**2. Transformation of inorganic contaminants in groundwater**

**Figure 1.** A river in the west of Johannesburg, South Africa, contaminated by acid mine drainage from disused under‐

Acid mine drainage (AMD) is formed via a cascade of reactions (Equations 1–4) when sulphide minerals are exposed to oxygen by mining [6]. The process begins when oxygenated water percolates through the finely divided tailings and pyrite is oxidised to ferrous iron (Equation 1) and then to ferric iron (Equation 2). Ferric iron which is soluble at pH below 3.5 then acts as an additional oxidising agent for pyrite (Equation 3). Above pH 3.5, ferric iron precipitates as Fe(OH)3 (Equation 4); a reaction that is able to buffer the pH of AMD at pH 2.5–3.5 [7].

<sup>7</sup> 2 2 FeS / O H O Fe 2SO 2H 2 22 2 <sup>4</sup>

2 1 3 1 Fe / O H Fe / H O 4 2 2 2

<sup>3</sup> 2 2 FeS Fe 12H O 12Fe 2SO 16H 22 4

( ) <sup>3</sup>

+ ++ + +®+ + (1)

+ ++ + +® + (2)

<sup>+</sup> + ++ ++ ® + + (3)

<sup>2</sup> <sup>3</sup> Fe 3H O Fe OH 3H + + +® + (4)

ground mines. Note the orange colour which is due to deposition of iron flocs on the river bed.

70 Groundwater - Contaminant and Resource Management

The transformation of inorganic contaminants in groundwater is influenced by a number of chemical and physical characteristics of aquifers such as pH, redox potential (Eh) and organic matter [8]. These properties have an influence over contaminant transformation reactions including precipitation/co-precipitation-dissolution, oxidation-reduction and acid–base/ sorption–desorption [9].

### **2.1. Precipitation/dissolution**

Precipitation/co-precipitation and dissolution reactions play an important role in determining elemental concentrations in groundwater. Jean and Bancroft [10] showed, using pyrite (FeS), sphalerite (ZnS), galena (PbS) and pyrrhotite (Fe*1-x*S), that sulphide minerals are excellent scavengers for dissolved Hg2+, Pb2+, Zn2+ and Cd2+ ions, acting as adsorbents for these metals. However, Özverdi et al. [11] reported that under acidic conditions (< pH 3), metal removal by pyrite was by precipitation of metal sulphides due to the presence of H2S. In contrast, metal hydroxides were precipitated under basic conditions. These observations were supported by the field studies of Muller et al. [12] in the Kristineberg mines in Sweden and Al et al. [13] in the Kidd Creek tailings impoundment in Ontario, Canada, which that both adsorption and precipitation reactions were responsible for metal attenuation.

Ferric iron is ubiquitous in AMD-contaminated environments. Its precipitation via various pathways (Equations 5–7) is a significant process in the attenuation of metal on concentrations in mine drainage via co-precipitation reactions [14]. Indeed, Fe- and Al-compounds are commonly used for the chemical precipitation of As(III) and As(V) in water treatment plants [15].

$$\text{Fe}^{3+} + 3\text{H}\_{2}\text{O} \rightarrow \text{Fe} \text{(OH)}\_{3} + \text{3H}^{+} \tag{5}$$

$$\text{Fe}^{3+} + 2\text{H}\_2\text{O} \rightarrow \alpha\text{-FeOOH} + 3\text{H}^+ \tag{6}$$

$$\mathrm{K}^{\cdot +} + \mathrm{Fe}^{3+} + 2\mathrm{SO}\_{4}^{2-} + 6\mathrm{H}\_{2}\mathrm{O} \rightarrow \mathrm{KFe}\_{3}\mathrm{(SO}\_{4}\mathrm{)}\_{2}\mathrm{(OH)}\_{6} + 6\mathrm{H}^{\cdot} \tag{7}$$

Ferrous iron, on the other hand, is controlled by the precipitation of siderite (FeCO3) (Equation 8), a reaction occurring mainly in shallow tailings.

$$\text{Fe}^{2+} + \text{HCO}\_3 \rightarrow \text{FeCO}\_3 + \text{H}^+ \tag{8}$$

Elemental concentrations are also controlled by simple precipitation reactions in response to solution pH through the formation of insoluble hydroxides such as Pb(OH)2 and Cu(OH)2 or following reaction with sulphides and carbonates to form insoluble compounds [11].

With respect to dissolution, McGregor et al. [16] found that Ca and Mg concentrations in the Copper Cliff tailings in Ontario, Canada, were controlled by the dissolution of carbonate and aluminosilicate minerals during pH buffering reactions. Similarly, Mn was derived from the dissolution of pyroxene, chlorite, amphibole or carbonates, Al and Si from weathering of biotite, orthopyroxene and feldspars and Ni and Zn concentrations from the oxidation and dissolution of pentlandite ((Fe,Ni)9S8) and sphalerite (ZnS). K and Na on the other hand were controlled by dissolution of aluminosilicate minerals although their concentrations were limited by equilibrium with respect to jarosite (KFe3(SO4)2(OH)6) and natrojarosite (NaFe3(SO4)2(OH)6). Dissolution may also be microbially-driven. Cummings et al. [17] reported the release of arsenate following the dissolution of scorodite (FeAsO4.2H2O) by an ironreducing bacterium, *Shewanella alga*.

## **2.2. Oxidation–reduction**

including precipitation/co-precipitation-dissolution, oxidation-reduction and acid–base/

Precipitation/co-precipitation and dissolution reactions play an important role in determining elemental concentrations in groundwater. Jean and Bancroft [10] showed, using pyrite (FeS), sphalerite (ZnS), galena (PbS) and pyrrhotite (Fe*1-x*S), that sulphide minerals are excellent scavengers for dissolved Hg2+, Pb2+, Zn2+ and Cd2+ ions, acting as adsorbents for these metals. However, Özverdi et al. [11] reported that under acidic conditions (< pH 3), metal removal by pyrite was by precipitation of metal sulphides due to the presence of H2S. In contrast, metal hydroxides were precipitated under basic conditions. These observations were supported by the field studies of Muller et al. [12] in the Kristineberg mines in Sweden and Al et al. [13] in the Kidd Creek tailings impoundment in Ontario, Canada, which that both adsorption and

Ferric iron is ubiquitous in AMD-contaminated environments. Its precipitation via various pathways (Equations 5–7) is a significant process in the attenuation of metal on concentrations in mine drainage via co-precipitation reactions [14]. Indeed, Fe- and Al-compounds are commonly used for the chemical precipitation of As(III) and As(V) in water treatment plants

<sup>2</sup> <sup>3</sup> Fe 3H O Fe OH 3H + + +® + (5)

+ + +® + (6)

+ + +® + (8)

4 2 34 2 6 K Fe 2SO 6H O KFe SO OH 6H ++ - <sup>+</sup> ++ + ® <sup>+</sup> (7)

( ) <sup>3</sup>

<sup>3</sup> Fe 2H O -FeOOH 3H <sup>2</sup> a

( ) ( ) 3 2

<sup>2</sup> Fe HCO FeCO H 3 3

following reaction with sulphides and carbonates to form insoluble compounds [11].

Ferrous iron, on the other hand, is controlled by the precipitation of siderite (FeCO3) (Equation

Elemental concentrations are also controlled by simple precipitation reactions in response to solution pH through the formation of insoluble hydroxides such as Pb(OH)2 and Cu(OH)2 or

With respect to dissolution, McGregor et al. [16] found that Ca and Mg concentrations in the Copper Cliff tailings in Ontario, Canada, were controlled by the dissolution of carbonate and aluminosilicate minerals during pH buffering reactions. Similarly, Mn was derived from the dissolution of pyroxene, chlorite, amphibole or carbonates, Al and Si from weathering of

precipitation reactions were responsible for metal attenuation.

8), a reaction occurring mainly in shallow tailings.

sorption–desorption [9].

[15].

**2.1. Precipitation/dissolution**

72 Groundwater - Contaminant and Resource Management

Oxidation–reduction reactions may be chemically- or biologically-driven [18]. Selenate (SeO4 2−) may be reduced to elemental selenium (Se0 ) by ferrous hydroxide [19] and zero-valent iron (ZVI) has been used for the reductive precipitation of As from contaminated water [20]. Microbial oxidation-reduction in some cases can be many times faster than abiotic reactions [21]. The oxidation of As(III) to As(V) by a *Thermus* species, for example, was found to be approximately 100 times faster than abiotic rates [22].

Redox states of As have environmental implications because of their effect on As speciation. As(V), the predominant form in aerobic environments, is more strongly sorbed to mineral surfaces and thus less mobile than As(III) which sorbs less strongly and is thus more mobile [23]. Routh et al. [24] conducted microcosm experiments to investigate As behaviour in mine tailings near the Adak mine in northern Sweden. They found that microbial reduction of As(V) to As(III) increased the concentrations of the latter in aqueous media and as such, enhanced As remobilisation from sediments. In contrast, As(V) concentrations increased in sediments and aqueous media of control experiments treated with formaldehyde and HgCl2.

Treatments applied to mine tailings may also have an effect on microbial activity and As behaviour in mining-impacted environments. Macur et al. [25] reported that the addition of lime (CaO), a common treatment applied to mine tailings to immobilise metal ions, stimulated As-reducing microorganisms (*Caulobacter-, Sphingomonas*- and *Rhizobium*-like populations) and in turn, enhanced As(V) reduction and mobilisation in tailings.

Fe, an important variable in mining-impacted environments, also influences As mobilisation. Han et al. [26] reported that Fe(II) significantly inhibited the removal of As(III) by MnO2 in acidic environments (pH 3). They postulated that ferric iron compounds formed a coating on MnO2 surfaces which inhibited access of As(III) ions to oxidation sites on MnO2. The inhibition was however dependent on how Fe(II) ions were introduced into the system. Where the MnO2 was pre-treated with Fe(II), As(III) diffused through the schwertmannite coatings that formed on the MnO2 and its oxidation to As(V) was possible. However, where Fe(II) and As(III) were introduced simultaneously, competitive oxidation of the two ions prevented the com‐ plete oxidation of As(III) due to the formation of FeOHAs or FeAsO4 coatings on the MnO2 surface.

Cr exists mainly as Cr(III) and Cr(VI). Cr(VI) is highly soluble and therefore more mobile while Cr(III), tends to precipitate as amorphous hydroxides e.g. Cr(OH)3 and ((Fe,Cr)OH3) in slightly acidic and alkaline environments. Cr(III) is commonly oxidised to Cr(VI) by manganese oxides. In fact, manganese oxides are the only naturally-occurring inorganic phases capable of this reaction [27, 28] which Weaver and co-workers [29] found proceeded in multiple stages. However, Eary and Rai [28] reported that Cr(III) oxidation by pyrolusite (*β*-MnO2) was slow in both acidic and slightly acidic solutions. In an acidic solution, slow oxidation was likely the result of the strong sorption of the oxidation product, Cr(VI), to the pyrolusite surface. Such sorption limited contact of unoxidised species with the pyrolusite surface, inhibiting additional oxidation [30]. In slightly acidic to basic media, slow oxidation was the result of the low solubility of Cr(OH)3. Clearly, the equilibrium favoured the trivalent ion. Nevertheless, as with As, a suite of reducing microorganisms exist for reduction of Cr(VI) to Cr(III). Native isolates of *Acinetobacter sp*. from the Sukinda Valley in Jaipur, India were able to reduce initial Cr(VI) concentrations of 5 mg L−1 by 80% in 7 hours [31]. Similar results were reported by Dhal et al. [32] using a *Bacillus sp*. bacterium from chromite mine soils in Boula-Nuasahi mine in Orissa, India. The strain reduced > 90% of 100 mg L−1 Cr(VI) in 144 hours at pH 7 and 35°C.

### **2.3. Acid–base/sorption–desorption**

Variable charge/amphoteric minerals such as crystalline and short-range ordered Fe-, Al-, and Mn- significantly influence the concentrations of elements in groundwater through adsorp‐ tion/ion exchange reactions. This is because of (i) their large surface areas (ii) the acid–base surface hydroxyl groups resulting from the dissociative chemisorption of water molecules on their surfaces [33, 34]. Surface functional groups on mineral oxides undergo protonation (Equation 9) and deprotonation (Equation 10) reactions depending on solution pH [35]. As such, oxide surfaces are positively charged and primed for anion adsorption at lower pH values and negatively charged and primed for cation sorption at higher pH values [35, 36].

$$\text{FeOH} + \text{H}^+ \rightarrow \text{FeOH}\_2^{\cdot -} \tag{9}$$

$$\text{FeOOH} \rightarrow \text{FeO}^{\cdot} + \text{H}^{\cdot} \tag{10}$$

Adsorption of ions from solution (**Figure 3**) is therefore a bid to maintain electric neutrality both on the oxide surface as well as in the solution (ion exchange) [36].

The adsorption of divalent ions to oxide surfaces including goethite, hydrous iron and manganese oxides has been reported in several studies. Borah and Senapati [37] investigated the factors influencing the adsorption of Cd2+ to natural pyrite. They found that metal uptake increased with decreased pyrite particle size and was maximal at 30°C and pH 6. At this pH, Cd2+ ions were the main ions in solution and metal uptake was thus an exchange between the H+ and Cd2+ ions on the pyrite surface. Similar findings were reported by Forbes et al. [38] in the adsorption of Cd2+, Co2+, Pb2+, and Zn2+ on goethite and by Gadde and Laitinen [39] in the adsorption of Pb2+, Cd2+, Zn2+ and Tl+ onto hydrous iron and manganese oxides. Ion exchange may also occur with fixed charge minerals such as zeolites where contaminated water comes into contact with clay minerals [40–42]. Adsorption is not always accompanied by proton loss. In the adsorption of arsenate to goethite, for example, FeAsO4H2 0 and FeAsO4H<sup>−</sup> were the dominant species at pH < 5 and pH 5–8 respectively, in reactions that were not accompanied by the loss of protons from the goethite surface [35].

In fact, manganese oxides are the only naturally-occurring inorganic phases capable of this reaction [27, 28] which Weaver and co-workers [29] found proceeded in multiple stages. However, Eary and Rai [28] reported that Cr(III) oxidation by pyrolusite (*β*-MnO2) was slow in both acidic and slightly acidic solutions. In an acidic solution, slow oxidation was likely the result of the strong sorption of the oxidation product, Cr(VI), to the pyrolusite surface. Such sorption limited contact of unoxidised species with the pyrolusite surface, inhibiting additional oxidation [30]. In slightly acidic to basic media, slow oxidation was the result of the low solubility of Cr(OH)3. Clearly, the equilibrium favoured the trivalent ion. Nevertheless, as with As, a suite of reducing microorganisms exist for reduction of Cr(VI) to Cr(III). Native isolates of *Acinetobacter sp*. from the Sukinda Valley in Jaipur, India were able to reduce initial Cr(VI) concentrations of 5 mg L−1 by 80% in 7 hours [31]. Similar results were reported by Dhal et al. [32] using a *Bacillus sp*. bacterium from chromite mine soils in Boula-Nuasahi mine in Orissa,

India. The strain reduced > 90% of 100 mg L−1 Cr(VI) in 144 hours at pH 7 and 35°C.

FeOH H FeOH2

both on the oxide surface as well as in the solution (ion exchange) [36].

In the adsorption of arsenate to goethite, for example, FeAsO4H2

Adsorption of ions from solution (**Figure 3**) is therefore a bid to maintain electric neutrality

The adsorption of divalent ions to oxide surfaces including goethite, hydrous iron and manganese oxides has been reported in several studies. Borah and Senapati [37] investigated the factors influencing the adsorption of Cd2+ to natural pyrite. They found that metal uptake increased with decreased pyrite particle size and was maximal at 30°C and pH 6. At this pH, Cd2+ ions were the main ions in solution and metal uptake was thus an exchange between the

 and Cd2+ ions on the pyrite surface. Similar findings were reported by Forbes et al. [38] in the adsorption of Cd2+, Co2+, Pb2+, and Zn2+ on goethite and by Gadde and Laitinen [39] in the adsorption of Pb2+, Cd2+, Zn2+ and Tl+ onto hydrous iron and manganese oxides. Ion exchange may also occur with fixed charge minerals such as zeolites where contaminated water comes into contact with clay minerals [40–42]. Adsorption is not always accompanied by proton loss.

+ + + ® (9)

0

and FeAsO4H<sup>−</sup>

were the

FeOH FeO H ® +- + (10)

Variable charge/amphoteric minerals such as crystalline and short-range ordered Fe-, Al-, and Mn- significantly influence the concentrations of elements in groundwater through adsorp‐ tion/ion exchange reactions. This is because of (i) their large surface areas (ii) the acid–base surface hydroxyl groups resulting from the dissociative chemisorption of water molecules on their surfaces [33, 34]. Surface functional groups on mineral oxides undergo protonation (Equation 9) and deprotonation (Equation 10) reactions depending on solution pH [35]. As such, oxide surfaces are positively charged and primed for anion adsorption at lower pH values and negatively charged and primed for cation sorption at higher pH values [35, 36].

**2.3. Acid–base/sorption–desorption**

74 Groundwater - Contaminant and Resource Management

H+

**Figure 3.** Acid and base hydroxyl sites on metal oxide surfaces and ion exchange reactions at the oxide-solution inter‐ face: ●=metal ions, ○=oxide ions, a=acid hydroxyl sites and b=base hydroxyl sites [36].

Determining the mode of contaminant binding is essential to predicting their behaviour in groundwater. Contaminants may be sorbed by electrostatic, hydrogen or covalent bonds. Electrostatic bonds are formed between charged hydrated species and oppositely charged mineral surfaces forming weak outer-sphere complexes. Contaminants sorbed in this way are easily desorbed by perturbations in solution parameters e.g. pH, ionic strength. Hydrogen bonds have intermediate strength while covalent bonding results in strong sorption of contaminants. Contaminants sorbed this way are harder to desorb and colloidal transport may play an important role in their transport through groundwater.
