitrification and Mn2+ concentrations were decoupled.
However, reducing the Fe/Mn oxides coupled with nitrification–denitrification under anaerobic conditions and replacing the need for O2 can be prevalent in these areas, where NH4+ oxidized and turned to NO3− or N2 (i.e., anammox) based on Equation (2) [78].
NH4+ + 1.31 NO2− + 0.066 HCO3− + 0.13 H+ → 1.02 N2 + 0.26 NO3− + 0.066 CH2O0.5N0.15 + 2.03 H2O
These findings highlight the importance of considering the biogeochemistry of trace elements, particularly Mn2+, in assessing and managing NO3− contamination in shallow groundwater. However, it is important to note that while denitrification mitigates NO3− excess, it cannot fully offset the increasing NO3− loading into the groundwater. Due to this process, there is a higher release of Mn2+, which, when combined with an excess of NO3−, poses a potential health risk to the residents.

4. Conclusions

In the present investigation, the hydrogeochemical parameters and multi-isotope tracers were used to identify the sources and transformation of NO3− in the shallow groundwater in the southeast of the Nile Delta region for the first time. Our findings reveal that the NO3− values ranged from 0.42 mg/L to 650 mg/L and the groundwater in the middle region was highly impacted by NO3− pollution. The NO3− pollution in the west/east regions was relatively weak due to a thick upper Holocene layer or as a result of the dilution process by the Nile River recharge in the west or the mixing with deep groundwater in the east. According to the WHO guidelines for drinking water, the criterion for NO3− and Mn2+ is exceeded in 23.6% and 65.5% of shallow groundwater samples, respectively. The values of δ18O-NO3 and δ15N-NO3 between −14.32‰ and −39.79‰ and between −23.4‰ and –75.42‰, respectively, implied that wastewater and SON and CF were the main NO3− sources. NO3− in the west/east regions can be derived from SON and CF associated with zonal nitrification. While in the middle region, NO3− was more derived from the wastewater, and the denitrification was prevalent. Remarkably, the denitrification might be closely coupled with the biogeochemical cycling of Mn. Finally, the data obtained from this study bear significant implications across several dimensions. Firstly, it establishes the first reference for NO3− isotope investigations of the groundwater in Egypt. Secondly, it is poised to furnish valuable insights to coupling Mn+2 and NO3− biogeochemical dynamics in the arid region, delineating pivotal denitrification and Mn reduction relationships.

Supplementary Materials

The following supporting information can be downloaded at: https://www.mdpi.com/article/10.3390/w16010022/s1. Figure S1: Geology of the southeastern region of the Nile Delta; Figure S2: Land use/Land cover (LU/LC) 10 m resolution (year 2022) of the study area; Figure S3: Hydrogeological map of the study area; Figure S4: Hydrogeological cross-section A-A′; Figure S5: Spatial distribution maps of the groundwater in the study area: (a) EC; (b) pH; (c) Na+; (d) HCO3−; Figure S6: Spatial distribution maps of the groundwater in the study area: (a) Ca2+; (b) Mg2+; (c) Mn2+; (d) Fe2+; Figure S7: Scatterplot of Mn2+ versus SO42−/Cl− of the groundwater in the study area; Figure S8: Scatterplot of (a) ln [NO3−] versus δ18O-NO3, (b) ln [NO3−] versus δ15N-NO3 of the groundwater in the middle region of the study area; Figure S9: Scatterplot of (a) ln [NO3−] versus δ18O-NO3, (b) ln [NO3−] versus δ15N-NO3 of the groundwater in the west/east regions of the study area; Table S1: Quality parameters of the groundwater samples in the study area; Table S2: Limit of Detection of the Inductively Coupled Plasma Optical Emission Spectrometer (ICP-OES, IRIS Intrepid II XSP, USA) and ion chromatography (IC, Dionex 120, USA). References [23,53,54,94,95,96] are cited in the Supplementary Materials.

Author Contributions

Conceptualization, A.M.K., Z.X. and W.L.; methodology, A.M.K., Z.X. and W.L.; formal analysis, A.M.K., H.J., W.L. and J.Z.; investigation, A.M.K. and A.M.N.; writing—original draft preparation, A.M.K.; writing—review and editing, A.M.K., Z.X., H.J., W.L., J.Z. and A.M.N.; supervision, Z.X.; funding acquisition, Z.X. All authors have read and agreed to the published version of the manuscript.

Funding

This research was financially supported by the National Key Research and Development Program of China (grant no. 2020YFA0607700), the National Natural Science Foundation of China (grant nos. 41730857 and 42273050), and the Key Research Program of the Institute of Geology & Geophysics, CAS (grant no. IGGCAS-202204). Wenjing Liu acknowledges support from the Youth Innovation Promotion Association CAS (2019067).

Data Availability Statement

Data are available on request.

Acknowledgments

Alaa M. Kasem would like to thank the CAS-TWAS President’s Fellowship Programme.

Conflicts of Interest

The authors declare that they have no known competing financial interest or personal relationship that could have appeared to influence the work reported in this paper.

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Figure 1. Maps of the study area and the sampling sites. (a) Landsat 8 and depth to the water table; (b) digital elevation model (DEM) and aquifer thickness; and (c) thickness of the upper semi-confined aquifer layer (modified from [53]) and hydraulic head.
Figure 1. Maps of the study area and the sampling sites. (a) Landsat 8 and depth to the water table; (b) digital elevation model (DEM) and aquifer thickness; and (c) thickness of the upper semi-confined aquifer layer (modified from [53]) and hydraulic head.
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Figure 2. Spatial distribution maps of hydrogeochemical parameters of the groundwater in the study area: (a) TDS; (b) NO3−; (c) SO42−; and (d) Cl−.
Figure 2. Spatial distribution maps of hydrogeochemical parameters of the groundwater in the study area: (a) TDS; (b) NO3−; (c) SO42−; and (d) Cl−.
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Figure 3. Scatterplot of water staple isotopes (δ2H-H2O and δ18O-H2O) in groundwater and endmember samples in the study area [55,65].
Figure 3. Scatterplot of water staple isotopes (δ2H-H2O and δ18O-H2O) in groundwater and endmember samples in the study area [55,65].
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Figure 4. Spatial distribution maps of the δ2H-H2O, δ18O-H2O, δ15N-NO3, and δ18O-NO3 of the groundwater in the study area.
Figure 4. Spatial distribution maps of the δ2H-H2O, δ18O-H2O, δ15N-NO3, and δ18O-NO3 of the groundwater in the study area.
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Figure 5. Plot of (a) NO3− + Cl− versus Cl−; (b) NO3− versus SO42−; and (c) TDS versus (NO3− + Cl−)/HCO3− of the groundwater samples in the study area.
Figure 5. Plot of (a) NO3− + Cl− versus Cl−; (b) NO3− versus SO42−; and (c) TDS versus (NO3− + Cl−)/HCO3− of the groundwater samples in the study area.
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Figure 6. Plot of the Pearson correlation matrix of hydrogeochemical and isotope data of the shallow groundwater in the (a) west/east regions and (b) middle region. * p < 0.05; ** p < 0.01; *** p < 0.001.
Figure 6. Plot of the Pearson correlation matrix of hydrogeochemical and isotope data of the shallow groundwater in the (a) west/east regions and (b) middle region. * p < 0.05; ** p < 0.01; *** p < 0.001.
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Figure 7. (a) Relationship between δ15N-NO3 and δ18O-NO3 values of the groundwater and endmember samples, with the ranges of potential nitrate sources. (b) δ15N-NO3 versus δ18O-NO3 of the groundwater samples in the study area.
Figure 7. (a) Relationship between δ15N-NO3 and δ18O-NO3 values of the groundwater and endmember samples, with the ranges of potential nitrate sources. (b) δ15N-NO3 versus δ18O-NO3 of the groundwater samples in the study area.
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Figure 8. Scatterplot of (a) Mn2+ versus δ15N-NO3; (b) Mn2+ versus δ18O-NO3; (c) Mn2+ versus NO3−; (d) Mn2+ versus δ18O-H2O; (e) Mn2+ versus Fe2+; (f) Mn2+ versus HCO3−; (g) Mn2+ versus Ba2+; and (h) Mn2+ versus SO42− of the groundwater in this study.
Figure 8. Scatterplot of (a) Mn2+ versus δ15N-NO3; (b) Mn2+ versus δ18O-NO3; (c) Mn2+ versus NO3−; (d) Mn2+ versus δ18O-H2O; (e) Mn2+ versus Fe2+; (f) Mn2+ versus HCO3−; (g) Mn2+ versus Ba2+; and (h) Mn2+ versus SO42− of the groundwater in this study.
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Table 1. Descriptive statistics of analyzed parameters of shallow groundwater and endmember samples in the study area.
Table 1. Descriptive statistics of analyzed parameters of shallow groundwater and endmember samples in the study area.
ParameterFresh Surface Water (n = 7)Agricultural Drain (n = 2)Polluted Lake (n = 2)
MeanSDMinMaxMeanSDMinMaxMeanSDMinMax
Total depth
(m)		––––––––––––
pH7.390.366.707.737.680.677.208.158.410.308.208.62
Temp (°C)18.310.8017.0019.2018.602.2617.0020.2021.000.0021.0021.00
EC (µS/cm)4982447554019806651510245017,555164816,39018,720
TDS (mg/L)324153093511287432982159311,411107110,65412,168
SiO2 (mg/L)1.030.590.452.0717.495.8513.3621.6313.5418.920.1726.92
K+ (mg/L)6.100.425.666.6024.582.8422.5726.6074.115.1770.4577.76
Na+ (mg/L)11.7412.456.4639.97112.2190.1348.48175.941650141.4215501750
Ca2+ (mg/L)41.092.0136.7743.05119.6368.4671.22168.03617316393840
Mg2+ (mg/L)16.121.3913.3117.5859.0534.9234.3683.74269.74.52266.5272.9
HCO3− (mg/L)186.7411.98173.17203.33254.66108.36178.04331.28550.07352.30300.95799.18
F− (mg/L)0.490.080.390.650.240.060.190.28––––
Cl− (mg/L)31.022.6027.3335.24263.39143.39161.99364.784272.91074.735135032.8
NO2− (mg/L)––––––––––––
NO3− (mg/L)1.861.330.804.5935.5813.8025.8245.330.680.030.660.70
SO42− (mg/L)34.31.432.436.3308.592.4243.2373.83700132.936063794
Mn (mg/L)0.0040.0050.0010.0140.1980.1980.0580.3380.0720.0990.0010.142
Fe (mg/L)0.0450.0050.0370.0530.1370.0930.0710.2030.1710.0040.1680.174
As (mg/L)0.00070.00050.00020.0020.0020.0010.0010.0030.0210.0010.0210.022
B (mg/L)0.0560.0390.0260.1340.0720.0060.0680.0762.4380.4772.1002.775
Li (mg/L)0.0010.00020.00050.0010.0020.00040.0010.0020.0330.0090.0270.039
Sr (mg/L)0.3240.0230.2980.3620.8590.3420.6171.1007.8450.1487.7407.950
Al (mg/L)0.1170.0980.0100.2660.0560.0570.0160.0970.0160.0010.0160.017
Cr (mg/L)0.0110.0010.0090.0130.0110.0020.0100.0120.0330.0230.0170.050
Co (mg/L)0.0000.0000.0000.0000.0010.0000.0010.0010.0020.0000.0010.002
Cu (mg/L)0.0020.0000.0010.0020.0030.0010.0030.0030.0100.0020.0080.011
Zn (mg/L)0.0060.0020.0030.0080.0080.0010.0080.0090.0110.0000.0110.011
Ba (mg/L)0.0180.0010.0170.0190.0320.0130.0230.0410.0190.0020.0170.020
Ni (mg/L)0.0060.0000.0050.0070.0090.0020.0080.0110.0270.0020.0250.029
δ18O-H2O (‰)2.000.131.742.131.290.650.831.756.310.595.896.73
δ2H-H2O (‰)18.491.1915.9419.2014.024.7210.6917.3631.164.3628.0834.25
δ15N-NO3 (‰)5.038.47−12.7611.5211.434.608.1814.686.452.514.688.23
δ18O-NO3 (‰)9.026.65−4.6914.245.511.674.336.703.281.522.204.35
ParameterWastewater Drain (n = 4)Groundwater (n = 55)
MeanSDMinMaxMeanSDMinMax
Total depth
(m)		––––32.1120.749.00156.00
pH7.240.486.567.707.080.366.207.75
Temp (°C)19.380.8518.5020.5020.296.6318.5025.30
EC (µS/cm)14102761080168014766887003650
TDS (mg/L)91717970210929604474552373
SiO2 (mg/L)8.993.654.1613.0263.5365.5718.70237.32
K+ (mg/L)17.976.797.7921.5011.646.504.6431.61
Na+ (mg/L)132.4588.857.87200.78147.27108.887.47579.77
Ca2+ (mg/L)62.3911.7346.1671.24102.6549.6919.07366.72
Mg2+ (mg/L)21.673.1017.0223.4735.5016.059.6385.81
HCO3− (mg/L)306.2566.54246.29367.02390.90120.78165.16753.35
F− (mg/L)0.280.010.270.290.250.190.141.38
Cl− (mg/L)208.1688.72129.54312.22170.34110.8633.31603.98
NO2− (mg/L)107.2463.7450.13176.01––––
NO3− (mg/L)18.2434.580.6770.1149.90109.220.42651.79
SO42− (mg/L)114.8857.7758.31165.67200.66175.322.47827.15
Mn (mg/L)0.2270.1500.1020.4290.8310.7200.00193.380
Fe (mg/L)0.0880.0200.0690.1080.0890.0600.0400.430
As (mg/L)0.0010.0010.0010.0020.0020.0010.0000.006
B (mg/L)0.1370.0380.0900.1770.0940.1120.0160.577
Li (mg/L)0.0050.0020.0040.0070.0030.0040.0000.023
Sr (mg/L)0.9280.5350.4251.4100.9860.9470.1286.590
Al (mg/L)0.0300.0050.0240.0360.0120.0060.0070.046
Cr (mg/L)0.0130.0010.0110.0130.0110.0030.0070.023
Co (mg/L)0.0010.0000.0010.0010.0010.0000.0000.003
Cu (mg/L)0.0020.0000.0020.0030.0030.0040.0010.025
Zn (mg/L)0.0200.0140.0100.0410.0220.0560.0030.404
Ba (mg/L)0.0230.0050.0190.0300.1220.0870.0110.341
Ni (mg/L)0.0110.0010.0090.0120.0090.0030.0050.023
δ18O-H2O (‰)2.080.151.932.291.231.46−1.863.32
δ2H-H2O (‰)19.100.8718.1920.2013.998.83−4.4626.61
δ15N-NO3 (‰)8.136.10−0.1613.868.9216.87−23.4075.42
δ18O-NO3 (‰)−1.6913.54−13.2513.2010.5312.85−14.3239.79
–: not detected.
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Kasem, A.M.; Xu, Z.; Jiang, H.; Liu, W.; Zhang, J.; Nosair, A.M. Nitrate Source and Transformation in Groundwater under Urban and Agricultural Arid Environment in the Southeastern Nile Delta, Egypt. Water 2024, 16, 22. https://doi.org/10.3390/w16010022

AMA Style

Kasem AM, Xu Z, Jiang H, Liu W, Zhang J, Nosair AM. Nitrate Source and Transformation in Groundwater under Urban and Agricultural Arid Environment in the Southeastern Nile Delta, Egypt. Water. 2024; 16(1):22. https://doi.org/10.3390/w16010022

Chicago/Turabian Style

Kasem, Alaa M., Zhifang Xu, Hao Jiang, Wenjing Liu, Jiangyi Zhang, and Ahmed M. Nosair. 2024. "Nitrate Source and Transformation in Groundwater under Urban and Agricultural Arid Environment in the Southeastern Nile Delta, Egypt" Water 16, no. 1: 22. https://doi.org/10.3390/w16010022

APA Style

Kasem, A. M., Xu, Z., Jiang, H., Liu, W., Zhang, J., & Nosair, A. M. (2024). Nitrate Source and Transformation in Groundwater under Urban and Agricultural Arid Environment in the Southeastern Nile Delta, Egypt. Water, 16(1), 22. https://doi.org/10.3390/w16010022

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