Arsenic (As) is abundant in the environment and can be found in both organic (e.g., methylated) and inorganic (e.g., arsenate and arsenite) forms. The source of As in the environment is attributed to both natural reactions and anthropogenic activities. As can also be released naturally to groundwater through As-bearing minerals including arsenopyrites, realgar, and orpiment. Similarly, agricultural and industrial activities have elevated As levels in groundwater. High levels of As in groundwater pose serious health risks and have been regulated in many developed and developing countries. In particular, the presence of inorganic forms of As in drinking water sources gained widespread attention due to their cellular and enzyme disruption activities. The research community has primarily focused on reviewing the natural occurrence and mobilization of As. Yet, As originating from anthropogenic activities, its mobility, and potential treatment techniques have not been covered. This review summarizes the origin, geochemistry, occurrence, mobilization, microbial interaction of natural and anthropogenic-As, and common remediation technologies for As removal from groundwater. In addition, As remediation methods are critically evaluated in terms of practical applicability at drinking water treatment plants, knowledge gaps, and future research needs. Finally, perspectives on As removal technologies and associated implementation limitations in developing countries and small communities are discussed.
Das et al. (2004) also proposed oxidation of arsenic-rich pyrite as one of the mechanisms that explain arsenic mobilization in the aquifer. 52 The reduction of iron oxyhydroxides and adsorption competition on the surface of iron oxides are other reasons for arsenic mobilization. 34 , 53 , 54 Liao et al. examined the relevant microbial community in arsenic-contaminated aquifers and indicated that a relatively diverse community of microorganisms can biotransform arsenic in the aquifer, 55 via detoxification and energy conservation reactions. 56 Most studies have identified methylation, demethylation, oxidation, and reduction as the primary microbial processes involved in arsenic transformation and mobilization. 37 , 57
Oremland and Stolz (2003, 2005) reported that certain prokaryotes use oxyanions of arsenic for their energy generation via As(III) oxidation or As(V) reduction. 37 , 51 These phylogenetically diverse microorganisms can mobilize aquifer arsenic from the solid to the aqueous phase ( ). These researchers advocated that the chemical and biological arsenic mobilization and speciation processes may be simultaneous. These microorganisms play important roles in the oxidation and reduction of not only arsenic but also iron mineral present in the aquifers. A conceptual model for the arsenic problem in Bangladesh aquifers has also been proposed. 37 , 51 It is important to note that aquifers in Bangladesh are considered highly contaminated with arsenic and more than one-quarter (40 million+) of its total population is exposed to arsenic-contaminated drinking water. The conceptual model developed for Bangladesh aquifers proposed that chemoautotrophic arsenite oxidizers (CAOs) and heterotrophic arsenite oxidizers (HAOs) first oxidized As(III)-containing minerals (e.g., arsenopyrite) leading to deposition (adsorption) of As(V) on the oxidized minerals (e.g., ferrihydrite). The overpumping of an aquifer to meet drinking and irrigation needs resulted in lowering of the groundwater tables and has led to the further oxidation of As(III) by oxidants (e.g., molecular oxygen and (agricultural) nitrate). Over time, the buildup of organic matter and biomass causes the aquifers to become anoxic. Such organics and forms of peats can lead to the dissimilatory reduction of earlier converted As(V) and dissolution of arsenic adsorbents like ferrihydrite, where dissimilatory arsenate-respiring prokaryotes (DARPs) participate. Thus, arsenic is released to the aqueous phase ( ).
The overall mobility of As in the aquifer is primarily driven by the adsorption and oxidation reactions of arsenic at the mineral water interface. 47 While As(V) is prevalent in the aqueous aerobic environment, an anoxic environment favors the presence of As(III). Adsorption, desorption, and biological transformation influence arsenic mobility in an aquifer. As(V) has affinity for some common elements and is easily bounded (adsorbed) to ferrihydrite and alumina. 37 In anoxic aquifer, As(III) is stable and dissolves in water as iron and manganese complexes. 48 A recent study reported that manganese (Mn) can impact iron and arsenic solubility in rice paddy soil. 49 As(III) often exists as neutral (uncharged) H 3 AsO 3 species at environmentally relevant pHs (pK a for H 3 AsO 3 is 9.2). For typical groundwater pH (i.e., pH lower than 9.2), As(III) is less likely to interact with mineral surfaces, making it more mobile in the environment and of concern. Additionally, any arsenic otherwise coprecipitated and sorbed onto metal oxides (iron or manganese) may be reintroduced to the aqueous phase when the metal oxides are dissolved. However, in anoxic zones and in the presence of sulfide, As(III) is immobilized due to the formation of orpiment, realgar, or arsenopyrite or is coprecipitated with iron pyrite. 50
Oremland and Stolz (2003) have proposed four possible ways for the mobilization of subsurface arsenic: 37 (i) arsenic oxidation in pyrites that contain As, (ii) the release of As(V) from iron oxide reduction by autochthonous organic matters (e.g., peat leading to the release of As(V)), (iii) iron oxides reduction by allochthonous organic matters present in, for example, recharging waters, and (iv) displacement of As(V) by chemical analogue phosphate (PO 4 3– ) present in fertilizers. Kurz et al. also reported that the arsenic mobilization in Vietnam’s groundwater is attributed to both oxidation and adsorption reactions, where arsenic is immobilized onto freshly formed iron hydroxides due to oxygen. 45 As mobilization has also been reported in the upper Jhelum basin, located in the western Himalayas, and is attributed to the reduction of Fe and Mn oxy/hydroxides. 46
Above a pH of 9, As(III) exists as an uncharged neutral molecule (H 3 AsO 3 ), whereas As(V) occurs as an anionic molecule (H 2 AsO 4 – , HAsO 4 2– ) above pH 2 ( ). 22 , 39 , 40 The transition from As(III) to As(V) involves a thermodynamically favorable process, but the transformation period can range from days to years, depending on the specific environmental conditions. For example, environmental parameters including strong acidity and alkalinity as well as higher temperatures can enhance the oxidation process. 20 Low groundwater redox potential favors the formation of As(III) species. On the other hand, As(V) is the most thermodynamically stable state of arsenic in oxidizing environments (redox potential above 100 mV). 41 Recently, Huq et al. reported that the high levels of arsenic in the shallow Holocene alluvial aquifers in Bangladesh are attributed to mobilized arsenic due to strong reducing and moderate alkalinity conditions. 42
In nature, As is present in organic, inorganic, and gaseous forms, but the inorganic form is relevant for immediate public safety. Inorganic arsenic can be present in the environment in four oxidation states (−3, 0, +3, +5) depending on the pH and redox conditions ( ). 8 , 20 As(V) or arsenate (H 3 AsO 4 0 , H 2 AsO 4 – , HAsO 4 2– , AsO 4 3– ) is the dominant species in aqueous aerobic environments, while As(III) or arsenite (H 3 AsO 3 0 , H 2 AsO 3 – , HAsO 3 2– ) is prevalent in anoxic environments. 20 , 36 − 38
The source of As-contaminated groundwater is predominately subsurface geologic formations containing arsenic. While As may find its way to the environment because of anthropogenic reasons, the significant contribution to water contamination comes from arsenic-bearing minerals in the subsurface of the earth. 22 , 30 − 32 Even though the earth’s crust only contains 0.0001% As, As is also found in virtually all rocks and sediments. 22 , 33 However, most geogenic As in contaminated groundwater comes from relatively recently deposited alluvial sediments. 27 The release of As depends on the redox and pH conditions, where under anoxic conditions, As leaches due to the dissolution of As-bearing iron minerals in the aquifer. As is released from iron and aluminum hydroxides under oxidizing and high-pH conditions. 19 , 27 , 34 , 35 In addition, low hydraulic gradient aquifers tend to have higher residence time leading to high concentrations of dissolved arsenic in groundwater. 27 Moreover, arsenic buildup in waters can be attributed to the release of arsenic from arsenic-bearing sulfide minerals (via oxidation)-enriched geothermal deposits. 27
Several techniques have been proposed to remove arsenic from groundwater including precipitation, coagulation/filtration, adsorption, ion exchange, lime treatment, oxidation, and membrane filtration. 8 However, the broader implementation of these technologies is challenged by cost and complex operation and maintenance, making them less desirable for remote and challenged communities. Yet, significant research efforts have been spent on identifying appropriate aqueous arsenic removal technologies that are cost-effective, easily operated, and require minimal experience, while having high arsenic removal rates. This paper reviews the geo-chemistry, occurrence, mobilization, and microbial interaction of arsenic in the aquifer. Various technologies potentially appropriate for small, challenged, rural, and remote communities are discussed, and the pros and cons of some of the technologies are highlighted. Finally, research gaps and potential research and development opportunities are discussed.
To protect human health, the World Health Organization (WHO) and the U.S. EPA have set a maximum contaminant level (MCL) of 10 μg/L for inorganic arsenic in drinking water. 15 , 16 To date, elevated arsenic levels in drinking water sources have been reported in more than 50 countries, affecting well over 200 million people. 17 , 18 Many aquifers worldwide have been identified with arsenic concentrations greater than 10 μg/L, 19 − 24 and these aquifer hotspots include Argentina, Bangladesh, Cambodia, China, India, Mexico, Nepal, and many parts of the U.S. ( ). 22 , 25 − 27 In many of these countries, groundwater was promoted as safe for drinking given the reduced likelihood of microbial contamination. 28 Even though microbial contamination was not present, groundwater sources in these countries were contaminated with naturally occurring subsurface arsenic. Further, arsenic is odorless, tasteless, and colorless, making arsenic detection in contaminated water a challenge even when present at high concentrations. 29
Arsenic (As) has been widely used in medicinal and industrial applications. Yet, the health risks associated with arsenic exposure have not been recognized until the 20th century. 1 As present in various environmental matrices (i.e., water, soil, and air) and food can be attributed to both natural and anthropogenic sources. 2 More importantly, arsenic is the most toxic naturally occurring groundwater contaminant. 3 , 4 In fact, arsenic, particularly in its inorganic form, has been reported as a carcinogen and identified as highly problematic in drinking water. 5 , 6 It has been estimated that at least 150 million people globally drink water with elevated levels of arsenic. 7 , 8 Yet, according to the International Agency for Research on Cancer (IARC), arsenic is a Group 1 human carcinogen. 9 , 10 The U.S. Environmental Protection Agency (U.S. EPA), the U.S. National Toxicology Program, and the American Conference of Industrial Hygienists have also classified arsenic in their list of cancer-causing agents. 11 Aside from lung and skin cancer, 12 , 13 arsenic exposure also leads to ailments of the stomach, intestine, skin, respiratory system, kidney, and central nervous system. 14
The conventional arsenic removal processes include chemical-aided sedimentation, coagulation and flocculation, adsorption, ion exchange, and membrane processes.57−59 The advantages and limitations of these existing technologies have been previously reported in detail.60 In fact, high arsenic removal rates can be achieved with these process (80–100%). While these processes work well for municipal water utilities and resourceful communities, they may be cost-prohibitive when implemented at a small scale for rural communities. Furthermore, these practices are only efficient in removing As(V), and sometimes a pretreatment step (i.e., peroxidation unit process) is needed to oxidize As(III) to As(V) for effective arsenic removal.58 As for the bench scale, the laboratory-based arsenic removal techniques include arsenic precipitation with Fe/Mn, lime softening, electrodialysis, ion exchange, adsorption on activated alumina/carbon, and others.61 In the next section, the state-of-the-art methods for the removal of arsenic from water are discussed based on their relevance for potential use in small and remote communities as well as potential sustainability aspects.
Arsenic may be mainly present in groundwater as As(III) and/or As(V) depending on solution pH and redox conditions.22 As(III) is generally more mobile than As(V), so arsenic immobilization can be facilitated by As(III) oxidation to As(V).20,62 The selection of appropriate oxidant materials for the oxidation of As is important and must consider the potential residuals of the oxidant used, oxidation byproducts, oxidation of other present water constituents, and the amount of the oxidant required (i.e., economics, worker safety, and storage requirement).
To date, the most feasible and common oxidants for As(III) are chlorine, potassium permanganate, ozone, manganese dioxide, hydrogen peroxide (in Fenton’s reagent and involving Fenton-like reactions), and persulfate.63 UV oxidizes As(III) to As(V) very quickly, generally within seconds, and UV irradiation can be achieved using high-pressure mercury arc discharge lamps with a 190–254 nm emission spectrum. Similarly, ozone-based oxidation is very fast and often used in surface water treatment for oxidation and disinfection.64 When used for arsenic oxidation ( ), ozone reportedly achieved more than 96% oxidation of As(III) to As(V) within 10 min. Both UV and ozone require high-energy inputs, which may not be cost-effective for small-scale water treatment plants. Given their abundance, oxygen, chloride, and permanganate deserve additional discussion (eqs 1–3).
1
2
3
Open in a separate windowWhile As(III) can be oxidized into As(V) by atmospheric oxygen (eq 1), the process is very slow and can take weeks.65 When pure oxygen was used, a maximum oxidation of 57% occurred in 5 days.66 Chlorine is a widely deployed disinfectant in most water systems across the globe and can oxidize arsenite to arsenate. Hypochlorite is a popular source for chlorine that oxidizes As(III) to negatively charged As(V) (eq 2). The use of chloramine and chlorine dioxide has also been reported.67 Similarly, KMnO4 (α-MnO2) can oxidize As(III) to As(V) (eq 3). Manganese dioxide has also been used by some small drinking water systems for As(III) oxidation.68
The Fenton’s reagent process has also been studied for oxidation of arsenic. The ferric ions formed in the process lead to precipitation of As(V). However, the residual hydrogen peroxide should not exceed the maximum permissible limit of hydrogen peroxide (i.e., 1 mg/L) to avoid potential toxicity for the consumers. Krishna et al. reported that adding 100 μL of H2O2 and 100 mg/L of Fe(II) (as ferrous ammonium sulfate) followed by adsorption (medium: zerovalent iron (ZVI)), with a contact time of 10 min, is capable of reducing As(III) from 2 mg/L to below 10 μg/L ( ).69
Open in a separate windowSome researchers have demonstrated that reactive oxidizing species (e.g., hydroxyl radical, ·OH) can be produced during the photolysis of nitrite (NO2–) or nitrate (NO3–) and can convert As(III) to arsenate As(V). UV photolysis of NO2– and NO3– promotes the breakage of the O–N bonds producing intermediate products that can serve a dual purpose: photochemically transform contaminants in an aqueous medium and oxidize chemical and microbial contaminants in water. The main reactions involved in this UV/As(III)/ (NO2–) or (NO3–) system have been described in detail by Kim et al.70 Briefly, the photolysis of NO2– or NO3– mainly generates nitrogen oxide radical and oxide radical anion, which in turn undergoes protonation to form ·OH. Even though As(III) was not directly oxidized via UV photolysis, adding 20 or 200 μM nitrite promoted the oxidation of As(III) under UV irradiation (λ > 295 nm), making NO2– a photosensitizer for As(III) oxidation.70 Another study reported the use of glow discharge plasma (GDP) to simultaneously oxidize and immobilize organoarsenic. The results showed that Roxarsone (ROX) (an organoarsenic compound) can be effectively oxidized to inorganic arsenic, but the process was energy-intensive. The addition of Fe(II) significantly enhanced the oxidation of ROX, mainly due to the additional ·OH generated via Fenton reaction in GDP coupled with simultaneous arsenic immobilization. The immobilization of arsenic was favorably obtained at a pH range of 4–6 and a 500–1000 μM Fe(II) concentration range.71
Future oxidation–reduction studies should focus on providing additional information about the rates of arsenic reactions and determining their specific rate constants.
Coagulation is commonly used in drinking water treatment to destabilize dissolved and suspended solids allowing their aggregation to form flocs, which are subsequently removed via sedimentation.72 The process can be described in three steps: (i) coagulant addition and colloid destabilization (rapid mix), (ii) interparticle collisions and colloid agglomeration (coagulation), and (iii) colloid removal via floc sedimentation (last phase of coagulation and then settling). To treat arsenic-laden water, coagulants are added, which, in turn, form arsenic-containing flocs of low solubility, which then settle and are removed by sedimentation. While calcium salts73 and lanthanum salts74 have been used in the water treatment industry, alum (aluminum sulfate),75−77 ferric chloride, and ferric sulfate are the most widely used coagulants for arsenic removal (eqs 4–6).78−83 Alum is the preferred choice by many water treatment units.
4
5
6
The possible reactions of arsenate with hydrous iron oxide (≡FeOH0 represents the oxide surface site)84 are listed below (eqs 7–9).
7
8
9
Immobilization of arsenic by hydrous iron oxide is another process of interest (eqs 7–9) that is highly influenced by the pH of the solution.72 In one study, using a high dosage of FeCl3 (i.e., 300 mg/L), arsenic was highly reduced to less than 5 μg As/L in the coagulated water.85 Using the same coagulant, only 50–60% As(III) removal was achieved.86,87 Under pH 5–6, As(V) adsorption onto hydrous iron oxide and goethite was higher than that of As(III), whereas, above pH 7–8, As(III) adsorption was more favorable to these adsorbents.88 In fact, in mineral processing effluents, lime neutralization and coprecipitation of arsenate with iron has been implemented for the removal and immobilization of arsenic.89 Arsenic coagulation with iron, aluminum, and zirconium salts has also been reported.90
A recent study reported an efficient and fast arsenic removal technology known as air cathode assisted iron electrocoagulation (ACAIE) enabled with in situ generation of hydrogen peroxide. During ACAIE operation, Fe(II) is readily oxidized and As(III) was completely removed. These authors demonstrated that within a short time frame (30 s), and at a charge dosage rate of 1200 Coulomb/L/min, ACAIE consistently outperformed iron electrocoagulation (FeEC) by reducing arsenic levels to below the 10 μg/L WHO level ( ).91
Open in a separate windowA recent study investigated a mixture salt composed of cactus mucilage and ferric (Fe(III)) salt to remove arsenic from water. Mucilage treatment improved As removal compared to only Fe(III) treatment, where the system achieved an As removal efficiency of 75–96% within 30 min.92
Arsenic can also be removed during lime softening processes using calcium hydroxide (Ca(OH)2) and soda (Na2CO3). Similar to other processes, the arsenic removal efficiencies are enhanced if arsenic in the water is oxidized from As(III) to As(V). A recent survey on 516 full-scale water treatment plants93 reported that hard water tends to have a higher concentration of arsenic compared to soft water. However, up to 90% of the soluble As(V) could be potentially removed if a high softening pH was used to precipitate magnesium. McNeill and Edwards (1997) reported that 90% arsenic removal can be achieved via precipitation softening with iron.94 pH and the presence of chlorine govern the removal of arsenic in the lime softening process. For example, only 15% arsenic removal was possible in nonchlorinated water containing 400 μg/L arsenic at a pH less than 10.5. However, when the pH was adjusted to 11, arsenic removal reached 80%.95 Sorg and Logsdon (1978) also demonstrated that pH impacts the removal of arsenic during lime softening. Low As(III) and As(V) removal efficiencies were achieved at pH values less than 10, but As(V) and As(III) removal approached 100% and 75%, respectively, at pH values greater than 10.5.96
Activated carbon, activated alumina, zerovalent iron, iron oxides, and clay are among the adsorbents that have been deployed to remove arsenic from water. Rapid small-scale column testing (RSSCT) procedures for arsenic were developed similar to those used for adsorption of organic constituents by granular activated carbon (GAC). However, it is important to note that there are significant differences between adsorption of organics by GAC and adsorption of arsenic by metal oxides. The most prominent differences are the nature of the bonds and the internal pore structure of the adsorbents. Hydrophobic bonds are dominant between organic solutes and GAC, whereas arsenic adsorption onto metal oxide surfaces is initially attributed to electrostatic attraction. RSSCT allows full-scale results to be obtained in a short time with a fragment of the processed water and samples collected and analyzed.97 Thomson et al. used RSSCT to investigate arsenic adsorption using activated alumina, ferric oxyhydroxides, and a proprietary medium developed by Sandia National Laboratories.98
Activated alumina (AA) or granular ferric oxides and hydroxides are the optimal adsorbents due to their abundance, high arsenic removal efficiency, and ease of operation.99 Research studies have also evaluated industrial byproducts, montmorillonite clay, and other activated carbon-based adsorbents to remove arsenic from aqueous solutions, which exhibited low adsorption capacities. Low-cost natural adsorbents such as Moringa oleifera were investigated with favorable arsenic removal results (>60% removal of As(III) and >80% of As(V)) and regeneration potential.100 Magnetic biochars prepared by chemical co-precipitation of Fe2+/Fe3+ on a biomass (i.e., water hyacinth) followed by pyrolysis showed that it is a highly efficient adsorbent for aqueous As(V) removal.101
Prasai et al. studied the interactions of organoarsenic compounds with ferric and manganese binary oxide (FMBO) to elucidate the mechanism that governs adsorption, and the adsorption performance was compared to that of ferric oxide (FeOOH) and manganese oxide (MnO2).102 The maximum adsorption capacities (Qmax) of FMBO toward aromatic organoarsenic compounds including p-arsanilic acid (p-ASA) and roxarsone (ROX) were 0.52 and 0.25 mmol/g, respectively, at pH = 7.5, which were higher compared to those of FeOOH and MnO2. The corresponding Qmax values of p-ASA were 0.40 and 0.33 mmol/g, and those for ROX were 0.08 and 0.07 mmol/g, respectively.102
Activated alumina (AA) is considered to be an adsorbent material even though ion exchange takes place.103 A series of cerium oxide composites of AA have been prepared and employed as adsorbents for arsenic removal from water. The adsorption capacity of these composites (Al2O3–CeO2) depended on the preparation conditions. The highest. adsorption capacity of arsenic was achieved when CeO2 was deposited on powdered AA (6 g) at a 0.5 mol ratio of H2O2/Ce. More specifically, arsenic adsorption was optimal when the nanosize CeO2 was deposited on the Al2O3 support, where a Langmuir adsorption isotherm model was followed with Qmax of 13.6 and 10.5 mg/g for arsenate and arsenite, respectively ( ).104
Open in a separate windowRegeneration of saturated alumina can be achieved by exposing the medium to a strong base (e.g., NaOH). Using pilot filters, Frank and Clifford (1986) demonstrated that As(III) is not removed by activated alumina but As(V) is adsorbed, and breakthrough may occur at 10 000–20 000 bed volumes (As initial concentration of 100 ppb and pH = 6).105 shows the removal kinetics on alumina at a pH of 6. Giles et al. reported the use of iron and aluminum-based adsorbents for As removal and further discussed the oxidation of As(III) to As(V) using various oxidizing agents to facilitate As adsorption (see Section 2.1).106
The adsorption of arsenic by iron oxides is a very important process because it occurs naturally in the environment and is highly favorable under oxidizing and slightly acidic conditions. It can also be engineered and configured as a removal mechanism during water treatment. Granular ferric oxide Bayoxide E33 and granular ferric hydroxide GFH are iron-based media frequently used as adsorptive media in small drinking water systems in the U.S.107 Unlike activated alumina, iron-based media are not typically regenerated and require periodic media replacement. A recent high-resolution imaging analysis demonstrated that arsenic atoms diffuse preferably along the grain boundary of iron oxides explaining the surface sorption or surface complex formation of arsenate on ferric hydroxide (FeOOH).108 Understanding the effectiveness of this removal technology across a wide range of environmental conditions can improve As removal in water treatment plants and provide an accurate prediction of As transport through soils and aquifers.
In situ chemical immobilization has been used to reduce arsenic contamination in soils by applying iron amendments to contaminated soils and, thus, reducing As mobility and bioavailability in the soil.109 While there have been many studies focused on arsenic adsorption using iron oxides, the specific mechanisms that govern how adsorption changes with pH and redox conditions remain unclear.
Surface complexation and electrostatic models can predict the effectiveness of As removal using iron oxides. The accurate prediction of adsorption performance is difficult because of assumptions incorporated into various models. Different approaches can be taken to model the adsorption phenomena occurring at the sorbent surface including the diffuse double-layer model of Dzombak and Morel that predicts arsenic adsorption by iron oxides.110 However, this model was derived for synthetic solutions and does not account for the competitive interactions found in real complex systems. More importantly, the database contains a limited amount of experimental data.110 On the other hand, the CD-MUSIC model of Hiemstra and van Riemsdijk111 is more promising but more complex. Additional work with model verification and validation is required to improve the output of these products.
The use of mineral oxides in small-scale water treatment systems may be feasible for arsenic removal, but additional studies are needed to evaluate their performance in treating highly concentrated arsenic-laden industrial effluents. Over the years, zerovalent iron (Fe0) has been used as a permeable reactive barrier medium for the removal of various groundwater contaminants,112 including various chlorinated hydrocarbon compounds through reductive dehalogenation.113,114 Su et al. also reported that Fe0 has been used to remove arsenic from groundwater and performance was impacted by reaction time, pH, and oxidation reduction potential.112 In another study, Pierce et al. reported that using 4.45 mg/L of Fe(OH)3 resulted in arsenite removal efficiency of at least 50%.65 A ferric-based layered double hydroxide intercalated with alpha-alanine was synthesized by a coprecipitation method to remove As(III) from water.115 Nanoscale iron–manganese binary oxides have also been tested in column experiments for the removal of both As(III) and As(V) from aqueous media.116
Zerovalent iron (ZVI) is one of the widely used adsorbents for the rapid removal of As(III) and As(V) in the subsurface environment.117−120 The large size, lower surface area, and lack of mobility of ZVIs have limited their applications shallow groundwaters. To overcome these limitations, nanoscale zerovalent iron (nZVI) has been studied due to its high surface area and reactive properties.121 nZVIs have shown great potential for As removal from groundwater sources.122,118,123 Both As(III) and As(V) were removed from real water matrices containing dissolved natural organic matter.118,123 As(III) may be adsorbed either onto nZVI or can be oxidized to As(V) and adsorbed on nZVI surface.124−126 Ezzatahmadi et al. reviewed the adsorption performance of clay minerals, zerovalent iron materials, and clay-supported nZVI composites in removing various heavy metals including arsenic from aqueous solutions. Clay-supported nZVI composites achieved the highest removal efficiency for various contaminants including arsenic.124
Recently an iron-anode enhanced sand filter was employed to remove arsenic with initial concentration of 196–472 μg/L to below 10 μg/L from tube well water in the Jianghan Plain, central China. A current of 0.6 A and a flow rate of about 12 L/h were used, and the residual As was about 110 μg/L without electricity.127
Metal oxides and hydroxides of iron or alumina128 have been widely studied for the removal of As from water and wastewater. However, there is an urgent need to develop cost-effective, efficient, and sustainable adsorbents that can be implemented in developing countries.129 Several industrial byproducts including red mud,129 neutralized red mud, fly ash, ferruginous manganese ore,130 Fe(III)/Cr(III) hydroxide waste,131 and steel slag117 have been evaluated for the removal of arsenic from water. Steel slag (basic oxygen furnace slag (BOFS) and blast furnace slag (BFS)) are steel industrial byproducts, having FeO and CaO as their main components, and are considered one of the cheapest widely available materials.132 More importantly, iron oxide and calcium oxide are important components, due to the strong bonding capacity and ability to stabilize sludge produced after As adsorption, making BOFS and BFS good candidates for As adsorption. From an environmental perspective, BOFS recovered from high temperature is safe since metals are well-stabilized and do not leach easily.133 shows a scanning electron microscopy (SEM) image of the steel slag and the As adsorption kinetics to steel slag, showing that total As can be effectively removed.117
Open in a separate windowIn addition to As adsorption being influenced by the concentration of the arsenic, the chemical and physical properties of aquifer materials affect As adsorption. The clay fraction has been identified as a major sink for arsenic contamination. The composition of the clay is also one of the fundamental factors affecting the biological toxicity of As.134 The pH and cation exchange capacity (CEC) of the subsurface material affect adsorption. The form of arsenic species impacts adsorption onto clay minerals as a function of pH. The maximum adsorption capacity of As(V) onto kaolinite and montmorillonite was found to be at pH 5, whereas the As(III) adsorption increased beyond this pH. For example, at pH = 8, more As(III) was adsorbed compared to As(V).47,135,136 The effect of pH can also be observed with other adsorbents, where the As(III) adsorption onto alumina, bauxite, and carbon decreases at pH greater than 7.137
As(V) adsorption onto clay minerals has been studied using batch experiments as a function of pH and NaCl. Four clay minerals (i.e., illite, kaolinite, montmorillonite, and chlorite) were selected due to their abundance in sediments and their different crystalline structure. Increasing NaCl concentration enhanced As(V) adsorption, indicating that surface complexes were formed, in which Na+ ions served as bridges between the clay surface and the As(V) anions.138 In a different study, the effects of dissolved calcium (Ca2+) and pH on As(V) reactivity and surface speciation were investigated using adsorption performance experiments and extended X-ray absorption fine-structure spectroscopy (EXAFS) analyses. EXAFS analysis revealed that As(V) tetrahedral molecules were coordinated on aluminum octahedral via bidentate binuclear bonding at pH 4.5–6.75.139
Sund et al. also studied the adsorption of arsenites with four different Indian soils with different CECs and pH values.140 The adsorption of arsenic was impacted by the fraction of sesquioxide, clay, and the concentration of exchangeable Ca2+ and Mg2+ in the soils. The amount of arsenic sorbed to soils, including clayey brown, degraded chernozem, slightly gray, sod-alluvial carbonate, and carbonate meadow was influenced by the soil solution pH. Galba et al. indicated that, in strongly acidic soils, arsenic sorption was localized mainly on the surface of clay minerals, while in weakly acidic soils, colloidal sesquioxides participated in the arsenic sorption. In extremely alkaline environments (pH > 11), calcium arsenate complexes were found.141 The quantity of arsenic adsorbed by soil increased with increasing initial concentration.141
Coexisting ions can impact arsenic sorption in the environment, as ions tend to compete for sorption sites or can form complexes with arsenic. Similar to arsenate, phosphate strongly competes for sorption sites on iron oxides, soils, and sediments, whereas anions including Cl– and SO42– have no or slight impact. On the other hand, organics can form complexes with arsenic, increasing its mobility and decreasing its sorption. Manning et al. reported the fractionation of added As(III) to three soils: Fallbrook, Panoche, and Indio.142 Fallbrook and Panoche soils displayed greater adsorption and oxidation compared to Indio soil. Yet, for all three soils, the total recoverable As(III) consistently decreased for 48 h, due to the slow intraparticle diffusion of both As(III) and As(V). Following adsorption, oxidation of As(III) to As(V) occurred, which then partitioned between the solid and solution phases. Soil can oxidize As(III) to As(V), which is very important in the cycling of arsenic species.143
Activated carbon has been widely deployed as a point-of-use treatment for waters containing metals.144 Arsenic removal by chemically treated activated carbon at ambient pH145 is more effective for As(V) removal than As(III) removal. A relatively large quantity of carbon is required to effectively remove As(V) from solution. Recently, arsenic removal from water by δ-MnO2 modified activated carbon has been reported. Within the first 9 h, approximately 90.1% and 76.8% removal efficiencies of As(III) and As(V), respectively, were achieved, and adsorption equilibrium was achieved within 48 h.146 The Qmax value of As(V) and As(III) at pH 4.0 was determined as 13.30 and 12.56 mg/g, respectively, using Langmuir adsorption isotherms.146
In a different study, Huang et al. studied arsenate adsorption by 15 different activated carbons147 and found that arsenate adsorption depended on carbon type, pH, and arsenic concentration. Contaminant removal has been investigated using virgin activated carbon, chemically treated activated carbon, and activated carbon impregnated with ferric hydroxide.148,149 High As adsorption capacity to chemically treated activated carbon has been recently reported.150 Activated carbon prepared from coconut and rice husks has been investigated for the removal of arsenic from water. Lee et al. reported the effect of dosages on uptake using carbon from quaternized rice husks with an optimal As removal efficiency of 86% and solid phase concentration of 3 mg As/g quaternized carbon at pH = 7.5.151 Manju et al. evaluated the performance of coconut husk carbon (CHC) and copper impregnated coconut husk carbon (CICHC) for the removal of As(III) from water and demonstrated significant removal with a 4 h reaction time.152 The As(V) adsorption kinetic process was similar to that of As(III), consistent with the results generated with the adsorption isotherm profiles of As(V) and As(III) by the δ-MnO2-modified activated carbon ( ).
Open in a separate windowDuring the ion exchange process, a reversible interchange of ions occurs between the solid and the liquid phase. The arsenic exchange and regeneration with common salt can be represented by the following equations
Arsenic exchange:
10
Regeneration:.
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where R = ion-exchange resin.
Recently, a tetrahedron ion-exchange resin filter assessed in Bangladesh has shown promise in removing arsenic from water. As expected, the bed volumes to arsenic breakthrough decreased as sulfate concentration increased from 0 to 150 mg/L . An additional increase in sulfate concentration (>150 mg/L ) exhibited a slight increase in bed volumes to arsenic breakthrough.
The applicability of ion exchange (IX) has been well-studied along with competing water constituents (e.g., SO42– for IX) and regeneration requirements.1153 Magnetically impregnated ion-exchange (MIEX) has also been used as a medium for arsenic removal via ion exchange. Incorporating iron into the matrix combines favorable iron–arsenic interactions with the ionic exchange mechanisms. As(V) was removed using an iron-impregnated ion exchange bed.154 A new class of environmentally friendly starch-bridged magnetite nanoparticles was developed and evaluated for removal of arsenate from spent IX brine. 0.049% (w/w) of the low-cost, “green” starch was used as a stabilizer to prevent the agglomeration of nanoparticles while preserving their high arsenic adsorption capacity ( ). When this adsorbent was used to treat a synthetic spent IX brine, complete arsenic removal was achieved within 1 h, compared to only 20% removal when bare magnetite particles were used. The Langmuir adsorption capacity was determined as 248 mg/g at pH 5. Increasing NaCl concentration from 0 to 10% (w/w) in the brine slightly impacted the adsorption capacity of arsenic.155
Open in a separate windowRecently, arsenic removal from geothermal water has been studied using a hybrid system consisting of novel 1JW, 2JW, and 2PTN resins along with Dowex XUS 43594.00 resin ( ).
Open in a separate windowSeveral membrane technologies have been used for the removal of arsenic from water (e.g., nanofiltration (NF), reverse osmosis (RO), and electrodialysis). Membrane separation offers several advantages including ease of operation, no addition of chemicals, and high performance, as many contaminants including bacteria, salts, and other heavy metals could be potentially removed. Clifford et al. found that reverse osmosis could effectively remove 98–99% of As(V), while As(III) was partially removed (46–75%) due to the neutral form of As(III) as H3AsO3.157 Yet, the low water recovery rates, the high transmembrane pressure requirements, the relatively high capital and operating costs, and membrane fouling hamper the wide spread of membrane technologies.
Nanomaterials have been increasingly used in membrane technologies for arsenic treatment. Recently, a TiO2-coated carbon nanotube (CNT) network membrane, prepared via a simple filtration–steam hydrolysis method, was evaluated for arsenic removal. The TiO2 coating uniformly covered the CNT network surface with an ∼2-fold greater specific surface area than that of the bare CNT network. The TiO2–CNT As sorption kinetics increased with increasing cell potential, as shown in .158
Open in a separate windowIn this procedure, arsenic waste (e.g., As-loaded or spent adsorbent) is stabilized by adding chemicals (e.g., cement, lime) to avoid costly disposal of hazardous waste. The stabilization of arsenic adsorbed by this process has been investigated by various researchers.159−161 In one study, Comba et al. investigated arsenic removal from solution by forming mimetite, a lead arsenate chloride mineral (Pb5 (AsO4)3 Cl).162 Aqueous arsenic concentrations were reduced from several mg/L to 0.2 μg/L, and a phosphate/arsenate solid solution (phosphomimetite) was formed. Cement or cement lime has also been used to stabilize various arsenic-bearing materials,163,164 calcium and iron arsenate contaminated soil, and ferric hydroxide precipitated soils in the presence of chloride and phosphate. In a different study, Diamadopoulos et al. studied As(V) removal from aqueous solution by fly ash and found that arsenic adsorption was almost irreversible with arsenic removal of more than 80% at a pH of 4.165
Cornwall et al. provided guidelines for disposing waste resulting from arsenic removal processes, which serve as a good reference for water utilities on how to dispose residuals containing elevated concentrations of arsenic.166 Sullivan et al. also reviewed the disposal of water treatment As wastes, with a particular emphasis on solidification/stabilization (S/S) technologies.167 As can be treated by Portland cement/lime mixes, oxidation of As(III) to As(V), and the formation of calcium–arsenic compounds using binders such as geopolymers and activated alumina.167 Arsenites can chemically bond with hydrated lime to form precipitates as follows.168
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Dutre et al. also investigated S/S of arsenic-containing waste and the behavior of arsenic in leachate.168 The mechanism that governs the S/S of As-contaminated soils including Portland cement (PC) and cement kiln dust (CKD) was explained using various analytical tools including X-ray powder diffraction (XRPD), X-ray absorption near edge structure (XANES), and extended X-ray absorption fine structure (EXAFS) spectroscopy. As immobilization after stabilization was evaluated using a 1 N HCl extraction according to the Korean Standard Test (KST). After 1 day and 7 days of treatment with 30 wt % PC and 50 wt % CKD, respectively, the level of As leaching from the amended soils was less than 3 mg L–1 (i.e., the Korean countermeasure standard).169
In situ S/S offers three main advantages including being a well-established, efficient, and cost-effective approach; requires no secondary treatment; and more importantly, poses no risk to site workers through As exposure.170 Influence of the addition of fly ash on S/S performance is reported in . It was found that addition of cement-stabilized arsenic waste stabilized the arsenic.171
Open in a separate windowNatural attenuation of As can occur when the dissolved phase is trapped by precipitation of secondary As minerals, coprecipitation/sorption reaction with Fe, Mn, and Al oxyhydroxides, carbonates, clay minerals, or complexed by organic matter.172 Redox reactions mediated by microorganisms play an important role in the fate of As, as the microbial oxidation of As (As(III) to As(V)) favors the stabilization of As in the solid phase.173 Arsenic sorption onto aquifer sediments has been investigated in anaerobic bench- and pilot-scale experiments and analyzed by X-ray absorption spectroscopy (XAS).174 Margarita et al. reported that arsenic in semiarid soils contaminated by wastes containing oxidized arsenic species is naturally stabilized.175 For areas with high acetate loadings and under sulfate-reducing conditions, arsenic release was higher. Lab-scale experiments revealed that, in the presence of the reduced precursors arsenite and sulfide, multiple thioarsenic formed.176 Iron hydroxide can bind better to arsenic, followed by clay minerals and then feldpars.177 The methylation of arsenic occurs via alternating reduction of pentavalent As to trivalent As and addition of a methyl group ( ).178,179
Open in a separate windowNatural attenuation can be a viable treatment technology for developing communities that lack resources to implement some of the previously discussed processes (e.g., metal-impregnated adsorbents, IEX, and fly ash incineration). Solar oxidation of As-laden waters is another low-cost remediation technology. A technology called SORAS—solar oxidation and removal of arsenic—has been shown to remove 75–90% of As180 from waters containing both arsenic and iron, which have been photochemically pretreated with the addition of lemon juice. The citric acid induces the precipitation of ferric iron, where arsenic then adsorbs to the solid iron compounds and clean water is decanted off of the top.180 Optimizing the process with the addition of steel wool to the source water, Cornejo et al. was able to reach a removal efficiency of at least 99.5% with final arsenic concentrations below 10 μg/L.181 The optimal molar ratios for arsenic, citrate, and iron were determined by Lara et al. to be 1:4.5:18.7, respectively, to achieve over 90% As removal efficiency after 4 h of irradiation.182 SORAS was not significantly affected by pH, as increasing the pH from 5 to 9 had no negligible impact on aresenic removal.183 Geroni et al. attributed that pH independency to at least 30 individual reactions and equilibria occurring simultaneously (many of which are competing), where at the end state, arsenic removal was mostly unaffected.183 With this positive performance over a wide pH range, SORAS could be applied to arsenic contamination due to mining activities (mining waters usually have a wide pH range). SORAS requires behavior modification by the community in order to be effective. For example, treated water bottles need to be left overnight in the vertical position to allow arsenic–iron compounds to settle out.184
Unlike other treatment technologies, subsurface arsenic removal (SAR) does not generate waste (e.g., As-containing sludge, waste from the regenerated resin or the spent adsorbent, or brine with high As concentrations) that needs to be handled and disposed in a safe manner. In addition, SAR requires no electricity or chemical additives. SAR is not a new technology but is a modified form of subsurface iron removal (SIR). With SIR, the oxidation of iron species is prompted, which leads to iron precipitation, and the solid iron compounds stay in the ground as iron-free water is extracted. Fortuitously, if arsenic is present it will react and combine with iron and become trapped in a precipitated iron–arsenic complex. Similar to technologies described earlier, oxidizing As(III) to As(V) is preferrable for optimal arsenic removal in situ. With SAR, water with dissolved oxygen concentrations of ∼8 mg/L is introduced into the anaerobic subsurface to accomplish the oxidation.6 With such an oxygen concentration, the reaction is slow—on the order of hours to days—but the residence times in the subsurface are of the same order of magnitude, and therefore the viability of technology remains. The benefits to using the aerated water as the oxidant are cost and convenience; no additional chemicals are needed to convert arsenite to arsenate.
SAR requires a source of aerated water to introduce into the subsurface, which is typically accomplished with aboveground storage tanks that incorporate paddles, plates, bubbling, or a similar methodology to achieve aeration. The aerated water is then introduced into the subsurface aquifer by a handpump, which is widely available in developing countries including Bangladesh and Bengal Delta areas. A simple diagram of the one pump SAR design is included in .
Open in a separate windowThe following section describes the chemical reactions during SAR. After the aerated water is introduced into the aquifer, arsenite is oxidized into arsenate as shown below.
Arsenite oxidixzing to arsenate:
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For SAR to be effective, iron is a requirement where oxidation of ferrous iron to ferric iron takes place at the same time as arsenic oxidation. Ferric iron exists in different complexes (e.g., ferric hydroxide or hydrous ferric oxide (also referred to as ferrihydrite and HFO) or goethite (FeOOH), etc.), but the majority of these complexes are in solid form. eqs 16 and 17 demonstrate these processes.
Ferrous oxidizing to Ferric:
16
Ferric to Ferric Hydroxide (or to HFO (FeH2O4), or Fe2O3, FeOH2+, Fe(OH)2+, FeOOH, etc.):
17
These solid forms of iron have relatively high surface areas and will sorb arsenate compounds out of solution. HFO has a surface area of 141 m2/g, and goethite has reported surface areas of 39 and 103 m2/g.99 HFO and goethite, in particular, have demonstrated significant arsenic removal.6eqs 18 and 19 below show the arsenic adsorption to the solid iron oxides. The various iron oxides are represented with the “S–” notation to reflect the “solid” form of the compound.
Arsenite adsorbs to iron oxides:
18
Arsenate adsorbs to iron oxides:
19
Arsenate to iron oxides:
20
Additionally, when (aerobic) water is extracted from the aquifer, ferrous iron in solution becomes adsorbed onto soil particles (eq 21). The adsorbed ferrous iron then undergoes oxidation to ferric iron by a pathway similar to eq 16 and thereby creates more sites that are available for arsenic adsorption during follow-on cycles of SAR aeration and extraction.
Adsorbing Ferrous
21
The oxidation zone formation and activation time is between 8 and 12 h, and, similar to SORAS, behavioral change is needed to allow for this time to completed before water is extracted from the well. Care and caution must also be exercised to avoid contamination of the injected aerated water from above-ground. Additional areas needing further research for optimization of SAR include determining the impact of the natural organic matter concentrations on the As removal efficiencies and further researching the remobilization of arsenic once fixed to the subsurface.6
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