Recovery of Metals

A metal is basically any chemical element that is considered as a good conductor of both electricity and heat (Geiger and Cooper, 2010). From the chemistry point of view, metals readily lose electrons to form cations and

FIGURE 9.1

Periodic table showing metal, metalloid, and non-metal elements.

ionic bonds with non-metals. Metals comprise the greater part of the periodic table of elements, while non-metallic elements are only found on the right-hand side of the periodic table (see Figure 9.1). A diagonal line, drawn from boron (B) to astatine (At), separates the metals from the non-metals (Geiger and Cooper, 2010). Most elements on the diagonal line are metalloids, which are called semiconductors sometimes because they display electrical characteristics that are common to both conductors and insulators (Geiger and Cooper, 2010). Amongst the 88 elements present in the periodic table, about 30 elements may be considered as toxic heavy metals, and it is generally accepted that all the heavy metals have a specific gravity (density) greater than 5.0 g/cm3 (Parker, 1989; Lozet and Mathieu, 1991; Morris, 1992; Duffus, 2002).

Generally, metals and their alloys are divided into two classes - ferrous and non-ferrous (Hall and Giglio, 2010). Ferrous metals consist of iron, steel, and alloys related to them that are magnetic in nature (Dunaway, 1984). Non- ferrous metals are those that contain either none, or very small amounts of ferrous metals, and they are generally divided into the aluminium, copper, magnesium, lead, and similar groups (Dunaway, 1984).

It is indisputable that metals are an integral part of human life and world development (Hall and Giglio, 2010). According to Norgate and Rankin (2002), metals are actually well suited for sustainable development goals as they have unlimited life span and the potential for unlimited recyclability. However, metal demand and usage around the globe are rising coupled with the depletion of high-grade reserves (Sethurajan, 2015). Norgate and Rankin (2002) also argue that the source of primary metals is finite. As a result, it has become increasingly important to consider other sources of metals. For example, AMD which is considered as an environmental contaminant is now viewed as a potential source of valuable metals (Naidu et al., 2019). Moreover, according to Garcia et al. (2013), AMD contains several metals and metalloids of particular interest such as copper, iron, manganese, aluminium, and zinc. Kwon et al. (2016) also argue that though most of the current remediation technologies consider AMD as a waste material, in actual fact, AMD contains several species of metals including valuable metals. In addition, the increases in prices of many metal commodities have revived questions as to whether metals can be economically recovered from AMD (Smith et al., 2013). Nordstrom et al. (2017) acknowledge that metal recovery from mine waters and effluents is not a new approach, but one that has occurred largely opportunistically over several millennia. The focus of this section of the chapter is to discuss and evaluate different techniques that facilitate the recovery of metals from AMD. The techniques are divided into four categories: settling and sedimentation, selective precipitation, selective adsorption, ion-exchange, and electrochemical treatment.

  • 9.2.1 Settling and Sedimentation
  • 9.2.1.1 Concepts of Settling and Sedimentation

Many approaches can be used for the treatment of mine waters, with the treatment solution dependent on the mine water contamination. Particle settling and sedimentation represent the simplest processes employed for treating mine impacted water (McCauley, 2011). The two processes have been credited for efficiently removing heavy metals associated with particulate matter in AMD particularly in natural and constructed wetlands (Hammer, 1997; Sheoran and Sheoran, 2006). However, despite the effectiveness of settling and sedimentation, the processes rely on a range of other processes like precipitation, co-precipitation, and adsorption that have to occur first in order to aggregate heavy metals into particles large enough to sink (Walker and Hurl, 2002; Sheoran and Sheoran, 2006). For example, chemicals may be added before the AMD flows into sedimentation ponds so that metals in the AMD can settle out or precipitate (Skousen, 2014). In fact, the processes can be accomplished by either active or passive techniques (Skousen et al., 2000; Waters et al., 2003), and see Chapter 7 for details of the two techniques. In view of the two techniques, Trumm (2010) has developed a selection criterion for the two techniques. However, in both passive and active treatment systems, neutralisation (or pH control) is the most commonly used approach (Taylor et al., 2005). By increasing the pH to create alkaline conditions (e.g., pH > 9.5), the solubility of most metals can be significantly decreased by precipitation (Taylor et al., 2005). Ideally, dissolved metals will precipitate from AMD as loose, open-structured mass of tiny grains called floes or sludge (Skousen, 2014).

This technique, conventional precipitation, has many disadvantages including being expensive, labour intensive, and ineffective when metal concentrations are below ppm levels (Brown, 1996). This conventional method of treating AMD by chemical neutralisation using lime or limestone also results in huge amount of sludge (Kefeni et al., 2015). In addition, it requires being pumped, dewatered, trucked, and processed or even buried, and subsequently monitored for any leachates (Brown, 1996). In view of these disadvantages, Simate and Ndlovu (2014) argue that a suitable method should be based on recovery and reuse of the heavy metals. In other words, to use AMD as a resource, the separation of metals as well as their removal from AMD is required (Bessho et al., 2017). Kefeni et al. (2015) also argue that in order to achieve sound environmental protection and sustainable remedy, it is imperative that the means of recovery of the valuable minerals and reuse of the resources from AMD should be developed. Therefore, there is no doubt that it is desirable to have an AMD remediation technique that would also lead to recycling of major and/or minor metal elements. Such methods are discussed in Sections 9.2.2-9.2.5.

9.2.7.2 Typical Studies of Settling and Sedimentation

Several chemicals are used in AMD settling and sedimentation processes. According to Skousen et al. (2000), each chemical has characteristics that make it more or less appropriate for a specific condition. In addition, Skousen et al. (2000) state that the best choice among alternatives depends on both technical and economic factors. The technical factors include acidity levels, flow, and the types and concentrations of metals in the AMD. The economic factors include prices of reagents, labour, machinery and equipment, the number of years that treatment will be needed, and the interest rate.

INAP (2009) and ITRC (2010a) state that raising the pH of AMD solution using alkaline agents causes certain dissolved metals (e.g., cadmium, copper, iron, lead, manganese, and zinc) to precipitate as hydroxides. A coagulant and/or flocculant may also be added to enhance flocculation, and the solution may be transferred to a clarifier in order to settle the solids and thus separate them from the cleaned overflow effluent. The resultant metal-hydroxide sludge extracted from the bottom of the clarifier usually contains a large percentage of bound water, limiting the potential for reuse, and is disposed of as a solid waste. The amount of sludge generated can be reduced by employing a high-density sludge (HDS) treatment technique. In HDS processes, the precipitated hydroxide sludge is recycled to a conditioning tank, where it is mixed with the alkali reagent. The sludge/alkali slurry is then added into the AMD to raise the pH and cause additional metal precipitation. The reconditioning of the sludge provides for precipitation sites for the dissolved metals to bond, thus increasing the overall density of the sludge in the clarifier underflow.

A two-step neutralisation process, leading to the formation of ferrites, was developed by Herrera et al. (2007) for the treatment of AMD. The process applied in the study used MgO and NaOH as the first and second neutralisers, respectively. In the first neutralisation step, MgO was used to raise the pH of AMD to around 4.5 so as to eliminate aluminium and to reduce the silica concentration. In the second neutralisation step NaOH was used to complete the neutralisation and to co-precipitate ferrous and ferric hydroxides, from which ferrite could be formed. The results of the study showed that the two-step neutralisation process is a very promising option in AMD remediation because it can reduce a sludge volume in the order of 20-80% compared to the conventional lime neutralisation process. Additionally, an industrial use of the ferrite sludge produced would reduce the amount of sludge to be disposed of by about 90%. The results of batch and continuous flow tests showed that MgO performed better than NaOH because of its higher neutralizing capacity per unit weight and because the use of MgO avoids the massive precipitation of Ca species in the second neutralisation step. Large amounts of Ca species would hinder the ferritisation of the precipitates.

A study by Kaur et al. (2018) compared the performance of Bayer liquor and Bayer precipitates with commercially available alkali commonly used in the treatment of AMD water in view of material requirements and discharge water quality. The research questions addressed were: (1) can the waste alkali materials raise the pH to the required levels to meet water discharge limits; (2) is it possible to reduce dissolved metal concentrations to satisfy regulations; and (3) what is scientific explanation for differences in performance for the various alkalis. In a study by Kaur et al. (2018), the treatment of mine pit water involved the addition of known amounts of lime, Bayer hydrotalcite, Bayer liquor, sodium carbonate, and sodium hydroxide to 25 mL of AMD water at ambient temperature. The resultant mixture was then agitated for 24 h before being centrifuged. With respect to the research questions, firstly, all investigated alkaline materials successfully raised the pH of mine pit water so as to meet discharge limits, i.e. pH 6.5-8.5. Secondly, the study found that lime and Bayer precipitates were more effective in removing the metals present in mine pit water than either sodium hydroxide, sodium carbonate, or Bayer liquor. Thirdly, the ability of the precipitates to encapsulate heavy metals was determined to be more important than surface area. Mechanistically, larger precipitates were found to positively influence the removal of heavy metals, with lime and Bayer precipitates forming the largest precipitates. Sludge produced after treatment with Bayer precipitates was more stable and showed minimum metal leaching as compared to sludge produced after treatment with other alkali. The mass of material required for attaining the desired pH was higher for Bayer precipitates compared to lime, but the capital cost for a system using lime was considered high due to its hydrophobic nature and the resultant extensive mixing required.

Olds et al. (2013) evaluated the effect of surface area of precipitates formed by neutralisation of AMD using three alkalinity reagents (NaOH, Ca(OH)2, and CaC03) on the sorption of Ni and Zn, and their subsequent removal. Jar stirrers were used for neutralisation of the AMD in 1-L beakers, with different mixing regimes for different alkaline reagents. The NaOH neutralised samples were dosed with NaOH followed by a 1-min rapid mixing phase (100 rpm) and a 25-min flocculation phase (20 rpm), as neutralisation by NaOH was instantaneous. By contrast, the Ca(OH)2 and CaC03 neutralised samples were dosed with slurry of reagents followed by a 60-min rapid mixing phase (100 rpm). At the end of the flocculation/mixing period, the beakers of neutralised AMD were allowed to settle for 2 h. The results of the study indicated that the removal of Ni and Zn by sorption onto AMD precipitates is influenced by the surface area of the floe formed during neutralisation. Neutralisation of AMD by NaOH and Ca(OH)2 produces large fluffy floe, with surface areas, an order of magnitude greater than floe formed by CaC03 neutralisation. As a result, significantly more Ni and Zn were sorbed and co-precipitated on the NaOH and Ca(OH)2 floe.

Bologo et al. (2009) proposed a process whereby Mg(OH)2 was used for the neutralisation of free acid and subsequent raising of the mine wastewater pH to above 7 to facilitate rapid iron (II) oxidation and precipitation as Fe(OH)3. In other words, the investigation aimed to demonstrate that Mg(OH)2 treatment, in combination with lime treatment, offers an attractive solution for the treatment of acid mine water that is rich in Fe(II) and other metals. In this approach, Fe(II) with a concentration of 900 mg/L was completely oxidised within 10-min reaction time and precipitated as Fe(OH)3 together with other metal hydroxides. The precipitated Fe(OH)3 together with other metals hydroxides was separated from the water. In the next stage, magnesium was precipitated with lime as Mg(OH)2 and was separated from the water together with gypsum. Bologo et al. (2009) suggested that Mg(OH)2 can be separated from the gypsum by treating it with CO, to form Mg(HC03)2 or with H2S04 to form MgS04. This study demonstrated that the integrated Mg(OH)2 and lime process can be applied for treating acid mine water effectively. In addition, by using Mg(OH)2 instead of Ca(OH)2 or CaC03, gypsum precipitation can be avoided and metal hydroxides can be precipitated separately from gypsum.

9.2.1.3 Recovery of Metals from the Sedimentation Sludge

The metals within the sludge that consist of particulate matter and other impurities can be economically extracted using a number of technologies (Sethurajan, 2015). For example, different pyrometallurgical and hydromet- allurgical processes have been developed for the extraction of metals from metallurgical sediments and/or sludge (Sethurajan, 2015). Pyrometallurgy employs high temperatures to carry out smelting and refining operations to extract metals, whereas hydrometallurgy uses aqueous solutions to separate the desired metals (Nassaralla, 2001). Hydrometallurgy has a number of advantages including low initial capital investment, low energy requirements, and high-purity grade of the metal produced compared to pyrometallurgy (Nassaralla, 2001); and hence it is mainly adopted.

Sethurajan (2015) proposed different approaches for the extraction of heavy metals from the metallurgical sludges and/or sediments. These processes include (1) thermal treatment coupled with high concentrated acid leaching, (2) acid leaching, (3) combination of pyrometallurgical (roasting) and hydrometallurgical processes (sulphuric acid, water and NaCl), (4) hydro- metallurgical process including leaching, cementation and refining, and (4) bioleaching with iron oxidizing bacteria (acidithiobacillus ferrooxidans, aci- dithiobacillus caldus), and archaea (sulpholobus metallicus), and many other techniques. After the metals have been extracted from the sludge, other processes discussed in Sections 9.2.3-9.2.5 may be used to recover them.

9.2.1.4 Summary

There are many techniques that can be used for the treatment of AMD, with the treatment option dependent on the degree of contamination. The simplest of all the techniques is generally termed settling and sedimentation (or conventional precipitation method). The traditional solution to treat AMD and the first of the settling and sedimentation methodology involves collecting and chemically treating acidified effluents in a centralised treatment plant. Alternatively, the second method involves routing effluents through natural or constructed wetlands within which microbial communities perform the same function. Both methods use pH control (or neutralisation) as a means of precipitating the metals in the AMD, which results in the formation of sludge. Thereafter, the metals can be economically extracted from the sludge using a number of pyrometallurgical and hydrometallurgical technologies.

  • 9.2.2 Selective Precipitation
  • 9.2.2.1 Concepts of Selective Precipitation

Selective precipitation is based on solubility differences among the metal compounds (Rodriguez-Galan et al., 2019 ). A lot of studies have shown that it is one of the most promising ways to overcome the problems of the conventional precipitation technique discussed in Section 9.2.1 (Wei et al., 2005; Simate and Ndlovu, 2014; Oh et al., 2016; Rodriguez-Galan et al., 2019). The main advantages of this method are the reduced volume of the produced sludge and the valorisation of metals (Oh et al., 2016; Rodriguez- Galan et al., 2019).

The most common reactive agents for metal precipitation are hydroxides or sulphides (Rodriguez-Galan et al., 2019). Amongst the two chemicals, sulphides have various advantages compared to using hydroxides (Lewis, 2010; Oh et al., 2016). Firstly, there is a high possibility of metal separation using sulphides because of the more distinct solubility of various metal sulphides (Sampaio et al., 2009; Oh et al., 2016). Secondly, metal sulphides also have other advantages including faster reaction rates, low solubility over a wide range of pHs, better settling properties, and higher potential for reuse by smelting (Gharabaghi et al., 2012; Oh et al., 2016; Uc;a, 2017). Furthermore, the use of sulphide not only allows producing effluents with metal concentrations in the order of magnitude of ppm and ppb, but also gives the possibility of precipitation at low pH and selective precipitation for metal reuse (Sampaio et al., 2009). Sulphide precipitation can be effected using either solid (FeS, CaS), aqueous (Na2S, NaHS, NH4S), or gaseous (H2S) sulphide sources (Lewis, 2010; Patil et al., 2016). There is also the possibility of using the degeneration reaction of sodium thiosulphate (Na2S203) as a source of sulphide for metal precipitation (Lewis, 2010).

9.2.2.2 Typical Studies of Selective Precipitation

This section discusses metal recovery from AMD and/or industrial wastewater using a selective precipitation process based on solubility characteristics of the major and minor metals in the wastewater. The examples discussed involve the use of various neutralisation reagents.

The experimental study performed by Wei et al. (2005) involved the treatment of AMD from a bond-forfeited coal mine site (Upper Freeport seam) that is located in north central West Virginia. The objective of the study was to recover iron and aluminium from AMD by selective precipitation. The untreated AMD water samples were collected and stored in closed high- density polyethylene bottles and kept at 4°C. Thereafter, the AMD water was removed from the fridge and bubbled with compressed air for at least 24 h to ensure the complete oxidation of Fe2+. The water was then filtered to remove debris and suspended solids. Thereafter, it was referred to as "raw" AMD and was used as a feed solution for metal solubility and recovery experiments. A two-step process was used for metal recovery tests: iron precipitation followed by aluminium precipitation. For iron precipitation, raw AMD water samples of 500 mL were neutralised with 10-N caustic soda (NaOH) solution to pH end points between 3.0 and 4.5, at 0.5 standard unit intervals to assess iron recovery at different pHs. After iron recovery (at pH 3.5), the filtered AMD water was used as a feed solution for aluminium precipitation. Samples of 500 mL each were then neutralised with 10-N NaOH to pH end points from 4.5 to 8.0 to determine the aluminium recovery performance over a range of pHs. In addition to NaOH, similar pH adjustment tests were conducted using soda ash (Na2C03), ammonia (NH4OH), quick lime (CaO), and hydrated lime (Ca(OH)2) for iron recovery at pH of 3.5 and aluminium recovery at pH of 6.5 so as to assess the performances of different neutralisation reagents. The NH4OH solution of 25-30%, as acquired, and 1-M Na2C03 solution were added for pH adjustment. The CaO and Ca(OH)2 were applied as fine powders.

The results of the study by Wei et al. (2005) showed that separate iron and aluminium hydroxide products with relatively high purity were successfully recovered via iron precipitation at pH 3.5-4.0 followed by aluminium precipitation at pH 6.0-7.0, while simultaneously meeting the National Pollutant Discharge Elimination System (NPDES) effluent discharge standards by the United States. In this study, iron precipitate recovery of >98.6% with a purity of >93.4% was achieved, while aluminium precipitate purity reached >92.1% at a recovery of >97.2%. In addition, the study found that during each metal recovery operation, other metals remained in solution, which ensured the relatively high purity of precipitate products. All of the five neutralisation reagents used in the study that are commonly used in AMD treatment were found to be suitable for iron and aluminium recovery. However, Wei et al. (2005) recommended the application of ammonia and caustic soda for metal recovery in the full-scale processes because unreacted hydrated lime might pose a threat to the purity of precipitate products if lime or hydrated lime was added. It was also suggested that oxidants should be added in the full-scale systems in order to enhance the oxidation of ferrous iron.

The aim of a study by Luptakova et al. (2010) was to precipitate heavy metals from AMD using bacterially produced hydrogen sulphide combined with intermediate steps of metals precipitation by sodium hydroxide at various pH values as shown in Figure 9.2. The experiments were conducted with AMD obtained from the abandoned and flooded deposit of Smolnik (Slovak Republic).

Briefly, the study comprised several process steps, which could be grouped into three main stages: (1) biological hydrogen sulphide production using sulphates reducing bacteria, (2) selective heavy metals precipitation by the bacterially produced hydrogen sulphide, and (3) three intermediate steps of metal hydroxide precipitation by 1-M NaOH at various pH values. The solids produced in each step were separated from the remaining solution by filtration. The results of the study indicated that the process by Luptakova et al. (2010) is able to sequentially precipitate Cu2+, Zn2+, and Fe3+ in the form of sulphides, Al3+, Fe2~, and Mn2+ in the form of hydroxides. Table 9.1 shows the results obtained by Macingova and Luptakova (2012) in a similar study of selective sequential precipitation process. As can be seen from the table, selective recovery of various metal precipitates was achieved.

Mulopo (2015) studied the sulphidation behaviour of Fe(II) using CaS derived from waste gypsum as a sulphidation agent, together with the possibility of selective precipitation of Pb, Zn, Ni, Co, and Fe(II) from various AMD solutions. The study is an appropriate case illustrating a simple strategy for integrated recycling of two mining waste streams (AMD and gypsum) and highlighted the need for the mining industry to break away from the traditional “linear" cul-de-sac disposal of wastes and think of new sustainable ways of waste management. A tubular muffle furnace consisting of a 750 mm long, 24 mm diameter mullite tube mounted horizontally and equipped with a temperature controller was used for the thermal reduction of waste gypsum to calcium sulphide. Basically, CaS was produced by carbothermal reduction of waste gypsum at a temperature of 1025-1030°C

FIGURE 9.2

Schematic illustration of the six-step precipitation process for recovery of metals from acid mine drainage. (From Luptakova et alv 2010.)

TABLE 9.1

Conditions and Results of Selective Sequential Precipitation Process

Step

1

2

3

4

5

6

pH

2.8

3.7

3.7

5.0

5.0

9.5

Reagent

H^O,

NaOH

H2S

NaOH

H,S

NaOH

Removed metals

Fe

Fe

Cu

Al, Zn

Zn

Mn

Proportion (%)

99.99

99.99

99.99

98.94:1.06

99.99

99.99

Source: Macingova and Luptakova, 2012.

for 45 min using a C/CaS04 molar ratio of 2. The CaS yield obtained was about 78%. In this study, the effect of sulphide addition to the AMD system was investigated using sulphide/total metal mole ratios of 0, 0.5, 0.75, 1.0, 1.5, 2.0, and 2.5. Appropriate amounts of CaS were added to the AMD to give a total of 1-L mixture and batch experiments were carried out in plastic beakers equipped with overhead stirrers fitted with radial turbine impellers. The experiments were run at appropriate pH values using a pH cascading approach. The metal removal in the batch experiments were carried out for at least 5 min at a particular pH or until a steady pH was attained. The results showed that sulphidation was dependent on the pH, sulphide dosage, and metal concentration. The selective sulphidation of metals also showed significant dependence on the respective metal sulphide solubility order as a function of pH. It was found that Pb, Zn, Ni, and Co could be removed as metal sulphides at lower pH values while Fe(II) remained in solution, thus enabling ferrous iron to be separated from the other metals, which is a great advantage for metal recovery. The results of the study clearly demonstrated that selective metal removal and recovery as metal sulphides may be achieved conveniently using CaS as the sulphidation medium. However, the purity of CaS obtained by the thermal reduction of waste gypsum and mass transfer limitations associated with the AMD-CaS system was found to be critical for the process development. In addition, the settling characteristics of the precipitates were poor, but this could probably be improved by the use of an anionic polymer at low pH.

Uc;a (2017) successfully precipitated metals separately in a pH-controlled system by using sulphide produced in an ethanol fed anaerobic baffled reactor. For simplicity, equation 9.1 shows the reaction of the sulphate reduction in the reactor using an organic source such as formaldehyde in the presence of sulphate reducing bacteria. Basically, sulphidogenic bacteria generate hydrogen sulphide primarily by using either sulphate or elemental sulphur as an electron acceptor, and an organic (e.g., ethanol) or inorganic (e.g., hydrogen) electron donor (Nancucheo and Johnson, 2012).

Simulated AMD (pH 2.5) contained approximately 120 ± 2.4 mg/L Cu2+ and 124 ± 3.1 mg/L Fe2+. In this study, the sulphide and alkalinity in the anaerobic baffled reactor effluent was used for the selective metal recovery experiments. Nitrogen gas was bubbled through the anaerobic baffled reactor effluent in order to transport the sulphide gas from the reactor to the metal mixture bottle. This method allowed the transportation of H2S only and leaving the alkalinity in the sulphide bottle. As a result, the pH in the bottle containing a mixture of metals was not increased, thus iron was not precipitated. Since the pH was low, only copper ions reacted with the hydrogen sulphide gas according to the following reaction:

Over 99% of the Cu2+ was removed in first 2 min. A complete removal of 120 ± 2.4 mg/L Cu2+ was achieved in 20 min. After filtration of the CuS precipitate, the metal mixture bottle was mixed with anaerobic baffled reactor effluent, which had high concentrations of sulphide and HC03~. As a result, Fe2+ was precipitated as FeS at elevated pHs. The results showed that the supernatant Fe2+ concentration 5 min after mixing was only 0.3 ± 0.02 mg/L, which implies that most of the Fe2+ ion had precipitated.

In a study by Sampaio et al. (2009), Cu was continuously and selectively precipitated from Zn using Na2S. Selective precipitation was based on the control of pS (= -log [S2 ]) and pH. Here, having the solubility product defined as KSP = (Me2+)(S2-), it means that different sulphide concentrations (S2_ potentials) are required to precipitate different metals. Therefore, the addition of sulphide to selectively precipitate heavy metals can be controlled using an ion selective electrode for sulphide (S2), a so-called pS electrode (Sampaio et al., 2009). In this study, selective precipitation of copper from zinc was achieved at pS and pH of 25 and 3, respectively.

Another example of a technique that uses biologically generated precipitating hydrogen sulphide gas was described by Huisman et al. (2006). In this process (Thioteq technology), just like other similar processes that biologically generate hydrogen sulphide, gaseous or dissolved H2S is produced on-site and on-demand in an engineered high rate bioreactor. According to Huisman et al. (2006), the Thioteq process consists of two stages: a biological and a chemical. The water to be treated only passes through the chemical stage. Sulphide is produced in the biological stage and transported to the chemical (precipitation) stage with a carrier gas. The properties of the sulphide gas from the bioreactor and the contactor design result in metal sulphides with good settleability and filterability. The chemical stage consists of a gas-liquid contactor. As already stated, the sulphide is transported to the contactor with the help of a carrier gas (e.g., a mixture of CO, and N2) and the metal-loaded water (e.g., AMD) is fed to the contactor. Metals like copper precipitate as sulphides according to reaction 9.2. In the Thioteq process, copper can be precipitated as a sulphide usually without pH adjustment and without significant precipitation of other heavy metals present in the water. The result is a product with a high copper sulphide content usually greater than 90%. Other metals such as zinc and nickel can be recovered as separate high-grade sulphide products when the number of precipitation stages are increased. In addition, a pH control using an alkali source might be required to meet the optimum precipitation conditions. The precipitated metal concentrates are recovered in a clarifier and then dewatered using a filter press. These metal sulphides can be transported to smelters as high quality concentrates.

Several precipitating agents have also been developed in the last few years for chemically and selectively precipitating divalent and univalent heavy metals from water and effluents (Blais et al., 2008). Selective metal precipitating agents include dithiocarbamate (Matlock et al., 2002a), Thio-Red (Matlock et al., 2002a), and dipropyl dithiophosphate (Ying and Fang, 2006). The principal advantages of using metal precipitating agents are (Blais et al., 2008):

  • (1) the formation of metal compounds which have a very low solubility, and
  • (2) the lesser production of metallic residue in comparison to the production of metallic sludge using common chemicals, like sodium hydroxyde or lime. However, the high cost of the metal precipitating reagents inhibits their use for different industrial applications (Meunier et al. 2002; Blais et al., 2008).

A good example of an alternative metal precipitating reagent is 1,3- benzenediamidoethanethiol dianion (BDET, known commercially as MetX). Research has shown that BDET can reduce the concentrations of a wide variety of divalent metals in water and sediments to below wastewater discharge limits. In fact the ligand has been found to selectively and irreversibly bind soft heavy metals from aqueous solutions (Matlock et al., 2002b). Furthermore, it has been demonstrated that the metal-BDET precipitates are insoluble in aqueous solution and in common organic solvents and are stable over pH ranges of 0.0-14.0 (Matlock et al., 2001). In order to explore the utilisation of BDET for iron and metal binding under AMD conditions, an abandoned coal mine was selected for study in Pikeville, Kentucky by Matlock et al. (2002b). The study by Matlock et al. (2002b) involved treating water within the coal mine as well as water being discharged from the mine using BDET. As a result of a variety of metals present in the AMD waters, multiple BDET-metal compounds were expected to be produced, and thus each of the metal-BDET precipitates was identified using NMR, IR, and XRD. The study found that BDET-Fe precipitates were predominant. Therefore, in addition to analysing the multiple BDET-metal precipitates, pure samples of BDET-Fe were prepared and analysed for stability using NMR, IR, Raman, XRD, and elemental analyses. The study found that BDET was able to remove > 90% of several toxic or problematic metals from AMD samples. For example, the concentrations of metals such as iron were reduced at pH of 4.5 from 194 ppm to below 0.009 ppm. In the leaching experiments conducted for stability tests at pH 0.0, 4.0, and 6.5, maximum leaching was seen for pH 0.0 solutions on day 7 (16.7 mg of Fe leached from 1000 mg BDET-Fe). During the 30-day leaching period, no additional leaching was seen after 7 days for samples tested at pH 0.0, 4.0, and 6.5.

Another developed method for heavy metal removal based on chelating precipitants is termed CH collector method. This is simply a solid material which binds heavy metals to its surface (Turhanen and Vepsalainen,

  • 2013). Typically, such chelating precipitants contain groups with replaceable hydrogen atoms such as carboxyl (-COOH), hydroxyl (-OH), mer- capto (-SH), or sulphonic (-S03H) groups, together with functional groups of basic character, such as amino (-NH2), amino (cyclic)(-NH-), carbonyl or thio keto with which the reacting metal is coordinated to form a four- five- or six-member ring. This invention is unique in that ion channels are formed inside the material in which metal ions are collected from the solution (Turhanen and Vepsalainen, 2013). The collection of the heavy metal ions does not require a separate precipitation step or any adjustments to the solution's pH. Unlike traditional methods, the CH collector method also allows the recovery of metals to occur from very small concentrations. The new method enabled the complete removal of uranium from the water collected from a Finnish mine (Turhanen and Vepsalainen, 2013). The study showed that there was no need to pre-treat the water even though it contained high concentrations of other metals. The efficiency of the method was also tested on scandium and it was removed from wastewater with 98% recovery (Turhanen and Vepsalainen, 2013).
  • 9.2.2.3 Summary

The large volumes of sludge that are produced through the active treatment of AMD place a huge burden on the mining industry because the sludge requires further processing and/or final disposal. Moreover, the AMD sludge contains a mixture of various metal oxides/hydroxides that are of little to no practical value. However, the recovery of individual metals has potential commercial value. This section of the chapter has shown that based on the solubility of the metals dissolved in AMD, selective precipitation process can be developed to recover high-purity metals. A variety of precipitation reagents are available. However, the most common reactive agents for metal precipitation are hydroxides and sulphides; and amongst the two chemicals, sulphides have various advantages compared to using hydroxides.

  • 9.2.3 Selective Adsorption
  • 9.2.3.1 Concepts of Selective Adsorption

In the past number of years, adsorption has been one of the most widely used techniques amongst the different methods developed for removal of toxic metals from polluted natural water or industrial wastewater mainly due to its low cost and environmental friendliness (Larsson et al., 2018). Patil et al.

(2016) define adsorption as the accumulation of one substance on the surface of another; and that the mechanism of adsorption can be one or a combination of several phenomena, including chemical complex formation at the surface of the adsorbent, electrical attraction (a phenomenon involved in almost all chemical mechanisms, including complex formation), and exclusion of the adsorbate from the bulk solution. Al-Rashdi et al. (2011) consider adsorption as a mass transfer process in which a substance is transferred from a liquid phase to the surface of a solid and becomes bound by physical and/ or chemical interactions.

Bessho et al. (2019) reiterated that adsorption is one of the effective techniques for recovery of metal ions from water, and that separation and recovery of metals by adsorption can achieve both purification of the AMD and metal recovery. Rodriguez-Galan et al. (2019) also state that adsorption is the most employed technique commercially since it can recover 99% of the metals.

9.2.3.2 Selected Typical Studies of Selective Adsorption

The adsorption process has evolved over the years to become an important method for AMD treatment and metal recovery technique (Bessho et al., 2017; Bessho et al., 2019; Rodriguez-Galan et al., 2019). This chapter deals with only a few examples of materials used in the adsorption process for recovery and/ or removal of metals from AMD. Most of the adsorbents, particularly, the newly developed sustainable adsorbents for industrial wastewater including AMD treatment, are discussed elsewhere in other references.

The possibility of using different types of cross-linked gelatin hydrogels for recovery and removal of metals from acidic wastewater by adsorption was investigated by Bessho et al. (2017). Gelatins isolated from a porcine skin by an acid process (Type A) and a bovine skin by an alkaline process (Type B) were used in the study; and glutaraldehyde was used for cross- linking of Type A and Type В gelatin hydrogels. Copper was used as a model substance in the study because it is a common AMD metal. Metal recovery by gelatin hydrogels mainly consists of adsorption to gelatin molecules and absorption into hydrogels. The results of the study showed that the cross- linked Type В gelatin hydrogels recovered more Cu than Type A gelatin which appears to indicate that Type В gelatin is a more suitable adsorbent material to recover cationic metal ions. It must be noted that the liberation of a proton from the carboxyl group is required for adsorption of metals to gelatin (Bessho et al., 2017). Therefore, in the case of Type В gelatin, it appears that the increased Cu recovery was mainly due to electrostatic adsorption of Cu2+ to the carboxyl group without a proton. In other words, the Cu recovery using Type В gelatin hydrogels was mainly affected by carboxyl groups on side chains of gelatin hydrogels at higher pH. On the other hand, Type A rarely had carboxyl groups as side chains. In fact, amongst the two metal recovery mechanisms, adsorption was rarely detected with Type A gelatin. It is possible that Type В gelatin hydrogels also recovered Cu by absorption, but at higher pH, adsorption surely contributed to Cu recovery Doubtless, in this study, the copper recovery by Type В gelatin hydrogels was dependent on pH. In order to develop gelatin hydrogels having a high- performance adsorption capacity for metal recovery from AMD, preparation of "mixed" gelatin hydrogels blended with other natural organic compound was suggested by Bessho et al. (2019). However, the study by Bessho et al. (2019) only investigated the metal recovery efficiency of chitosan with a view of preparing, in future research, "mixed" gelatin hydrogels including some organic solids with a high metal adsorption capacity. Thus, metal recovery efficiency of chitosan was investigated in a study by Bessho et al. (2019) using 2 mM of three kinds of simulated metal solutions (i.e., copper, zinc, and manganese). The results showed that at pH 5,0.1 g of chitosan recovered approximately 90% of Cu from 50 mL of the simulated solution. At relatively higher pH (>3.0), over 90% of Cu recovery was achieved. However, Cu recovery was not detected at pH 2.0. Thus, it was considered that adjustment of solution pH allowed the recovery of Cu from acidic wastewater. Chitosan mainly has lots of amino and hydroxyl groups. This implies that Cu recovery using chitosan was mainly induced by formation of chelate compounds. Ideally, the Cu recovery using chitosan was affected by pH. In contrast, lower Zn was recovered in comparison to Cu. The Mn recovery was hardly detected within the pH range of 2-5. From these results, and in particular, because chitosan had a high performance for Cu recovery, Bessho et al. (2019) suggested that "mixed" gelatin hydrogel blended with chitosan had a potential for the high-performance adsorbent for Cu recovery.

In 1997, Tavlarides and Doerkar developed a new class of adsorbents for selective separation and recovery of metal ions from dilute aqueous solutions (Tavlarides and Doerkar, 1997a, b, c, d). According to Doerkar and Tavlarides

  • (1998), selectivity of these materials is a function of immobilised ligands on the ceramic supports and the condition under which the metal ion solutions are treated. These adsorbents differ from polymeric resins/ion-exchange resins and have the following advantages: (1) selectivity for separation of desired metal ions, (2) reduced or no interference from the accompanying cations (Na+, Ca2+, Mg2+), anions (N03~, S042', Cl), and complexing agents,
  • (3) easy and selective regeneration of the adsorption bed which yields concentrated solutions for recovery, (4) high mechanical strength for fixed bed applications, (5) higher metal ion uptake due to open pore structure, higher porosity, and non-swelling characteristics, and (6) no irreversible adsorption of organics. Following the development of the adsorbents, Doerkar and Tavlarides (1998) undertook a study to (1) evaluate the performance of adsorbents for separation and recovery of iron, copper, zinc, cadmium, and lead from Berkeley Pit simulated waters, and (2) propose an integrated process with fixed bed set-ups and specify operating conditions. The adsorbents studied include ICAA-A, ICAA-B with a substituted quinoline group (Tavlarides and Doerkar, 1997d), ICAA-C with a substituted oxime, ICAA-D

TABLE 9.2

Target Metals and Other Constituents in Berkeley Pit Water

Ions

Concentration (mg/L)

Other Constituents

Aluminium

2.60

pH = 2.85

Cadmium

2.14

[Fe3*]/[Fe2*] = 0.15

Calcium

456

Copper

172

Iron

1068

Lead

0.031

Magnesium

409

Manganese

185

Sodium

76.5

Zinc

550

Nitrate

<0.1 (as N)

Sulphate

7600

Source: Doerkar and Tavlarides, 1998.

with the thio/amine group (Tavlarides and Doerkar, 1997a), ICAA-E with the phosphoric acid group (Tavlarides and Doerkar, 1997a), and ICAA-F with a thiophosphinic group (Tavlarides and Doerkar, 1997a). The batch shakeout tests and breakthrough curve studies were executed using simulated Berkeley Pit water. Simulated Berkeley Pit water containing targeted metal ions (Fe, Cu, Zn, Cd, Pb), sulphate ions, and a pH of 2.8 was prepared according to the compositions shown in Table 9.2.

The results showed that the prepared adsorbents ICAA-A, ICAA-B, or ICAA-C and ICAA-D have the potential to remove and recover Fe3+, Cu2+, Zn2+, Cd2~ and Pb2+ from Berkeley Pit waters. Given the selectivity of the adsorbents and the potential to synthesise an adsorbent for Fe2+, an integrated adsorption process was devised. In one scheme the first bed comprised of ICAA-A for Fe3+ removal, the second bed comprised of ICAA-C for Cu2+ removal, and the third bed comprised of ICAA-D for removal of Zn2+, Cd2+, Pb2*, and some Fe2+. In a second scheme, another adsorbent can be synthesised to remove Fe2+ and a bed of this material can be used between the second and third bed. The study showed that the effluent from the last bed of either scheme was not acidic and can be discharged in an environmentally safe manner if other toxic metal ions are removed. Furthermore, the processes do not require adjustment of the pH of the feed stream. A number of the adsorbents were found to be able to retain a stable adsorption capacity after 20 cycles of adsorption and stripping.

The main objective of a study by Mohan and Chander (2006) was to remove and recover metal ions (Fe2+, Fe3*, Mn2+, and Ca2+ ions) from AMD using lignite as a low-cost sorbent in single and multiple column set-ups operating in downward flow modes. The results showed that the lignite usage rate was higher in a single column (0.981 g/L) compared to multiple columns (0.085 g/L), three columns in this study. In this study lignite usage was defined as follows:

Some studies have shown that using multi-stage treatments of heavy metal solutions with lignite could reduce the pollutants to acceptable discharge limits at a lower cost than using conventional heavy metal treatment processes (Simate et al., 2016). Mohan and Chander (2006) used a batch mode to examine the effect of pH. The sorption of Fe2+, Mn2+, and Fe3+ on lignite was found to increase with an increase in pH of the test solution. However, it must be noted that for sorption studies, the pH of solution must be less than the pH for precipitation of respective metal ions (Mohan and Chander,

2006). In their study, Mohan and Chander (2006) observed that the sorption of Fe2+ was very low at pHin < 2, but it increased from 6% to 84% at pH of 4.0. The insignificant removal of metal ions at low pH is attributed to the competition between the protons and the metal ions for the same binding sites (Schiewer and Volesky, 1995). Furthermore, the increase in the positive surface charges at low pH results in a higher electrostatic repulsion between the surface and metal ions (Reddad et al., 2002; Wang et al., 2006). However, at pHin > 4.0, the removal of Fe2+ was considered to have taken place by sorption as well as precipitation, i.e., the OH ions from the solution formed some complexes with Fe2~ (e.g., Fe2+ + 20H ^ Fe(OH),) (Kuhr et al., 1997; Arpa et al., 2000; Karabulut et al., 2000; Butler et al., 2007; Simate et al., 2016). Similarly, the removal of Mn2+ was also negligible at pHin < 2, but increased with an increase in pHin though precipitation only occurred at pH„, > 8. These results show that solution pH is a significant factor that determines the degree of metal adsorption. Furthermore, the equilibrium solution pH is a major parameter governing the extent of metal adsorption (Simate et al., 2016). Coals that generated higher solution pHs were found to exhibit the largest metal adsorption.

Stankovic et al. (2009) presented the results of the batch and column adsorption for copper and some associated ions by utilizing sawdust of deciduous trees (i.e., linden and poplar) as a low-cost adsorbent. The AMD from an abandoned copper mine, as well as synthetic solutions of the ions (Cu2+, Zn2+, Mn2+, and Fe2+) that are the main constituents of the AMD were both used as a model-system in the study. The adsorption of heavy metal ions strongly depended on the process time, the initial pH of the aqueous phase and the kind of ions adsorbed. The adsorption process was fast and after ten to twenty minutes of contact time, the system reached equilibrium. More specifically, the adsorption capacity of the studied sawdust was significantly affected by the initial pH of the solution and the kind of metal ions adsorbed. At lower pH of solutions the adsorption percentage decreased leading to a zero adsorption percentage at pH < 1.1. Maximum adsorption percentage was achieved at 3.5 < pH < 5. It was found that both poplar and linden sawdust have almost equal adsorption capacities against copper ions. The highest adsorption percentage (=80%) was achieved for Cu2+, while for Fe2+ it was slightly above 10%. The other considered ions (Zn2+ and Mn2+) were within this interval. The following gives the ranking of the ability of the considered ions to be adsorbed on sawdust: Cu2+ > Zn2+ > Mn2+ > Fe2+. The used sawdust had shown certain selectivity in the adsorption of heavy metal ions. Calculated selectivity coefficients over ferrous ions were: PCu2+_Fe2+ = 22; Pzn2+-Fe2+ “ 11; pMn2+-Fe2+ “ 2. The results obtained in the batch mode were validated in the column experiments using the real mine water originating from an AMD of a copper mine. Very high degree of copper ions adsorption was achieved in the column adsorption (>99.7%) experiments before the breakthrough point. Both kinds of sawdust used had an equal ability to adsorb copper ions. The pH of solution increased slightly during the adsorption which implied that there was a co-adsorption of protons contained in a treated solution. After completing the adsorption, instead of desorption, the loaded sawdust was drained, dried, and burned; the copper bearing ash was then leached with a controlled volume of sulphuric acid solution to concentrate copper therein. The obtained leach solution had the concentration of copper higher than 15 g dm'3 and the amount of H2S04 was high enough to serve as a supporting electrolyte suitable to be treated by the electrowinning technique for recovery of copper.

With the advent of nanotechnology, various types of nanomaterials with large surface area and small diffusion resistance have been developed and are now receiving considerable attention in water treatment (Zhang, 2003; Hu et al., 2006; Simate, 2012; Simate et al., 2012). For example, nanoscale iron particles have been established as effective reductants and catalysts for a variety of contaminants including heavy metals (Zhang, 2003). Iron nanoparticles possess the advantages of large surface area, high number of surface active sites, and high magnetic properties, which lead to high adsorption efficiency, high removal rate of contaminants, and easy and rapid separation of adsorbent from solution via magnetic field (Hu et al., 2006). In addition, it is possible that after magnetic separation by the external magnetic field, the harmful components can be removed from the magnetic particles, which can then be reused (Ponder et al., 2000; Oliveira et al., 2003; Hu et al., 2006).

The adsorption studies by Hu et al. (2006) showed that the nanoscale maghemite (y-Fe203) synthesised using a sol-gel method was very effective for selective removal of Cr(VI), Cu(II), and Ni(II) from wastewaters. The removal efficiency was highly pH dependent, which also governed the selective adsorption of metals from the solution. The optimal pH for the selective removal of Cr, Cu, and Ni was found to be 2.5, 6.5, and 8.5, respectively. Regeneration and readsorption studies demonstrated that the maghemite nanoparticles could be recovered efficiently for the readsorption of the metal ions, and metals could be highly concentrated for recycling.

Several studies have also been carried out to assess the technical feasibility of various kinds of raw and surface oxidised carbon nanotubes (CNTs) for sorption of various metals from aqueous solutions (Rao et al., 2007). The CNTs are carbon nanomaterials that were re-discovered by Iijima (Iijima, 1991). These materials have shown exceptional adsorption capabilities and high adsorption efficiencies for various organic pollutants (Lu et al., 2005), inorganic pollutants (Li et al., 2003a), and heavy metals (Li et al., 2003b; Li et al., 2006; Li et al., 2007). The CNTs are particularly attractive as sorbents because, on the basis of mass, they have larger surface areas than bulk particles, and that they can be functionalised with various chemical groups to increase their affinity towards target compounds (Savage and Diallo, 2005; Simate et al., 2012; Simate, 2012). The CNTs also have small size and are hollow with layered structures (Wu, 2007), which are important attributes for adsorption.

The studies have shown that the sorption capacities of metal ions by raw CNTs are very low, but significantly increased after oxidisation by HNO,, NaOCl, and KMnOT solutions (Rao et al., 2007). In fact, this is another advantage of CNTs in that they can be functionalised (or oxidised) with various kinds of chemical agents depending on the adsorption objective. The removal efficiency was also highly pH dependent, thus controlling the selective adsorption of metals from the solution. The sorption/desorption studies showed that CNTs could be regenerated and reused consecutively several times without significant loss in adsorbent capacity which signify the appropriateness of CNTs for commercial applications. Therefore, the superior sorption capacity and effective desorption of heavy metal ions suggest that the CNTs are promising sorbents for environmental protection applications (Rao et al., 2007).

The aim of a study by Rios et al. (2008) was to evaluate the use of low-cost sorbents like coal fly ash, natural clinker, and synthetic zeolites to clean up AMD generated at the Parys Mountain copper-lead-zinc deposit, Anglesey (North Wales), and to remove heavy metals and ammonium from AMD. Coal fly ash is a by-product of coal combustion that has been regarded as a problematic solid waste (Skousen et al., 2013; Ram and Masto, 2014), mainly due to the presence of potentially toxic trace elements (e.g., Cd, Cr, Ni, Pb) and organic compounds (e.g., polychlorinated biphenys, polycyclic aromatic hydrocarbons) (Shaheen et al., 2014). Natural clinker is a product of coal-bed fires ignited by natural processes (Rios et al., 2008). Zeolites are naturally occurring alumino-silicates with a three-dimensional framework structure bearing АЮ4 and Si04 tetrahedra (Motsi et al., 2009). These are linked to each other by sharing all of the oxygen to form interconnected cages and channels (Englert and Rubio, 2005; Motsi et al., 2009) where exchangeable cations are present which counter-balance the negative charge on the zeolite surface generated from isomorphous substitution (Barrer, 1978; Dyer, 1988; Motsi et al., 2009). The manufacture of synthetic zeolite is also possible (Simate et al., 2016). In this study synthetic zeolites were prepared by two methods, namely, (1) classic hydrothermal synthesis using natural clinker, and (2) alkaline fusion prior to hydrothermal synthesis using both coal fly ash and natural clinker. The sorption of Cu, Pb, Zn, Ni, Cr, Fe, As, and ammonium onto coal fly ash, natural clinker, and synthetic zeolites was studied in laboratory-batch experiments, which were carried out at room temperature to investigate the efficiency of the sorbents for removing heavy metals and ammonium from AMD.

With a rise in pH values as the sorbent dosage was increased, the results in a study by Rios et al. (2008) suggest that pH is strongly affected by the sorbent material rather than the AMD composition and particularly a higher sorbent dosage. It was noted from the study that there is a possibility of two competing reactions, namely, (1) release of alkalinity from sorbents, and (2) the removal of acidity from AMD components. The study found that at higher sorbent dosage the acidity from AMD components is overrun and pH is bound to increase whereas with the lower sorbent dosage the alkalinity from the sorbent is exceeded by the acidity from the AMD components and the pH remains low. In other words, the pH played a very important role in the sorption/removal of the contaminants and a higher adsorbent ratio in the treatment of AMD promoted an increase of the pH and vice versa. The results also revealed that the heavy metal removal was depended on the sorbent material and the applied dosage. For example, coal fly ash and natural clinker did not show good efficiency as sorbents to neutralise the AMD, but their synthetic products (e.g., coal fly ash-based faujasite; natural clinker-based faujasite; natural clinker-based Na-phillipsite) were effective as ion exchangers in removing acidity, Fe, Zn, and Cu from AMD. Amongst the two variants of the zeolite (i.e., faujasite and Na-phillipsite), faujasite was more effective. In fact, the results of the adsorption experiments suggest that faujasite can be applied in wastewater treatment as an immobiliser of pollutants, and the selectivity of faujasite for metal removal was as follows in decreasing order: Fe > As > Pb > Zn > Cu > Ni > Cr. In general, the results of the study showed that different sorbents contain considerable amounts of accessory phases that partly dissolve during the batch reaction, which may explain the sudden increase or decrease in metal concentration and, therefore, the release rate of the metal elements is controlled by the dissolution of the sorbent.

In addition to cation-exchange reactions, precipitation of hydroxide species (mainly of Fe) also played an important role in the sorption and coprecipitation, and thus led to the immobilisation of metals in the batch experiments. The efficiency in the removal of ammonium by coal fly ash and natural clinker was poor. However, the reaction between synthetic zeolites (coal fly ash-based faujasite; natural clinker-based faujasite; natural clinker- based Na-phillipsite) and AMD after 24 h of contact time produced lower ammonium concentrations in the solution. In fact, the study indicated that natural clinker-based faujasite produced a complete removal of ammonium after 24 h of contact time when a dosage of 1 g was used.

The adsorption behaviour of natural zeolite (clinoptilolite) was studied by Motsi et al. (2009) in order to determine its applicability in treating AMD containing 400, 20, 20, and 120 mg L"1 of Fe3+, Cu2+, Mn2‘, and Zn2+, respectively. Batch experimental tests for single and multi-component solutions were performed to ascertain both the rate of adsorption and the uptake at equilibrium. The optimum conditions for the treatment process were investigated by observing the influence of pH levels, the presence of competing ions, varying the mass of zeolite and thermal modification of the natural zeolite (calcination and microwaves). The adsorption studies showed rapid uptake, in general, for the first 40 min, and after the initial rapid period, the rate of adsorption decreased. The study found that about 80%, 95%, 90%, and 99% of Fe3+, Mn2+, Zn2+, and Cu2+, respectively, were adsorbed from single component solutions in the first stage.

For multi-component solutions in a study by Motsi et al. (2009), only the adsorption of Fe3+ was significantly unaffected by the presence of competing ions. This may be because the main mechanism responsible for Fe3+ removal from solution is believed to be precipitation. The adsorption of the other three cations was affected significantly. For example, the amount adsorbed from multi-component solutions of concentration of 40 mg L"1 decreased by 33%, 41%, and 39% for Cu2+, Zn2+, and Mn2+, respectively, compared to their respective single component solutions. When the solution concentration was increased from 40 mg L"1 to 120 mg Lr1, the relative decrease in the amount adsorbed between the multi-component and single component cases increased further.

The study by Motsi et al. (2009) also indicated that the removal of the heavy metal ions was not only due to ion exchange, but also due to precipitation of metal hydroxides from the solution. This observation is similar to other previous studies (Khur et al., 1997; Arpa et al., 2000; Karabulut et al., 2000; Butler et al., 2007; Rios et al., 2008; Simate et al., 2016). It is noted that natural zeolites are generally weakly acidic in nature and that sodium form exchangers are selective for hydrogen (R-Na + H20 RH + Na+ + OH-), which leads to high pH values when the exchanger is equilibrated with relatively dilute electrolyte solutions (Erdem et al., 2004; Motsi et al., 2009) making metal hydroxide precipitation feasible. In other words, as reaction proceeds the solution pH increases which promotes metal precipitation. The rate of adsorption was also found to be directly proportional to the pH value of the solution. Adsorption decreased in more acidic solutions, due to hydrogen ion competition. The rate of adsorption and capacity also depended on the mass of the adsorbent and heat treatment by either microwaves or heating in a furnace. For zeolite exposed to microwave radiation, the adsorption rate increased with exposure time, but only up to a certain limit. The adsorption rate of the zeolite exposed to microwave radiation began to decrease as exposure time approached 30 min. The rate of adsorption by calcined zeolite was also found to be faster compared to untreated zeolite, but the efficiency decreased for zeolite exposed to very high temperatures (e.g., 800°C). The increase in rate of adsorption capacity as a result of thermal treatment may be attributed to the removal of water from the internal channels of natural zeolite which leaves the channels vacant and hence increases the adsorption capacity of the zeolite (Turner et al., 2000; Ohgushi and Nagae, 2003; Motsi et al., 2009). The removal of water also resulted in a change in the surface area of the samples after thermal treatment (Motsi et al., 2009). The samples that were exposed to extreme thermal conditions had lower surface areas due to thermal runaway, whereby the zeolite structure collapses (Ohgushi and Nagae, 2005; Akdeniz and Ulku, 2007; Motsi et al., 2009). When the structure collapses the porosity of natural zeolite decreases and thus the adsorption capacity is reduced (Motsi et al., 2009). The study also found that efficiency of metal removal from solution by natural zeolite is inversely proportional to the initial solution concentration. According to the equilibrium studies, the selectivity sequence of metals by natural zeolite can be given as Fe3+ > Zn2+ > Cu2+ > Mn2+, with good fits being obtained using Langmuir and Freundlich adsorption isotherms. The significance of this study is that preliminary tests using AMD samples from Wheal Jane Mine, UK, showed that natural zeolite has great potential as an alternative low-cost adsorbent in the treatment of AMD.

Petrilakova and Balintova (2011) utilised five different types of natural and synthetic adsorbents for the removal of Fe, Cu, Al, Mn, and Zn from real AMD (shaft Pech, Smolnik locality, Slovakia) with a pH of 4.2. The following adsorbents were employed in the study: (1) inorganic composite sorbent SLOVAKITE, (2) active carbon (granularity < 1 mm), (3) turf brush PEATSORB, (4) universal crushed sorbent ECO-DRY (REO AMOS Slovakia), and (5) zeolite (granularity 0.5-1 mm, 2.5-5 mm, 4-8 mm). The chemical composition of the AMD was as follows in mg/L: Fe (338), Cu (1.16), Al (44.4), Mn (26), and Zn (5.81). Batch experiments were carried out by mixing various amounts of adsorbents into 100 mL of raw AMD over a period of 24 h. Active carbon was found to be the most efficient for the removal of Fe at 99.98% efficiency. The efficiency of Cu removal from AMD using active carbon and inorganic composite sorbent SLOVAKITE was 98.3% in both cases. Active carbon and inorganic composite sorbent SLOVAKITE adsorbents at 99.98% efficiency were the two most efficient for Al removal. For the Mn, the most effective adsorbent was active carbon (93.08%) and Turf brush PEATSORB (87.69%).

9.2.3.3 Summary

The typical examples given in the use of the adsorption processes to remove metallic pollutants from the AMD clearly demonstrate the feasibility of the adsorption techniques. In other words, the efficiencies of various adsorbents for the removal of heavy metals from AMD were illustrated in a number of studies. Several studies have also evaluated various techniques for recycling of used adsorbents and recovery of the heavy metals from the desorbing agents (Lata et al., 2014).

For regeneration and reuse of adsorbents, various possible regenerating agents such as acids, alkalis, and chelating agents such as ethylene diamine tetraacetic acid (EDTA) have been studied by many researchers. Lata et al. (2014) reviewed and summarised the performance of various desorbing agents for removal of adsorbed metals and regeneration of the saturated adsorbents. A study by Lata et al. (2014) made the following conclusions: (1) the alkalis are efficient desorbing agents for desorption of heavy metal(s) from chemical adsorbents or chemically modified adsorbents, (2) acids are efficient for desorbing bio-adsorbents, and (3) the chelating agent, EDTA, is the most efficient desorbing agent for biomass desorption. The review by Lata et al. (2014) found that many of the adsorbents can be reused effectively after regeneration.

After the solution is separated from the adsorbents, the metal laden solution is often subjected to various processes of purification and concentration before the valuable metals can be recovered either in their metallic state or as chemical compounds. The methodologies may include precipitation, distillation, adsorption, and solvent extraction (Roto, 1998; Parnell, 2019), and the final recovery step may involve precipitation, cementation, or electrometallurgical processing (Parnell, 2019).

  • 9.2.4 Ion Exchange
  • 9.2.4.1 Concepts of Ion Exchange

According to Hubicki and Kolodynska (2012) ion exchange is the exchange of ions between the substrate and surrounding medium. In addition, Patil et al. (2016) regard ion exchange as a physical treatment technique in which ions dissolved in a liquid or gas interchange with ions on a solid medium. Patil et al. (2016) state further that the ions on the solid medium are associated with functional groups that are attached to the solid medium, which is immersed in the liquid or gas. Typically, ions in dilute concentrations replace ions of like charge that are of lower valence state, but ions in high concentration replace all other ions of like charges (Patil et al., 2016). EPA (2014) define ion exchange as the reversible exchange of contaminant ions with more desirable ions of a similar charge adsorbed onto solid surfaces known as ion-exchange resins. Despite the myriad of definitions for the ion-exchange process, it is also noted that positively charged molecules bind to cation-exchange resins while negatively charged molecules bind to anion-exchange resins (Thermo Scientific, 2007). Flowever, it must be noted that Dinardo et al. (1991) classify ion-exchange resins (i.e., insoluble matrixes or support structures that act as a medium for ion exchange) as anionic, cationic, and chelating. Anionic and cationic resins are used extensively in water purification processes. More specifically, anionic resins are significant in extracting amphoteric elements such as arsenic and metals that form sulphate complexes such as uranium (Dinardo et al., 1991). Cationic resins can have either sulphonic or carboxylic functionality and are essentially non-selective and thus can extract most polyvalent cations including magnesium, iron, calcium, and many others (Dinardo et al., 1991). Chelating resins are relatively new and have high selectivity for some specific metals.

Ion exchange is considered as the most energy efficient and economical technology of all the recovery techniques (Patil et al., 2016). It is the only process that can efficiently treat very dilute solutions in parts per million (ppm) levels on a once-through basis (White and Asfar-Siddique, 1997; Patil et al., 2016). In other words, the technology allows efficient removal of traces of contaminants from solutions (Dqbrowski et al., 2004). Hubicki and Kolodynska (2012) also emphasise that ion-exchange process is designed to remove traces of ionic impurities from water and process streams and give a product of desired quality. According to Hardwick and Hardwick (2016) the advantage of ion exchange over other methods such as solvent extraction or even precipitation is that the technique can still be viable when feed concentrations have dipped below the economic threshold of the other technologies. Furthermore, according to Dqbrowski et al. (2004), the technique is specifically convenient when there is a need to treat large volumes of diluted solutions. According to EPA (2014), the process is appropriate for desalination and the removal of a number of pollutants including hardness, alkalinity, radioactive waste, ammonia, and metals.

9.2.4.2 Selected Typical Studies of Ion Exchange

Ion exchange has been successfully tested on wastewater from mining operations and, generally, works more effectively for waters in the pH range of 4 to 8 that has low suspended solids and low concentrations of iron and aluminium (EPA, 2014). The more complex the mixture, the harder it is to remove all metals effectively, and the capacity of any resin to remove contaminants is limited by the type of resin, the number of available exchange sites and the chemistry of input water (EPA, 2014).

Hardwick and Hardwick (2016) provided an overview of the potential for the recovery of value from contaminated mine waters and waste streams. The study discussed various metal ions commonly found in AMD and other mine waters and evaluated the levels of concentration the metals would become attractive for recovery using ion-exchange technology. Based on the capital and running costs, Hardwick and Hardwick (2016) determined the capital payback period for various contamination levels for a number of valuable metals that are likely to be found in mine waters. Firstly, the study categorised the constituents of AMD as follows: (1) value containing elements, (2) hazardous elements, (3) low value elements that will co-load, (4) ion-exchange poisons, and (5) elements of little concern with regard to the ion-exchange process. Secondly, the study discussed the viability and performance of the resins and economics of the process based on metal concentrations, presence of impurities (or competing/poisonous ions), flow rate, and price of metals.

Hardwick and Hardwick (2016) suggested that the first influencing factor on the economics to consider is the concentration at which a particular metal is present in a possible feed source. When the concentration of metals in a stream is very low, efficiency of removal by ion-exchange resins is relatively high because at that point it is film diffusion rather than particle diffusion that limits the kinetics. In such a condition, metal leakage is very low, and the operating capacity of the resin increases. However, it is important to note that where the concentrations of the desired metals are high, the rough guideline is that for levels above lg/L, it may become more economical to investigate another technology, such as solvent extraction for recovery. The study also found that the concentration at which a metal becomes economically viable for recovery is heavily dependent on the sales price of the product produced. If there is a depression in metal pricing, the operation is more likely to become less economical even for relatively high concentration streams.

Undoubtedly, the resin choice is often made on how suitable it is for the recovery of a specific element, but it is possible for the wastewater to contain trace amounts of other metals that have an even higher affinity to the functional groups on the resin, thus taking up active sites and reducing the capacity of the resin for the element of value (Hardwick and Hardwick, 2016). Iron is a very good example. For instance, most common ion-exchange resins have a high selectivity for trivalent iron. The iron may displace the desired element and poison the resins over time. Aminophosphonic resins, for example, may lose useful capacity over a period of time and thus become uneconomical. Therefore, it is imperative to use a resin with a lower selectivity for iron upfront of such resins (e.g., aminophosphonic resins) in order to protect them. Alternatively, iron may be removed by precipitation by raising the pH above 3 and, thereafter, use the filtration process. Radioactive elements when present in the feed solution also may make the recovered metals unsaleable. A very good example is thorium that may be loaded onto resin during the recovery of rare earths (Hubicki and Kolodyriska, 2012). However, thorium may be removed by precipitation or selective elution. Uranium can be recovered separately by a strong base anion resin (Botha et al., 2009), whereas radium may be removed using a strong acid cation resin (Clifford, 2004).

As discussed in Chapter 3, a number of reaction pathways, in the course of AMD generation, lead to the production of aqueous hydrogen cation (H30+) thus resulting in very low pH for AMD (Dold, 2010; Garland, 2011). Studies on recovery of Cu(II) ions have shown that due to low pH values of wastewaters (pH < 2) such as AMD, conventional chelating ion exchangers of functional iminodiacetate and aminophosphonic groups practically do not adsorb Cu(II) ions. Therefore, special chelating ion exchangers characterised by much greater affinity for Cu(II) than for other metal ions have been used instead (Melling and West, 1984; Bolto and Pawlowski, 1987; Dorfner, 1991). For example, Cu(II) ions can be removed from highly acidic solutions (pH > 1.5) using Dowex XFS-4196 chelating ion exchanger which can be

TABLE 9.3

Physical Properties and Specifications of the Ion-Exchange Resins

Parameter

Indion 820 (Strong Base Anion Exchange Resin)

Indion 850 (Weak Base Anion Exchange Resin)

Physical form

Spherical opaque beads

Moist beads

Ionic form as supplied

Chloride

Free base

Moisture holding capacity

49-56%

40%

Particle size

0.3-1.2 mm

0.3-1.2 mm

Uniformity coefficient

1.7 max

1.7 max

Total exchange capacity

1.0 meq mL'1

1.10 meq mL-1

pH range

0-14

0-7

Source: Gaikwad et al., 2009.

easily regenerated by means of sulphuric acid with a concentration of 100 g H2S04 dm 3 (Dqbrowski et al., 2004). Amphoteric ion-exchange fibres can also be applied for the removal of Cu(II) ions from acidic wastewaters. Various other types of ion exchangers have also been used for selective removal of Cu(II) ions (Hubicki and Jusiak, 1978; Hubicki and Pawlowski, 1986; Hubicki et al., 1999).

Gaikwad et al. (2009) performed a laboratory scale investigation to remove copper from AMD using ion-exchange resins - Indion 820 and Indion 850. The physical properties and specifications of the resins are given in Table 9.3. The concentrations of copper used in the study ranged from 50 to 250 mg L 1 and the dosage of resin from 25 to 700 mg L_1. The effect of the initial concentration of copper ions, dosage of resin and pH on exchange capacities of ion-exchange resins was studied in a batch mode. The mixture of resins and copper solutions was agitated for a predetermined period at room temperature. Thereafter, the resins were separated and the filtrate was analysed by an atomic absorption spectrometer (Chemito, AAS-3000) for the residual concentration of copper.

The ion-exchange process, which is pH dependent, showed maximum removal of copper in the pH range of 2-6 for an initial copper concentration of 50-250 mg Lr1 and with a resin dosage of 25-700 mg Lr1. Five isotherm models (Freundlich, Langmuir, Redlich-Peterson, Temkin, and Dubinin- Radushkevich) were tested and the equilibrium data fitted to all the sorption isotherms very well. The study found that the uptake capacity of Indion 850 was larger than that of Indion 820 due to the intrinsic exchange capacity. Therefore, of the two different ion-exchange resins studied, Indion 850 was considered to be the most efficient in removing copper ions. The uptake of copper by the ion-exchange resins was reversible which showed that the resins have a good potential for the removal/recovery of copper from AMD. In other words, the study showed that ion-exchange resins such as Indion 820 and Indion 850 can be used for an efficient removal of copper from mine wastewater.

In a study by Gaikwad et al. (2010), using the ion-exchange technique, a factorial experimental design method was used to examine the removal of Cu2+ ions from AMD wastewater. The strongly acidic cation-exchange resins, Indion 730, were used and the following four factors were varied at three different levels: initial concentration of Cu2+ ions (100,150, 200 mg L'1), pH (3, 5, 6), flow rate (5, 10, 15 L h'1), and dosage of the resin or bed height (20, 40, 60 cm). A matrix was established according to the high, middle, and low levels of experimental parameters, coded as +1, 0, and -1, respectively, and, thereafter, 81 experiments with all possible combinations of variables were conducted. The factors were coded so as to simplify the calculations (Simate and Ndlovu, 2008). For each run, 25 L of Cu2+ solution was passed through the column (height = 100 cm; diameter = 5 cm) made of glass at 5 L h_1 for 100 min. The results of the study clearly showed that ion-exchange process was effective in removing Cu2+ ions and can provide a solution for removal of such metals from AMD waste. The results were analysed statistically using Student's t-test, analysis of variance, F-test, and lack of fit so as to establish the most important process variables affecting the removal of Cu2+ ions by Indion 730 resins. In this study, pH was found to be the most important variable. It is noted from previous studies that the cation- exchange capacity and selectivity both increase with increasing pH, while the anion-exchange capacity and selectivity decrease with increasing pH (Churms, 1966). The interaction between the resin bed height and initial concentration of Cu2+ ions and that between the pH and initial concentration of Cu2+ ions were also found to be highly important.

Gordyatskaya (2017) explored the selective removal of toxic metal (Cu, Ni) ions present in AMD with excess ferrous ions using a chelating resin with a di-(2-picolyl)amine functional group (Lewatit TP 220) and the prospect of subsequent recovery of metallic copper with electrowinning. Simulated AMD solution used in the study contained 2 g/L of iron together with manganese, zinc and copper or nickel ions. Metals were added as sulphates and the pH was adjusted to 2 using sulphuric acid. Column dynamic experiments were used to determine the sorption and desorption efficiencies of Cu and Ni ions, separately. The chelating resin TP 220 demonstrated a very high affinity for copper, while manganese, zinc, and iron were not taken up during the experiments. Similar experiments performed related to the removal of nickel from the same AMD matrix, showed that the sorption of nickel was less efficient than that of copper. During the experiments for the removal of copper or nickel, small quantity of iron in the form of hydrated ferric oxide flakes formed a layer on the surface of the column. Therefore, a two-stage desorption process using sulphuric acid in the first step and ammonia solution in the second step was used. Unlike the Dowex XFS 43084 which is not produced anymore, Lewatit TP 220 (analogue of Dowex XFS 4195) could not be regenerated directly with sulphuric acid. In this process, the necessity to strip copper from the loaded resin by a complexation reaction with ammonia solution is the main drawback of the resin. This is because the resulting ammonia solution of copper is not suitable for the electrowinning of Cu. Therefore, a weak base anion exchanger having tetraethylenepentamine (TEPA) functional group on macroporous polyacrylate matrix Purolite A 832 was used as a chelating resin to take up copper from ammonia solution. Thereafter, once copper has been stripped off from TEPA using sulphuric acid, it can easily be recovered by electrolysis.

In view of the high toxicity of lead, its content in water and industrial waste- waters must be reduced to a minimum within the ppb level (D^browski et al.,

2004), thus several ion-exchange studies have been carried out to remove Pb(II) ions. Of significant importance is the simultaneous removal of Pb(II) and Cd(II) ions as well as organic ligands using anion exchangers of various types (Dudziriska and Clifford, 1991; Dudziriska and Pawlowski, 1993). Some results showed that the anion exchangers of weakly basic functional groups are characterised by higher affinity for the complexes of Pb(II) and Cd(II) with EDTA than strongly basic anion exchangers and that the anion exchangers of polyacrylic skeleton are characterised by greater affinity for the complexes (i.e., Pb (II) and Cd (II) with EDTA) than the anion exchangers of the same type of polystyrene skeleton (Dqbrowski et al., 2004). Other studies have shown that higher selectivity, great exchangeability as well as reversibility of the sorption-elution process towards Pb (II) ions are characteristics of the sulphine cation exchanger exhibiting-S042~ groups (Bogoczek and Kociolek-Balawejder, 1988). On the other hand, chelating ion exchanger with functional iminodiacetate groups, Lewatit TP 207, has been recommended by the Bayer Company for selective removal of metal ions, particularly Pb(II) ions (Bayer, 2000).

The study by Ladeira and Gonsalves (2017) investigated the separation of uranium from the other anions present in the acid water under batch and column mode using the ion-exchange technique. Two strong base resins (Dowex Marathon A and IRA-910U) were compared in the study for their effectiveness in the removal of uranium from high sulphate AMD. The AMD sample was collected nearby the uranium mine in the southeast of Brazil and consisted of acid water generated at waste rock piles. The acid water pH was around 2.7, the uranium concentration was in the range of 6-14 mg L'1, sulphate concentration was nearly 1400 mg L ', fluoride concentration of 140 mg Lr1, and iron concentration of 180 mg L-1. The influence of ions, commonly found in acid waters like sulphate and fluoride, was also assessed in the ion-exchange process. The resins showed a significant capacity for uranium uptake which varied from 66 to 108 mg/g for IRA-910U and 53 to 79 mg/g for Dowex A. This shows that IRA-910U performed significantly better than Dowex A. However, the results showed that both resins performed at only about 40-60% of their theoretical value (1 equiv. g"1 for IRA-910U and 1.3 equiv. g 1 for Dowex A) probably because of the interference of other anions. The effect of pH on the loading capacity for column experiments was more accentuated than that for batch tests. The results also showed that S042 was the most interfering ion and it had a deleterious effect on the uranium recovery in the pH range studied. Fluoride did not affect uranium removal. Based on the fact that the study was carried out with a real acid water sample it was demonstrated that, although loadings were not considered so high, uranium can be removed efficiently, and elevated recoveries may be achieved.

9.2.4.3 Summary

The ion-exchange methodology has been found to be technologically simpler compared to other techniques and enables efficient removal of even traces of impurities from solutions. In fact, according to Hardwick and Hardwick (2016) when the concentration of impurities in a waste stream is very low, efficiency of removal by ion-exchange resins is relatively high because at that point it is film diffusion rather than particle diffusion that limits the kinetics. By its nature the undesirable ions in waste streams are replaced by the ones on ion-exchange resins, for example, that do not contribute to contamination of the environment (Dqbrowski et al., 2004). In other words, ion exchange enables replacing the undesirable ion by another one which is neutral within the environment. At the moment new types of ion exchangers with specific affinity to specific metal ions or groups of metals are available as an effort to enhance selectivity. To sum it up, the importance of ion exchange with respect to AMD treatment is characterised by two basic approaches according to Dinardo et al. (1991): (1) the selective removal of heavy metals from the AMD solution, and (2) water recovery to produce potable water, which involves the total removal of both anions and cations from AMD.

  • 9.2.5 Electrochemical Treatment
  • 9.2.5.1 Concepts of Electrochemical Treatment

The application of electricity in the treatment of contaminated water, generally referred to as the electrochemical techniques, has been in existence for over a century, and since then the methods are still highly reliable for wastewater treatment (Tran et al., 2017a). In fact, electrochemical technologies have the potential to provide selective and measurable recovery of base metals from dilute mining influenced water (Figueroa and Wolkersdorfer,

2014). However, the technology is not widely used in the treatment of AMD (Figueroa and Wolkersdorfer, 2014).

The nature of the electrochemical process is premised on the usage of electricity to pass a current through an aqueous metal bearing solution, which also contains a cathodic plate and an insoluble anode (Tran et al., 2017b). In other words, an electrochemical system consists of at least two

FIGURE 9.3

Schematic illustration of the (a) electrolysis process and (b) the galvanic element.

electrodes - an anode and a cathode - and an intermediate space filled with electrolyte (Muddemann et al., 2019). The electrical circuit is closed through electrical wires either with a voltage source (electrolysis cell) or an electrical load (galvanic element). In many applications, a separator (membrane or diaphragm) separates the reactor into anode and cathode compartments. The electrolyte surrounding the anode is named anolyte and the electrolyte on the cathode side is called catholyte. Figure 9.3a shows the general set-up of an electrolysis cell, while the function of a galvanic element is shown in Figure 9.3b.

There are a number of possible mechanisms involved in electrochemical processes including electrocoagulation, electroflotation, and electrodeposition (Chen, 2004; Fu and Wang, 2011; Figueroa and Wolkersdorfer, 2014; Liu et al., 2016; Tran et al., 2017a). Other electrochemical techniques are available including electrochemical oxidation (and reduction), electrochemical precipitation, electrokinetic, and emulsion splitting electrolysis (Chen, 2004; Drogui et al., 2007; Muddemann et al., 2019), but will not be covered in this chapter. Only electrocoagulation, electroflotation, and electrodeposition will be discussed in the chapter.

9.2.5.1.1 Electrocoagulation

The concept of electrocoagulation is based on the use of sacrificial metal anodes to stimulate electrolytic metal precipitation in undivided cells (Bejan and Bunce, 2015). Ideally, electrocoagulation (or sometimes called electrofloculation) involves the generation of coagulants in situ by dissolving either aluminium or iron ions electrically from aluminium or iron electrodes, respectively (Chen, 2004; Fu and Wang, 2011). Flowever, in electrocoagulation, iron is the most widely used electrode followed by aluminium (Singh and Mishra, 2017). Chen (2004) outlined the following reactions as the ones taking place at the anode under alkaline and acidic conditions:

For the aluminium anode:

For iron anode:

In addition, oxygen is also liberated at the anode through the following reaction:

During electrocoagulation, water electrolysis at the cathode also takes place and generates hydrogen (Equation 9.11) in form of microbubbles (Muddemann et al., 2019).

In addition, the dissolved H‘ ions are also reduced at the cathode as follows:

From reactions 9.1 to 9.12, it can be summarised that during the electrocoagulation process, the generation of metal ions takes place at the anode and the hydrogen gas is released from the cathode (Chen, 2004).

In the electrocoagulation process, co-precipitation of the sacrificial metal (reactions 9.4 and 9.7) and the metals present in the AMD take place due to pH increase because of the cathodic reduction of H~ ions (Equation 9.12) (Bejan and Bunce, 2015) or due to the OH formed at the cathode through reaction 9.11 (Muddemann et al., 2019). Ideally, the reduction of dissolved hydrogen ion to hydrogen gas as shown in reaction 9.12 results in a pH increase in solution (Jenke and Diebold, 1984). In addition, Muddemann et al. (2019) as well as

Bejan and Bunce (2015) postulate that the OH ions formed at the cathode enable the coagulation or precipitation of the metal ions present in the AMD and metal ions generated from sacrificial metal anodes as well as flocculation of dissolved or colloidal constituents in water. Mollah et al. (2004) also argue that during electrocoagulation, iron and aluminium almost instantly become polymeric hydroxides, which are excellent coagulating agents. Drogui et al.

(2007) propose that in an electrocoagulation process, aluminium or iron ions are produced and they combine with hydroxyl ions (generated by electrolysis of water at the cathode) to form a polymeric coagulant, aluminium hydroxide, or iron hydroxide, which adsorbs colloidal material present in the solution to form insoluble floes which are then carried away by the hydrogen gas bubbles generated at the cathode to the surface of the liquid. According to Heidmann and Calmano (2008), in the case of aluminium, the metal ions are removed from the solution by several mechanisms including direct reduction at the cathode, hydroxide formation by the hydroxyl ions formed at the cathode, and co-precipitation with the aluminium hydroxides.

In the electrocoagulation process, the hydrogen gas formed at the cathode (reaction 9.12) helps to float the flocculated particles out of the water (Chen, 2004). Muddemann et al. (2019) also re-emphasise that the hydrogen bubbles formed are often used for flotation (i.e., electroflotation) and separation of the formed aggregates simultaneously. Therefore, according to Muddemann et al. (2019) electroflotation is often combined with electrocoagulation. Feng et al. (2016) reiterate that in practice, an electrocoagulation process will be often followed by an electroflotation process and this combined system can be considered as electrocoagulation-flotation process. In other words, the combination of electrocoagulation and electroflotation is called electrocoagulation-flotation process (Azimi et al., 2017). This method achieves better removal percentages.

Chen (2004) suggests that the advantages of electrocoagulation include high particulate removal efficiency, compact treatment facility, relatively low cost, and possibility of complete automation. In addition, Vasudevan et al. (2010) state that some of the advantages of electrocoagulation are its generally low cost, reduced sludge production, and is easy to operate. According to Rodriguez et al. (2007), electrocoagulation may prove to be not only feasible and economically friendly, but also technically and economically superior to conventional technology like chemical precipitation.

9.2.5.1.2 Electroflotation

Electroflotation is a separation process in which hydrophobic particles in water or particles generated by other processes (e.g., electrocoagulation) are carried to the aqueous surface by adhering to gas bubbles (Muddemann et al., 2019). In the context of this chapter, electroflotation is considered as a simple solid/liquid separation process that floats pollutants to the surface of a water body by tiny bubbles of hydrogen and oxygen gases generated from water electrolysis (Raju and Khangaonkar, 1984; Fu and Wang, 2011;

Chen, 2014). In this process, hydrogen and oxygen gases are liberated from the electrochemical reactions at the cathode (reaction 9.11) and anode (reaction 9.10), respectively (Chen, 2014). It is also possible to combine electroflotation with electrocoagulation (see the discussion on electrocoagulation) if one of the electrodes is dissolved by electric current during electrolysis. The released metal ions cause coagulation of colloidal molecules, which adhere to the gas bubbles formed by the water electrolysis (Muddemann et al., 2019). According to Kraft (2004) electroflotation is one of the most effective and versatile methods of electrochemical water purification, as micro gas bubbles are produced and the size distribution of the gas bubbles is very narrow. In addition, Srinivasan and Subbaiyan (1989) state that the advantages of electroflotation technique are its high degree of gas saturation and the dimensional uniformity of the generated bubbles which subsequently lead to the highest recovery in a short time span.

9.2.5.1.3 Electrodeposition

The recovery of metals using electrochemical techniques has been practiced in the form of electrometallurgy for a very long time (Dubpernel, 1978; Chen, 2004). Figueroa and Wolkersdorfer (2014) also state that electrochemical deposition of metals is widely used in metallurgical processing and treatment of high metal waste streams (e.g. electroplating waste) to recover high- purity metallic forms. In the process of electrochemical deposition of metals, the metal ions are electrochemically reduced and, in contrast to electrocoagulation, removed as metals of valence 0 (Muddemann et al., 2019). The electrochemical mechanism for metal recovery is very simple and is basically the cathodic deposition process (Chen, 2004). In the process, the positively charged metal ions move in the electric field between the electrodes to the negatively polarised cathode, where they are reduced to an element and deposited on the electrode surface according to the following equation (Muddemann et al., 2019):

In other words, when the dissolved metals migrate towards the cathode, they are reduced and deposited on the cathode (Figueroa and Wolkersdorfer, 2014). In general, the anode composition (effectively inert) is selected to limit its oxidation and thus promote the oxidation of water at the anode (Figueroa and Wolkersdorfer, 2014).

The metals extracted from the wastewater such as AMD using electrodeposition technique can be of a very high purity and thus the recovery of valuable metals is possible (Muddemann et al., 2019). The process is used not only for metal separation from wastewater, but also for large-scale production of metals such as copper and zinc. Indeed, electrodeposition which is commonly referred to as electrowinning is an alternative to conventional smelting technologies (Bejan and Bunce, 2015 ).

Э.2.5.2 Selected Typical Studies of Electrochemical Treatment

This section discusses typical examples of studies performed using electrochemical techniques for the treatment and/or recovery of metals from waste- waters including AMD. Electrochemical treatment of AMD offers possible advantages in terms of operating costs and the opportunity to recover metals (Chartrand and Bunce, 2003; Gaikwad and Gupta, 2008; Figueroa and Wolkersdorfer, 2014; Park et al., 2015). The technique can be operated at ambient temperature and pressure and has a robust performance and capability to adjust to variations in the influent composition and flow rate (Tran et al., 2017b). Furthermore, Tran et al. (2017a,b) including Fu and Wang (2011) emphasised that electrochemical processes are known to be very efficient methods for the treatment of industrial wastewaters, particularly, for the removal of heavy metal ions.

9.2.5.2.1 Electrocoagulation

Nariyan et al. (2017) used electrocoagulation to investigate the removal of copper, silicon, manganese, aluminium, iron, and zinc as well as sulphate from real mine water. Batch experiments with monopolar iron anode and stainless steel cathode as well as monopolar aluminium anode and stainless steel cathode were conducted separately to identify the best electrocoagulation conditions. The removal efficiency in mine water increased with increasing reaction time and increasing current density and the type of electrodes affected the metals and sulphate removal as could be shown by the adsorption isotherms. Based on kinetic modelling, the aluminium electrode was found to be more efficient for metal removal than the iron electrode. The removal behaviour of the metals can be explained by the Eh-pH and the metal's stability diagram. Both the free sorption energy calculation from the Dubinin-Radushkevich isotherm and the к-parameter for the kinetics showed that the removal of copper and silicon were influenced by physical interaction with the two electrodes, while zinc was merely being influenced by physical interaction with the iron electrode in the electrocoagulation process. All of the contaminants, except manganese and sulphate, obeyed a Langmuir isotherm when an iron electrode was applied. However, when an aluminium anode was utilised, the metals presented a different behaviour compared to the iron electrodes. Specifically, silicon, copper, and manganese were obeying a Freundlich isotherm, whereas zinc and sulphate were obeying a Langmuir isotherm. Sulphate was better removed by an aluminium electrode compared to the iron electrodes with a maximum removal rate of 40.5% and 28.9%, respectively.

Mamelkina et al. (2017) studied the possibility of treating mining water using 1 L and 70 L electrocoagulation reactors and pressure filtration. The results indicated that metals were almost completely removed after 1 h of operation. The highest sulphate removal was obtained using aluminium electrodes after 5 h of treatment, while for nitrate after 3 h. Basically, nitrates were also almost completely removed just like metals. No significant effect of current density, reactor configuration and electrode material on metal removal was observed within the variable range investigated. However, sulphate and nitrate removal declined with the increase in treated volume. Cake formed during filtration had high porosity and moisture content.

Venkatasaravanan et al. (2016) investigated the efficiency of electrocoagulation process on the removal of heavy metals from synthetically prepared AMD. The electrocoagulation studies were performed in batch mode using vertically positioned aluminium electrodes (anode and cathode) in a 1 L reactor connected to a DC supply of 0-30 V and 5 A. The reactor was mixed at 200 rpm to avoid the mass transport over the potential of electrocoagulation reactor. As current density and pH were increased from 10 to 25 mA/cm2 and 5 to 7, respectively, the removal efficiency of Cu and Zn increased drastically from 57.6% to 95.4% and 53.4% to 86.2%, respectively. The energy consumption in the electrocoagulation process varied from 4.7 kWh/т3 to

29.4 kWh/т3 for current densities of 10 mA/cm2 to 25 mA/cm2. Clearly, this study proved that the electrocoagulation is an efficient process.

A study by Nariyan et al. (2016) investigated the removal of cadmium from real mine water by the electrocoagulation process using iron-stainless steel anode/cathode electrode combinations as well as aluminium-stainless steel anode/cathode electrode combinations. The effects of time, current density, and the type of electrode on the performance of electrocoagulation process were investigated. It was found that the current density had a direct effect on the removal of cadmium. In particular, cadmium was removed better at 70 mA/cm2 than at 10 mA/cm2. In addition, the reaction time had a direct effect on cadmium removal. For example, by increasing the time, cadmium was removed at higher removal rates compared to the beginning of the reaction. The type of electrode was also found to have an influence on the removal of cadmium. For example, cadmium was removed much better by an iron-stainless steel anode/cathode combination than by a combination of aluminium-stainless steel anode/cathode electrodes. The removal efficiency of the aluminium-stainless steel anode/cathode combination reached 82%, whereas the cadmium removal efficiency by iron-stainless steel was 100% at 120 min of reaction and 70 mA/cm2. The best condition in which 100% of cadmium was removed was obtained by using an iron-stainless steel anode/ cathode electrodes combination with a current density and reaction time of 70 mA/cm2 and 120 min, respectively.

The aim of the study by Orescanin and Kollar (2012) was to develop and apply the purification system suitable for the treatment of the AMD accumulated in the "Robule" Lake, which represents the part of the Bor copper mining and smelting complex in Serbia. The study was undertaken in order to minimise adverse effects on the environment caused by the discharge of untreated AMD, which was characterised with low pH value (2.63) and high concentration of heavy metals (up to 610 mg/L) and sulphates (up to 12 000 mg/L). The treatment of the effluent included pre-treatment/pH adjustment with CaO followed by electrocoagulation using iron and aluminium electrode sets. The results showed that the application of the electrochemical method for the treatment of the AMD pre-treated/pH adjusted with CaO resulted in extremely high removal efficiencies of heavy metals from the waste effluent of above 99% in most cases. The removal efficiency increased with increasing initial metal concentration. High degree of the removal of sulphates (over 70%) was also achieved. The concentration of iron in the treated effluent was reduced from 610 to 0.010 mg/L, copper from 82.5 to 0.006 mg/L, manganese from 59 to 0.336 mg/L, and zinc from 41.6 to 0.024 mg/L. The concentrations of other heavy metals in the final effluent were below 0.005 mg/L. The removal efficiencies for the metals and sulphates from the studied AMD are summarised as follows: S042- = 70.83%, Hg = 98.36%, Pb = 97.50%, V = 98.43%, Cr = 99.86%, Mn = 97.96%, Fe = 100.00%, Co = 99.96%, Ni = 99.78%, Cu = 99.99%, and Zn = 99.94%. Since the concentrations of heavy metals in the electrochemically treated AMD (ranging from 0.001 to 0.336 mg/L) are very low, the negative impact of this effluent on the aquatic life and humans is not expected. The waste sludge from the combined treatment process could be reused for the pH adjustment/pre-treatment of the AMD instead of CaO, and afterwards, due to its inertness, it could be used as an overlaying layer of the flotation waste heap during its recovery work. From the presented results, it could be concluded that electrochemical treatment is a suitable approach for the treatment of AMD.

A study by Oncel et al. (2013) compared the removal of heavy metals such as Fe, Al, Ca, Mg, Mn, Zn, Si, Sr, B, Pb, Cr, and As from coal mine drainage wastewater at a laboratory scale using chemical precipitation and electrocoagulation. The chemical precipitation was performed with NaOH, whereas the electrocoagulation process was evaluated via an electrolytic cell using iron plate electrodes. In the chemical precipitation process, the optimum pH for removal of most of the heavy metals from coal mine drainage wastewater was 8 except for Ca, Sr, and В (pH 10 or higher). The removal efficiencies at the optimum pH varied from 28.4% to 99.96%. Influence of current density and operating time in the electrocoagulation process was explored on the removal efficiency and operating cost. Results from the electrocoagulation process showed that the removal of metals present in coal mine drainage wastewater increased with increasing current density and operating time. The electrocoagulation process was able to achieve higher removal efficiencies (>99.9%) at an electrocoagulation time of 40 min, a current density of 500 A/m2 and pH of 2.5 as compared to the results obtained with the chemical precipitation at pH 8. The residual metal ion concentrations which varied from 0.00001 to 0.104 mg/L in the electrocoagulation process were below the limiting value for coal mine drainage wastewater discharge. The operating costs at the optimum operating conditions were also determined to be 1.98€/m3 for the electrocoagulation and 4.53€/m3 for the chemical precipitation. There is no doubt from the results that the electrocoagulation process was more effective than the chemical precipitation with respect to the removal efficiency, amount of sludge generated, and operating cost. The study by Oncel et al. (2013) has clearly shown that electrocoagulation has the potential to extensively eliminate disadvantages of the classical treatment techniques in order to achieve a sustainable and economic treatment of polluted wastewater.

Chartrand and Bunce (2003) performed the electrolysis of synthetic AMD solutions containing iron, copper, and nickel both singly and mixtures of the metals using a flow-through cell divided with an ion-exchange membrane. The anode used for all experiments was a dimensionally stabilised anode (DSA) consisting of titanium metal coated with iridium dioxide (=7.2 cm2), and the cathode was made of platinum. Each set of experiments was carried out at constant current densities at the highest value of which the steady- state pH of the exiting catholyte was about 10 for iron only solutions or 12 for all other solutions. The results showed that iron was successfully removed from a synthetic AMD solution composed of FeS04/H2S04 via Fe(OH)3 precipitation outside the electrochemical cell after oxygenation of the catholyte. The experiments with copper and nickel were only partly successful due to the fact that metal removal occurred more by hydroxide precipitation than electrocoagulation. The work was extended to an authentic AMD sample containing principally iron and nickel. Electrolysis of authentic AMD was successful in removing iron from solution, but quantitative removal of nickel required re-electrolysis or chemical precipitation. In the study, Chartrand and Bunce (2003) noted that the development of an electrolytic technology for AMD remediation requires more work on the chronology of electrolysis, aeration, and sludge separation, and on cell design so as to optimise mass transfer and permit the in situ separation of the sludges formed when the original AMD contains significant quantities of Fe3+.

Both mine water and industrial effluents are known antecedent of heavy metals (Singh and Mishra, 2017). Therefore, other studies that are important in the use of electrocoagulation technique involved the removal and/or recovery of heavy metals from wastewaters of other industries. For example, a study by Akbal and Camci (2011) investigated the applicability of an electrocoagulation method in the treatment of metal plating wastewater under various conditions. The results indicated that electrocoagulation can effectively reduce metal ions to very low levels. The metals (Cu, Cr, and Ni) were removed by precipitation as hydroxides by the hydroxyl ions formed at the cathode via water electrolysis (see reaction 9.11) and by co-precipitation with aluminium and iron hydroxides. The results also showed that Cu, Cr, and Ni removal efficiency increased with increasing current density. This was because higher current density increased the production rate of aluminium and iron hydroxide floes, and that the hydroxide floes subsequently acted as adsorbents for metal ions and thus removed the metal ions from the waste- water. The highest removal rate for Cu, Cr, and Ni was achieved at a pH of 9.0. The high efficiency of metal removal at higher pH levels was attributed to the precipitation of their hydroxides at the cathode. The Fe-Fe and Fe-Al electrode combinations were more effective for the removal of Cu, Cr, and Ni from the wastewater. The results also indicated that electrocoagulation with Fe-Al electrode pair was very efficient and was able to achieve removal rates of 100% Cu, 100% Cr, and 100% Ni at a current density of 10 mA/cm2 and pH of 3.0 after an electrocoagulation time of 20 min with energy and electrode consumptions of 10.07 kWh/т3 and 1.08 kg/m3, respectively.

A systematic study to ascertain the performance of an electrocoagulation system with aluminium electrodes for removing heavy metal ions (Zn2+, Cu2+, Ni2+, Ag+, Cr2072") on laboratory scale was carried out by Heidmann and Calmano (2008). Several parameters - such as initial metal concentration, numbers of metals present, charge loading, and current density - and their influence on the electrocoagulation process were investigated. The study was able to establish the removal mechanisms of Zn, Ni, Cu, Ag, and Cr. The results showed that Zn, Ni, Cu, and Ag are removed by direct reduction at the cathode surface, as hydroxides by the hydroxyl ions formed at the cathode via water electrolysis (see reaction 9.11) and by co-precipitation with aluminium hydroxides. It was proposed that Cr(VI) was reduced first to Cr(III) at the cathode before precipitating as a hydroxide, Cr(OH)3. The results from experiments with five metals indicate a co-precipitation of Cr with the other metals.

Kabda§li et al. (2009) experimentally investigated the treatability of a metal plating wastewater (concentration range, 230-280 mg/L) containing com- plexed metals originating from the nickel and zinc plating process through electrocoagulation technique using stainless steel electrodes. In this study, nickel and zinc were removed by hydroxide precipitation and incorporation in the colloidal material generated by the formation of Fe(OH)3 floes. The study by Kabda§h et al. (2009) demonstrated that the highest TOC abatement (66%) as well as nickel and zinc removals (100%) were achieved with an applied current density of 9 mA/cm2 to the original electrolyte (chloride) concentration and original pH of the composite sample used.

A study by Shafaei et al. (2015) investigated the removal of Mn2+ ions from solutions by an electrocoagulation process with aluminium electrodes. The study found that Mn2+ ion was removed by direct reduction at the cathode surface, as hydroxides by the hydroxyl ions formed at the cathode via water electrolysis and by co-precipitation with the aluminium hydroxides. It was found that the optimum initial pH for the removal of Mn2+ ions was 7.0. The results from the study also showed that increasing the current density and electrolysis time has a positive effect on the Mn2+ removal efficiency. The removal of Mn2+ ions was not influenced by the solution conductivity whilst the electrical energy consumption decreased with an increase in the solution conductivity. In addition, the study also found that as the initial concentration of the Mn2* increased, the rate of removal of the contaminant decreased.

A study that investigated electrocoagulation treatment using aluminium sacrificial electrodes for a synthetic water containing Cu, Cr, and Zn heavy metals was conducted by Singh and Mishra (2017). The effects of operational parameters, such as current density, inter-electrode distance, operating time, and pH, were studied and evaluated for maximum efficiency. This study showed that experimental results as well as kinetic modelling data gave high removal rate for all metals and total suspended solids at higher current density except Cu in which the same results were obtained at lower current density. In the case of energy consumption, it was concluded that 0.459 kWh/m3 was sufficient for the removal of 99% Cu, 59% Cr, and 71% of Zn up to 30 min of treatment time for which 0.450 A current was required. In the context of sludge generated, the study showed that the sludge that accumulated on the top layer of solution was more in comparison to amount of sludge at the bottom of the reactor. In addition, sludge generated was less at bottom at high current density.

9.2.5.2.2 Electroflotation

According to Feng et al. (2016) and many other studies (Azimi et al., 2017; Muddemann et al., 2019), an electrocoagulation process is often followed by an electroflotation process. A combination of electrocoagulation and electroflotation, which actually achieves better removal percentages, is called electrocoagulation-flotation process (Azimi et al., 2017). As already discussed, electroflotation is a simple process where pollutants are floated to the surface of a water body with the aid of tiny bubbles of hydrogen and oxygen gases generated from water electrolysis. Therefore, most of the examples in this section of the chapter pertain to studies where electroflotation was combined with other water treatment systems. In other words, electroflotation is mainly integrated into a process train with other technologies, particularly, electrocoagulation.

The possibility of the removal of metal ions from mining wastewaters through ion flotation was investigated by Alexandrova et al. (1994). Wastewater of an opencast mine containing about 50 mg/L copper ions underwent precipitation with xanthates forming chelate complexes with high hydrophobicity. Electroflotation was used to generate a gaseous phase with sufficient volume and high dispersity so as to effect precipitate flotation. Alexandrova et al. (1994) argue that effective precipitation and adsorbing colloid flotation require the presence of a gaseous phase with sufficiently large area and maximal dispersity due to the sensitivity of precipitated and coprecipitated particles to hydrodynamic conditions. Electroflotation proved appropriate in this study since it combined the electrocoagulation effect and electrolytic gas separation.

Khelifa et al. (2005) used electroflotation to reduce the concentrations of copper and nickel found in real wastewater. Through the electro-generation of gas bubbles (hydrogen and oxygen) at the electrodes and the variation of pH, this technique allowed the precipitation of hydroxides of the polluting metals by alkalisation and subsequent transport by flotation to the surface of the solution (Srinivasan and Subbaiyan, 1989; Alexandrova et al., 1994; Muddemann et al., 2019). The effects of the following parameters were examined: current density, pH, heavy metal concentration, supporting electrolyte concentration, and the nature of the electrodes. By optimizing the operation, heavy metal removal reached 98-99% and maintained final and global concentration to a value lower than the World Health Organisation standard, which is 1 mg/L for nickel and copper. In a later study, Khelifa et al. (2013) demonstrated the feasibility of simultaneous removal of heavy metals and EDTA in an electrolytic undivided cell equipped with Ti/Ru02 as anode and stainless steel as cathode. In the absence of EDTA, results showed that nickel and copper removal by electroflotation process is pH sensitive; and nickel and copper were substantially removed by electroflotation with removal efficiencies of 99.6% and 97%, respectively. In the presence of EDTA, the metal removal by the electroflotation process was inhibited. The inhibition rate was found to be dependent on EDTA/metal molar ratio. In the study, Khelifa et al. (2013) also used a one-step process, involving the combination of two techniques, i.e., electrochlorination and electroflotation. Active chlorine generated in situ allowed the decomplexation of metal-EDTA. As a result, free metal ions were removed by precipitating and subsequent floating to the surface by rising electro-generated bubbles. The obtained results revealed that, with 0.6 EDTA/metal molar ratio, removal efficiencies were 77% and 78% for nickel and EDTA, respectively, in the case of nickel-EDTA solutions. Removal efficiencies were 89% and 96% for copper and EDTA, respectively, in the case of copper-EDTA solutions. Furthermore, heavy metal removal efficiency by the combined process showed that it was affected by chloride content and current intensity.

The objective of a study by da Mota et al. (2015) was to remove Pb, Ba, and Zn ions from solutions containing 15 mg dm4 of each metal representing a typical concentration of wastewater from washing soil contaminated by drilling fluids from oil wells by electrocoagulation/electroflotation using stainless steel mesh electrodes. The effects of different parameters, including the pH, the electrolysis time, the current density, and the supporting electrolyte dosage were evaluated. The results of the study indicated that it is possible to remove Pb, Ba, and Zn metals by electrocoagulation/ electroflotation and achieving up to 97% removal efficiency with a power consumption of 14 kWh m~3. The optimal conditions of the treatment process were 0.1% sodium dodecyl sulphate (as a foamy) in a molar ratio against the heavy metals of 3:1, current density of approximately 350 Am 2, ionic strength of 3.2 x 10~3 M, pH of 10.0, and 20-min operation time. The results of the study indicated that the proposed electrocoagulation/electroflotation was adequate to simultaneously treat the common heavy metals found in the drilling fluids from oil wells.

9.2.5.2.3 Electrodeposition

Electrowinning, as a form of electrodeposition, is the deposition of a metal from solution due to an applied electrical potential. The dissolved metals migrate towards the cathode where they are reduced and deposit on the cathode. In general, the anode composition (effectively inert) is selected to limit its oxidation and thus promote the oxidation of water at the anode (Figueroa and Wolkersdorfer, 2014). Most importantly, electrochemical recovery processes, particularly electrowinning, have been applied primarily to high metal concentration solutions (»1000 mg/L) at low pH (<1) in waste streams where competitive metals tend to be at much lower concentrations than the target metal (e.g., electroplating waste) (Figueroa and Wolkersdorfer, 2014). Nordstrom et al. (2017) also reiterate that electrowinning for Cu is generally not done on AMD because concentrations of nearly 1000 mg/L are necessary.

In a study by Gorgievski et al. (2009), the removal of copper from AMDs originating from a closed copper mine "Cerovo" RTB Bor in Serbia containing approximately 1.3 g dm 3 of copper and a very small amount of Fe2+/ Fe3+ ions was successfully performed by direct electrowinning using either a porous copper sheet or carbon felt as the cathode. The cells used in the electrowinning experiments were compared in terms of cell voltage, pH, and copper concentration. A high degree of electrowinning, higher than 92% copper removal rate, a satisfactorily good current efficiency (>60%) and a good, dense metal deposit was obtained with both cathodes. However, the cell with the porous copper cathode had better features than that with the carbon felt cathode in terms of current efficiency and specific energy consumption. Depending on the process time and the applied current, a final copper concentration in the remaining solution of less than 0.1 g dm-’ was achieved. The specific energy consumption was approximately 7 kWh kg4 of deposited copper. A dense copper deposit was obtained when a three- dimensional electrode was used. The cell voltage decreased with time due to decreasing pH as a consequence of oxygen evolution as an anode reaction which increases the acid content. When the cells were compared with respect to the achieved cell voltage, it showed that the cell with the porous copper sheet cathode is the most suitable, having the lowest cell voltage. When using the carbon felt cathode, high cell voltage was registered due to the increased potential drop within the cell as well as the higher overvoltage of copper deposition onto carbon. The decrease in pH of the treated mine water with time due to the anodic oxygen evolution causes an increase in conductivity and acidity of the mine water. This fact may have a beneficial effect if the treated water is being recycled in a leaching stage, if present, leading to a decrease in the consumption of sulphuric acid. The opposite effect is an elevated consumption of chemicals needed to neutralise sulphuric acid formed during the electrowinning process prior to its release into a receiving water course.

Stankovic et al. (2008) ran experiments with the aim of generating data which would be needed in designing the mine waters' treatment process by direct electrowinning. The mine waters from the open pit mine were characterised with a reasonably high copper concentration (=1 g dnr3) and with a very low iron ion content allowing the consideration of a direct electrowinning as a method for the copper recovery from the mine waters. Zinc ions also existed in an unexpected amount of about 25 mg dm-3, but do not affect the electrowinning process. The removal of copper from mine waters was performed using two types of electrochemical cells: a cell with inert turbulence promoters and a cell with copper foam both operating in galvanostatic mode at different current densities and at different hydrodynamic conditions. There was no pre-treatment of the mine waters prior to the electrowinning process. Copper concentration was monitored periodically and the cell voltage was recorded thus allowing energy consumption in the process to be determined. The results of the study clearly showed that it was possible to remove copper successfully from the mine waters by direct electrowinning. The process achieved high degree of electrowinning and satisfactory current efficiency. Dense metal deposit was obtained in both electrolytic cells. Cell voltage was considerably high in both cells, but was much higher in the cell with inert turbulent promoters up to 12 to 14 V at the very beginning. However, the voltage fell down over time due to the lowering of pH with time as a consequence of the evolution of oxygen as an anode reaction that increases acid content in the water. Both cells used in the study were compared in a view of their applicability for treatment of mine water. Copper foam allows the application of higher geometrical current densities because of its developed internal surface. This allowed the application of higher operating currents on cells with copper foam and thus achieving lower final copper concentrations in the outlet stream compared to the cell with inert turbulence promoters.

The removal of copper and nickel from aqueous solutions using a simple tank cell, an improved mass transfer inert fluidised bed cell (Chemelec) and a three-dimensional high surface area cell (i.e., graphite packed bed cell) were examined by Campbell et al. (1994). A series of electrolysis experiments were carried out to examine the effect of current density and flow rate on both the efficiency of the process and the quality of the deposit. The packed bed cells were found to be extremely effective in reducing metal concentrations in the remaining solution to below 1 ppm. The most economic concentration range in which each cell should operate was also determined. The study found that the electrochemical recovery of nickel and copper over the concentration range of 20 g/L to less than 1 ppm can be carried out at high current efficiency if the concentration of metal ions is reduced in stages using the three types of electrochemical cells. At concentrations from 20 g/L to 2 g/L a tank cell should be used, 2 g/L to 0.2 g/L a Chemelec cell is required and below 0.2 g/L a graphite packed bed should be used followed by a high surface area cell to reduce the metal level to below 1 ppm. These results show that suitable combinations of cells can be chosen to remove the metals from high concentrations (20 000 ppm) to very low concentrations (<1 ppm).

Ubaldini et al. (2013) demonstrated the technical feasibility of electrowinning as a remediation process for toxic metals removal from AMD. In this study, high recoveries of metals were achieved of about 99% and 93% for Zn and Mn (as Mn02), respectively, with relatively low power consumptions (i.e., 118 kWh/kg for Zn and 619.05 kWh/kg for Mn02). The other metals such as Cu and traces of Cd, Ni, Mn (as metallic Mn) co-deposited with the Zn such that the levels of the remaining metals in solution were within the recommended limits suggested from the Peruvian law. The degree of purity of Zn was over 90%. Although all the studied metals deposited on the cathode, manganese deposited on the anode as Mn02 except traces of metallic Mn that co-deposited with Zn on the cathode. According to Veglio et al. (2003), the following reactions occur at the electrodes during Mn deposition in the electrowinning process:

In a comparative study of manganese removal from pre-treated AMD, Macingova et al. (2016) used three methods - precipitation using sodium hydroxide alkaline, oxidative precipitation using potassium permanganate, and electrowinning for anodic Mn recovery in MnO, form.

The results showed that the three methods are effective and manganese was removed from AMD with levels of the remaining metals in solution complying with environmental requirements. However, when sodium hydroxide was used as a reagent, co-precipitation of manganese and magnesium present in AMD was observed. Therefore, the alkalisation process did not show sufficient selectivity. Oxidative precipitation by potassium permanganate resulted in an enhanced selectivity and purity of the obtained manganese precipitates was achieved. In the process of electrowinning, over 95% of Mn was deposited as MnO, with a high-grade degree of purity of about 99% having been attained. With reference to reaction 9.14 and reaction 9.15, in the electrowinning process, manganese was deposited on the anode as Mn02 as already stated, while only a small amount was deposited on the cathode as Mn.

A number of studies (Marracino et al., 1987; Stankovic et al., 2008) have tested the attractiveness of metal foam electrodes for electrolytic treatment of dilute solutions containing metal ions. Panizza et al. (1999) used copper foams as high surface area cathodes in order to remove copper (II) from an industrial effluent that came from the filtration unit in the plant of copper phthalocyanine, which is produced by a discontinuous process from phthalic anhydride, urea, and cupric chloride in Italy. The electrochemical reactor used was an undivided monopolar cell. A series of experiments in a batch recycle mode were performed in order to find the trend of copper removal and current efficiency over time at different flow rate conditions. The influence of initial metal concentration on the removal of copper (II) and current efficiency was also evaluated. The results confirmed that copper foams can be successfully used as cathode materials for copper (II) removal from an industrial effluent. The study found that the best performance was obtained with the higher flow rate which is evident of a process under mass transport control. Working with a flow rate of 1000 L/h, a recovery of copper ions from feed solution greater than 98% was achieved. To obtain such a high copper removal, a very low current efficiency was used due to the great amount of organics present in the wastewater that could lead to the fouling of the electrodes. Furthermore, the study results indicated that the initial copper concentration had no influence on the rate at which metal ions were removed, but the current efficiency increased with initial copper concentration. In a similar and/or follow-up study, Panizza et al. (1999) used a cell that contained a series of 50 cathodes and 51 anodes (48 x 36 cm). The cathodes were vertical polymeric foam (0.6 cm thick) covered with a thin outer layer of copper, and the anodes were 02-evolving DSA® coated titanium mesh. The cell also had a provision for an air-sparging system to improve mixing of the solution. The cell can be considered as consisting of a cascade of 50 continuous stirred tank reactors (CSTR). The solution flowed through a porous cathode and the electrochemical reactions took place. Afterwards, the solution arrived in the chamber, between two cathodes, where it was mixed by the air sparging, before passing through the following electrode. A cascade with such a high number of CSTR can be well approximated with an ideal plug flow reactor (PFR).

The results of the study by Panizza et al. (1999) have shown that a tank reactor with metal foam cathodes can be approximated successfully to a PFR in order to predict its performance for copper removal from an industrial effluent. High fractional conversion was reached in a single pass, but it was necessary to use moderate flow rate and in this way there was a low cathodic current efficiency (e.g., at 100 L/h, the fractional conversion of copper (XA) was 86% and the current efficiency (r|) was 3.5%). An acceptable compromise was obtained at 500 L/h, having a current efficiency of 16.5% and a fractional conversion of 76.5%. Moreover, the presence of microcrystals of phthalocya- nine and sulphates ions in the effluent that was being treated in the study caused the poisoning of the cathode surface, thus increasing the cell voltage and consequently decreasing the current efficiency due to the secondary reactions. In addition, the quality of the copper deposit was poor since it contained copper salts and phthalocyanine that compromised the exploitation of the fully loaded cathodes and anodes in the plating processes. The study indicated that the foam cathodes could be regenerated for reuse by acid chemical stripping. Panizza et al. (1999) suggested that in order to improve the process, more attention should be addressed to the filtration unit before the electrochemical cell, so as to reduce the concentration of phthalocyanine and sulphates in the effluent that were the major causes of low current efficiency and the poor quality of the copper deposit. The study also showed that the presence of chloride ions did not influence copper removal, and no evolution of toxic chlorine was detected because of the use of 02-evolving anodes. This fact was confirmed as chloride concentration remained almost constant during the electrolysis.

9.2.5.3 Summary

Chen (2004) observed that electrochemical techniques have been used for wastewater treatment for over a century. More specifically, electrochemical processes have been utilised for many years to extract metals from aqueous wastes for reuse (Ali, 2011), and in the recent past electrochemical metal recovery has attracted increasing attention as compared to normal electrochemical metal removal (Jin and Zhang, 2020). There are a lot of electrochemical techniques being used for various purposes. However, this chapter focused particularly on electrocoagulation, electroflotation, and electrodeposition for the treatment and/or recovery of metals from wastewaters including AMD. According to Ali (2011), metal recovery via electrochemical techniques has the advantages that: (1) additional chemicals are not required, (2) selective metal recovery is possible taking into consideration thermodynamic and kinetic requirements of each species involved, (3) the metals can be recovered in their metallic form, and (4) the processes tend to operate at low temperature and pressure, specifically at room temperature and pressure of 25°C and 1 atm. In addition, electrochemical methods have attracted considerable attention because of their environmental compatibility, high efficiency, selectivity, versatility, feasibility, and cost effectiveness via the employment of the green redox reagent "electron" (Meunier et al., 2006; Jin and Zhang, 2020). Despite several advantages, there are also many bottlenecks associated with electrochemical metal recovery such as the concentration polarisation of dilute metal ions, the production of dendrites and spongy deposits, the sluggish kinetics of ions transportation, and side- reactions from the hydrogen evolution and oxygen reduction reactions (Jin and Zhang, 2020).

 
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