**4. Prevention and remediation of groundwater contamination by AMD**

Conventional methods of AMD treatment involve the hydroxide precipitation using quicklime (CaO), hydrated lime (Ca(OH)2), caustic soda (NaOH) or soda ash (Na2CO3) [9] or sulphide precipitation agents such as NaS, NaHS, BaS, FeS or H2S [61]. This latter approach is superior because metal sulphides are generally less soluble that hydroxide counterparts, and as such, allows more complete precipitation (although the objectionable odour and toxicity of H2S has to be considered).

Iron, in its reduced form, either as ferrous salts or zero valent form, has also been used for the chemical treatment of contaminated water. The primary advantages associated with ZVI use include low cost, simplicity in handling and the formation of strong complexes between contaminants and the iron oxides [62]. Farrell et al. [20] and Melitas et al. [63] applied ZVI for the removal of arsenate ions from solution. Adsorption was resolved as the removal mecha‐ nism because no As(III) was detected in solution. ZVI was also effective for the removal of CrO4 2−, TcO4 − , MoO2 2− and UO2 2+ ions by reductive precipitation [26, 62, 64–67]. The rate of removal followed the order CrO4 2− >TcO4 <sup>−</sup> >UO2 2+ >>MoO2 2− with CrO4 2− concentrations decreasing from 10,000 μg L−1 to < 5 μg L−1 in 1 hour [65]. Cationic contaminants e.g. Ag+ and Hg+ may also be reduced to their zero valent metallic forms [68]. The use of ZVI exploits redox reactions and the strong reducing properties of iron to convert contaminants into less soluble immobile forms [63]. In this process, ZVI is first oxidised by water to Fe(II) which then acts as an electron donor for the reduction of dissolved ions (Equation 11) [20, 69].

**Figure 4.** Schematic representation of colloid facilitated transport in a subsurface water-saturated medium. Contami‐

For colloid-facilitated contaminant transport to be efficient, three criteria must be met: (i) colloids must be generated, (ii) a strong association must be formed between contaminants and the colloids, (iii) the colloid-contaminant composites must be transported through groundwater [50, 51]. Colloids are particles in the 1 nm to 1 μm size range [52]. They may be organic e.g. humic acids and microbes or inorganic e.g. metal oxy(hydr)oxides, carbonates, silicates and phosphates [52–54]. Inorganic colloids are particularly important in miningcontaminated groundwater. In these environments, colloids may be formed biogenically, as a result of ore processing or by precipitation from supersaturated solutions [53, 55]. Webster et al. [54] suggested that colloids formed in AMD were more effective sorbents than pure minerals due to the presence of sulphates and the influence of bacterial activity on their synthesis. Colloids may also be mobilised by perturbations to groundwater properties including pH, ionic strength and flow velocity (e.g. flow through fractures or variable infiltration following rainfall events). pH shifts are especially important in AMD-impacted environments as pH influences the formation of Fe and Al colloids [53, 56]. It also influences colloidal surface charges, the affinity of contaminants for colloid surfaces and the suspension or precipitation

Solution ionic strength also influences colloid mobilisation due to its effect on the electric double layer of ions as put forward in the DVLO (Derjaguin-Landau-Verwey-Overbeek) theory. According to this theory, colloid mobilisation increases with decreasing ionic strength because the electrostatic double layer around colloids expands resulting in greater repulsion between like-charged colloids. Thus, 137Cs by kaolinite through quartz found that transport was substantially increased at low ionic strengths because kaolinite colloids were more mobile and bound more 137Cs. [57]. Increases in ionic strength on the other hand, lead to compression of the double layers, hence a decrease in repulsive forces and colloid aggregation/coalescence [58, 59]. Thus, contaminant transport may be retarded due to colloidal sedimentation. Kimball et al. [56] found that while Fe colloids aided the transport of As, Cd, Cu, Mn, Pb and Zn from mining flows, the colloidal load decreased by half after the first 50 km due to aggregation and sedimentation of colloids in the stream bed. Retardation may also be due to colloidal plugging/

nants (●) are either dissolved in solution, adsorbed to mobile phases (colloids) or to stationary phases [50].

76 Groundwater - Contaminant and Resource Management

of colloids.

blockage of pores [60] within transport matrices.

$$\text{Fe}^{0} + 2\text{H}\_{2}\text{O} \rightarrow \text{Fe}^{2+} + \text{H}\_{2} + 2\text{OH}^{-} \tag{11}$$

With the advent of nanotechnology, nanoscale ZVI (nZVI) use has also been attempted in laboratory and field studies [70, 71]. Li et al. [72] reported that for Cr(IV) degradation, reaction rates for nZVI were at least 25–30 times faster and the sorption capacity was much higher compared with granular iron. In another study, 25% of As(V) was reduced to As(III) after 90 days [73]. Despite this, the environmental and human health risks associated with nanoparti‐ cles have meant that larger scale application of these materials has been approached with caution. Tratnyek and Johnson [74], however, pointed out that the mobility of nanoscale ZVI was less than a few metres under almost all relevant environmental conditions and thus, human exposure was likely to be minimal.

Sulphate reducing bacteria (SRB) have been increasingly investigated for mine drainage remediation since Tuttle et al. [75] documented sulphate reduction in an AMD-contaminated stream. Samuel et al. [31] reported the remediation of Cr(VI) by indigenous isolates of *Bacillus, Acinetobacter* and *Escherichia spp*. from chromite mines in the Sukinda Valley of Orissa, India. SRB utilize organic carbon or hydrogen to reduce sulphates to sulphides (Equation 12) which then facilitate the precipitation of metal sulphides (Equation 13). This reaction also increases the alkalinity and pH of solutions, further promoting metal precipitation [76]. As such, the use of SRB for AMD treatment reduces sulphate concentrations, precipitates metal ions from solution and raises solution pH.

$$2\text{CH}\_2\text{O} + \text{SO}\_4^{2-} + 2\text{H}^\* \rightarrow \text{H}\_2\text{S} + 2\text{CO}\_2 + 2\text{H}\_2\text{O} \tag{12}$$

$$\text{Fe}^{2+} + \text{H}\_2\text{S} \rightarrow \text{FeS} + 2\text{H}^+ \tag{13}$$

Cardenas et al. [77] reported the biological in situ remediation of uranium contaminated groundwater. The growth of denitrifying, Fe(III)-reducing and SRB including *Desulfovibrio, Geobacter, Anaero- myxobacter, Desulfosporosinus, Acidovorax, Ferribacterium* and *Geothrix spp* through weekly injections of ethanol into the subsurface. After 2 years, U concentrations were reduced from 60 mg L−1 to < 30 μg L−1. Sulphate concentrations also decreased when ethanol was injected and rebounded when injection stopped, indicating SRB activity in the subsurface.

Such in situ approaches for the treatment of mine drainage contamination have attracted much attention in recent times. Designed to intercept contaminants in the subsurface with reactive materials, in situ treatment has the advantage of treating contaminated groundwater prior to the oxidation of Fe2+ (Reaction 2), thus preventing the generation of additional acidity and mobilisation of additional metal ions [64]. Sulphate reduction is also optimised in the nearneutral pH characteristic of many aquifers and the process is less costly because the volumes of water to be treated are lower than in pump-and-treat systems [78]. Two possible approaches exist for *in situ* remediation. The first involves injecting reactants into the subsurface to form a reactive treatment zone in which reactants are adsorbed onto aquifer materials (**Figure 5a**). The procedure by Cardenas et al. [77] described above falls in this category. Alternatively, permeable reactive barriers may be keyed in to underlying bedrock in the flow path of contaminants (**Figure 5b**).

A number of studies have documented the use of PRBs for the remediation of AMD-contami‐ nated groundwater [64, 78, 79]. The reactive materials within PRBs may be chemical or biological [80]. Baker et al. [81] used a mixture containing 50 wt% silica sand, 45 wt% crushed limestone and 5 wt% metal oxide for the removal of phosphates while ZVI, FeCO3, FeS were investigated for Cr(VI) removal [82]. Biologically-driven PRBs (biobarriers) mostly comprise of SRB [61, 83] and require (i) an anaerobic environment (redox potential of ~200 mV), (ii) pH values greater than 5, (iii) a sulphate species to be reduced and (iv) an energy source (electron donor), mostly short chain organic substrates e.g. ethanol although a variety of natural substrates including leaf mulch, vegetal compost and sawdust [84] have been tested. Benner et al. [85] installed a PRB 20 m long, 4 m thick and 3.5 m tall into the Nickel Rim aquifer downstream of a tailings impoundment. They recorded, after 22 months, dramatic changes in the concentrations of several contaminants. Concentrations of sulphates, Fe and Ni decreased by 2000–3000 mg L−1, 270–1300 mg L−1 and 30 mg L−1, respectively. In addition, alkalinity increased by 800–2700 mg L−1 and the populations of SRB were 10, 000 times greater than before the installation of the PRB. Column experiments by Waybrant et al. [86] showed similar results: iron concentrations decreased from 300–1200 mg/L to <0.01–220 mg/L while Zn and Ni decreased from 0.6–1.2 mg L−1 to 0.01–0.15 mg L−1 and 0.8–12.8 mg L−1 to <0.01 mg L−1, respec‐ tively. The pH increased slightly from 5.5–6.0 to 6.5–7.0 and alkalinity from <50 mg/L to 300– 1300 mg L−1. Biobarriers are also effective for the attenuation of Cr(VI) [87] radionuclide ions [88] and sulphates [89]. Sulphate concentrations were decreased from 1800 to < 250 mg L−1 and the mine waters neutralised using only bacterially-mediated alkalinity. Natural treatment as

of SRB for AMD treatment reduces sulphate concentrations, precipitates metal ions from

Cardenas et al. [77] reported the biological in situ remediation of uranium contaminated groundwater. The growth of denitrifying, Fe(III)-reducing and SRB including *Desulfovibrio, Geobacter, Anaero- myxobacter, Desulfosporosinus, Acidovorax, Ferribacterium* and *Geothrix spp* through weekly injections of ethanol into the subsurface. After 2 years, U concentrations were reduced from 60 mg L−1 to < 30 μg L−1. Sulphate concentrations also decreased when ethanol was injected and rebounded when injection stopped, indicating SRB activity in the subsurface. Such in situ approaches for the treatment of mine drainage contamination have attracted much attention in recent times. Designed to intercept contaminants in the subsurface with reactive materials, in situ treatment has the advantage of treating contaminated groundwater prior to the oxidation of Fe2+ (Reaction 2), thus preventing the generation of additional acidity and mobilisation of additional metal ions [64]. Sulphate reduction is also optimised in the nearneutral pH characteristic of many aquifers and the process is less costly because the volumes of water to be treated are lower than in pump-and-treat systems [78]. Two possible approaches exist for *in situ* remediation. The first involves injecting reactants into the subsurface to form a reactive treatment zone in which reactants are adsorbed onto aquifer materials (**Figure 5a**). The procedure by Cardenas et al. [77] described above falls in this category. Alternatively, permeable reactive barriers may be keyed in to underlying bedrock in the flow path of

A number of studies have documented the use of PRBs for the remediation of AMD-contami‐ nated groundwater [64, 78, 79]. The reactive materials within PRBs may be chemical or biological [80]. Baker et al. [81] used a mixture containing 50 wt% silica sand, 45 wt% crushed limestone and 5 wt% metal oxide for the removal of phosphates while ZVI, FeCO3, FeS were investigated for Cr(VI) removal [82]. Biologically-driven PRBs (biobarriers) mostly comprise of SRB [61, 83] and require (i) an anaerobic environment (redox potential of ~200 mV), (ii) pH values greater than 5, (iii) a sulphate species to be reduced and (iv) an energy source (electron donor), mostly short chain organic substrates e.g. ethanol although a variety of natural substrates including leaf mulch, vegetal compost and sawdust [84] have been tested. Benner et al. [85] installed a PRB 20 m long, 4 m thick and 3.5 m tall into the Nickel Rim aquifer downstream of a tailings impoundment. They recorded, after 22 months, dramatic changes in the concentrations of several contaminants. Concentrations of sulphates, Fe and Ni decreased by 2000–3000 mg L−1, 270–1300 mg L−1 and 30 mg L−1, respectively. In addition, alkalinity increased by 800–2700 mg L−1 and the populations of SRB were 10, 000 times greater than before the installation of the PRB. Column experiments by Waybrant et al. [86] showed similar results: iron concentrations decreased from 300–1200 mg/L to <0.01–220 mg/L while Zn and Ni

2 4 2 22 2CH O SO 2H H S 2CO 2H O - + + +® + + (12)

<sup>2</sup> Fe H S FeS 2H + + +®+ (13)

2

2

solution and raises solution pH.

78 Groundwater - Contaminant and Resource Management

contaminants (**Figure 5b**).

**Figure 5.** In situ treatment of groundwater may be by a reactive treatment zone (a) or a permeable reactive barrier (b) (After [74]).

well as permeable reactive barriers therefore both hold promise for the treatment of ground‐ water contamination. Future research should look into the incorporation of nanomaterials e.g. embedded in polymers, into PRBs to facilitate faster reaction times and more efficient removal of contaminants.
