Both enhanced coagulation and carbon adsorption have been identified as Best Available Technologies by the USEPA, for organics removal from drinking water sources. Enhanced coagulation is a coagulation process optimized for TOC removal. However many water quality conditions affect TOC removal. Those parameters include alkalinity, pH, turbidity, TOC concentration, nature of NOM and temperature. The design and operation conditions include coagulant dose and type, pH, pre-oxidation, coagulation aids, mixing and mixing time, sedimentation process, and sludge handling.
The use of alternative disinfectants or combinations of disinfectants can effectively reduce the formation of regulated DBPs. However, as already stated, each disinfectant generates its own suite of DBPs and has its own limitations. The regulated DBPs have been historically selected as indicators of total DBPs formed when disinfecting water with chlorine (Richardson et al., 2007) but these only capture a fraction of all by-products and miss many others including those associated with alternative disinfectants.
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Anion exchange resins such as MIEXïƒ’ have been successfully employed in recent years for NOM removal, often prior to a coagulation step (e.g. Wanneroo GWTP, WA)(Smith et al., 2001). Anionic NOM is exchanged for chloride ions at the exchange sites of the resin, which can be easily regenerated once exhausted. Studies have reported improved removal of NOM using MIEXïƒ’ than traditional coagulation methods (Bond et al., 2011). A recent 2 year study found that pre-treatment with MIEXïƒ’ before either traditional coagulation or microfiltration produced water with a consistently lower DOC and SUVA than water without MIEXïƒ’ pre-treatment (Drikas et al., 2011). The inclusion of MIEXïƒ’ into the treatment train allowed a more consistent water quality that was less affected by changes in the raw water DOC, over the 2 years of the study.
Carbon adsorption is a process that can remove both DBPs and DBP precursors. Both GAC and PAC are commonly used in the water industry. Biological treatment such as use of biologically activated carbon can remove NOM from water sources by enzyme-controlled microbial degradation as well as adsorption. Importantly, DBP precursor removal by biological treatment is dependent on the biodegradability of the NOM present. A water with a highly biodegradable NOM content will be most suitable for biological treatment (Bond et al., 2011).
Advanced oxidation processes are those in which hydroxyl radical is produced in situ. Examples of these are O3/UV, UV/H2O2, O3/H2O2 and Fenton's reactions (Bond et al., 2011). All of these processes produce hydroxyl radical, which indiscriminately and quickly reacts with organic compounds, thereby degrading NOM. Advanced oxidation processes can in principle completely mineralise NOM to carbon dioxide, however, in practice, partial oxidation is the more economically feasible mode of treatment. Although NOM is degraded by these processes, specific DBPs can still be increased by these treatments, due to their precursors being formed by the degradation process.
The presence of bromide (Brâˆ’) and iodide (Iâˆ’) in source waters can result in the formation of brominated and/or iodinated DBPs upon exposure to NOM and disinfectant, which are often more toxic than their chlorinated analogues (Richardson, 2003b; Von Gunten, 2003; Magazinovic et al., 2004; Plewa et al., 2004b; Richardson et al., 2007). Both natural processes, including seawater intrusion and dissolution of geologic sources, and anthropogenic activities, such as seawater desalination, generation of mining tailings, chemical production, production of sewage and industrial effluents, may contribute to bromide concentrations in drinking water sources(Von Gunten, 2003; Magazinovic et al., 2004; Richardson et al., 2007; Valero, 2010). Similarly, seawater intrusion, seawater desalination and dissolution of geologic sources contribute to iodide concentrations in drinking water sources (Von Gunten, 2003; Hua et al., 2006; Agus et al., 2009), although biological activity of microorganisms and marine algae can contribute to iodide removal from water sources through specific metabolic processes (Suzuki et al., 2012). Error: Reference source not found shows typical bromide and iodide concentration in different source waters.
Current drinking water treatment schemes are challenged to effectively remove ambient bromide and iodide before final disinfection, in order to produce acceptable levels of the suspected carcinogen bromate (Kurokawa, 1990), when using ozone or advanced oxidation processes (von Gunten, 1995), and brominated/iodinated DBPs when disinfecting with chlorine or chloramines (Hua et al., 2006; Hua and Reckhow, 2007).
It is well known that bromide and iodide present in water may react differently with different disinfectants (Trofe, 1980; Kumar et al., 1986; Kumar and Margerum, 1987; Nagy, 1988). The kinetic rate constants of bromide and iodide with chlorine and chloramine in the formation of HOBr and HOI, respectively, are shown .
Always on Time
Marked to Standard
The reaction of bromide with free chlorine is five orders of magnitude faster than with chloramine (Trofe, 1980). On the other hand both disinfectants react relatively quickly with iodide ion (Kumar et al., 1986; Nagy, 1988). HOBr, formed from bromide, may then react with NOM and generate bromine-containing DBPs. In the presence of excess free chlorine, HOI is largely oxidized to iodate (IO3-), the desired sink for iodide, but in the presence of chloramines it is relatively stable, as shown by the relevant rate constants (Bichsel and Von Gunten, 1999). Therefore, in the presence of chloramines, HOI will react with certain organic precursors producing iodine substitution and the generation of iodinated analogues of many of the chlorine- and bromine-containing DBPs.
Although some brominated DBPs are regulated in the ADWG (NHMRC, 2011) (in particular; bromate, bromodichloromethane, dibromochloromethane and bromoform) the highly variable nature of NOM and its reactivity with different disinfectants means that there may be many other brominated and/or iodinated DBPs species formed in any given treated water in which these halides are present. Strategies for DBP minimisation vary, but can be broadly classified into three categories; DBPs precursor removal (halides and NOM), optimising disinfection to minimise DBPs formation, and DBPs removal prior to water distribution (for example, air-stripping of volatile DBPs such as THMs) (Wu and Wu, 2009). One advantage that DBPs precursor removal has over other DBPs minimisation strategies is that it is not specific to removing/lowering a particular suite of DBPs, it can broadly minimise all DBPs, both known and unknown, potentially creating greater trust in the quality of the water produced. All disinfection methods produce their own suite of DBPs, however, minimising the precursors available for this to occur is applicable regardless of disinfection method employed. As water regulation become increasingly stringent and salinity impacted water sources are increasingly utilised there may be a need for effective bromide and iodide removal to control the formation of emerging DBPs.
The objective of bromide and iodide removal is to control the formation of brominated and/or iodinated DBPs (both organic and inorganic). Bromide and iodide removal techniques can be broadly classified into three categories, namely; membrane, electrochemical and adsorptive techniques.
Over the past 50 years, membrane technologies have become a distinguished separation technology with significant commercial applications in the water industry. With increasing water demands and diminishing water supplies due to escalating populations, environmental degradation and climate change, membrane technologies are being employed to produce high quality potable water from impaired and alternative water supplies. Membrane techniques comprising reverse osmosis (RO), nanofiltration (NF), ion exchange membranes, electrodialysis (ED) and electrodialysis reversal (EDR) are discussed in the following sections.
RO is a process whereby water is forced through a semi-permeable membrane under pressure to remove organic contaminants and salts, producing purified water (Figure : Schematic of the RO process. Note that while this illustrates the desalination process, other contaminants present in water including NOM and small organic molecules can also be rejected by RO membranes.). There are 2 main types of RO membranes; cellulose acetate and more recently thin-film composite membranes, each with differing water flux, rejection and physiochemical characteristics (Escobar, 2010). Most commercially successful RO membranes are thin-film composites with a top ultra-thin active filtration layer that consists most commonly of cross-linked polyamide (50-200 nm) but other polymers such as piperazine and others are employed as well. This thin layer is backed by an intermediate porous polysulfone support and a grid of polyester fibres to provide the desired mechanical stability (Robert J, 1993). Some membranes also have a surface coating that yields a more hydrophilic, neutral and fouling-resistant surface (Petersen, 1993; Hachisuka, 2002; Tang et al., 2007b; Coronell et al., 2008). The active layer of RO membranes is the main barrier against the permeation of salt and contaminants which results from a combination of equilibrium partitioning at the water/active layer interfaces, and diffusive and advective transport (Petersen, 1993; Wijmans and Baker, 1995; Mulder, 1996; Urama, 1997; Paul, 2004; Tang et al., 2007b).
Figure : Schematic of the RO process. Note that while this illustrates the desalination process, other contaminants present in water including NOM and small organic molecules can also be rejected by RO membranes.
The broad spectrum of solute rejection of RO membranes allows the utilisation of seawater, brackish water and reclaimed water as alternative potable water sources (Xu, 2005; Asano, 2007). A comprehensive review of the early history of RO membrane development is available (Glater, 1998). Recent developments in drinking water treatment applications of RO have been discussed by many including Greenlee et al. (Greenlee et al., 2009), Fritzmann et al. (Fritzmann et al., 2007), Pearce et al. (Pearce et al., 2004), Veza (Veza, 2001), Antrim et al. (Antrim et al., 2005), Koutsakos and Moxey (Koutsakos and Moxey, 2007), Redondo and Lomax (Redondo and Lomax, 1997, 2001), Wolf et al. (Wolf et al., 2005), Vial et al. (Vial et al., 2003) and Laborde et al. (Laborde et al., 2001).
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RO membranes typically have a salt rejection of 98-99%, although several recently developed membranes can achieve 99.7-99.8% salt rejection under standard conditions (Bates and Cuozzo, 2000; Reverter et al., 2001; Wilf, 2004; Asano, 2007; Fritzmann et al., 2007; Hydranautics, 2007; Reverberi and Gorenflo, 2007; Tang et al., 2007a; Colquhoun, 2010). The current body of literature indicates that RO membranes reject monovalent halide ions to differing degrees. The rejection of ionic solutes by RO membranes has been observed to follow the Hofmeister series, that is, rejection increases with increasing hydrated radius. Ions with high hydration energy (determined by its hydrated radius and charge density) are superiorly rejected due to their low partition coefficient with the hydrophilic membrane (Naaktgeboren et al., 1988; Pontié et al., 2003a). The halide rejection occurs in the following sequence: F > Cl > Br > I. This is identical to the order of decreasing hydration energy and decreasing hydrated radius of the ions.
A study by Magara et al. (Magara et al., 1996) investigating bromide removal from seawater using RO with spiral modules of aromatic polyamide, achieved over 99% bromide rejection. Similarly, several other studies of RO have found high bromide rejection, ranging from 93-99.3%, depending on the membrane used, and operating conditions (Drage et al., 2000; Mays, 2004; Khan et al., 2005; Bartels et al., 2009). Khan et al. found that RO membranes display a high rejection of iodide ions (71-90%) (Khan et al., 2005). RO can reject over 89% of the total iodide from seawater (Duranceau, 2010). The viability of several RO membranes in treating oilfield-produced brackish water for water reuse and iodide recovery has been investigated (Xu et al., 2008a; Drewes et al., 2009). The aromatic polyamide membranes were ranked according to the rejection of iodide as: Koch TFC-HR (92%) > Toray TMG-10 (87%) > Koch TFC-ULP (82%) > Dow/Filmtec XLE (80%). Furthermore, Koch TFC-HR, Koch TFC-ULP and Toray TMG-10 membranes rejected between 98 and 100% of both chloride and bromide ions (Drewes et al., 2009). Dow/Filmtec FT-30 membranes achieved 99.55% rejection of bromide at 2 MPa and 99.50% rejection of iodide at 5 MPa (Pontié et al., 2003b).
Table : Physical characteristics of the halide ions.
Electronegativity (Pauling scale)
Molar hydration enthalpies
(kcal/mol) (Weiss, 1995)
Non-hydrated radius (Å) (Nightingale, 1959)
Hydrated radius (Å) (Nightingale, 1959)
RO is a proven technology for the removal of an extensive range of contaminants, and this was the most effective bromide and iodide removal technique of all investigated. Importantly, this technique can remove both organic and inorganic DBP precursors simultaneously making it invaluable in DBP minimisation. However, RO remains relatively expensive, requires extensive pre-treatment, has high energy consumption due to high operating pressures, and is susceptible to scaling, as well as brine disposal difficulties. The capital and operational expenses of RO as well as the disposal of the generated concentrate currently restricts the widespread application of this technique in drinking water treatment plants.
NF is a pressure driven membrane process, which is an intermediate between RO (non-porous diffusion) and ultrafiltration (porous-sieving) membrane processes, and exhibits features of both (Wang et al., 2008). Compared to RO, NF generally runs at somewhat lower pressures, thereby reducing energy costs. It also has less resistance to the flow of both solvents and solutes (Sarrade et al., 1998; Water Environment Federation, 2008; Mohapatra et al., 2009). Although there are many types of membranes, most applications of NF use polyamide thin-film composite membranes in a spiral wound configuration (Asano, 2007). Like RO membranes, most NF membranes are charged, thus electrostatic interactions also influence the transport and selective rejection behaviour of NF membranes (Wang et al., 2008). Consequently, depending on the molecular weight cut-off of the membrane, many NF membranes can effectively separate both NOM, small organic molecules (such as pesticides, pharmaceuticals and endocrine disrupting compounds) (Grib et al., 2000; Jiraratananon et al., 2000; Kiso et al., 2001; Lee et al., 2002; Mänttäri et al., 2002; Agenson et al., 2003; Hong and Bruening, 2006; Benitez et al., 2009; Acero et al., 2010) and salts from water. Salts rejected include bromide and iodide ions, although in general, polyvalent ions may be more easily retained than monovalent ions (Bougen et al., 2001; Tanninen and Nyström, 2002; Qin et al., 2004; Hilal et al., 2005; Al-Zoubi et al., 2007).
Harrison et al. (Harrison et al., 2007) investigated the ion rejection of two commercial NF membranes (NF-90 and NE-90) in a seawater desalination application (Error: Reference source not found). The NF-90 membrane achieved 94-96% bromide rejection and 84-91% iodide rejection. Similarly, NE-90 membrane accomplished 94-97% bromide rejection and nearly 90% iodide rejection. The authors found that salt rejection generally increased with pressure, until their maximum rejection threshold was reached, beyond which, rejection started to decline with increasing pressure. This is to be expected and holds for all RO and NF membranes. A maximum rejection threshold for bromide was observed at 3.4 MPa. Anion rejection was not significantly affected by water temperature under the conditions tested
Lhassani et al. (Lhassani et al., 2001) investigated the selective demineralization of water by NF with particular emphasis on monovalent anions. The study achieved a maximum rejection of iodide (80%) with NF70 membrane at a pressure of 1200 kPa. Although the membrane selectivity favoured chloride over iodide ion, it was demonstrated that this can be reversed at pressures over 800 kPa, thus showing that operating conditions can be adjusted to selectively remove individual ions of the same valence. Pontie et al. (Pontié et al., 2003a) and Diawara et al. (Diawara et al., 2003) demonstrated that tight NF membrane NF-45 can selectively separate monovalent halides, following the Hofmeister series of halide ions: F> Cl> I. The membrane reflection coefficient and solute permeability were observed to be correlated with the hydration energy of halide ions. Tight NF-45 membrane was shown to have mass transfer properties similar to RO membranes.
NF membranes with a molecular weight cut off of 150-300 Da have been observed to reject up to 50% of bromide and bromate from low-turbidity source waters (Amy and Siddiqui, 1998). Prados-Ramirez et al. (Prados-Ramirez et al., 1995) saw a 63% rejection of bromide upon conducting NF on bromide-spiked river water. Pontié et al. (Pontié et al., 2003b) found NF-70 membranes reject approximately 94% of bromide at a pressure of 1.5 MPa. The study demonstrated that halide ions are transferred across NF membranes by two mechanisms; convective and solubilisation-diffusion. Pontié et al. (Pontié et al., 2003b) also demonstrated that under low pressure, retention of halide ions follow the Hofmeister series, however under high pressure, chloride and bromide order is inverted. Drewes et al. (Drewes et al., 2009) investigated the salt and iodide rejection from oilfield-produced brackish water utilising three NF membranes: NF-90 (Dow/Filmtec), TFC-S (Koch) and ESNA (Hydranautics). The membranes were ranked according to the rejection of iodide, as: NF-90 (78%) > TFC-S (69%) > ESNA (55%). The NF-90 membrane rejected over 80% of both chloride and bromide (Drewes et al., 2009). In contrast, Listiarini et al. (Listiarini et al., 2010) and Chellam and Krasner (Chellam, 2000; Chellam and Krasner, 2001) found that NF was inefficient for bromide removal, however, the operating conditions investigated, including pressure, permeate flux, anion concentrations and water type, varied significantly from those used for successful applications of NF for bromide removal.
When compared to RO, NF has slightly lower capital costs, significantly lower operational costs due to lower operating pressures, can be operated at a higher water recovery, which means a smaller waste concentrate stream, while achieving comparable bromide and iodide removals (Harrison et al., 2007). Due to these advantages, the application of NF has increased, especially in industrial applications and drinking water treatment. This membrane technique still experiences (although to a lesser extent) the same limitations as RO. NF requires extensive pre-treatment, has medium to high energy consumption, and is susceptible to scaling and brine disposal difficulties. Again, like RO, NF has the advantage of having capacity to remove both organic (NOM) and inorganic (halide) DBP precursors simultaneously.
Electrodialysis and electrodialysis reversal
The ED process uses a driving force of direct current (DC) to transfer ionic species through cell pairs of oppositely charged membranes, allowing their separation from the source water (Error: Reference source not found) (AWWA, 1995; Valero, 2010). The degree of salt removal is directly proportional to the current and inversely proportional to the flow rate through each cell pair (Scott, 1995). Thus, hydraulic and electrical staging are used to achieve the desired salt removal. The hydraulic and electrical staging in the membrane stack array configuration used in ED water treatment systems is determined by the source water quality and the level of water quality required (Valero, 2010). Membranes are composed of a polymer matrix with charged groups attached, with pores that allow ions to permeate. EDR is a modification of the ED process, where electrode polarity is periodically reversed during the treatment process to reduce scaling and clean membrane surfaces (AWWA, 1995; Valero, 2010). EDR has been successfully used for desalination, waste treatment, treatment of boiler feed and process water, and hardness removal, providing a reliable and economical alternative to RO of brackish waters (van der Hoek et al., 1998; Spiegler and El-Sayed, 2001). The EDR process is used as a 'finishing' treatment for desalting brackish water with TDS concentrations of up to 4000 mg/L. After 4000 mg/L the energy costs dramatically increase, reducing its competitiveness with other membrane techniques such as RO (Seneviratne, 2007). The halide rejection of ED and EDR has been summarised (Error: Reference source not found).
Recently, Valero and Arbós (Valero and Arbós, 2010) conducted a 28 month pilot plant study of EDR, as well as a study of the implementation of EDR in Llobregat's desalination drinking water treatment plant (DWTP) in Barcelona, Spain. The pilot plant study utilised the same source water as is treated at Llobregat's DWTP. Anthropogenic activities severely impact the quality of this source water, which consequently has substantial concentrations of salts, NOM and micropollutants, causing elevated concentrations of DBPs during the drinking water treatment process (160 Â± 40 Âµg/L THMs). EDR was incorporated into the water treatment system after GAC treatment, in both pilot and full-scale studies. On average the pilot study achieved over 75% bromide, 60% chloride, 65% electrical conductivity (EC), 58% total alkalinity (TAC), 30% TOC, 75% NO3âˆ’, 70% SO42âˆ’, 80% Ca2+, 70% K+ and 80% Mg2+ reduction, while maintaining over 90% water recovery. The improved chemical quality of the water resulted in a decrease in tTHMs formation potential (THM-FP) to 64 Â± 60 Î¼g/L, which was below the regulated concentration of 100 Î¼g/L. During the trial the systems' performance was highly robust and reliable with no operational problems experienced (Valero and Arbós, 2010). Llobregat's DWTP (with full-scale application of EDR) has the capacity to produce an average of 200,000 m3/day, making it the world's largest desalination plant using this technology. The plant obtained water quality parameters and THM-FP consistent with the pilot study values. Optimisation of the plant achieved a maximum of 80% bromide and 80% EC reduction, while maintaining over 94% water recovery. Furthermore, Valero and Arbós showed EDR treatment allows the control of THM speciation as a result of Br- removal, resulting in a reduction of brominated species in favour of chlorinated species.
Van der Hoek et al. (van der Hoek et al., 1998) investigated the possibilities of applying EDR as an alternative for RO filtration in three integrated membrane systems (IMS) that use ozone for disinfection. The IMS-1 system used pre-treated (coagulation/sedimentation/filtration) Rhine River water and further treated sequentially with ozonation, biologically activated carbon filtration, slow sand filtration and RO; IMS-2 treatment was the same as IMS-1, however the RO step was replaced with an EDR step; IMS-3 took pre-treated Rhine River water then treated sequentially with EDR, ozonation, biologically activated carbon filtration and slow sand filtration. IMS-1 and IMS-3 were able to achieve compliance with bromate regulation, however, IMS-2 could not. IMS-3 reduced bromide concentrations by 72% (prior to the disinfection step), thus reducing bromate formation to 5 Âµg/L after ozonation (van der Hoek et al., 1998), in compliance with the European Union drinking water standard (Lenntech Water, 2009 ). The authors found IMS-3 (using EDR) had lower energy and chemical consumption and thus lower operational costs when compared to IMS-1 (using RO) however, RO provided a dual barrier for disinfection and removal of organic compounds such as NOM and pesticides, which is an advantage over the EDR process (van der Hoek et al., 1998). EDR membranes AR204-SZRA (anion) and CR67-HMR (cation) both manufactured by Ionics were used in this study.
Permselectivity of ions is determined by ion exchange selectivity and mobility selectivity (steric and membrane density effects) with a membrane (Sata, 2004). Generally, ions with higher valence and a smaller hydrated radius have a higher permeability in an ion exchange membrane than ions with lower valence and larger hydrated radius. Hann et al. (Hann et al., 1983) found the permselectivity counterion exchange sequence of an anion-exchange membrane containing quaternary ammonium groups as fixed charges was: I- > NO3- > Br- > Cl- >SO42- > F-. Sata and colleagues have also conducted several studies on the effect of hydrophilicity of anion-exchange membranes on the permselectivity of specific anions in ED (Sata, 1994; Sata et al., 1998a; Sata et al., 1998b; Sata, 2000). The group showed that permselectivity of specific anions is mainly dependent on the balance between hydration energy of anions with hydrophilicity of the membranes. Specifically, the more hydrophobic membranes have a higher permselectivity for less hydrated (higher hydration energy) ions. In keeping with this, increasing hydrophobicity of the strongly basic anion-exchange membrane enhanced the permeation of bromide and iodide, while decreasing the permeation of fluoride ions. As expected, increasing the hydrophilicity of the membranes reversed this trend.
Inoue et al. has carried out several investigations into the removal of radioactive iodide ion (125I) from wastes using ED with anion-exchange paper membranes (Inoue and Kagoshima, 2000; Inoue, 2001, 2002, 2003, 2004; Inoue et al., 2004). These membranes were found to be electroconductively more permeable to iodide than to chloride ions, allowing iodide to be concentrated from the feed stream (Inoue and Kagoshima, 2000). In a further project, covalent linking of glucose and urea to membranes was reported to increase the membrane/solution distribution of iodide, however the diffusion process of iodide within the membrane was not significantly altered, relative to the unmodified membrane (Inoue, 2001). Inoue et al. also worked on membrane separation control using three different anion-exchange groups bound to a pulp/cellulose fibre matrix: trimethylhydroxypropylamino, diethylaminoethyl and 50% saturated quaternary diethylaminoethyl (Inoue, 2002). It was observed that the membrane permeability for iodide was higher than chloride in all three cases. Notably, trimethylhydroxypropylamino groups improved the iodide diffusion process, whereas 50% quaternary diethylaminoethyl groups improved the iodide solution/membrane distribution process. High iodide permselectivity was achieved in membranes with the trimethylhydroxypropylamino quaternary amine anion-exchange group, due to electrostatic effects (Inoue, 2003; Inoue et al., 2004).
Relative advantages and disadvantages of the membrane techniques
The ED/EDR processes are not commonly used in drinking water treatment plants although examples of successful implementation do exist (Pequignot and Rigaudeau, 2007; Valero and Arbós, 2010). The main advantages of ED compared to other membrane techniques are; minimal pre-treatment of feed water is required, higher water recovery can be achieved than for RO, although process recoveries for EDR should be similar to NF (Valero, 2010). EDR membrane life would be expected to be higher than RO membrane life (7-10 years for EDR membranes versus 5-7 years for RO membranes) (Valero, 2010). EDR has several technical and economic limitations, including high energy consumption and high capital cost (Strathmann, 2010). ED/EDR has the potential to be widely applied to brackish and anthropogenically-impacted waters to enable the utilisation of alternative sources of water, however, more research is needed in the membrane development and in the optimization of ED/EDR for large scale drinking water treatment plants. A limitation of ED/EDR processes is that unlike RO and NF, they do not remove neutral, organic DBP precursors, they exclusively remove ionic species, i.e., hardness, halides, and other salts. Therefore, the use of this technology without an additional organic matter removal step may decrease the level of bromination/iodination of organic DBP precursors during disinfection, but may not decrease the total concentration of DBPs (e.g. tTHMs) formed. Furthermore, ED and EDR do not provide any disinfection, unlike NF and RO, however, these techniques can be used in the presence of a chlorine residual, thereby enabling disinfection and reducing biological fouling of membranes.
Electrochemical techniques have been successfully used to remove contaminants from various industrial wastewaters and environmental waters (Pletcher and Walsh, 1990; Rajeshwar and Ibanez, 1997). Electrochemical techniques comprising electrolysis and capacitive deionization (CDI) are briefly discussed in the following sections.
Electrolysis is a process that uses the passage of an electric current through a solution to induce chemical decomposition (Hale, 2008). Historically, electrolysis processes have been used to produce bromine from brines containing bromide (Kimbrough and Suffet, 2002; Society for Mining Metallurgy and Exploration US, 2006; Black & Veatch Corporation, 2010). Kimbrough and Suffet (Kimbrough and Suffet, 2002) examined the feasibility of using electrolysis to remove bromide from drinking water sources (Error: Reference source not found). Carbon rod cathodes and dimensionally stable anodes (DSA) were used in the electrolysis cell. Electrolysis of raw water oxidised bromide to a mixture of hypobromite, hypobromous acid and bromine gas. Degassing this solution with carbon dioxide caused a decrease in the solution pH, driving the conversion of hypobromite to hypobromous acid, which was then volatilised, along with bromine gas, leading to a decrease in the solution concentration of bromide. The rate of bromide removal is dependent on the applied current and effectiveness of air stripping. Upon chlorination the electrolysed water produced lower THMs concentrations and decreased the proportion of brominated THMs formed. The authors suggest that the removal of bromide should also reduce the formation of other brominated DBPs.
An electrolytic process that combines disinfection with the removal of bromide from raw water was patented by Bo (Bo, 2008). Electrolysis cells used silver cathodes and DSA arranged in parallel fashion. During electrolysis chloride was oxidised to chlorine gas, providing disinfection, while bromide ions were oxidised to bromine gas which volatilised without requiring stripping. High bromide removal efficiency is achieved due to the high silver conductivity and large electrode surface contact area coupled with several passes of the solution through the electrolysis cell. The effectiveness of this process is dependent upon the chloride:bromide ratio, the magnitude of the electrical current, ionic strength of the water, distance between electrodes, water residence time in the cell and electrode material. Bromide removal efficiency decreased with lower influent bromide concentrations. At low bromide concentrations (<125 Âµg/L) the process achieved between 48%-62% removal, but at high bromide (>200 Âµg/L) concentrations bromide removal increases to between 47%-79%. Brominated THMs concentrations (generated by THM-FP)) were reduced by 27% after the initial bromide concentration was reduced from 461 Âµg/L to 48 Âµg/L during the treatment. This reduction in Br-THMs may be expected to be greater, given the excellent bromide removal reported, however, the authors report losing chlorine residual during the THM FP of the high bromide sample, so they may not have formed a maximum of Br-THMs during the experiment. Additionally, several electrolysis processes that remove bromide by producing bromine gas from different aqueous sources, including brines, bittern and waste effluents have been patented (Nidola, 1978; Sharma, 1978; De Nora, 1983; Williams, 1993; Hawley, 1994; Howarth, 1995; Blum, 1999; Bonnick, 2002; Kroon, 2003; Ramachandraiah, 2004).
Although electrolysis has been shown to remove bromide reasonably on a small scale the feasibility of large scale application to drinking water treatment has not been assessed. Further development of electrodes would be required as part of working toward larger scale treatment.
CDI is a recently developed electrolysis technology for removing ionic species from aqueous solutions using porous activated or aerogel carbon electrodes (Error: Reference source not found) (Gabelich et al., 2002; Ying et al., 2002). The deionisation process occurs by an induced electrical potential difference across an aqueous solution, which flows in between oppositely charged porous electrodes. As a result of the applied electrical potential, ions are adsorbed in the electrodes, deionising the product stream (Oren, 2008). Although CDI technology is in its infancy, it has the potential to develop into a feasible low-cost alternative to membrane and thermal desalinisation of brackish waters (Error: Reference source not found) (Welgemoed and Schutte, 2005; Xu et al., 2008b; TDA Research, 2010). The CDI process operates at ambient conditions of temperature and pressure, requires minimum pre-treatment, does not require chemicals for scaling control or chemical cleaning, has low voltage requirements and a low fouling/scaling potential (Xu et al., 2008b). The adsorption capacity of carbon aerogels is dependent upon the surface characteristics of the electrodes, including; surface area, size and microstructure of pores, electrical conductivity, chemical composition and electrical double-layer capacity (Yang et al., 2001; Gabelich et al., 2002; Ying et al., 2002). Halides are usually removed from solution by electrostatic attraction within the electrode, whereas large polyvalent oxyanions, heavy metals, and colloidal impurities can be removed by means of chemi/physisorption, electrodeposition, electrophoresis, double-layer charging and possibly faradaic reactions, as well as simple electrostatic interactions (Pekala et al., 1994; Yang et al., 2001). Electrode polarity is reversed after saturation to regenerate carbon aerogel electrodes (Tran, 2004; Welgemoed and Schutte, 2005; TDA Research, 2010).
Gabelich et al. (Gabelich et al., 2002) and Ying et al. (Ying et al., 2002) report that an ion's hydrated radius may regulate the ionic species' sorption capacity into carbon aerogel electrodes. Monovalent ions with smaller hydrated radii were preferentially removed from solution over multivalent ions. Additionally, counterion valency appears to have a strong influence on an individual ion's sorption capacity (Gabelich et al., 2002). Experiments conducted with natural waters showed the sorption capacity of carbon aerogels was significantly lower when a high concentration of NOM was present. The authors suggest pre-treatment for NOM removal would increase effectiveness of CDI in treating natural waters (Gabelich et al., 2002). CDI was found to increase in total capacity for anion removal in the order of Clâˆ’ <Brâˆ’< Iâˆ’, thus this technique could be used most effectively to selectively recover iodide from solutions (Ying et al., 2002). Ying et al. (Ying et al., 2002) theorised this effect was due to iodide ions having a partial charge-transfer coefficient larger than bromide and chloride ions. The adsorption capacity of the carbon aerogel electrodes was shown to increase with increasing solution concentration of ionic species, voltage, and surface area of the electrodes. Due to its selectivity, this technique has the potential (with further development) to be applied in the removal of halides from drinking water sources.
Welgemoed and Schutte (Welgemoed and Schutte, 2005) developed an industrial CDI bench scale prototype (MK-8A) and evaluated the module's performance for coal-bed methane brackish water desalination. The module achieved high ionic species reduction in artificial brackish water, reducing the feed stream conductivity from 1000 ÂµS/cm to 23.4 ÂµS/cm at a flow rate of 50 mL/min. Interestingly, in the artificial system bromide had the highest reduction percent of all ions monitored (86.11%). The prototype was then tested on naturally occurring brackish water from the natural gas industry in Wyoming, USA. It reduced the feed stream conductivity from 2095 ÂµS/cm to <1000 ÂµS/cm at approximately 70% water recovery rate. Rinse brine was recycled, reducing the volume of waste brine produced. Furthermore, the authors compared the costs of RO, EDR and CDI for brackish water desalination to a potable water standard. CDI could be significantly cost effective compared to RO for brackish water applications (CDI: US$0.11/1000 L, RO: US$0.35/1000 L). CDI could also reduce brackish water desalination costs by 70% when compared to existing EDR technologies. A water recovery rate of 70% while still retaining quite a high conductivity (2095 ÂµS/cm to <1000 ÂµS/cm) is, however, quite inferior to what could be achieved with RO, i.e. organic precursor removal and disinfection, as well as halide removal.
The viability and ion selectivity of CDI technology in treating brackish water generated during natural gas mining for water reuse and iodide recovery has been investigated (Xu et al., 2008b). CDI testing units used in this study were provided by CDT Systems, Inc. The anion sorption capacity of the carbon aerogel (in mol/g aerogel) was dependent upon initial ion concentrations in the feed water. That is, the ions present in the greatest concentration were adsorbed to a greater extent, following the order Cl- â‰¥ Br- > I-. However, the maximum percentage of removal for these anions followed the opposite trend; I- > Br- > Cl-, which the author's attributed to iodide's higher partial charge-transfer coefficient compared to the other anions, and intermolecular interactions with the carbon-aerogel electrodes, resulting in a higher sorption capacity. In concurrence with the findings of Ying et al. (Ying et al., 2002), bench scale tests showed preferential sorption of iodide from brackish water even in the presence of dominant coexisting ions. The removal of iodide reached 69% and removal of bromide reached 50% in artificial water (Xu et al., 2008b). During the regeneration phase, 77-107% of sorbed iodide was recovered from the carbon aerogels (a recovery >100% was attributed to the desorption of iodide from the previous run). The authors found bench scale and pilot scale CDI cells exhibited similar sorption capacities. During the pilot study the maximum removals observed were 83% of UV254, 77% of I-, 62% of Br-, 40% of Ca2+, 40% of alkalinity (as CaCO3), 34% of Mg2+, 18% of Na+ and 16% of Cl-. However, field experiments employing a three stage CDI treatment could not meet the water quality standards for reuse due to high total dissolved solids (TDS) concentrations. To resolve this issue additional CDI stages were applied to simulate a multi-stage desalination treatment. Ten CDI stages were needed to reduce TDS to an acceptable level for reuse.
Shiue et al. (Shiue et al., 2005) have enhanced the efficiency of CDI by using spiral wound electrodes (activated carbon coated on titanium foil) in combination with online electrolytic ozonation. This allows the reduction of uncharged constituents and disinfection of water by ozone, in addition to the removal of charged species by CDI. Ozone was produced by low-voltage electrolysis of water. The electricity retrieved at the discharging of CDI operation could be harnessed for use in producing ozone, thus increasing energy efficiency.
CDI technology is a promising alternative for brackish water desalination, although the operational performance and sorption capacity of the electrodes may need further development before the technology becomes economically feasible (Xu et al., 2008b).
Membrane Capacitive Deionization
Membrane Capacitive Deionization (MCDI) is a modification of the CDI process, in which ion-exchange membranes are added onto a CDI system (Figure : Schematic showing the membrane capacitive deionisation (MCDI) process.) (Li, 2008; Biesheuvel, 2009). Ion-exchange membranes are positioned in front of their corresponding charged electrodes. MCDI has several advantages over CDI namely; the membranes inhibit ions from leaving the electrode region, thereby increasing the salt removal efficiency of the process, and ion release from the electrode region (during electrode regeneration) is more efficient (Biesheuvel, 2009). Both Lee et al. and Li found that MCDI had a higher salt removal rate than traditional CDI systems, with 19% and 49% higher rejection found, respectively (Lee et al., 2006; Li, 2008). Although no specific studies on bromide and iodide removal were found using this process, MCDI would be expected to produce improved halide removal efficiency to that experienced with CDI, and is thus an area of potential future research.
Figure : Schematic showing the membrane capacitive deionisation (MCDI) process.
Relative advantages and disadvantages of the electrochemical techniques
The CDI and MCDI processes are not currently used in the treatment of drinking water. The CDI process is both robust and energy efficient, although MCDI may be expected to be able to remove halides with greater efficiency than CDI. With further development CDI has the potential to be applied to drinking water, wastewater, boiler water and coal seam gas water deionisation, as well as brackish water desalination (Oren, 2008). Conversely, the potential for large-scale application of electrolysis to halide removal during water treatment may be limited, due to difficulties in scaling up the process, although this has not yet been explored. Areas in which further development would be expected prior to wide-spread use of CDI technology are related to the optimisation of deionisation, commercial development of aerogels, full scale application and commercialisation of the technique.
Although membrane techniques can successfully reduce bromide and iodide concentrations in water, surface sorption methods form a major component of halide reduction research and application because of their ease of application and low cost. Recent research has continued to explore the development of low cost, effective, bromide and iodide adsorbents. Sorption techniques comprising; hydrous oxides, activated carbons, silver-doped activated carbons and carbon aerogels, ion-exchange resins, aluminium based adsorbents and soils, are briefly discussed in the following sections.
Layered double hydroxides
Layered double hydroxides (LDHs) (also called hydrotalcite-like compounds (HTCs)) are attracting considerable attention for their ability to selectively remove contaminants in aqueous systems (Villa et al., 1999; Pavan et al., 2000; Inacio et al., 2001; You et al., 2001; Barriga et al., 2002; Seida and Nakano, 2002; Das et al., 2003; Das et al., 2004; Yang et al., 2005; Duan and Evans, 2006; Lv et al., 2007; Mandal and Mayadevi, 2008). LDHs consist of positivity charged metal hydroxide layers, with interstitially located anions and water molecules (Reichle, 1986; Goh et al., 2008; Mandal and Mayadevi, 2008). They have large surface areas and numerous sites for anion exchange, making them ideal ion-exchangers and adsorbents. Additionally, LHDs are produced from low-cost precursors that can be easily regenerated (Pavan et al., 2000; Das et al., 2003; Mandal and Mayadevi, 2008). A summary of studies investigating halide removal using LDH's is reported (Error: Reference source not found).
Curtius and Kattilparampil (Curtius and Kattilparampil, 2005) studied the application of Mg-Al-Cl LDH for 135I- removal from radioactive wastes. It was found that the adsorption of iodide was independent of pH between 3.5 and 8.5. In the test parameters studied the sorption capacity for iodide decreased with increasing chloride concentrations. A Kentjono et al. (Kentjono et al., 2010) study using Mg-Al-(NO3) LHD found that the optimum pH for iodide removal was 9.0-9.2. The maximum iodide adsorption capacity achieved was 10.1 mg/g at a LDH dose of 20 g/L and pH of 9.2. The optimum pH for iodide adsorption coincides with that of boron adsorption, however they do not compete with each other's adsorption, indicating that this LDH could be used to simultaneously remove both boron and iodide. The iodide removal capacity of thiosulfate intercalated Zn-Al LDH has been reported by Thomas and Rajamathi (Thomas and Rajamathi, 2009). Approximately 60% of iodide ions were intercalated in the interstitial layer of the LDH, making it a potentially useful treatment for the removal of iodide from drinking waters.
The Lv group investigated the influence of LDH calcination temperatures and the Mg:Al molar ratio on the adsorption of bromide from water by Mg-Al LDHs (Lv and Li, 2007; Lv et al., 2008). The bromide and iodide adsorption capacity of calcined LDHs is higher than that of uncalcined LDHs. Adsorption capacity of bromide and iodide increased with increasing calcination temperature between 200 oC and 500 oC, however it drastically decreased with calcination temperatures from 500 oC to 800 oC. A maximum bromide and iodide adsorption capacity of 94.0 mg/g and 96.1 mg/g, respectively, was achieved with LHD calcination at 500 oC. It was found that a Mg:Al molar ratio of 4 had the highest capacity to remove bromide and iodide from aqueous solution. Additionally, the group found that increasing adsorbent concentrations from 0.2 g/L to 1.0 g/L significantly increased bromide removal from 73.4% to 91.6%, after which the removal plateaus off to a maximum of 94% using 5 g/L (Lv et al., 2008). Similarly, adsorbent concentrations from 0.2 g/L to 1.0 g/L significantly increased iodide removal from 39.4% to 96.5%, after which the removal plateaus with a maximum of 97.6% using 4 g/L (Lv and Li, 2007). LHDs maintained similar bromide removal capacities after five regeneration cycles (Lv and Li, 2007; Lv et al., 2008).
A preliminary study on the effect of layered hydroxides metal composition on iodide sorption was conducted by Pless et al. (Pless et al., 2007) Uncalcined layered hydroxides containing Cu2+ exhibited the highest sorption for iodide, followed by Ni2+ and Co2+. The authors found calcination decreased the sorption of iodide, however calcination temperatures were only 550oC for 1 or 24 hours and only a limited number of layered hydroxides were tested.
Echigo's group investigated the removal of bromide from a real water matrix by LDH's to control the formation of brominated DBPs in the drinking water treatment process (Echigo et al., 2007). Two LDH's, Mg-Fe-LDH (ratio 4:1) and Mg-Al-Fe-LDH (ratio 8:1:1) were compared to a commercially available gel-type polystyrene-divinylbenzene quaternary amine anion-exchange resin; Diaion SA10A. An approximately 60% reduction in bromide using a real water matrix was achieved using both LDH's, whereas an approximately 73% reduction in bromide was achieved with the Dianion resin. The LDH's selectivity sequence was shown to be HCO3- >> NO3- > Br- > SO42- and hence LDH's were found to be better for bromide removal in a water matrix with high sulfate ion and low bicarbonate and nitrate concentrations due to their selectively for bromide under these conditions and faster ion exchange reactions in comparison to Diaion SA10A. However, in the presence of bicarbonate, bromide removal was impaired due to the LDH's preferential adsorption of this anion rather than bromide. This may limit the extent to which these LDH's could be used in bromide removal from drinking water sources, since they would not be expected to be useful under high bicarbonate conditions. Organic carbon (measured as TOC) was not removed by either LDH, indicating it was rejected by the ion-sieve effect of the LDH's. This is important since organic carbon comprises many DBP precursors, so although bromide may be efficiently removed (in low alkalinity conditions), the organic DBP precursors would be expected to remain. The performance of LDH's was found to be comparable to Diaion SA10A in terms of the treatment volume and the bromide uptake before breakthrough for low alkalinity waters. The authors conclude that the application of LDH in the drinking water treatment process would provide similar performance to organic resins such as Dianon SA10A without the potential for secondary contamination.
A Zn-Al LDH adsorbed approximately 14% of iodide from a deionised solution but almost none from a mineralised solution containing 1 mM Clâˆ’, 15 mM SO42âˆ’, and 5 mM HCO3- at an initial KI concentration of 0.01M at both pH 7 and 10 (Kaufhold et al., 2007). Thus, in a multi-ion solution LDH exhibited preferential selectivity of bicarbonate and sulfate over iodide, in agreement with several other studies (Miyata, 1983; Israeli et al., 2000; Kaufhold et al., 2007). The usefulness of LDHs for halide removal from water is therefore dependent on the nature of the other anionic species present in solution.
Despite this limitation, LDHs have been shown to effectively remove bromide and iodide from real water matrices. Several LDHs' anion sorption performance was comparable to commercially available resins in terms of the treatment volume and anion adsorption before breakthrough. Additionally, ion exchange reactions were shown to be faster using LDH's than commercial ion-exchange resins, and no concerns about secondary contamination when using LDH's have been raised, unlike many commercial ion-exchange resins. The preferential adsorption of bicarbonate over bromide ions, and both sulfate and bicarbonate over iodide ions, for particular LDH's is an important limitation, however, further investigation into the large scale application of promising LDH's is warranted, given they are a prospective low-cost treatment technology, which has the potential for widespread application in drinking water treatment.
Sol-gel double hydrous oxide
In a recent development, the sol-gel method was used to synthesize an inorganic ion exchanger based on a double hydrous oxide (Fe2O3Â·Al2O3Â·Ã-H2O) (Chubar et al., 2005). Adsorption behaviour of fluoride, chloride, bromide, and bromate ions was investigated by varying experimental parameters including time, pH and adsorbant concentrations. The ion exchanger exhibited both cation and anion-exchange capacity, which reached values of 1.38 and 1.8 mEq/g, respectively. Adsorption of selected anions was observed over the pH range 3 - 8.5, with maximum bromide adsorption occurring at pH < 5. Kinetic data on bromide sorption fit a pseudo-second-order model, with a rate coefficient of 0.16 minâˆ’1. Within the first 10 min of treatment 50% of the bromide ions were adsorbed. The maximum bromide sorption capacity achieved was 80 mg/g at bromide concentrations of 200 mg/L (Error: Reference source not found). Investigation into competitive adsorption of Br- and BrO3- at equal concentrations found that at concentrations over 40 mg/L bromide and bromate, competition for adsorption sites favoured bromide, whereas for lower concentrations of bromide and bromate, bromate dominated adsorption. A number of other novel sol-gel double hydrous oxides have recently been developed, and have been shown to have bromide removal capability (Chubar, 2011).
Further investigation into the ion-exchange interactions with organic matter and complex water matrices under field conditions is needed before the suitability of these adsorbents for drinking water treatment can be more thoroughly evaluated. Further research should also include the development and study of new types of sol-gel double hydrous oxides.