Hydrochemical facies
Chadha’s classification (Chadha 1999) was proposed to identify the groundwater type. As shown in Fig. 7, the main hydrochemical facies in the study area are Na–Cl and mixed Ca–Mg–Cl. Only one of data points fall in field 5 (Ca–Mg–HCO3 water type) which indicates alkaline earths and weak acidic anions exceed both alkali metals and strong acidic anions, respectively. Such water has temporary hardness. Thirteen data points fall in the field 6 (Ca–Mg–Cl water type) which represent alkaline earths that exceed alkali metals and strong acidic anions exceed weak acidic anions. Such water has permanent hardness and does not deposit residual sodium carbonate in irrigation use. Sixteen data points fall in field 7 (Na–Cl water type) which represent alkali metals exceed alkaline earths and strong acidic anions exceed weak acidic anions. Such water generally creates salinity problems both in irrigation and drinking uses (Sutharsiny et al. 2012).
Spatial variability of groundwater quality parameters
The EC values of the groundwater samples range from 693 μS cm−1 in the close the northern slopes of the Sahand Mountains to 12,500 μS cm−1 in the southwest region of the aquifer. Low EC (sample 1) is located in the recharge areas originating from coarse-grained alluvial fans, while individual elevated EC values (e.g., sample 7) should have probably variable origins, such as the low depth of water table and subsequent increased evaporation, the elevated salinity of Aji-Chay River (average EC = 9500 μS cm−1) which is hydraulically connected to groundwater (Barzegar et al. 2016a) and, the excessive use of salts from industrial tanning activities. Spatial distribution of the EC is shown in Fig. 8a. Generally, EC values increase from recharge areas in eastern and southern edges towards discharge areas in west and southwest. However, anomalies can be observed in this trend due to geological formations, groundwater extraction amount, size and type of sediments, depth of groundwater and, interactions between aquifer systems in the area. It should be noted that the depth of sampling wells can affect the EC values, because the groundwater salinity of the confined and unconfined aquifers varies significantly. The critical threshold of 1500 μS cm−1 (World Health Organization (WHO) 2011) for drinking water purposes is exceeded in all samples except three of them (1, 8 and 9); a fact which denotes a significant environmental pressures of salinity.
The pH of samples varies between 7.5 and 8.4. Generally, samples which are in the range of pH between 6 and 9 are in their natural state (Stumm and Morgan 1996). This is due to the natural waters generally containing dissolved carbon dioxide and hydrogen ions and carbonates (Mokhtar et al. 2009) which forms a buffer system (Eq. 4 and 5).
$${\text{H}}_{2} {\text{O}} + {\text{CO}}_{2} \to {\text{H}}_{2} {\text{CO}}_{3}$$
(4)
$${\text{H}}_{2} {\text{CO}}_{3} \to 2{\text{H}}^{ + } + {\text{CO}}_{3}^{2 - } .$$
(5)
The distribution map of pH values (Fig. 8b) shows that pH is high in the eastern part of the study area due to high bicarbonate levels in groundwater, while pH levels are decreased towards the central parts due to low levels of bicarbonate and high salinity. As well as, groundwater pH increases with water residual time because the interaction of water–rock will consume H+ (Hinkle and Polette 1999). Therefore, pH is increased in the central parts of plain due to low hydraulic gradient and high water residual time.
The calcium content of the samples varies between 39 and 460 mg L−1. According to World Health Organization (WHO) (2011) standards, its permissible range in drinking water is 75 mg L−1. Calcium in the groundwater can originate from Miocene and Pliocene formations, e.g. limestone, sandstone, conglomerate, gypsum in the study area. Concentration of calcium against sodium is low, revealing a lack of soluble calcium minerals, as calcium of groundwater is replaced by sodium through ion exchange reaction (Sharma and Rao 1997).
Sodium ion appears an elevated concentrations; the amount of Na+ ion increases from recharge to discharge areas, likewise Cl− denoting their relationship. The origin of Na+ in groundwater of the study area may be attributed to the dissolution of sodium-bearing minerals (especially halite), plagioclase weathering (Eq. 6), ion exchange reactions (replacing the calcium and magnesium with sodium in clays), irrigation water return flow and municipal and/or industrial wastewater.
$$2{\text{NaAlSi}}_{3} {\text{O}}_{8} + 2{\text{CO}}_{2} + 11{\text{H}}_{ 2} {\text{O}} \to 2 {\text{Na}}^{ + } {\text{ + 2HCO}}_{ 3}^{ - } {\text{ + 4H}}_{ 4} {\text{SiO}}_{{ 4\left( {\text{aq}} \right)}} {\text{ + Al}}_{ 2} {\text{Si}}_{ 2} {\text{O}}_{ 5} \left( {\text{OH}} \right).$$
(6)
The potassium concentration is variable ranging between 13.8 and 148.5 mg L−1 (Fig. 8f), while the optimal level of potassium in drinking water is 12 mg L−1 which indicates potassium is excessive in groundwater (World Health Organization (WHO) 2011). Potassium in the groundwater of the study area can originate from weathering K-bearing minerals such as feldspars (Eq. 7), industrial activities and agricultural fertilizers potash which is used by farmers. Additionally, irrigation water return flow may increase overall K concentrations,
$$2 {\text{KAlSi}}_{ 3} {\text{O}}_{ 8} {\text{ + 2CO}}_{ 2} {\text{ + 11H}}_{ 2} {\text{O}} \to 2 {\text{K}}^{ + } {\text{ + 2HCO}}_{ 3}^{ - } {\text{ + 4H}}_{ 4} {\text{SiO}}_{{ 4\left( {\text{aq}} \right)}} {\text{ + Al}}_{ 2} {\text{Si}}_{ 2} {\text{O}}_{ 5} \left( {\text{OH}} \right)_{ 4} .$$
(7)
Spatial distribution of bicarbonate ion is shown in Fig. 8g. Bicarbonate concentrations range between 234.1 and 663.1 mg L−1. Elevated Ca concentrations may be attributed to dissolution of limestones, Ca-rich feldspars, and potentially due to biodegradation or organic matter in upper soil horizons.
Sulfate concentrations vary between 62.2 and 847.6 mg L−1 (Fig. 8h). Sulfate concentration greater than 50 mg L−1 causes a bitter taste in water and in higher concentration of 400 mg L−1 with calcium and magnesium can cause frailty in the body (Shankar et al. 2008). Anomalies of sulfate spatial shows compliance with calcium and magnesium distribution maps which indicates these ions originate from dissolution of Miocene and Pliocene formations which contain gypsum and anhydrite. The concentration of sulfate shows a high anomaly in Tabriz city which is affected by contamination from municipal and industrial wastewater. Also, irrigation water return flow can be another source of the SO4
2− in the groundwater of the study area.
Chloride concentrations range between 49.7 and 2041.2 mg L−1 in the groundwater of the aquifers (Fig. 8i). According to World Health Organization (WHO) (2011), concentration of chloride should not exceed 250 mg L−1. The origin of chloride in groundwater of the study area can originate from evaporative deposits of the Miocene formations, contamination from municipal waste and urban waste and irrigation water return flow. This statement can be confirmed with moderate-to-high correlations between Cl− and Na+ (r = 0.965), Cl− and K+ (r = 0.843), Cl− and SO4
2− (r = 0.617). The chloride increase in the plain groundwater is justified by evaporative conditions due to shallow water level, fine-grained sediments and low hydraulic gradient and thus more groundwater residual time and more dissolution.
Nitrate concentrations vary between 4 and 243.7 mg L−1 with an average concentration of 55.6 mg L−1 (Fig. 8j). Many parameters such as the amount of fertilizer used, surface water quality, land use type, depth of groundwater level, land drainage by the river and sediment type have resulted in variations of nitrate concentrations in different parts of the plain (Barzegar 2014). The lowest nitrate concentration is obtained from the northwest of the plain (sample 16) because of low agricultural activity, confined aquifers in these areas and fine-grained sediments in comparison with other parts of the area. The highest nitrate concentration (sample 29) is in the west of Tabriz city, which is attributed to urban and domestic sewages and dense farming in this area (Barzegar 2014). Nitrate concentration is much higher in the southern part of Aji-Chay River; it would be due to intensive agricultural activity, industrial concentration, type of aquifer and shallow groundwater. The presence of high nitrate concentration in the drinking water increases the incidence of gastric cancer and other potential hazards to infants and pregnant women (Nagireddi Srinivasa Rao 2006).
Barzegar et al. (2016b) studied the intrinsic vulnerability of Tabriz plain aquifer using the DRASTIC method. They indicated that the least vulnerable area was located in the west where the aquifer system is confined, while the most vulnerable areas were located in the east and south where the aquifer is unconfined and the recharge is high. The vulnerability map obtained an R
2 of 0.83 between DRASTIC index and NO3
− concentrations (Barzegar 2014). The DRASTIC vulnerability map is shown in Fig. 9. The existence of a confining clay layer in the western part of the plains provides protection, reducing the risk of contamination. As shown in Fig. 9, the nitrate concentrations are high in vulnerable area and low in less vulnerable area. This indicates that there must be a special consideration in management of groundwater in the study area.
Identification of the hydrochemical processes
The scatter plots of different ions were analyzed to identify the interrelationships between the ions and possible chemical reactions which may occur in the Tabriz plain. Major ion concentrations as a function of TDS values were plotted to determine the contributing ions to groundwater mineralization. Figure 10 shows that Na+ and Cl− are strongly correlated with TDS with an R
2 of 0.9374 and 0.9739, respectively, which represent that these ions are the most effective in the mineralization and salinization of the groundwater of the study area. As well as, cross-plot of K+ with TDS (Fig. 10d) shows a good correlation with an R2 of 0.7254, indicating that potassium contributes, with both Na+ and Cl−, in groundwater mineralization.
As mentioned previously, dissolution of minerals such as gypsum (CaSO4.2H2O), anhydrite (CaSO4) and halite (NaCl) occurs frequently across the study area. The results of the computed SI values for these mineral phases show that groundwater is significantly undersaturated with respect to halite (−7.32 ≤ SI ≤ −3.13) and, to a lesser extent, with respect to anhydrite (−2.61 ≤ SI ≤ −0.7) and gypsum (−2.39 ≤ SI ≤ 0.86), indicating that these minerals can dissolve into the groundwater of the plain. Some groundwater samples have elevated Ca2+ and Mg2+ concentrations relative to HCO3
− which are balanced in charge by SO4
2−. These sulfate-rich waters probably derive from dissolution of gypsum and/or anhydrite. The well-defined relationship in the plot of Ca2+ versus SO4
2− (Fig. 12a) confirms this deduction (Barth 2000). Also, it can be understood by the parabolic trend observed in the correlation of the negative saturation indices, with regard to the gypsum and anhydrite minerals, versus the sum of calcium and sulfate ions (Fig. 11a, b) resulting from the CaSO4 dissolution (Hamed et al. 2011). The plot of Ca2+ versus SO4
2− can be evidence of the predominance of the cation exchange (Tarki et al. 2011). As shown in Fig. 11a, the excess of Ca2+ in groundwater samples indicating that Ca2+ releases into water to compensate the adsorption of Na+ while samples exhibiting an excess of SO4
2− reflects a probably oxidizing environment of the Tabriz plain. Barzegar et al. (2015) indicated the oxidizing environment of the unconfined aquifer of the Tabriz plain by measuring of Eh between 234 and 284 mv. The computed SI values show that groundwater is supersaturated with respect to calcite (0.1 ≤ SI ≤ 1.25) and dolomite (1.13 ≤ SI ≤ 2.91), indicating that these minerals precipitate in groundwater. This condition brings solution to equilibrium with respect to these carbonates and subsequently transported into a different environment where a higher pH or an apparent condition caused by the failure of the measured pH to accurately represent the actual equilibrium pH of the water in the aquifer (Kortatsi 2007; Christian et al. 2014).
The plot of Ca2+ + Mg2+ versus Na+ (Fig. 12b) was used to identify the ion exchange process. This plot shows that data points fall in the both sides of the 1:1 line indicating that ion exchange as well as some reverse ion exchange is taking place in the study area. Reverse ion exchange normally occurs in the presence of clays, As previously mentioned aquifer materials contain clays, where the Na+ is removed from the system with the release of Ca2+ + Mg2+.
The plot of Na+/Cl− versus EC was used to characterize the impact of the evaporation on the groundwater chemistry. This plot would give a horizontal line, which would then be an effective indicator of the concentration by evaporation, evapotranspiration and halite dissolution. Theoretically, the Na+/Cl− ratio approximately equal to one is attributed to halite dissolution, whereas a ratio greater than one is typically indicative for Na+ release due to silicate weathering (Meybeck 1987). Generally the molar ratio of Na+/Cl− for groundwater samples ranges from 0.38 to 2.1 (Fig. 12b). Most of the samples have Na+/Cl− molar ratio below one, indicating that halite dissolution was the major process. Scatter plot of EC versus Na+/Cl− shows an inclined trend line, which indicates that evaporation may not be the major geochemical process controlling the chemistry of the groundwater. As well as, halite dissolution indicates by the well-defined relationship in the correlation of the negative saturation indices versus the sum of ions resulting from the NaCl dissolution (Fig. 11c).
The impact of silicate weathering on groundwater system can be found by plot of Na+ + K+ versus total cations (TC). As shown in Fig. 11d, most of the data points are plotted above the Na+ + K+ = 0.5 TC. This indicates the involvement of silicate weathering in the groundwater system, which contributes Na+ and K+ to the groundwater (Stallard and Edmond 1983; Rajmohan and Elango 2004; Senthilkumar and Elango 2013).
As previously mentioned, agriculture and industry are the main human activities in the study area. Therefore, the chemistry of the groundwater can be influenced by anthropogenic activities. Gillardet et al. (1999) suggested that variation in TDS of groundwater may be attributed to land use and as a result of contamination. It is well known that in rural areas NO3
−, SO4
2−, Na+ and Cl− ions are mostly derived from agricultural fertilizers, animal waste, and municipal and industrial sewage (Jalali 2009; Nagaraju et al. 2014). Han and Liu (2004) and Jalali (2009) suggested that the high correlation between TDS and (NO3
− + Cl−)/HCO3
− molar ratios reveals the influence of anthropogenic activities on water chemistry. As shown in Fig. 13, there is a positive correlation with an R
2 of 0.8603 between TDS values and (NO3
− + Cl−)/HCO3
− molar ratios which indicates that human activities affect the chemistry of the groundwater.
Gibbs (1970) plots of log TDS against Na+/(Na+ + Ca2+) and Cl−/(Cl− + HCO3
−) were used to identify the groundwater interaction, with precipitation (rainfall), rock and evaporation, as the mechanisms controlling the groundwater chemistry. The ratios Na+/(Na+ + Ca2+) of as well as anions Cl−/(Cl− + HCO3
−) are plotted from the rock domain towards the evaporation domain (Fig. 14a, b), which reflect that rock–water interaction, as a major source of dissolved ions, and evaporation occur in the groundwater system. Some of the plotted data points in the evaporation domain show a dense clustering. As suggested by Li et al. (2013), the dense clustering of points in the evaporation domain can be an indication of anthropogenic activities.