1. Introduction
Arsenic is globally considered as one of the major pollutants in drinking water sources and a worldwide concern because of its toxicity and carcinogenicity [
1]. The presence of arsenic at elevated concentrations in natural environments can be attributed to both natural and anthropogenic inputs [
2]. Arsenic pollution is primarily caused by natural processes, such as the weathering of rocks and minerals, followed by leaching and industrial activities that lead to the pollution of soil and groundwater [
3]. The discharge of arsenic polluted waters from mining or mining-related activities, the pharmaceutical industry and agricultural activities plays an important role in anthropogenic arsenic pollution in Asia [
4]. However, the introduction of arsenic into groundwaters is expected to occur mainly as a result of its natural geological presence in rocks [
5].
Arsenite As(III) and arsenate As(V) are considered as the main oxidation states of inorganic arsenic found in natural waters. As(V) anions are predominant and stable in oxygen-rich environments, whereas the As(III) anions prevail in moderately reduced environments (i.e., anaerobic or anoxic groundwaters). Therefore, arsenic speciation mostly depends on pH and redox potential (Eh) conditions. Under oxidizing conditions and at pH values relevant to drinking water treatment, H
3AsO
4 is present as an oxyanion in the forms of H
2AsO
4− and/or HAsO
42−, whereas at low Eh values, arsenic becomes dominant as H
3AsO
3. Up to pH 9, H
3AsO
3 does not dissociate and, therefore, is present in most natural waters as the uncharged arsenious acid [
6]. Therefore, As(III) species are considered as much more mobile in aquifers and cannot be easily adsorbed (and removed) onto the usually co-existing mineral surfaces, such as those of iron oxides. Moreover, As(III) is more toxic for the biological systems, as compared to As(V) [
3,
7].
The pollution of drinkable water sources by arsenic has been reported in more than 70 countries, where more than 150 million inhabitants are under high health risk [
8]. Due to its high toxicity to humans, the World Health Organization [
9] lowered the guideline value for arsenic in drinking water from 50 to 10 µg/L in 2004, aiming to minimize the health-related problems associated with arsenic pollution. The same standards also apply in the European Commission, as well as the US Environmental Protection Agency. Among other countries, the arsenic pollution of groundwater is considered as a particularly serious health-related problem in Pakistan, as a recent survey reveals [
10]. Approximately 50 to 60 million people relying on groundwater as a source of drinking water in the Indus Valley are at high health risk [
11]. In Punjab, more than 3% of the inhabitants are exposed to arsenic concentrations higher than 50 µg/L in drinking water, and 20% of the population is exposed to concentrations higher than 10 µg/L, while 16% and 36% of inhabitants in Sindh areas are exposed to arsenic pollution of concentrations higher than 50 µg/L and 10 µg/L, respectively [
10].
Several treatment technologies to remove arsenic from drinking water have been applied worldwide [
12], including adsorption using activated alumina [
13] or iron oxide-based adsorbents, such as tetravalent manganese feroxyhyte [
14], bayoxide [
15,
16,
17], granular ferric hydroxide (GFH) [
18], etc. Other treatment methods include the application of oxidation and arsenic removal using zero-valent iron (especially in Bangladeh) [
19], coprecipitation of arsenic with iron or aluminum salts [
2], preliminary arsenic oxidation by ozonation or biological oxidation [
19], ion exchange [
20], high pressure membrane separation [
21,
22] and electrocoagulation [
23]. According to Tresintsi [
14], chemical precipitation by ferric coagulation has significantly higher arsenic removal efficiencies in comparison to adsorption by iron oxy-hydroxides, and a drinking water regulation limit can be achieved at an affordable price, with operational costs estimated between 0.09 and 0.16 €/m
3 for initial arsenic concentrations, ranging between 19 and 208 μg/L. However, the major part (>90%) of treatment costs was attributed to the management of produced sludge, since the coagulant costs are estimated to be ≤0.01 € [
15]. Previous studies have identified high pressure membrane processes as an emerging technology, due to their high removal efficiencies and easy operation features [
21,
22], but these high pressure membrane processes are rather energy (and cost) intensive, and subjected to the fouling of membranes. Moreover, the disposal of produced brine (high salt concentrations) is a considerable challenge.
On the other hand, for the treatment of waters with moderate arsenic concentrations, i.e., slightly higher than the regulation limits, adsorption onto iron oxide-based adsorbents has been proved to be the most economically efficient procedure [
15]. The two mostly commercially applied adsorbents are the Granular Ferric oxy-Hydroxides (GFH) and the Bayoxide E33 (GFO), which are favorable in terms of cost, removal efficiency, simplicity of design, operation, maintenance and minimizing the (secondary) waste production [
24]. The GFH has been tested to remove arsenic from drinking water sources under both laboratory-scale and full-scale water treatment plants [
15,
18,
25,
26]. Arsenic adsorption onto GFH is usually performed in a column filtration mode, which is a rather simple process and can be continuously operated, but the production of this material is relatively cost intensive. The cost (on dry basis) for GFH and Bayoxide materials was estimated to be 9 €/kg and 12.5 €/kg, respectively [
15]. Currently, the small sized fraction (dust ferric hydroxide, DFH) generated during the industrial production of GFH cannot be employed in the common column filtration mode, since the small sized adsorbents can rapidly clog the fixed beds in filter columns, causing an increased pressure head, thereby increasing energy costs and maintenance and, hence, reducing the system performance.
Adsorption combined with the application of low-pressure membrane filtration is considered as a newly developed hybrid water treatment process. Low-pressure membrane processes, such as microfiltration or even ultrafiltration, have a reasonable energy demand and produce superior quality treated water with a rather controllable fouling of membranes and incurring quite low capital and operational costs [
27]. The low-pressure membrane processes are not able to remove mono- and polyvalent ions, i.e., arsenic species, from water sources, although they can efficiently remove suspended solids, colloids, bacteria, viruses and micro-particles [
28]. If the cost-effective small sized GFH adsorbent (having a substantially lower commercial price of only 1.6 €/kg on a dry basis) has the potential to remove arsenic species from drinking water sources, it might then be employed in the adsorption-microfiltration (MF) hybrid treatment scheme to economically and efficiently remove arsenic. The idea of a submerged membrane filtration adsorption hybrid system could be exploited in this regard, which allows the pollutant to be in contact with adsorbents for longer time.
The objectives of the study were: (i) To assess the adsorption potential/performance of the smaller fraction of GFH material with a particle size of <0.250 mm, which is abbreviated as DFH henceforth, for removing As(V) and As(III) species from different water matrixes; (ii) to determine the kinetics of arsenic adsorption on the studied material; (iii) to examine the effect of a water matrix on arsenic removal; and (iv) to compare the efficiencies of both major inorganic arsenic species, As(V) and As(III), with the established, conventionally applied adsorbents, such as GFH. Badruzzaman [
29] studied the use of small sized GFH in packed bed columns, but investigated the adsorption potential of this material only in the case of As(V) and ultra-pure water and has found promising results. However, to the best of our knowledge, no comprehensive study concerning the application of DFH material, systematically studying the arsenic adsorption efficiency of both arsenic species and different water matrices, such as the tap water of Hamburg (HH tap water) and the NSF (National Sanitation Foundation) challenge water, used to simulate typical arsenic-containing groundwater, has yet been performed.
2. Materials and Experimental Methods
2.1. Reagents
For the preparation of As(III) or As(V) 100 mg/L stock solutions, the standard solution of As(III), as As2O3 in 2% HNO3, and As(V), as H3AsO4 in HNO3, with a concentration of 1 g/L, were used, obtained from Carl Roth GmbH + Co. KG (Karlsruhe, Germany) and Merck Chemicals GmbH (Darmstadt, Germany), respectively. The pH buffer solution, N,N-Bis-(2-hydroxyethyl)-2-aminoethane sulphonic acid (BES), used in the experiments focusing on arsenic removal from deionized (DI) water, was obtained from Carl Roth GmbH + Co. KG.
2.2. Material Characterization
The DFH material, with a particle size of <0.250 mm, was supplied by GEH–Wasserchemie GmbH & Co. KG, Osnabrück, Germany. The material is predominantly akaganeite, a specific form of an iron oxy-hydroxide [
16]. DFH is mainly characterized by a relatively large specific surface area (252 m
2/g) [
29] and surface charge density (
Table 1).
Particle size distribution was determined by EyeTechTM instrument (combi, AmbiValue, Nijerdal, The Netherlands), ranging between 7.4 and 250 µm. The liquid flow cell of EyeTech was filled with 1 L of deionized water, and approximately 100 mg of material was added. Mechanical shaking was provided in the liquid flow cell, which keeps the material particles in suspension. Then, suspension was supplied to the optical cell and circulated through it for 5 min at a pump speed of 0.674 L/min. Three cycles of the suspension were performed to determine the particle size distributions.
To determine the surface charge of DFH in the suspension, the Isoelectric Point (IEP) and the Point-of-Zero Charge (PZC) were quantified. IEP was determined by a zeta-potential curve at 20 ± 1 °C of adsorbent dispersion in 0.01 M NaNO
3, with the respective pH of solution, using a Micro-electrophoresis Apparatus (Mk II device, Rank Brothers Ltd, Cambridge, England), while PZC and the surface charge density were defined by the application of acid/base potentiometric mass titration in suspensions of the adsorbent and for various ionic strengths [
31].
2.3. Water Matrix
The test solution was initially prepared using deionized water (DI), spiked with either As(III) or As(V) species, at an initial concentration of 190 µg/L. 2 mM of
N,
N-Bis(2-hydroxyethyl)-2-aminoethanesulfonic acid (BES) was added to the test solution, made of DI water, for pH control at pH 7.9. In addition to DI water, As(V) and As(III) test solutions were prepared in HH tap water and NSF water with the same initial arsenic concentration, as used in the case of DI water, in order to study the effect of different water matrixes on the arsenic adsorption capacity. The major physicochemical parameters of the HH tap water and of NSF challenge water are listed in
Table 2.
The NSF challenge water was prepared according to the National Sanitation Foundation (NSF) international and contains the following: 252 mg NaHCO
3, 12.14 mg NaNO
3, 0.178 mg NaH
2PO
4·H
2O, 2.21 mg NaF, 70.6 mg NaSiO
3·5H
2O, 147 mg CaCl
2·2H
2O and 128.3 mg MgSO
4·7H
2O in 1 L of DI water. Prior to adsorption experiments, the pH was adjusted to 7.9 by adding either NaOH or HCl standard solutions (0.1 N) [
32].
2.4. Batch Adsorption Procedure
Batch equilibrium and kinetic adsorption tests were performed to study the adsorption potential of DFH for removing arsenic species from the different examined test solutions/water matrixes. To derive the adsorption isotherms, the method of adding various quantities of adsorbent to a constant solution volume (500 mL), having the same initial concentration of As(V) or As(III) species, was adopted. Additionally, As(III) test solutions were preliminary bubbled for 30 min with pure N2 gas at 0.1 bar (flowrate 11.25 mL/min) to minimize the influence of dissolved oxygen on As(III) potential oxidation and adsorption, and the flasks were immediately closed and placed on the platform shaker in darkness in the thermostate cabinet (20 ± 0.5 °C) to insure the stability of As(III) species during and after adsorption onto the examined iron oxide-based adsorbents.
The evaluation of the examined adsorbent material focused on its ability to decrease the residual arsenic concentration below the drinking water regulation limit of 10 µg/L (termed Q
10 hereafter), rather than studying the (more convenient) maximum capacity (Q
max) at higher residual arsenic concentrations. If efficiency of the adsorbent was evaluated through Q
max, which usually points to high residual concentrations and, indeed, brings high adsorption capacities, but provides marginal information on its ability to reach low concentrations, such as the regulation limits [
33].
Different adsorbent dosages were placed in flasks for the three different water matrixes, while only adsorbent dosages, ranging between 5–40 mg/L, 6–50 mg/L and 10–80 mg/L, provided equilibrium As(V) concentrations between 1 and 120 µg/L in DI water, HH Tap water and NSF water, respectively. Adsorbent dosages of 10–60 mg/L, 15–100 mg/L and 40–200 mg/L were found to provide the same range of equilibrium concentrations in the experiments focusing on the removal of As(III) in DI water, HH Tap water and NSF water, respectively. For comparison, batch adsorption isotherm studies were also conducted with the GFH material, using DI water to compare the adsorption characteristics of GFH with those obtained when using DFH, i.e., to examine the efficiency of both particle size fractions of this adsorbent. GFH dosages, ranging between 10–80 and 20–120 mg/L, were carefully placed in flasks for the removal of As(V) or As(III) species, respectively. For each experimental test focusing on adsorption isotherm, a reference blank sample (i.e., without the presence of an adsorbent) was filled. The flasks were stirred using a platform shaker for 96 h at 20 ± 0.5 °C. The equilibration time was determined for the corresponding kinetic experiments. At the end of the equilibration time, the suspensions were immediately filtered through a 0.45 µm membrane syringe filter (PVDF, Carl Roth GmbH + Co. KG), and the filtrates were collected and stored for the subsequent analytical determination of residual (still dissolved and removed) arsenic.
In the kinetic studies, the initial arsenic concentration and adsorbent quantity was kept at 190 µg/L and 50 mg/L, respectively. The initial concentrations of either only As(V) or only As(III) species were the same as in the respective isotherm studies. Batch adsorption kinetics tests were conducted at the initial pH value of 7.9. Unlike the isotherm studies, a magnetic stirrer (100 rpm) was used in the kinetic studies experiments. The samples were collected at regular time intervals and the residual arsenic in the solution was analyzed. Each set of adsorption batch isotherm and kinetics experiments was replicated at least twice, and the average values are reported.
2.5. Chemical Analytics
Initial and residual arsenic concentrations were determined by Graphite Furnace Atomic Absorption Spectrophotometry (Perkin-Elmer 4100ZL, Baesweiler, Germany), using a Perkin-Elmer 4100ZL instrument [
34]. The limit of detection was 0.5 µg/L. Prior to analysis, As(III) water samples from the isotherm experiments were acidified (2 < pH < 4) and passed through a 30 mL column (with ID = 2 cm), containing an anion exchange resin (Dowex
® 1 × 8–100, mesh size 50–100, Sigma-aldrich Chemie GmbH, Taufkirchen, Germany), which retained As(V), whereas the total arsenic concentration of water samples from the adsorption kinetics experiments were analyzed. This method of arsenic speciation needs approx. 50 mL of water sample, noting that, in the kinetics experiments, only small volumes (~7 mL) of water samples were collected at regular intervals; accordingly, arsenic speciation using this method was not possible. Therefore, only the total arsenic concentration of the water samples from adsorption kinetics was analyzed, presenting the concentration of individual arsenic species in the water samples. The initial concentration of phosphate in HH tap water was measured using ICP-MS (NexION 300D, PerkinElmer, Baesweiler, Germany).