Hydrogeochemical processes controlling

Journal of Geochemical Exploration 170 (2016) 58–71

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Hydrogeochemical processes controlling groundwater quality around Bomboré gold mineralized zone, Central Burkina Faso Aboubakar Sako a,⁎, Ousmane Bamba b, Aridjima Gordio b a b

Université Ouaga 1 Pr Joseph Ki-Zerbo, Centre Universitaire Polytechnique de Dédougou, BP. 139 Dédougou., Burkina Faso Université Ouaga 1 Pr Joseph Ki-Zerbo, Département des Sciences de la Terre, 09 BP 848 Ouagadougou 09, Burkina Faso

a r t i c l e

i n f o

Article history: Received 6 March 2016 Revised 28 July 2016 Accepted 16 August 2016 Available online 22 August 2016 Keywords: Borewells Dug wells Arsenic Trace metals Groundwater quality

a b s t r a c t This study investigates the processes that control dug well and borewell water chemistry in crystalline basement − aquifers in semi-arid environment. Changes in SO2− 4 and NO3 were observed in both dug well and borewell samconcentrations relative to SO2− ples. Six dug wells and seven borewells had high NO− 3 4 suggesting prevailing ox2− concentrations were found in few idizing redox conditions in these samples. The highest NO− 3 and SO4 borewells. These high concentrations could be attributed to low recharge rates and chemical weathering of sulfide minerals. The acidic conditions may have promoted AsT adsorption in the majority of dug wells. The saturation indices and correlation coefficients showed that most dug wells and some borewells were supersaturated with respect to Fe-bearing minerals, implying that trace elements such as AsT, Cu, Cr, Ni and Zn were likely to coprecipitate with residual Fe minerals. The low mobility of AsT in dug wells could be also explained by the limited carbonate mineral abundance in the weathered layer. By contrast, the elevated AsT concentrations observed in the majority of borewells reflected an extended water–rock interaction that had led to deportonation of surface charges of the aquifer minerals, thereby inhibiting AsT adsorption. The average concentration of FeT and total coliform content of dug well samples exceeded the World Health Organization (WHO) limits for drinking waters. Likewise, the average concentrations of FeT, AsT (~55% of samples) and Pb in borewells were higher than the WHO permissible limits. The study showed that the local groundwater resources are exposed to both anthropogenic and geogenic sources of pollution. © 2016 Published by Elsevier B.V.

1. Introduction In Central Burkina Faso, where surface water availability is limited, groundwater is the main source of water supply for the local communities. Much of the groundwater resources in the area occur in crystalline basement aquifers which is subdivided into a shallow weathered layer and deep fractured bedrock (Wright, 1992; Gamsonré, 2003). Traditional hand-dug wells are used to extract groundwater from the saturated weathered layers, while borewells equipped with handpumps are used to source groundwater from discrete horizontal fractures within the basement (Acworth, 1987; BGS, 2002; Wyns et al., 2004). The weathered layer and fractured bedrocks are products of the combined effects of prolonged in-situ chemical weathering of the parent rock and tectonic movement, respectively (Wyns et al., 2004; Dewandel et al., 2006). Therefore, the groundwater chemistry of the crystalline basement aquifers is directly related to the mineralogical composition of the parent rock and the various weathering processes (Hem, 1985). Elevated concentrations of potentially toxic trace elements, often ⁎ Corresponding author. E-mail addresses: [email protected], [email protected] (A. Sako).

http://dx.doi.org/10.1016/j.gexplo.2016.08.009 0375-6742/© 2016 Published by Elsevier B.V.

encountered, in crystalline bedrock aquifers have been mostly attributed to chemical weathering (BILAN D'EAU, 1993; MEE, 1998; Smedley et al., 2007). Once in the aquifer, trace element distribution and mobility are controlled by several geochemical processes such as oxidation-reduction reactions (Smedley and Edmunds, 2002), dissolution-precipitation and cation exchange (Ayotte et al., 2003). However, the mobility of trace elements in aquifers varies with individual elements. For example, Pb, Cd, Fe and Mn are preferentially mobilized under acidic conditions, whereas As, along with Fe and Mn, are more soluble under reducing conditions. High population density and intensive agricultural and artisanal gold mining activities in the study area are also likely to exacerbate trace element mobility, putting enormous pressure on dug well and borewell water quality and availability. In contrast to groundwater availability, groundwater quality has received little attention in Burkina Faso, and the available data indicate that groundwater resources are frequently exposed to various sources of contamination (e.g., Groen et al., 1988; Yaméogo and Savadogo, 2002). By far, arsenic is the groundwater contaminant that poses the most serious health hazard to humans across the country (Somé et al., 2011). Few studies have reported the occurrence of high arsenic concentrations (N 10 μg/L) in groundwater around gold mineralized zones

A. Sako et al. / Journal of Geochemical Exploration 170 (2016) 58–71

in the north-central region of Burkina Faso (de Jong and Kikietta, 1981; COWI, 2004; Smedley et al., 2007). This so-called localized arsenic pollution has been attributed to sulfide mineral oxidation (Lipfert et al., 2006; Smedley et al., 2007) and, to a lesser extent, dissolution of westerveldite (FeAs) and scorodite (FeAsO4·2H2O) in granitic rocks (Peters and Blum, 2003; Utsunomiya et al., 2003). However, the mechanisms that control As and other trace element mobilization within mineralized crystalline basement aquifers have not been well studied in the country. Understanding the spatial distribution and mobility of trace elements in mineralized basement aquifers is an important factor for a long-term dug well and borewell water quality monitoring program. In the present study, geochemical modeling and multivariate statistical techniques were used to: (1) provide insight into the geochemical processes that control dug well and borewell water chemistry (2); determine the extent of water–rock interaction; and to (3) assess the groundwater quality for human consumption. The underlying hypothesis is that groundwater from shallow dug wells and deep borewells from two different aquifers with a similar bedrock lithology, or that have experienced similar geochemical processes and anthropogenic activities would be geochemically different. 2. Geological and hydrogeological context The study area lies within a strongly arcuate volcanosedimentary northeast-trending belt in the central domain that is bounded to the east by the Tiébélé-Dori-Markoye Fault (Fig. 1). The geological formations of the area are composed of mafic to intermediate metavolcanic rocks intruded by ultramafic rocks, diorite and granodiorite of Birimian age (2300–2150 Ma). The oldest formations date back to Proterozoic and later Eburnean age (2150–2095 Ma), and they consist mainly of crystalline (plutonic and metamorphic) rocks such as amphibolite, diorite, granodiorite, granite, metagabbro, metasediment and green schists (Sattran and Wenmenga, 2002; Castaing et al., 2003). These rocks, often

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associated with gold mineralization and a network of quartz veins and sulfide minerals, were mostly deformed and metamorphosed during the Eburnean orogeny. The gold mineralization can also be associated with disseminated sulfide minerals in alteration zones. In the latter case, gold may be either free-milling or refractory (Milési et al., 1992; Bamba et al., 1997). Depending on the nature of the parent rock, other sulfide minerals such as arsenopyrite in metasedimentary units or chalcopyrite and pyrrhotite within biotite schists may be associated with pyrite (Béziat et al., 2008). Oxidation of the sulfide minerals through natural processes (geogenic sources) or anthropogenic activities may generate acid drainage that will enhance trace element mobility in the environment. Artisanal gold miners in the area have been targeting the oxidized ore, and hence the risks related to the occurrence of acid drainage are limited. However, if the sulfide wall rocks are reached, these risks will be high with a possible mobilization of potentially toxic trace elements within the groundwater systems (Bamba et al., 2013). In July 2014, Orezone Gold Corporation, a Canadian gold mining company, carried out a heap-leaching test on three cores of extractable ores in the Bomboré gold mineralized zone, and concluded that mobility of As, Cu and Fe was of great concern from water resources perspectives. The fate of As, Cu, Fe and other trace elements in the local crystalline basement aquifers will depend primarily on lithological and hydrogeological characteristics of weathered and the fractured bedrock layers (Dewandel et al., 2006). Together the weathered and fractured layers form the weathering profile of the crystalline basement aquifers with each layer having specific hydrogeological properties (Groen et al., 1988; Detay et al., 1989; Taylor and Howard, 2000). Thus, the weathered layer is capped by a thin lateritic layer, rich in Fe and/or Al. Based on its mineralogy and the degree of weathering, the weathered layer is subdivided into upper and lower layers (Chilton and Foster, 1995). The upper layer has a high proportion of secondary clay minerals (e.g., kaolinite), whereas the lower layer is characterized by high content of primary weathering materials associated with earlier

Fig. 1. Simplified geological map of Burkina Faso (modified from Castaing et al., 2003) and the Bomboré Sub-basin within the Nakambé River basin. 1) Bomboré gold mineralization area; 2) fault; 3) Neoproterozoic to Paleozoic sedimentary basin; 4) Eburnean granitic rock; 5) Birimien volcanosedimentary rock.

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forms of secondary clay minerals (smectite). Because of its clayey-sandy composition, the weathered layer is characterized by a relatively low permeability and high water retention capacity (Wyns et al., 1999). The thickness (10–100 m) and hydraulic properties of the weathered layer are a function of the combined effects of the parent rock lithology and local climatic and geomorphological conditions (Wright, 1992). In granitic gneiss contexts, widely encountered in the study area, the weathered layer is thin, and thus has a low hydraulic potential (BILAN D'EAU, 1993). In contrast, in the volcanosedimentary contexts or Birimian sequences, the second largest geological formation in the area, weathered layer has thick and clayey layers, intercalated with schist, sandstone and quartz veins. Therefore, the weathered layer is a capacitive layer, and mostly ensures the storage function of the aquifer. The weathered layer is underlain by a less weathered layer called fractured layer with vertical fractures that ensure the transmissive function within the aquifer (Courtois et al., 2010). The permeability within the fractured layer can be high where a significant network of fractures exists (Chilton and Foster, 1995; Taylor et al., 2010). With low permeability and shallow depth, the fresh basement has a reduced groundwater storage capacity (Maréchal et al., 2004). During rainy seasons, shallow weathered aquifers and deep fractured aquifers can be continuous. In rainy seasons, surface water is used for garden market and watering

livestock, but following surface water scarcity in dry seasons the farmers rely heavily on dug well and borewell water for their various activities. Artisanal gold mining also requires large volume of water for ore processing. As a result, access to freshwater resources is particularly critical for farmers and artisanal gold miners in this semi-arid environment. The main factors that are likely to affect the dug well and borewell water quality and quantity in the area are low rainfall, high population growth, intensive agricultural practices and artisanal gold mining. 3. Environmental setting The Bomboré gold mineralized zone is located in the province of Ganzourgou (Central Plateau Region) 85 km east of Ouagadougou, the capital of Burkina Faso, and close to the town of Zorgho (Fig. 1). The local climate is of the Sudano-Sahelian type with a single rainy season from May to September and has a mean annual rainfall of about 700 mm (MEE, 1998). The zone belongs to the Bomboré river subbasin, a tributary of the Nakambé River, with catchment area of approximately 500 km2 (Fig. 1). The zone is situated across two agro-ecological zones. The southern sector or the “valley area” is drained by the Nakambé River and its tributaries. To develop the valley, in the 1970s, an irrigation system was set up by the government, which allowed the

Fig. 2. Groundwater sampling locations, superimposed on the local lithological unit.


settlement of numerous farmers, making the area the breadbasket of the province (MEE, 1998). The northern sector known as the “plateau area” receives lower rainfall, and therefore has less fertile soils (Jacob et al., 2002). In this proven agricultural setting, artisanal gold mining has been taking place since 1990. The mining activity involves about 700 households or approximately 3000 individuals. 4. Materials and methods 4.1. Sample collection Thirty groundwater samples were collected in dug wells (W1–W11) and borewells (B1–B19; Fig. 2) around the mineralized zone during the beginning of rainy season in 2014. The well depths ranged from 10 to 30 m for the dug wells and about 40 to 80 m for borewells. Before the sampling, borewells were purged (4–5 min depending on depth). The sampling was carried out by pumping water through previously cleaned Teflon tubing using a peristaltic pump, and subsequently through inline filter capsule (0.45 μm diameter) attached to the tubing. The samples were split, and one split was acidified with ultrapure nitric acid (pH ~ 3), whereas the second split was unacidified. Each pre-cleaned polyethylene bottle was triple rinsed with the filtered water sample before the actual sample collection (e.g., Horowitz et al., 1994; Koterba et al., 1995). To avoid photocatalyzed oxidation of As (III), all filtered water samples were transferred into 500 mL opaque bottles (McCleskey et al., 2004). The samples were immediately placed within an ice-box, and returned to the laboratory for major and trace element analysis. 4.2. Sample analysis Bicarbonate, pH and electrical conductivity (EC) were measured on 2− 3− the site. In the laboratory, major anion (Cl−, NO3−, NO− 2 , SO4 , PO4 + − and F ) and NH4 concentrations were determined on unacidified samples by aquakem analyzer, whereas major cation (Ca2+, Mg2+, K+ and Na+) and total iron (FeT) concentrations were determined on acidified samples using inductively coupled plasma-optical emission spectroscopy (ICP-OES). Total trace element concentrations were determined on acidified samples by inductively coupled-mass spectroscopy (ICP-MS). The analyses were carried out at the SGS laboratories in Accra, Ghana. Detection limits for ICP-MS and ICP-OES analyses were based on standard deviation of three individual runs of the blank concentrations. Concentrations of duplicate and external calibration data of multi-element standards were also analyzed to determine sampling and analytical errors, and the results were within 10% error. Chemical oxygen demands (COD) were estimated using colometric method (APHA5220D), whereas the quantitative method was used to investigate total and fecal coliform content of the samples (APHA9221D). A brief description of the analytical methods and their detection limits are provided in Table 1. 4.3. Geochemical modeling The interaction between water and the minerals within an aquifer controls the chemical composition of the groundwater. To predict mineral activities and trace element mobility without collecting and analyzing the solid phase (Deutsch, 1997; Zhu et al., 2008), visual MINTEQ version 3.1 was used to perform saturation index (SI) calculation of mineral phases that are suspected to be responsible for the chemical composition of dug wells and borewells. The program contains a thermodynamic database of the WATEQ3 that contains equilibrium constants for simple aqueous and complex species as well as solubility products (Jenne, 1979; Melchior and Bassett, 1990; Gustafson et al., 2003). Several physico-chemical variables including pH, HCO–3 and major and trace element concentrations were used to calculate the SI values. The SI calculations are based on an equilibrium model, so their

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outputs are only indicative, as kinetic factors may prevent the equilibrium to be reached.

4.4. Multivariate analysis Multivariate statistic techniques such R-mode and Q-mode factor analyses have been widely used to decipher associations between groundwater samples and/or physico-chemical variables (e.g., Seyhan et al., 1985; Join et al., 1997; Hernandez et al., 1991; Farnham et al., 2003; Boruvka et al., 2005; Kuppusamy and Grirdhar, 2006). R-mode is entirely based on eigenanalysis of the correlation and covariance, and thus the normality is not imposed on the multivariate data (Meglin, 1991). When the correlation matrix is used, each variable is normalized to unit variance and therefore contributes equally to the variance (Farnham et al., 2003). In the present study, major element concentrations are in mg/L range, whereas those of trace elements are in μg/L range. As a result, R-mode was applied to the correlation matrix (Reyment and Jvreskog, 1996; Everitt et al., 2001). Eigenvalues of the principal components (PC) are a measure of their associated variance, the most participation of the original variables in the PC is shown by the loadings, and the individual transformed observations are called scores. That is, the eigenvalues describe the amount of variance explained by each PC. To reduce the overlap between original variables in each PC, a varimax rotation can be conducted (Helena et al., 2000). In the present study, the dataset for R-mode included 18 variables (physico-chemical parameters) for both dug well and borewell samples, thereby providing a simultaneous analysis of the whole hydrogeochemical dataset. The first principal component has the maximum variance followed by the second principal and so on (Hotelling, 1933; Gnanadesikan, 1977). The number of significant principal components is selected on the basis of the Kaiser criterion, and only the principal components with eigenvalues greater than or equal to 1 are taken into account (Kaiser, 1960). The measure of how well the variance of a given variable is described by a given set of factors is called the communality (Drever, 1982). Table 1 Summary of analytical method and detection limits (DL). Parameter

Method

Description

pH EC Total coliforms COD

APHA2540Ba APHA2510 APHA9221D APHA5220D

pH meter Calculation

APHA4500

Aquakem Analyzer

APHA2510 APHA2340B

Colorimetric method Calculation

APHA2130B

ICP-OES

EPA_200.8c

ICP-MS

Anions Cl− NO− 3 SO2− 4 3− PO4 HCO3− TH Cations Ca2+ Mg2+ K+ Na+ Fe Total trace metals As Ba Cr Cu Mn Ni Zn a b c

The American Public Health Association. Most probable number. The US Environmental Protection Agency.

DL 0.5 mS/m 1 MPN/100 mLb 25 mg/L

0.1 mg/L 0.06 mg/L 0.02 mg/L 0.02 mg/L 2 mg/L 5 mg/L

1 mg/L 0.5 mg/L 0.1 mg/L 0.1 mg/L 0.1 mg/L

0.0005 mg/L 0.001 mg/L 0.001 mg/L 0.001 mg/L 0.002 mg/L 0.001 mg/L 0.005 mg/L

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Table 2 Physico-chemical parameters plus total coliform numbers of dug well waters from Bomboré Sub-basin, Central Burkina Faso. Samples W1, W6 and W7 with high charge balance error (%) were excluded from the data interpretation. Parameter pH EC CODa TDS HCO–3 Ca2+ Mg2+ Na+ K+ F− Cl− SO2− 4 NO− 3 NO2− NH+ 4 PO3− 4 Al AsT Cu FeT MnT NiT Cr Pb Zn Na/Cl Na/Ca Tcc a b c d

Unit μS/m mg/L mg/L mg/L mg/L mg/L mg/L mg/L mg/L mg/L mg/L mg/L mg/L mg/L mg/L mg/L μg/L μg/L μg/L μg/L μg/L μg/L μg/L μg/L

MPN/mL

W1

W2

W3

W4

W5

W6

W7

W8

W9

W10

W11

6.4 475 NDb 356 270 43 15.9 26.2 10.8 0.6 0.2 40 4.3 ND 0.03 0.12 ND 2 5 500 9 65 82 1.1 ND 131 0.60 60

5.8 100 71 58 43 8 2 5.5 3.7 0.5 6.1 4 3.8 0.13 0.06 0.06 0.07 0 30 22,400 270 26 55 9.2 85 0.90 0.68 180

5.6 143 135 106 72 8 2.8 12 2.6 ND 0.8 1 3.4 ND ND 0.22 ND 11 3 13,300 830 21 41 9.1 160 15 1.5 45

5.9 132 ND 76 57 9 3.7 7 3.6 0.6 2.7 4 6.9 0.09 0.07 ND ND 1 3 400 240 3 2 0.6 56 2.59 0.77 21

6 58 64 94 23 4 1.3 1.9 3.5 0.6 5.5 ND 1.8 ND ND 0.14 0.08 19 8 17,500 300 21 45 7.1 85 0.34 0.47 7

5.7 106 ND 63 39 8 2.1 4 4.2 ND 7.1 28 ND ND ND 0.32 ND 26 11 8400 350 12 24 4.4 110 0.56 0.50 40

5.4 113 ND 74 30 8 2.1 4.2 3.0 0.6 4.4 9 3.7 ND ND 0.15 ND 1 1 3200 38 12 16 3.0 90 0.95 0.52 ABd

5.6 79 80 54 30 6 2.5 2.5 3.2 0.6 6.4 2 3.9 0.08 0.1 0.06 0.04 2 26 13,700 220 24 42 7.5 84 0.37 0.41 AB

6.3 280 41 156 149 16 5.7 33.5 1.0 0.8 4.0 4 7.6 0.29 0.02 0.09 0.25 0 15 15,100 600 17 31 6.1 130 8.37 2.09 AB

6.3 146 97 84 65 10 4.6 5.1 5.3 ND 5.7 3 5.5 0.09 0.03 0.09 0.16 3 21 13,900 700 25 40 5.4 130 0.89 0.51 AB

6.8 536 ND 324 332 50 16.4 29.6 5.9 ND 0.6 11 0.7 ND ND ND ND 5 2 ND 16 7 2 ND 69 49.33 0.59 AB

Chemical oxygen demand. Not detected. Total coliforms. Absence.

Q-mode factor analysis was also used to evaluate the interrelationships among the collected samples, using the 18 variables. This multivariate technique is similar mathematically to R-mode analysis, but it evaluates relationships among the samples rather than among the variables. The statistical package for Social Science Software (IBM SPSS Statistics 20) was used to perform statistical analyses.

5. Results and discussion 5.1. Groundwater constituents The physico-chemical parameters including pH, electrical conductivity (EC), total dissolved solids (TDS) and major ion and trace element

Table 3 Physico-chemical parameters and total coliform numbers of borewell waters from Bomboré Sub-basin, Central Burkina Faso. Parameter pH COD EC TDS HCO3 Ca 2+ Mg2+ Na+ K+ F− Cl− SO2− 4 NO3 NO− 2 NH+ 4 3− PO4 Al AsT Cu FeT MnT NiT Cr Pb Zn Na/Cl Na/Ca T. colif.

Unit mg/L μS/m mg/L mg/L mg/L mg/L mg/L mg/L mg/L mg/L mg/L mg/L mg/L mg/L mg/L mg/L μg/L μg/L μg/L μg/L μg/L μg/L μg/L μg/L

cfu/mL

B1

B2

B3

B4

B5

B6

B7

B8

B9

B10

B11

B12

B13

B14

B15

B16

B17

B18

B19

6.5 ND 394 288 245 42 14.5 12.7 5.6 0.7 0.1 ND 0.5 ND ND 0.12 ND 2 1 ND 9 6 2 ND 60 127 0.30 AB

7.2 ND 634 396 325 59 20.4 23 3.6 0.7 2.2 ND 30.4 ND ND ND ND 13 3 ND 19 9 2 ND 12 10.4 0.38 AB

6.7 ND 594 342 386 51 18.5 38.8 2.9 0.1 0.5 4 1.3 ND 0.06 ND ND 71 3 100 36 6 2 ND 150 77.6 0.76 AB

6.6 ND 878 502 578 84 52.4 22.4 0.9 0.5 2.4 15 11.8 0.05 0.02 ND ND 24 11 ND 88 8 3 2.2 100 9.3 0.26 300

6.9 ND 412 252 279 37 15 22.8 2.5 ND 0.2 19 2.2 ND 0.04 ND ND 100 3 100 62 6 2 6.0 11 114 0.61 AB

7.1 ND 627 398 257 54 19.9 34.9 4.4 0.6 3.2 100 ND ND ND ND ND 4 7 ND 130 9 2 9.0 10 10.9 0.64 AB

7 ND 813 470 534 65 23.9 50.3 5.0 ND 2.8 19 1.1 ND ND ND ND 1 7 200 45 11 3 11.0 18 17.9 0.77 6

7.1 ND 457 262 284 48 10.1 22.4 2.2 0.9 ND 10 ND 0.04 ND ND ND 5 4 ND 120 10 1 10.0 10

6.4 ND 549 332 305 60 20 18.4 1.9 ND 2.2 18 28.9 ND ND 0.2 ND 85 4 ND 11 7 2 1.3 63 8.3 0.30 AB

6.5 ND 744 412 475 71 30.2 36.7 2.6 ND 0.5 21 ND ND ND ND ND 27 2 ND ND 7 3 8.0 41 73.4 0.51 AB

7 ND 707 400 427 51 22.3 45.3 6.5 1.0 1.3 18 2.1 ND ND 0.05 ND 10 6 ND 53 8 3 0.6 110 34.8 0.88 AB

6.8 45 651 346 433 59 26.8 25.4 6.5 ND 0.8 5 ND ND ND ND ND 9 66 19,400 910 36 71 9.9 140 31.7 0.43 AB

7 105 796 456 503 64 26.8 42.5 10.1 ND 0.8 9 ND ND ND 0.14 ND 10 23 44,400 760 71 161 4.4 150 53.1 0.66 AB

6.7 ND 946 588 541 75 44.1 45.7 7.7 ND 2.1 60 30.2 ND ND ND ND 49 8 7400 260 24 36 2.0 130 21.7 0.60 AB

6.6 ND 322 192 194 34 13.2 6.6 4.2 ND 0.8 ND 5.51 ND ND 0.09 ND 41 38 500 230 6 2 ND 87 8.2 0.19 AB

6.,9 ND 621 376 357 54 20.5 44.8 2.2 ND 0.4 20 2.1 ND ND ND ND 4 14 9900 1250 17 23 ND 78 21.6 0.81 AB

6.9 ND 563 346 474 45 15.4 36.8 4.0 ND 1.7 21 5.5 ND ND ND ND 14 15 19,500 950 21 14 ND 100 112 0.82 AB

6.9 ND 854 452 543 72 31,9 55.2 5.5 ND 1.1 8 15.9 ND ND ND ND ND 25 100 110 10 2 ND 42 50.2 0.76 AB

6.6 ND 240 162 136 26 6 7.4 1.0 ND 0.6 ND 5.8 0.07 ND 0.03 ND 7 4 ND 10 4 ND ND 55 12.3 0.28 AB

0.46 AB


concentrations in water samples varied between shallow dug wells and deep borewells (Tables 2–3). The extent of this variability was further illustrated by box plots of selected parameters, which showed that − 2− and AsT have the largest dispersions between the HCO− 3 , NO3 , SO4 median and maximum concentrations in borewell samples compared to those of dug wells (Fig.3). The borewell samples also showed high sporadic concentrations of four oxygen-sensitive species (i.e., NO− 3 , SO2− 4 , Mn and FeT), indicating the effects of possible mixed reduction/ oxidation (redox) conditions and variable recharge patterns on the borewell water chemistry. Overall, borewell samples were circumneutral with pH ranging from 6.5 to 7.2 and a mean value of 6.8 (Table 4), whereas those of dug wells were acidic with pH varying from 5.4 to 6.8 with a mean value of 5.9. The lower pH values observed in dug well waters could be attributed to the presence of organic matter, as evidenced by the high COD content. That is, decomposition of organic matter involves a progressive loss of dissolved oxygen and release of CO2, leading to a decrease in pH. The low pH values can also be related to the low carbonate mineral content in the weathered layer. A substantial increase in pH values

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in the borewells is attributed to proton (H+) consumption (i.e., deprotonation), following hydrolysis of carbonate and silicate minerals (Snellman et al., 1998; Luukkonen et al., 1999). Electrical conductivity (EC) in borewells varied from 240 to 946 μS/ cm (mean = 616 μS/cm) compared to 58 to 475 μS/cm (mean = 163.2 μS/cm) in dug wells. The samples from borewells also showed the highest concentrations of TDS varying from 162 to 588 mg/L with a mean concentration of 364 mg/L. The elevated EC and TDS in borewells with respect to dug wells reflected a longer residence time of the groundwater, suggesting an extended water–rock interaction. Mineralization was therefore greater in borewells compared to dug wells. Both EC and TDS concentrations for all samples were below the permissible limits set by the World Health Organization for drinking water (WHO, 2006). The anion abundance orders in the groundwater samples were 2 N NO3− N Cl− in dug wells and HCO3– N SO− N Cl− HCO3– N SO2− 4 4 N F− in borewells. Chloride concentrations in the investigated samples were low (from below detection limit to 6.1 mg/L), with the lowest concentrations recorded in borewells (mean = 1.2 mg/L) compared to dug

Fig. 3. Box plots of selected physico-chemical parameters of dug well and borewell water samples. The tops and bottoms of the boxes represent the 75th and 25th percentiles, respectively. The horizontal line across the boxes indicates the median. The vertical lines from the tops and bottoms of the boxes extend to 90th and 10th percentiles, respectively.

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wells (mean = 4.3 mg/L). In primary crystalline rock forming minerals such as biotite and amphibole, Cl− can substitute for OH−, and under intensive chemical weathering, Cl is released into the groundwater system (Edmunds et al., 1984a). So it was postulated that the low Cl− abundance in the present samples reflects the limited Cl abundances in the parent rock as well as in the continental rainfall (Smedley et al., 2007). However, the relative higher Cl− concentrations in dug wells with respect to borewells could be attributed to a possible contamination from agricultural sources such as fertilizers (KCl) and livestock manure. Nitrate was present in all dug well samples (0.66–7.58 mg/L). Six borewell samples had NO− 3 concentrations below the detection limit. Furthermore, SO2− 4 was not detected in four borewells, and yet some inconcentrations (up dividual borewells exhibited relatively higher SO2− 4 to 100 mg/L) compared to dug wells (1.0–11.0 mg/L). 2− distribuVariations in redox conditions may explain NO− 3 and SO4 tions in dug well and borewell samples. In anoxic or suboxic waters, chemolithotropic bacteria may use sulfide minerals such as pyrite to re2− 2+ (Rödelsperger, duce NO− 3 and produce SO4 , gaseous N2 and Fe 1989; Frind et al., 1990; Postma et al., 1991; Böttcher et al., 1990). In and NO− the present study, SO2− 4 3 contents in individual dug wells and borewells revealed mixed redox conditions. Thus, six dug well (W3, W4, W5, W8, W9 and W10) had high NO− 3 concentrations relative to compared to seven borewells (B1, B2, B4, B9, B15, B18 these of SO2− 4 and B19), implying that autotrophic denitrification did not occur in these samples. That is, in these wells, NO− 3 is not being used as an electron acceptor and redox conditions should remain oxic (Winograd and Robertson, 1982; Edmunds et al., 1984b). However, the majority of borewell samples (12 samples) and two dug well samples showed concentrations relative to NO− high SO2− 4 3 , suggesting that these wells were exposed to autotrophic denitrification in pyrite-bearing fractures under suboxic conditions. This biogeochemical process could also explain the high Fe concentrations observed in some borewell samples. The highest NO− 3 concentrations were measured in five borewells (B2, B4, B9, B14 and B18; Table 3). This may suggest the absence of new recharge water inputs, leading to low dilution rates within the aquifer. In fractured bedrock aquifers, contaminant concentrations are often spatially heterogeneous, and they may increase due to recharge

patterns (Flint et al., 2002). According to Levison and Novakowski (2009), the high NO− 3 concentrations, often measured in the shallowest sections of wells with respect to deep sections, reflect the influence of dilution. The authors also demonstrated that the low mean NO− 3 concentrations in wells were strongly correlated with high rainfalls. While 2− Cl−, NO− 3 and SO4 concentrations for all samples were below the maximum contaminant level recommended by the WHO, the data suggested that the groundwater resources in the area are vulnerable to contamination from point and nonpoint sources such as manure trough seepage and agricultural runoff. Minor amounts of F− were detected in a few samples with the highest concentration (1.8 mg/L) measured in sample B11 (Table 2). The occurrence of F− in the groundwater samples can be attributed to granite intrusions in the local lithology (BGS, 2002). Due to their shallow nature and proximity to domestic and agricultural contaminant content (0.06 to sources, dug well samples had relatively high PO3− 4 was only measured in five borewells (0.03– 0.22 mg/L), whereas PO3− 4 0.20 mg/L). Major cation abundance trends in dug wells and borewells were identical and dominated by Ca2 + and Na+ followed by Mg2 + (Tables 2–3). Both dug well and borewell samples were classified as Ca-Mg-HCO3 dominant type (Fig. 4). The abundance of major cations in the samples was highly controlled by water–rock interaction, and the relative high concentrations of these cations in borewells are an indication of a longer residence time of borewell waters in the aquifers. On the Na+ versus Cl− scatter plot borewell samples plotted above the 1:1 line implying that the excess Na+ was due to silicate weathering, whereas dug well samples (except two samples) were plotted below the 1:1 line reflecting the influence of salt dissolution (Stallard and Edmond, 1983; Kortatsi, 2007; Fig. 5a). Four dug well samples (W2, W5, W8 and W10) had Na/Cl ratio below 1, suggesting that halite dissolution was an important geochemical process in these wells. The scatter plots of SO4 + HCO3 versus Ca + Mg (Fig. 5) showed that the chemistry of three borewell samples was strongly influenced by carbonate weathering. Trace element concentrations varied in individual dug wells and borewells. Thus, trace amounts (from below detection limit to

Table 4 Average abundances of physico-chemical parameters from dug well and borewell waters plus standard deviations (±SD). Dug well pH EC (μS/cm COD (mg/L) TDS (mg/L) HCO–3 (mg/L) Ca2+ (mg/L) Mg2+ (mg/L) Na+ (mg/L) K+ (mg/L) F− (mg/L) Cl− (mg/L) (mg/L) SO2− 4 NH+ 4 (mg/L) NO− 3 (mg/L) NO− 2 (mg/L) (mg/L) PO3− 4 Al (mg/L) AsT (μg/L) Cu (μg/L) FeT (μg/L) Mn (μg/L) Ni (μg/L) Cr (μg/L) Pb (μg/L) Zn (μg/L) Na/Cl Na/Ca TC

n 8 8 6 8 8 8 8 8 8 5 8 7 5 8 5 8 5 8 8 7 8 8 8 8 8 8 8 5

Borewell Min. 5.6 58 41 54 23 4 1.3 1.9 1 0,5 0,6 1 0.02 0.66 0.08 0 0.04 0 2 400 16 3 2 0.6 56 0.14 0.42 7

Max. 6.8 536 135 324 332 50 16.4 33.5 5.9 0,8 6,4 11 0.1 7.6 0.29 11 0.25 19 30 22,400 830 26 55 9.2 160 1.5 2.09 180

Mean ± SD 6.0 ± 0.4 184 ± 157 81 ± 32 119 ± 88 96 ± 102 14 ± 15 5 ± 4.9 12 ± 12.4 3.6 ± 1,5 0,6 ± 0.1 4,0 ± 2,4 4.1 ± 3.2 0.1 ± 0.0 4.2 ± 2.4 0.13 ± 0.1 3,6 ± 3,3 0.12 ± 0.1 5,1 ± 6,7 13,5 ± 11,1 13,757 ± 6697 397 ± 279 18 ± 8.6 32.2 ± 19.8 5.8 ± 3.2 99.9 ± 35.6 0.70 ± 0.3 0.9 ± 0.3 62 ± 0.5

n 19 19 2 19 19 19 19 19 19 12 18 15 5 12 2 8 ND 18 19 10 18 19 13 18 19 19 3 19

Min. 6.4 24 45 162 136 26 6 6.6 0.9 0.5 0.1 4 0.02 0.5 0.05 0.02 ND 0.0 0.0 100 ND 0.0 0.0 0.0 10 8,2 1 0.19

Max. 7.2 9 105 588 578 84 52,4 60.1 10.1 1.8 3.2 100 0.06 30.4 0.07 0.7 ND 100 70 44,400 1250 70 40 160 150 127 300 0.88

Mean ± SD 6.8 ± 0.22 62 ± 19 75 ± 42 367 ± 106 383 ± 131 55 ± 15 22.7 ± 11 34.7 ± 16 4.2 ± 2.4 0.8 ± 0.35 1.3 ± 0.9 23.1 ± 24 0.03 ± 0.15 10.1 ± 12.2 0.06 ± 0.01 0.23 ± 16 ND 26.3 ± 30 11.1 ± 16 10,160 ± 14,330 253 ± 382 14.5 ± 15.8 5.1 ± 9.44 18.6 ± 39.7 71.9 ± 49 44,1 ± 40 102 ± 171 0,55 ± 22

WHO (2006) 6.5–9.2 1500 – 1000 – 75–200 50–150 200 200 1.5 250 250 – 50 3

10 200 300 400 20 50 10 3000 1


Fig. 4. Piper diagram showing the main hydrochemical facies of dug well and borewell samples.

0.25 mg/L) of Al were detected in half of the dug well samples, but not in borewells (Tables 2–3). Likewise, FeT concentrations in dug wells varied between 400 and 22,400 μg/L with an average value of 10,840 μg/L,

Fig. 5. a–b) Relationships between major ion concentrations ratios, highlighting the main geochemical processes controlling dug well and borewell chemistry.

65

whereas Fe T was only detected in nine samples out of 19 in borewells. Although the concurrent high FeT concentrations is partially due to Fe-rich lateritic materials of the weathered layer, it is consistent with the acidic conditions of dug wells and suboxic nature of some borewells as this element is more soluble under reducing and low pH conditions (Nath et al., 2008). The means FeT concentrations found in both dug well and borewell samples largely exceeded the WHO guideline for drinking water. Manganese was detected in all dug well samples, and fell below detection limit in one borewell. Four dug well and three borewell samples had Mn concentrations that were higher than the WHO recommended limit for drinking water (400 μg/L). In borewell samples, AsT concentrations ranged from b0.5 to 100 μg/L with an average concentration (24.05 μg/L) exceeding the WHO limit for drinking water (10 μg/L). Total arsenic was detected in seven dug well samples with two samples showing concentrations higher than 10 μg/L. In contrast, nine borewell samples exceeded the WHO guideline limit for drinking water with maximum concentrations ranging from 49 to 100 μg/L. Other trace element such as, Ni, Cu, Cr and Zn and Pb were also detected in the both dug wells and borewells, but their concentrations (except Pb) were below the WHO permissible limits.

5.2. Processes controlling arsenic and trace element distribution Saturation indices (SI) of groundwater are used to investigate different forms of mineral phases such as precipitated, dissolved and adsorbed phases. The accuracy of SI and chemical analysis in general is evaluated through the percent charge balance errors (CBE). Samples with small CBE values are considered to be more accurate or containing all the important ionic species. In the present study, the highest CBE are negative, implying an excess of anions in the samples. Three dug well samples (W1, W6 and W7) with extremely high CBE were excluded from the data interpretation (Table 5). The modeling data showed that most dug wells (W3, W4, W8 and W9) were strongly supersaturated with respect to Fe-oxide (hematite) and hydroxides (goethite and ferrihydrite (Table 5; Fig.6a and b). As a result, the low AsT concentrations in moderately oxidizing and acidic dug wells could be attributed to the tendency of As-laden Feoxyhydroxide minerals to dissolve, remobilizing As in solution. The released As would be readily readsorbed onto residual oxyhydroxide surfaces (Ravenscroft et al., 2001). The results also revealed that carbonate and silicate mineral dissolution may play a central role in controlling AsT and trace element abundance in borewells. Relatively high calcite concentrations and extended water–rock interaction in borewells tend to raise pH values, with saturation indices of calcite approaching equilibrium (Fig.6c). This greater association between pH and calcite saturation may lead to a deprotonation of the aquifer mineral surfaces. As pH values increase, the surface charges become less positive and As adsorption is less favorable. This observation is consistent with several studies that demonstrated that arsenate, the dominant As species in oxygenated groundwater, adsorption increases at low pH values and decreases with increasing pH (e.g., Bowell, 1994; Hsia et al., 1992; Jain and Loeppert, 2000; Manning et al., 1998; Meng et al., 2000; Pierce and Moore, 1982; Wilkie and Hering, 1996). In the present study, most borewell samples, with pronounced low Mn and high NO− 3 concentrations, exhibited the highest AsT concentrations (Fig. 7). Contrary to dug wells, pH fluctuations could be the primary factors controlling AsT mobilization in borewell waters. The saturation indices were consistent with R-mode factor analysis. Two separate R-mode were carried out on dug well and borewell samples. A varimax rotation generated three PC loadings for dug well samples and five PC loadings for borewell samples with eigenvalues higher than 1, and the communalities N0.500 for all variables.

66


The three PC loadings extracted for dug well samples accounted for 41.5%, 28.9% and 17.4% for PC1, PC2 and PC3, respectively (Table 6). 2+ , Ca2+, PC1, with high positive loadings for pH, HCO− 3 , EC, TDS, Mg K+, Na+ and SO24 −, is the most controlling factor on the dug well water chemistry. Electrical conductivity reflects the extent of mineral dissolution in dug wells. This geochemical process reflects H+ consumption, and thereby increases pH values (Pacheco and Van der Weijden, 1996). Therefore, the PC1 can be attributed to chemical weathering in dug wells. However, a moderate negative loading of PC1 for FeT (−0.371) can be explained by a limited precipitation of Fe, scavenging Cu (− 0.442) and Cr (− 0.433) and, to a lesser extent, Ni (− 0.336) and Pb (− 0.302) in the process. The high correlation be(r = 0.77) is consistent with their mixed origin tween Na+ and SO2− 4 (Table 7). Thus, Na+ and SO2− 4 in groundwater are derived from mineral dissolution, chemical fertilizer and atmospheric input (Edmunds et al., 2003; Valdes et al., 2007, Brenot et al., 2008). The strong positive correand HCO–3 indicated that these lations between Mg2+, Ca2+, Na+, SO2− 4 ions are derived from silicate and carbonate mineral weathering as demonstrated by major ion chemistry. These elements also showed high positive correlations with pH, indicating that an increase their concentrations will ultimately raise pH values. PC2 had high positive loadings for Pb (0.876), Cr (0.856), Cu (0.852), FeT (0.819) and AsT (0.731). This component also showed moderate positive correlations between AsT and FeT (r = 0.6), Cu (r = 0.50), Cr (r = 0.56), Pb (r = 0.58) and Ni (r = 0.40). These trace elements are also negatively correlated with SO24 − (except AsT, r = − 0.41; Table 8), implying that abundance of these chalicophile and siderophile elements in dug wells was partially controlled by sulfide and iron mineral oxidation. Several studies have hypothesized that oxidation of sulfide minerals such as arsenopyrite is the main source of As, Fe and SO2− 4

abundance in oxic groundwaters (e.g., Mandal et al., 1996; Schreiber et al., 2003). This can be expressed as: ‐

‐

2FeAsS þ 6:5O2 þ 3H2 O ¼ 2Fe2þ þ 2SO4 2 þ 2HAsO4 2 þ 4Hþ

ð1Þ

Although AsT, SO2− 4 and FeT may be derived from the same geogenic sources, as illustrated by Eq. (1), the lack of apparent correlation between AsT and SO24 − in dug wells could be attributed to the nonconservative behavior of these elements in aquatic environments (Table 7). After sulfide mineral dissolution, a series of disproportionate biogeochemical (e.g., sulfate and iron reduction) processes must have association (Smedley et al., 2007). As expected, decoupled the As-SO2− 4 negative correlations were observed between pH, Cr, Cu, Ni, Pb and, to a lesser extent, FeT and Mn. A decrease in pH would enhance the abundance of trace element concentrations in dug wells and vice versa, but the low pH does not have a discernable effect on AsT concentrations in these waters. As demonstrated by SI, liberated AsT following oxyhydroxide dissolution under low pH and mildly reducing conditions is likely to be readsorbed onto residual oxyhydroxide minerals. As a result, pH is unlikely to have a direct effect on AsT distribution in dug wells. PC3 is characterized by high positive loadings for Mn (0.843), + Zn (0.815), NO− 3 (0.673) and moderate positive loadings for Na (0.378) and Pb (0.306). Although Mn is largely derived from geologic + material, NO− 3 , Pb and, to lesser degree, Na and Zn tend to have anthropogenic sources. Therefore, PC3 represented anthropogenic contributions, such as agricultural activities and artisanal gold mining, to dug well chemistry. The five PC extracted for physico-chemical parameters of borewell samples accounted for 85.20% of the total variance (Table 8). Electrical conductivity, Ca2+, Mg2+, TDS, HCO–3 and Na+ have strong positive

Table 5 Saturation indices (SI) calculated using visual MINTEQ for shallow dug well waters. An SI value b0 is considered as undersaturated, whereas an IS N0 is regarded as supersaturated. Mineral phase CBE (%)b IS (M)c pCO2 (atm)d Al(OH)3 (a) Diaspore Gibbsite Boehmite Arsenolite Claudetite Calcite Cerrusite Aragonite Dolomite Vaterite CuCO3(s) Cupric Ferrite Malachite Langite Lepidocrocite Goethite Fe(OH)2·7Cl.3(s) Ferrihydrite Hematite Maghemite Magnesioferrite Siderite Strengite Rhodochrosite MnCO3 (am) MnHPO4(s) Ni(OH)2 (am) Ni4(OH)6SO4(s)

Formula

Al(OH)3 γ-AlOOH Al(OH)3 γ-AlO(OH) As4O6 As2O3 CaCO3 PbCO3 CaCO3 CaMg(CO₃)₂ μ-CaCO3 CuFe2O4 Cu2CO3(OH)2, Cu4(SO4)(OH)6·2H2O γ-FeO(OH) α-FeO(OH) (Fe3+)2O3·0.5H2O Fe2O3 γ-Fe2O3 Mg(Fe3+)2O4 FeCO3 FePO4·2H2O MnCO3 MnHPO4(S) Ni(OH)2

Bold are supersaturated. a Not calculated. b %Charge balance error. c Ionic strength. d Partial CO2 pressure.

W1

W2

W3

W4

W5

W6

W7

W8

W9

W10

W11

−19.8 0.008 0.10 n.c n.c n.c n.c −13.8 −13.8 −1.0 −2.2 n.c −2.0 n.c −2.6 n.c −3.6 n.c n.c n.c n.c n.c n.c n.c n.c −0.9 n.c −2.3 −2.8 −0.8 −63 n.c

−3.5 0.002 0.070 −2.4 1.6 0.7 −0.1 −3.2 −12.1 −3.0 −2.1 −3.2 −6.3 −3.6 −2.7 n.c −3.6 −12.7 n.c n.c n.c n.c n.c n.c n.c −0.62 n.c n.c n.c n.c n.c n.c

3.2 0.0001 0.0019 n.ca n.c n.c n.c −12.3 −12.3 −2.9 −2.2 −3.1 −6.0 −3.5 −2.7 14.1 −4.1 −14.7 6.3 7.2 7.7 5.0 16.8 9.0 5.7 n.c 3.7 −1.6 −2.1 1.3 −83 −15.7

−9.3 0.0002 0.14 n.c n.c n.c n.c −14.4 −14.4 −2.7 −3.2 −2.9 −5.5 −3.3 −3.5 11.3 −5.2 −16.1 5.1 6.0 6.5 3.8 14.4 6.6 4.0 n.c n.c −1.9 −2.4 n.c −85 n.c

2.6 0.0007 0.14 −1.9 2.0 1.2 0.3 −11.8 −11.9 −3.3 −2.2 −3.5 −6.8 −3.9 −2.8 15.9 −3.4 n.c 6.9 7.7 8.3 5.5 17.9 10.1 7.3 n.c 3.6 −2.1 −2.6 1.1 −74 n.c

−46,9 0,003 0.079 n.c n.c n.c n.c −11.5 −11.6 −3.2 −2.6 n.c −6.5 −3.7 −3.1 13.8 −4.6 n.c 6.2 7.1 7.8 4.9 16.6 8.8 5.6 n.c 3.2 −2.1 −2.6 0.7 -83 n.c

−37,1 0,002 0,123 n.c n.c n.c n.c −14.4 −14.4 −3.5 −3.1 n.c −7.3 −4.1 −3.6 11.7 −5.7 n.c 5.5 6.4 n.c 4.2 15.2 7.4 3.6 n.c 2.5 −3.5 −4.0 −0.9 −89 n.c

−4.0 0.001 0,08 −2.4 1.6 0.7 −0.2 −12.1 −12.2 −3.0 −2.1 −3.2 −6.3 −3.6 5.1 21.4 −9.1 n.c 6.8 7.6 8.3 5.4 17.7 9.9 6.8 n.c 3.3 −2.1 −2.6 0.4 −776 −15.7

−0,34 0.0043 0.075 0.06 3.9 3.1 2.2 −12.8 −12.9 −1.7 −1.6 −1.88 −3.6 −2.3 −2.3 n.c −2.7 −12.1 n.c n.c n.c n.c n.c n.c n.c n.c n.c n.c −1.29 1.29 −4.95 −25.3

−0.7 0.002 0.058 n.c 3.7 2.8 1.9 −2.8 −13.4 −2.6 −2.1 −2.8 −5.2 −3.2 0.4 18.6 2.7 17.1 8.2 7.6 8.2 5.4 7.6 9.9 7.5 n.c 3.1 0.4 −1.9 1.0 55 5.8

8.4 0.0076 0.0511 n.c n.c n.c n.c −12.8 −12.8 −0.4 −1.8 −0.6 −1 −1 −2.5 n.c −3.2 −13.3 n.c n.c n.c n.c n.c n.c n.c n.c n.c −2.1 n.c n.c n.c n.c


67

Fig. 6. a–b) Relationships between Fe-oxyhydroxide saturation indices and total arsenic concentrations in dug well and c) relationships between pH values and calcite saturation indices in dug well and borewell samples.

loadings for PC1. Strong relationship between PC1 loadings and the correlations between Ca2+, Mg2+, HCO− 3 , EC and TDS are illustrated in Tables 9. PC1 was interpreted as representing the weathering products of carbonate and silicate minerals from metasedimentary bedrock, consistent with the hydrogeochemical facies and the dominant geochemical process in borewell waters (i.e., water–rock interaction). PC2 had high positive loadings for K+, Cr, FeT, Ni and Zn, and it can be interpreted as lithological

contribution to the borewell chemistry. This component grouped the elements that are commonly found in the local rock units such as ultramafic rocks, biotite and amphibolite. PC3 was characterized by high positive loadings for Pb and Mn. The strong correlations between Pb and Mn indicated that these two elements were, at least in part, originated from a common anthropogenic source such as artisanal gold mining and agricultural activities in the vicinity of Bomboré gold mineralized zone.

Fig. 7. a) Relationship between pH values and total arsenic distribution in dug wells and boreholes; b) relationships between two oxygen-sensitive elements (Mn, NO− 3 ) and AsT distribution in dug well and borewell samples. Dashed lines correspond to 10 μg/L limit of WHO guideline for drinking water.

68


Table 6 Principal component loadings of physico-chemical parameters for dug well samples and their corresponding commonalities (Significant coefficient loadings ≥0.60 are in bold).

Table 8 Principal component loadings of physico-chemical parameters for borewell waters and their corresponding communalities (Significant coefficient loadings ≥0.60 are in bold).

Variable

PC1

PC2

PC3

Communalities

Variable

PC1

PC2

PC3

PC4

PC5

Communalities

pH HCO− 3 EC TDS Ca2+ Mg2+ K+ Na+ SO2− 4 NO− 3 AsT Cu Cr FeT Mn Ni Pb Zn Variance % Cumulative %

0.858 0.958 0.954 0.957 0.919 0.905 0.275 0.849 0.793 −0.375 −0.083 −0.442 −0.433 −0.371 −0.151 −0.336 −0.302 0.091 41.50 41.50

−0.265 −0.243 −0.276 −0.174 −0.238 −0.303 −0.075 −0.230 −0.372 −0.498 0.731 0.852 0.856 0.819 0.253 0.840 0.876 0.429 28.97 70.48

−0.200 −0.137 −0.093 −0.185 −0.296 −0.257 −0.708 0.378 −0.333 0.678 −0.144 0.159 0.208 0.265 0.843 0.212 0.306 0.815 17.46 87.94

0.846 0.996 0.995 0.981 0.989 0.978 0.583 0.916 0.878 0.848 0.563 0.946 0.964 0.879 0.797 0.863 0.953 0.856

pH HCO− 3 EC TDS Ca2+ Mg2+ K+ Na+ SO2− 4 NO− 3 AsT Cu Cr FeT Mn Ni Pb Zn Variance% Cumulative%

−0.071 0.919 0.971 0.947 0.942 0.901 0.276 0.738 0.327 0.176 −0.124 0.103 0.129 0.071 0.012 0.180 −0.009 0.236 29.36 29.36

0.189 0.207 0.189 0.179 0.069 0.082 0.771 0.144 −0.085 −0.077 −0.106 0.855 0.971 0.920 0.515 0.955 0.172 0.606 27.18 56.54

0.002 0.174 −0.017 −0.005 −0.057 −0.033 −0.183 0.024 0.080 −0.100 −0.101 0.410 0.040 0.343 0.722 0.146 0.935 0.226 10.28 66.83

−0.393 −0.127 −0.007 0.094 0.115 0.221 −0.143 −0.310 0.241 0.853 0.671 −0.122 −0.041 −0.089 −0.182 −0.064 −0.074 0.153 9.35 76.18

0.729 −0.115 0.127 0.215 0.005 −0.127 0.219 0.267 0.747 0.167 −0.304 0.019 −0.024 −0.009 0.033 0.047 0.019 −0.448 9.03 85.20

0.726 0.947 0.995 0.984 0.910 0.885 0.772 0.734 0.737 0.802 0.579 0.926 0.964 0.977 0.822 0.972 0.909 0.698

PC4 had high positive loadings for two oxyanions (AsT and NO− 3 ) with a weak positive correlation between them (r = 0.37). As a result, PC4 described the effects of redox process on AsT and NO− 3 distribution in the borewells rather than their common origin. A moderate inverse relationship between AsT and pH (r = −0.39) indicated that AsT concentrations in the borewells were partially controlled by pH fluctuation. In groundwaters with high pH values (8–9), As desorption is enhanced (Smedley and Kinniburgh, 2002; Sracek et al., 2004). However, the mean pH value in the present borewells was b8 (mean pH = 6.8), suggesting that AsT remains adsorbed onto the aquifer solids. 3− 2− Moreover, anions such as HCO− 3 , PO4 and SO4 can compete with As for binding sites on Fe-oxyhydroxide, Mn-oxides and clay minerals (Hingston et al., 1971; Jain and Loeppert, 2000; Sracek et al., 2004; Esteller et al., 2015). The lack of correlations between HCO–3 and SO2− 4 and AsT implied that the elevated AsT concentrations in borewells were not controlled by competitive adsorption between AsT and these two anconcentrations were ions for the binding sites (Table 9). Similarly, PO3− 4 too low (b0.5 mg/L) and undetectable in most borewells, and it was unlikely to compete with AsT for binding sites on mineral surfaces. Another geochemical process that may inhibit As adsorption in groundwater is cation exchange (Masue et al., 2007; Scanlon et al., 2009). Because of its positive charge and high abundance in natural waters, Ca2 + can promote adsorption of negatively charged arsenate

(Wilkie and Hering, 1996). However, an exchange of Na+ for Ca2 + may change surface charges and reduce As adsorption onto aquifer minerals, leading to high dissolved As concentrations in groundwater (Scanlon et al., 2009; Raychowdhury et al., 2014). All borewell samples had Na/Ca ratios b1 (Table 3), implying that Na+ exchange for Ca2+ did not occur in the wells, and cation exchange could not account for the high AsT concentration observed in most borewells (Raychowdhury et al., 2014). An increase in pH values appeared to be the main factor controlling AsT concentrations in borewells. A substantial increase in pH (i.e., OH−) following silicate mineral dissolution is likely to decrease AsT affinity for aquifer mineral binding sites. The influence of pH on AsT behavior in borewell waters was further supported by the absence of FeT in most borewells. At low Fe concentrations, the adsorption of As on Feoxyhydroxide minerals is strongly controlled by pH changes due to surface affinity for the available minerals (Smedley et al., 2005). At low pH, Fe-oxyhydroxides tend to attract As anions. As pH increases through the circumneutral range, as observed in the present borewells, the surface changes of the minerals become less positive thereby decreasing the affinity of Fe-oxyhydroxide for As adsorption. PC5 had high positive load(0.747) and pH (0.729). PC5 represented the contribution ings for SO2− 4 concentrations to borewell of sulfide mineral weathering to SO2− 4 samples.

Table 7 Pearson's correlation matrix for physico-chemical parameters of dug well waters (Correlation coefficients ≥0.6 are in bold).

pH HCO− 3 EC TDS Ca2+ Mg2+ K+ Na+ SO2− 4 NO− 3 AsT Cu Cr FeT Mn Ni Pb Zn

pH

HCO− 3

EC

TDS

Ca2+

Mg2+

K+

Na+

SO2− 4

NO− 3

AsT

Cu

Cr

FeT

Mn

Ni

Pb

Zn

1.00 0.89 0.90 0.90 0.88 0.88 0.38 0.75 0.82 −0.24 −0.14 −0,69 −0.61 −0.49 −0.39 −0.54 −0.61 −0.26

1.00 0.99 0.98 0.99 0.98 0.37 0.82 0.91 −0.33 −0.26 −0.64 −0.65 −0.59 −0.33 −0.55 −0.53 −0.13

1.00 0.97 0.98 0.98 0.35 0.84 0.91 −0.27 −0.30 −0,66 −0.66 −0.60 −0.31 −0.56 −0.55 −0.11

1.00 0.97 0.96 0.37 0.78 0.84 −0.44 −0.13 −0.62 −0.62 −0.57 −0.34 −0.54 −0.51 −0.13

1.00 0.99 0.50 0.71 0.94 −0.42 −0.27 −0.63 −0.66 −.62 −0.45 −0.55 −0.56 −0.25

1.00 0.52 0.71 0.93 −0.36 −0.35 −0.67 −0.69 −0.67 −0.41 −0.57 −0.62 −0.23

1.00 −0.19 0.53 −0.54 −0.19 −0.28 −0.31 −0.41 −0.42 −0.17 −0.44 −0.40

1.00 0.64 0.09 −0.21 −0.52 v0.48 −0.34 0.01 −0.43 −0.31 0.18

1.00 −0.24 −0.41 −0.67 −0.68 −0.63 −0.54 −0,59 −0.63 −0.40

1.00 −0.48 −0.13 −0.06 0.04 0.43 −0.07 −0.12 0.21

1.00 0.50 0.56 0.60 0.07 0.40 0.58 0.13

1.00 0.98 0.90 0.39 0.95 0.96 0.45

1.00 0.97 0.42 0.97 0.95 0.46

1.00 0.39 0.91 0.91 0.43

1.00 0.44 0.47 0.94

1.00 0.90 0.51

1.00 0.57

1.00


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Table 9 Pearson's correlation matrix for physico-chemical parameters of borewell waters (Correlation coefficients ≥0.6 are in bold).

pH HCO3EC TDS Ca2+ Mg2+ K+ Na+ SO2+ 4 NO− 3 AsT Cu Cr FeT Mn Ni Pb Zn

pH

HCO− 3

EC

TDS

Ca2+

Mg2+

K+

Na+

SO2− 4

NO− 3

AsT

Cu

Cr

FeT

Mn

Ni

Pb

Zn

1.00 0.01 0.09 0.07 −0.09 −0.20 0.29 0.29 0.23 −0.18 −0.37 0.22 0.15 0.18 0.20 0.22 0.09 −0.26

1.00 0.92 0.87 0.85 0.81 0.39 0.74 0.10 0.03 −0.14 0.37 0.32 0.33 0.21 0.39 0.25 0.37

1.00 0.98 0.93 0.87 0.43 0.79 0.37 0.18 −0.16 0.26 0.31 0.24 0.11 0.36 0.01 0.27

1.00 0.91 0.86 0.44 0.76 0.47 0.28 −0.15 0.25 0.28 0.22 0.09 0.34 0.02 0.25

1.00 0.91 0.23 0.58 0.31 0.28 −0.10 0.14 0.22 0.12 0.01 0.26 −0.09 0.17

1.00 0.23 0.44 0.27 0.30 0.00 0.12 0.20 0.11 0.02 0.22 −0.07 0.34

1,00 0.46 0.16 −0.07 −0.26 0.58 0.71 0.63 0.27 0.72 0.07 0.46

1.00 0.36 −0.13 −0.19 0.22 0.22 0.21 0.17 0.28 0.08 0.17

1.00 0.21 −0.12 −0.03 −0.05 −0.06 −0.01 0.00 0.01 −0.16

1.00 0.37 −0.20 −0.11 −0.17 −0.25 −0.11 −0.14 0.00

1.00 −0.22 −0.15 −0.20 −0.27 −0.21 −0.15 0.10

1.00 0.88 0.97 0.69 0.92 0.55 0.54

1.00 0.94 0.56 0.99 0.17 0.57

1.00 0.72 0.97 0.48 0.58

1.00 0.62 0.68 0.45

1.00 0.29 0.55

1.00 0.28

1.00

Elevated AsT concentrations in dug wells and borewells do not seem to be a function of geochemical background concentrations, but dependent on the ability or inability of bedrock rocks to mobilize As. For example, granitic rocks may have minerals containing high As content, but they are characterized by low carbonate minerals, and hence do not favor As mobilization during water rock–interaction. In contrast, metasedimentary rocks, the predominant rock type in the study area, may not contain high As concentrations, but, with high carbonate mineral content, are more susceptible to mobilize As within the aquifer (Cook et al., 1991). Similarly, highly weathered layers are characterized by low calcite content compared to less weathered fractured rocks. Consequently, the fractured bedrock aquifers are more prone to favor As mobilization than weathered aquifers. It is interesting to note that the samples with elevated AsT are also located near the mineralized zone, highlighting contribution of Au mineralization to As mobility in dug wells and borewells. Granite and metagabbro bedrock aquifers with lower pH values and lower As concentrations (e.g., W2, W4, W8, B6, B7, B8 and B18; Fig. 3) are also distant from the mineralized zone. Weathered layers are also rich in secondary minerals such as clay (semectite and kaolinite), Fe-oxyhydroxides, Mn-and Al-oxides that are excellent trace element adsorbents. High residence time of groundwater in borewells also increases water–rock interaction, and thereby enhancing rate of As mobilization in the aquifers. Q-mode factor analysis produced a single component that explains 96% of the total variance (data not presented), meaning that about 4% of variation was not high enough to produce two factors describing dug well and borewell samples. Despite the compositional differences between individual samples, the two hydrogeological systems have comparable geochemical signatures. This supports the idea that dug wells and borewells may be interconnected through vertical fractures. More studies on connectivity between shallow weathered aquifers and deep fractured bedrock aquifers are required.

Due to their closeness to the surface and constant recharge, dug well waters were characterized by relatively low residence time and shorter water–rock interactions. Furthermore, the relatively high Cl− and PO3− 4 concentrations and coliform presence suggested that dug wells were more exposed to anthropogenic activities than borewells. Although 2− concentrations were less than the WHO recommended NO− 3 and SO4 values, the pronounced abundance of these anions in water samples suggests that the local groundwater resources are vulnerable to agricultural activities. The average concentrations of FeT and Mn in dug wells and the average concentrations of Pb and AsT in borewells were higher than the WHO permissible limits for drinking water. Dug wells and borewells are therefore exposed to both geogenic and anthropogenic sources of pollution. Supersaturation of Fe bearing minerals in dug wells, suggested that trace element mobilization in dug wells was likely limited by their possible adsorption onto Fe-oxyhydroxides. Elevated Fe concentrations in conjunction with low As concentrations in dug wells suggested As being co-precipitation/readsorbed with residual with residual Fe. In contrast, relatively high pH values in most borewells, following chemical weathering, might have inhibited AsT adsorption onto the aquifer solids, leading to elevated AsT concentrations (N10 μg/L). The R-mode factor analysis suggested that Fe-oxyhydroxide precipitation and chemical weathering were the geochemical processes controlling dug well and borewell chemistry. The PC loadings and correlation coefficients revealed that most important factor that contributes to the mobility of trace and major elements in dug wells and borewells is the chemical weathering of the parent rocks. The high FeT concentrations in dug wells may also be related to the mineralogical composition of the weathered layers, to the low pH values and possible Fe-oxyhydroxide reduction. This study demonstrated that regardless to aquifer locations, As concentrations in groundwater is primarily controlled by the mineralogical composition of the aquifers and geochemical processes that may or not favor As mobilization.

6. Conclusion Acknowledgments Physico-chemical data, total coliform content, saturation indices and R-mode factor analysis provided insight into borewell and dug well water quality around Bomboré gold mineralized zone in central Burkina Faso. Hydrogeochemical facies of dug well and borewell waters was a Ca-Mg-HCO3 types. Scatter plots of major elements illustrated that dug well and borewell water chemistry was controlled by silicate behavior in samples indiand carbonate weathering. Nitrate and SO2− 4 cated prevailing mixed redox conditions in dug wells, whereas the majority of borewells appeared to be under suboxic conditions.

This publication is an integral part of Environmental and Social Impact Assessment of the Bomboré gold mining project carried out by Orezone Gold Corporation. The authors extend their thanks to Pascal MARQUIS the Vice-president of the company's exploration unit for his valuable assistance for sample collections. Geological and sample location map compilation by Adama Traoré is greatly appreciated. We express thanks to two anonymous reviewers for their thoughtful comments and suggestions that had greatly improved the manuscript.

70


References Acworth, R.I., 1987. The development of crystalline basement aquifers in a tropical environment. Q. J. Eng. Geol. 20, 265–272. Ayotte, J.D., Montgomery, D.L., Flanagan, S.M., Robinson, K.W., 2003. Arsenic in groundwater in eastern New England: occurrence, controls, and human health implications. Environ. Sci. Technol. 37, 2075–2083. Bamba, O., Béziat, D., Bourges, F., Debat, P., Lompo, M., Parisot, J.C., Tollon, F., 1997. Nouveau type de gisement aurifère dans les ceintures des roches vertes birimiennes du Burkina Faso: les albites de Larafella. J. Afr. Earth Sci. 25, 364–381. Bamba, O., Pelédé, S., Sako, A., Kagambéga, N., Miningou, Y.W.M., 2013. Impact de l'artisanat minier sur les sols d'un environnement ménagé au Burkina Faso. J. Sci. 13, 1–11. Béziat, D., Dubois, M., Debat, P., Nikiéma, S., Salvi, S., Tollon, J., 2008. Gold metallogeny in the Birimian Craton of Burkina Faso (West Africa). J. Afr. Earth Sci. 50, 215–233. BGS (British Geological Survey), 2002. Groundwater Quality: Burkina Faso 1–4 pp. BILAN D'EAU, 1993. Carte Hydrogeologique du Burkina Faso. Boruvka, L., Vecek, O., Jehlicka, J., 2005. Principal component analysis as a tool to indicate the origin of potentially toxic elements in soil. Geoderma 128, 289–300. Böttcher, J., Strebel, O., Voerkelius, S., Schmidt, H.L., 1990. Using isotope fractionation of nitrate-nitrogen and nitrate-oxygen for evaluation of microbial denitrification in a sandy aquifer. J. Hydrol. 114, 413–424. Bowell, R.J., 1994. Sorption of arsenic by iron oxides and oxyhydroxides in soils. Appl. Geochem. 9, 279–286. Brenot, A., Baran, N., Petelet-Giraud, E., Negrel, P., 2008. Interaction between different water bodies in a small catchment in the Paris basin (Breville, France): tracing of multiple Sr sources through Sr isotopes coupled with Mg/Sr and Ca/Sr ratios. Appl. Geochem. 23, 58–75. Castaing, G., Billa, M., Milési, J.P., Thieblemont, D., Le Metour, J., Egal, E., Donzeau, M., Guenot, C., Cocherie, A., Chevremont, P., Tegyey, M., Itard, Y., Zida, B., Ouédraogo, I., Koté, S., Kaboré, B.E., Ouédraogo, C., Zunino, J.C., 2003. Notice explicative de la carte géologique et minière a 1/100000 du Burkina Faso, Ouagadougou, Burkina Faso 58 pp. Chilton, P.J., Foster, S.S.D., 1995. Hydrogeological characterisation and water-supply potential of basement aquifers in tropical Africa. Hydrogeol. J. 3, 36–49. Cook, J.M., Edmunds, W.M., Robins, N.S., 1991. Groundwater contribution to an acid upland lake (Loch Fleet, Scotland) and the possibilities for amelioration. J. Hydrol. 111–128. Courtois, N., Lachassagne, P., Wyns, R., Blanchin, R., Bougaïré, F.D., Somé, S., Tapsoba, A., 2010. Large-scale mapping of hard-rock aquifer properties applied to Burkina Faso. Ground Water 48 (2), 269–283. COWI, 2004. Etude sur l'arsenic dans l'eau souterraine de la zone du PEEN. COWI Engineering, Programme Eau et Environnement – Région du Nord (PEEN). Ministère de l'Agriculture, de l'Hydraulique et des Ressources Halieutiques, Burkina Faso. de Jong, S.J., Kikietta, A., 1981. Une particularité bien localisée heureusement: présence d'arsenic en concentration toxique dans un village près de Mogtédo (Haute Volta). Bulletin de Liaison du Comite Interafricain d'Etude Hydrauliques, No. 42–43. Detay, M., Poyet, P., Emesellem, Y., Aubrac, G., 1989. Development of the saprolite reservoir and its state of saturation: Influence on the hydrodynamic characteristics of drillings in crystalline basement (in Franch). C.R. Acad. Sci. Paris Serie II 309, 429–436. Deutsch, W.J., 1997. Groundwater Geochemistry: Fundamentals and Application to Contamination. Lewis Publisher, New York 221 pp. Dewandel, B., Lachassagne, P., Wyns, R., Maréchal, J.C., Krishnamurthy, N.S., 2006. A generalized 3-D geological and hydrogeological conceptual model of granite aquifers controlled by single or multiphase weathering. J. Hydrol. 330, 260–284. Drever, J.I., 1982. The Geochemistry of Natural Waters. Prentice-Hall. Wiley and Sons Ltd., Hoboken, NJ, USA 330 pp. Edmunds, W.M., Andrews, J.N., Burgess, W.G., Kay, R.L.F., Lee, D.J., 1984a. The evolution of saline and thermal groundwaters in the Carnmenellis granite. Mineral. Mag. 48 (348), 407–424. Edmunds, W.M., Miles, D.L., Cook, J.M., 1984b. A comparative study of sequential redox processes in three British aquifers. In: Eriksson, E. (Ed.), Hydrochemical Balances of Freshwater Systems. Proc. Symp. Uppsala, Sept. 1984. IAHS Publication No. 150, pp. 55–70. Edmunds, W.M., Guendouz, A.H., Mamou, A., Moulla, A., Shand, P., 2003. Groundwater evolution in the continental intercalcaire aquifer of southern Algeria and Tunisia: trace element and isotopic indicators. Appl. Geochem. 18, 805–822. Esteller, M.V., Domiguez-Mariani, E., Garrido, S.E., Avillés, M., 2015. Groundwater pollution by arsenic and other toxic elements in an abandoned silver mine, Mexico. Environ. Earth Sci. 015, 4315-4315. Everitt, S.E., Landau, S., Leese, M., 2001. Cluster analysis. Arnold, London 237 pp, Gamsonré, P.E., 2003. Notice explicative de la carte géologique du Burkina Faso à 1/ 200,000. Farnham, I.M., Johannesson, K.H., Singh, A.K., Hodge, V.F., Stetzenbach, K.J., 2003. Factor analytical approaches for evaluating groundwater trace element chemistry data. Anal. Chim. Acta 490, 123–138. Flint, A.L., Flint, L.E., Kwicklis, E.M., Fabryka-Martin, J.T., Bodvarsson, G.S., 2002. Estimating recharge at Yucca Mountain, Nevada,USA: comparison of methods. Feuille Ouahigouya, First edition. Hydrogeol. J Vol. 10(1), pp. 180–204 Burkina Faso Ministère des Mines, des Carrières et de l'énergie–30. Frind, E.O., Duynisveld, W.H.M., Strebel, O., Böttcher, J., 1990. Modeling of multicomponent transport with microbial transformation in groundwater: the Fuhrberg case. Water Resour. Res. 26, 1707–1719. Gamsonré, P.E., 2003. Notice explicative de la carte géologique du Burkina Faso à 1/ 200,000. Feuille Ouahigouya. First edition. Burkina Faso Ministère des Mines, des Carrières et de l'énergie.

Gnanadesikan, R., 1977. Methods for Statistical Data Analysis of Multivariate Observations. Wiley, New York 347 pp. Groen, J., Schuchmann, J.B., Geirnaert, W., 1988. The occurrence of high nitrate concentration in groundwater in villages in northwestern Burkina Faso. J. Afr. Earth Sci. 7, 999–1009. Gustafson, J.P., Pechová, P., Berggren, D., 2003. Modelling metal binding to soils: the role of organic matter. Environ. Sci. Technol. 37, 2767–2774. Hem, J.D., 1985. Study and interpretation of the chemical characteristics of natural water. third edition. U. S. Geol. Surv. Water-Supply Pap. 2254 p. 272. Helena, B., Pardo, R., Vega, M., Barrado, E., Fernandez, J.M., Fernandez, L., 2000. Temporal evolution of groundwater composition in an alluvial aquifer (Pisuerga River, Spain) by principal component analysis. Water Res. 34, 807–816. Hernandez, M.A., Gonzalez, N., Levin, M., 1991. Multivariate analysis of a coastal phreatic aquifer using hydrochemical and isotopic indicators, Buenos Aires, Argentina, Proceedings of the International Association on Water Pollution Research and Control's International Seminar on Pollution, Protection and Control of Groundwater. Water Sci. Technol. 24, 139–146. Hingston, F.J., Posner, A.M., Quirk, J.P., 1971. Competitive adsorption of negatively charged ligands on oxide surfaces. Discuss. Faraday Soc. 52, 334–342. Horowitz, A.J., Demas, C.R., Fitzgerald, K.K., Miller, T.L., Rickert, D.A., 1994. US Geological Survey protocol for the collection and processing of surface-water samples for the subsequent determination of inorganic constituents in filtered water. Open-File Report 94. US Geological Survey 57 pp. Hotelling, H., 1933. Analysis of a complex of statistical variables into principal components. J. Educ. Psychol. 26 (6), 417–441. Hsia, T.H., Lo, S.L., Lin, C.F., 1992. As(V) adsorption on amorphous iron oxide: triple layer modelling: Chemosphere, v. 25, p. 1825–1837. J. Educ. Psychol. 24, 498–520. Jacob, J.-P., Ouédraogo, S., Paré, L., 2002. Etude des systèmes locaux de gestion foncière dans la zone d'intervention du Plan Foncier Rural/Ganzourgou, rapport final. Ministère de l'Agriculture. Jain, A., Loeppert, R.H., 2000. Effect of competing anions on the adsorption of arsenate and arsenite by ferrihydrite. J. Environ. Qual. 29, 1422–1430. Jenne, E.D., 1979. Chemical modeling in aqueous systems. ACS Symposium Series No. 93; Am. Chem. Soc. Symp., Washington, DC 147–180 pp. Join, J.L., Coudray, J., Longworth, K., 1997. Using principal components analysis and Na/Cl ratios to trace groundwater circulation in a volcanic island: the example of reunion. J. Hydrol. 190, 1–18. Kaiser, H.F., 1960. The application of electronic computers to factor analysis. Educ. Psychol. Meas. 20, 141–151. Kortatsi, B.K., 2007. Hydrochemical framework of groundwater in the Ankobra Basin, Ghana. Aquat. Geochem. 13, 41–74. Koterba, M.T., Wilde, F.D., Lapham, W.W., 1995. Groundwater data collection protocols and procedures for the National Water Quality Assessment Program: collection and documentation of water-quality samples and related data. Open-File Report. US Geological Survey Vol. 312, pp. 95–399. Kuppusamy, M.R., Grirdhar, V.V., 2006. Factor analysis of water quality characteristics including trace metal speciation in the coastal environmental system of Chennai Ennore. Environ. Int. 32, 174–179. Levison, L., Novakowski, K., 2009. The impact of cattle pasturing on groundwater quality in bedrock aquifers having minimal overburden. Hydrogeol. J. 17, 559–569. Lipfert, G., Reeve, A.S., Sidle, W.C., Marvinney, R., 2006. Geochemical patterns of arsenic enriched ground water in fractured, crystalline bedrock, Northport, Maine, USA. Appl. Geochem. 21, 528–545. Luukkonen, A., Pitkanen, P., Ruotsalainen, P., Leino-Forsman, H., Snellman, M., 1999. Hydrochemical conditions at the Hastholmen site. Report POSIVA 99–26. Posiva Oy, Helsinki, Finland 64 p. Mandal, B.K., Chowdhury, T.R., Samanta, G., Mukherjee, D.P., Chanda, C.R., Saha, K.C., Chakraborti, D., 1996. Impact of safe water for drinking and cooking on five arsenic affected families for 2 years in West Bengal, India. Sci.Total Environ. 218, 185–201. Manning, B.A., Fendorf, S.E., Goldberg, S., 1998. Surface structures and stability of arsenic(III) on goethite: spectroscopic evidence for inner-sphere complexes. Environ. Sci. Technol. 32, 2383–2388. Maréchal, J.C., Dewandel, B., Subrahmanyam, K., 2004. Contribution of hydraulic tests at different scales to characterize fracture network properties in the weatheredfissured layer of a hard rock aquifers. Water Resour. Res. 40, 1–17. Masue, Y., Loeppert, R.H., Kramer, T.A., 2007. Arsenate and arsenite adsorption and desorption behavior on coprecipitated aluminum: iron hydroxides. Environ. Sci. Technol. 41, 837–842. McCleskey, R.B., Nordstrom, D.K., Maest, A.S., 2004. Preservation of water samples for arsenic (III/V) determinations: an evaluation of the literature and new analytical results. Appl. Geochem. 19, 995–1009. MEE, (Ministry of Water and the Environment), 1998. Politique et stratégies en matière d'eau Ouagadougou. Meglin, R.R., 1991. Examining large databases: a chemometric approach using principal component analysis. J. Chemomet. 5, 163–179. Melchior, D.C., Bassett, R.L., 1990. Chemical modeling of aqueous system II. ACS Symp. Ser. No. 416. Am. Chem. Soc., Washington DC 556 pp. Meng, X., Bang, S., Korfiatis, G.P., 2000. Effects of silicate, sulfate, and carbonate on arsenic removal by ferric chloride. Water Res. 34, 1255–1261. Milési, J.P., Ledru, P., Feybbesse, J.L., Dommanget, A., Maraux, E., 1992. Early proterozoinc ore deposits and tectonics of Birimian orgenic belt, West Africa. Precambrian Res. 58, 305–344. Nath, B., Jean, J.-S., Lee, M.-K., Yang, H.-J., Liu, C.-C., 2008. Geochemistry of high arsenic groundwater in Chia-Nan plain, Southwestern Taiwan: possible sources and reactive transport of arsenic. J. Contam. Hydrol. 99, 85–96. Pacheco, F., Van der Weijden, C.H., 1996. Contribution of water–rock interactions to the composition of groundwater in areas within a sizable anthropogenic input: a case

A. Sako et al. / Journal of Geochemical Exploration 170 (2016) 58–71 study of the waters of the Fundao area, Central Portugal. Water Resour. Res. 32, 3553–3570. Peters, S.C., Blum, J.D., 2003. The source and transport of arsenic in a bedrock aquifer, New Hampshire, USA. Appl. Geochem. 18, 1773–1787. Pierce, M.L., Moore, C.B., 1982. Adsorption of arsenite and arsenate on amorphous iron hydroxide. Water Res. 16, 1247–1253. Postma, D., Boesen, C., Kristiansen, H., Larsen, F., 1991. Nitrate reduction in an unconfined sandy aquifer: water chemistry, reduction processes, and geochemical modeling. Water Resour. Res. 27, 2027–2045. Ravenscroft, P., McAthur, J.M., Hoque, B.A., 2001. Geochemical and palaeonhydrological controls on pollution of groundwater by arsenic. In: Cheppell, W.R., Abernathy, C.O., Calderon, R. (Eds.), Arsenic Exposure and Health Effects IV. Elsevier Sceinces Ltd., Oxford, pp. 53–78. Raychowdhury, N., Mukherjee, A., Bhattacharya, P., Johannesson, K., Bundschuh, J., Sifuentes, G.B., Nordberg, E., Martin, R.A., Storniolo, A., 2014. Provenance and fate of arsenic and other solutes in the Chaco-Pampean plain of the Andean foreland, Argentina: from perspectives and hydrogeochemical modeling and regional tectonic setting. J. Hydrol. 518, 300–316. Reyment, R.A., Jvreskog, K.G., 1996. Applied Factor Analysis in the Natural Sciences. Cambridge University Press, Cambridge 383 pp. Rödelsperger, M., 1989. Natural denitrification processes in the aquifer. In: Kobus, H., Kinzelbach, W. (Eds.), Proceedings of Symposium on Contaminant Transport in Groundwater. A.A. Balkema, Rotterdam, Netherlands, pp. 159–163. Sattran, V., Wenmenga, U., 2002. Geologie du Burkina Faso/Geology of Burkina Faso. Czech Geological Survey, Prague, Czech Republic 136 pp. Scanlon, B.R., Nicot, J.P., Reedy, R.C., Kurtzman, D., Mukherjee, A., Nordstrom, D.K., 2009. Elevated naturally occurring arsenic in a semiarid oxidizing system, Southern High Plains aquifer, Texas, USA. Appl. Geochem. 24, 2061–2071. Schreiber, M.E., Gotkowitz, M.B., Simo, J.A., Freiberg, O.G., 2003. Mechanisms of arsenic release to ground water from naturally occurring sources, Eastern Wisconsin. In: Welch, A.H., Stollenwerk, K.G. (Eds.), Arsenic in Ground Water. Kluwer Academic Publishers, Dordrecht, pp. 259–280. Seyhan, E., van de Griend, A.A., Engelen, G.B., 1985. Multivariate analysis and interpretation of the hydrochemistry of a dolomitic reef aquifer, northern Italy. Water Resour. Res. 21 (7), 1010–1024. Smedley, P.L., Edmunds, W.M., 2002. Redox patterns and trace element behavior in the East Midlands Triassic Sandstone Aquifer, UK. Ground Water 4, 44–58. Smedley, P.L., Kinniburgh, D.G., 2002. A review of the source, behaviour and distribution of arsenic in natural waters. Appl. Geochem. 17, 517–568. Smedley, P.L., Kinniburgh, D.G., Macdonald, D.M.J., Nicolli, H.B., Barros, A.J., Tullio, J.O., Pearce, J.M., Alonso, M.S., 2005. Arsenic associations in sediments from the loess aquifer of La Pampa, Argentina. Appl. Geochem. 20, 989–1016. Smedley, P.L., Knudsen, J., Maiga, D., 2007. Arsenic in groundwater from mineralized Proterozoic basement rocks of Burkina Faso. Appl. Geochem. 22, 1074–1092.

71

Snellman, M., Pitkanen, P., Luukkonen, A., Ruotsalainen, P., Leino-Forssman, H., Vuorinen, U., 1998. Summary of recent observations from Hastholmen hydrogeochemical studies. Work Report 98–44. Posiva Oy, Helsinki, Finland 37 pp. Somé, I.T., Sakira, A.K., Ouedraogo, M., Ouedraogo, T.Z., Traoré, A., Sondo, B., Guissou, P.I., 2011. Arsenic, levels in tube-wells water, food, residents' urine and the prevalence of skin lesions in Yatenga Province, Burkina Faso. Interdiscip. Toxicol. 5, 38–41. Sracek, O., Bhattacharya, P., Jacks, G., Gustafsson, J.P., Von Brömssen, M., 2004. Behavior of arsenic and geochemical modeling of arsenic enrichment in aqueous environments. Appl. Geochem. 19, 169–180. Stallard, R., Edmond, J.M., 1983. Geochemistry of the Amazon. 2. The influence of geology and weathering environment on the dissolved load. J. Geophys. Res. 88, 9671–9688. Taylor, R., Howard, K., 2000. A tectono-geomorphic model of the hydrogeology of deeply weathered crystalline rock: evidence from Uganda. Hydrogeol. J. 8 (3) 279–274. Taylor, R.G., Tindimugaya, C., Barker, J.A., Barrett, M., Macdonald, D.M.J., Kulabako, R., Nalubega, M., Edwardmartin, R., 2010. Convergent radial tracing of viral and solute transport in gneiss saprolite. Ground Water 48 (2), 284–294. Utsunomiya, S., Peters, S.C., Blum, J.D., Ewing, R.C., 2003. Nanoscale mineralogy of arsenic in a region of New Hampshire with elevated As-concentrations in the groundwater. Am. Mineral. 88, 1844–1852. Valdes, D., Dupont, J.-P., Laignel, B., Ogier, S., Leboulanger, T., 2007. A spacial analysis of tructural controls on karst groundwater geochemistry at a regional scale. J. Hydrol. 34, 244–255. WHO, 2006. Guidelines for drinking water quality. 3rd edition. World Health Organization, 20 Avenue Appia, 1211 Geneva 27, Switzerland, pp. 488–493. Wilkie, J.A., Hering, J.G., 1996. Adsorption of arsenic onto hydrous ferric oxide: effects of adsorbate/adsorbent ratios and co-occurring solutes. Colloids Surf. A Physicochem. Eng. Asp. 107, 97–110. Winograd, I.J., Robertson, F.N., 1982. Deep oxygenated ground water: anomaly or common occurrence? Science 216, 1227–1230. Wright, E.P., 1992. The hydrogeology of crystalline basement aquifers in Africa. In: Wright, E.P., Burgess, W.G. (Eds.), The Hydrogeology of Crystalline Basement Aquifers in Africa. Geol. Soc. Special Pub Vol. 66, pp. 1–27. Wyns, R., Gourry, J.-C., Baltassat, J.-M., Lebert, F., 1999. Caractérisation multiparamètres des horizons de subsurface (0–100 m) en contexte de socle altéré. 2ème Colloque GEOFCAN. I. BRGM. UPMC, Orléans, France 105-110 pp. Wyns, R., Baltassat, J.M., Lachassagne, P., Legchenko, A., Vairon, J., Mathieu, F., 2004. Application of proton magnetic resonance soundings to groundwater reserve mapping in weathered basement rocks (Brittany, France). Bulletin de la Société Géologique de France 175, no. 1: 21–34. Geol. Soci. Special Pub. No.66, pp. 1–27. Yaméogo, S., Savadogo, A.N., 2002. Les Ouverages de Captage de la ville de Ouagadougou et Leur Vulnérabilité a la Pollution. In: Maiga, A.H., Pereira, L.S., Musy, A. (Eds.), Sustainable Water Resources Management: Health and Productivity. Zhu, G.F., Su, Y.H., Feng, Q., 2008. The hydrochemical characteristics and evolution of groundwater and surface water in the Heihe River Basin, northwest China. Hydrogeol. J. 16, 167–182.