Arsenic removal from naturally arsenic contaminated ground water by packed-bed electrocoagulator using Al and Fe scrap anodes

Accepted Manuscript Title: Arsenic removal from naturally arsenic contaminated ground water by packed-bed electrocoagulator using Al and Fe scrap anodes Authors: Philip Isaac Omwene, Meltem C ¸ elen, Mehmet Salim ¨ Oncel, Mehmet Kobya PII: DOI: Reference:

S0957-5820(18)30247-7 https://doi.org/10.1016/j.psep.2018.10.003 PSEP 1533

To appear in:

Process Safety and Environment Protection

Received date: Revised date: Accepted date:

6-6-2018 3-10-2018 5-10-2018

¨ Please cite this article as: Omwene PI, C ¸ elen M, Oncel MS, Kobya M, Arsenic removal from naturally arsenic contaminated ground water by packed-bed electrocoagulator using Al and Fe scrap anodes, Process Safety and Environmental Protection (2018), https://doi.org/10.1016/j.psep.2018.10.003 This is a PDF file of an unedited manuscript that has been accepted for publication. As a service to our customers we are providing this early version of the manuscript. The manuscript will undergo copyediting, typesetting, and review of the resulting proof before it is published in its final form. Please note that during the production process errors may be discovered which could affect the content, and all legal disclaimers that apply to the journal pertain.

Arsenic removal from naturally arsenic contaminated ground water by packed-bed electrocoagulator using Al and Fe scrap anodes

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Philip Isaac Omwene, Meltem Çelen, Mehmet Salim Öncel, Mehmet Kobya*

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Gebze Technical University, Department of Environmental Engineering, 41400, Gebze, Turkey

*Corresponding author. Address: Department of Environmental Engineering, Gebze Technical University, 41400 Gebze-Kocaeli. E-mail address: [email protected] (M. Kobya)

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Both Al and Fe scrap anodes reduced the residual As level to WHO standard Overal, Fe scrap anodes performed much better than Al electrodes Electrogenerated Al or Fe-hydroxides showed satisfactory As removal capacity

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  

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Highlights

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Abstract:

In this work, feasibility of electrocoagulation (EC) process with Al and Fe scrap anodes for

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treatment of groundwater contaminated with arsenic (As) was examined as a cheaper treatment alternative for affected remote communities. EC experiments were carried out in a batch packed-bed EC reactor and the effect of applied current (0.010-0.100 A), type of scrap

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electrode (Fe and Al), packed-bed density (0.1-0.4 kg/m3 for Fe and 0.02-0.08 kg/m3 for Al) and EC time were investigated. Optimum operating conditions to obtain maximum contaminant level (MCL) of 10 μg /L for total As (> 93% removal) in groundwater samples

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were determined as 8 min and 0.05 A for Fe scrap anodes. Whereas for Al scrap anode, 30 min and 0.10 A were the optimums. The operating cost, energy and electrode consumptions at these optimums were calculated as 0.017 US $/m3, 0.070 kWh/m3 and 0.052 kg/m3 for Fe anodes and 0.181 US $/m3, 0.876 kWh/m3 and 0.067 kg/m3 for AL anodes respectively. The As removal slightly decreased with decrease in anode bed density. Moreover, Fe scrap anodes exhibited better As removal than the Al scrap anodes at all tested conditions. The scanning electron microscopy (SEM) of the electro-coagulated sludge revealed irregular and porous 1

particles with amorphous structure. The Fourier-transform infrared spectroscopy (FTIR) showed bonding between Fe(III) - As(V), and As-O bond, confirming As removal by coprecipitation and adsorption, respectively in the EC process.

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Keywords: Arsenic, Electrocoagulation, Al and Fe scrap anodes, Groundwater

1. Introduction

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Naturally arsenic-contaminated water resources in various parts of the world have adversely affected safe drinking water supply sources. About seventy countries worldwide including Turkey, have reported serious health hazards such as cancers, skin lesions, cardiovascular and

neurological effects as a result of using water contaminated with As (Gunduz et al., 2015;

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Ravenscroft et al., 2009; Çöl and Çöl, 2004). Due its high toxicity, environmental agencies

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have put stringent regulations towards the maximum permissible As levels in water. Total As

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concentration of 10 μg /L in potable water has been established as maximum contaminant

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level (MCL) by the European Union (EU), the United States-Environmental Protection Agency (US-EPA) and the World Health Organization (WHO) (US-EPA, 2005; EU, 1998;

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WHO, 2001). Major sources of arsenic pollution in natural waters are attributed to weathering of rocks, volcanic emissions, biological activities and anthropogenic inputs (Ng et al., 2003). Especially, sulfide minerals such as realgar, orpiment and arsenopyrite may contain very high

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As concentrations. Mineral dissolution in mineralized regions may be enhanced by mining activities; hence the mine wastes and drainages in these areas can have very high As levels

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(Pawlak et al., 2008). Natural As contamination in groundwaters of over 70 countries reported in literature widely varied from < 0.5 - 5000 μg/L (Ravenscroft et al., 2009). According to recent reports, inorganic arsenic varies from 0.5 – 10,700 μg/L for natural waters of Turkey

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(Simsek, 2013; Gemici et al., 2008; Çolak et al., 2003; Helvaci, 1995; Çöl and Çöl, 2004). Particularly, arsenic contamination around borate mines and deposits (Balıkesir-Bigadiç, Kütahya-Emet-Hisarcık, Eskişehir-Kırka, and Bursa-Kestelek) in Western Anatolia of Turkey are attributed to the dissolution of naturally occurring arsenic from a borate bearing clay zone, which is rich in arsenic bearing minerals such as realgar and orpiment. The affected communities need to attain a safe water supply source for both direct consumption and crop irrigation purposes. Therefore, technologies for treatment of water supplies contaminated with 2

As to enable affected communities access safe water is an urgent issue at present. To achieve compliance with MCL, the US-EPA has proposed best available treatment technologies such as reverse osmosis, ion-exchange, modified lime softening, modified coagulation/filtration, oxidation/filtration, electrodialysis and activated alumina for arsenic removal from waters (US-EPA, 2000). Many municipalities and industries have resorted into research on electrochemical methods as

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an effective As removal technology, due to the drawbacks in other technologies (Nidheesh and Singh, 2017; Vasudevan and Oturan, 2014; Amrose et al., 2014; Garcia-Lara and

Montero-Ocampo, 2010; Martinez-Villafane et al., 2009). EC is one of the most promising

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treatment technologies for removal of arsenic in solutions. The EC process has advantages

like; simple and robust equipment, higher removal efficiency, no or less addition of chemicals, less retention time, no chemicals used for pre-oxidation of As(III) to As(V), no pH

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adjustment, relatively cost-effective and simple to use (Song et al., 2017; Kobya et al., 2016; Banerji et al., 2016; Alcacio et al., 2014; Li et al., 2014; Molgora et al., 2013; Vasudevan et

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al., 2010; Parga et al., 2005). Sacrificial metal anodes used in EC process for As and other

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pollutants removal from contaminated solutions are usually of iron or aluminium plate, rod,

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mesh, and ball types (Hakizimana et al., 2017; Kabdasli et al., 2012; Khandegar and Saroha, 2016). However, conventional anodes are associated with high fabrication costs, besides the

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high energy costs, making the EC process economically unviable. Furthermore, the shape of the electrodes may also affect the performance of EC process. Each year, a huge amount of metal scraps (metal chips, filings, shavings, or turnings), an abundant by-product from metal

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planning machines, is generated from metalworking shops as wastes from lathes and Computer Numerical Control (CNC) machining. Accumulation of these waste metal scraps is

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of an environmental threat, and their removal is also tedious (Vignesh et al., 2017). In this perception, the use of waste iron and aluminum scraps as sacrificial anodes in EC process may be advantageous both in terms of cost economic feasibility and environmental point of view. The cost of waste iron or aluminum scraps is about one-third of the market price for

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fresh metal. Also, the scrap anodes have a larger electrode contact surface area than the other shape of the anodes, thereby providing a high contact area between the electrodes and pollutants in the EC reactor, leading to overall increase in EC process efficiency. A few studies have focused on the use of aluminum and iron scrap anodes in EC reactors (Vignesh et al., 2017; Ye et al., 2016; Ardhan et al., 2014; Tezcan-Un and Aytac, 2013; Wei et al., 2012). Even the available literature on EC process for As removal by plate, rod and ball electrodes, 3

generally used synthetic As solutions, and very limited studies with real groundwater samples contaminated with As have been performed (Kobya et al., 2016; Amrose et al., 2014; GarciaLara and Montero-Ocampo, 2010). Therefore, EC process using different electrodes for treatment of real As contaminated groundwater samples needs to be further investigated and discussed. In this study, removal of As from naturally contaminated groundwater samples in an EC batch

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reactor using metal iron and aluminum scrap anodes is investigated. Operating conditions like applied current, anode packed-bed density, and EC time on efficiency of pollutant (As) removal were explored for both iron and aluminum scrap anodes, and the optimum

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operational conditions were determined. Also, the operating costs for treatment of As contaminated groundwater samples by EC were computed to give an insight of the economic viability of the process.

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2.1. Groundwater samples and anode material

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2. Material and methods

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Arsenic contaminated groundwater samples were obtained from two spring wells, the first

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spring well is in Osmanca village, hereinafter referred to as ‘’SW-1’’. The second spring well is in Iskele town, hereinafter referred to as ‘’SW-2’’. Both the springs are around Balikesir-

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Bigadiç borate mine deposits in Western Anatolia in Turkey. The constituents of the groundwater samples are presented in Table 1.

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The aluminum and iron chips (length: 1.8-15 mm; width 0.5-3.0 mm) were collected from a metal machining shop in Gebze-Turkey. The chips were cleaned in an alkali solution (10%) to

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remove oil and grease and subsequently submerged (2%) in HCl solution to oxide films and wash rusts. Finally, the chips were rinsed with distilled water and dried.

2.2. Experimental set up procedure

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Fig.1 shows the set-up of experiment used in the present study. A Plexiglas cylindrical packed-bed EC batch reactor (volume of 500 mL, 80 mm internal diameter and 100 mm height) was equipped with cylindrical perforated titanium cathode (internal diameter: 65 mm, height: 80 mm, thickness: 2 mm). The perforations were of 2 mm diameter, drilled all over the cylindrical cathode wall. The anode comprised of Al or Fe scrap packed-bed centrally placed in the cathode (without physical contact). A stainless-steel rod of length 150 mm and 2 mm 4

diameter was inserted centrally into the Al or Fe scraps packed-bed to provide connection points; a DC power supply (galvanostatic mode, Agilent 6675A model, 120 V and 18 A) was connected to the system. 0.40 L of the groundwater sample was placed in the EC reactor.

The desired current value (i = 0.010 - 0.10 A) was adjusted in the DC power supply for each experimental run. Samples for analysis were collected from the reactor at designated time

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intervals. The influence of anode packed-bed density on systems performance was investigated at 0.10 - 0.40 kg/m3 for Fe scrap and 0.02 - 0.080 kg/m3 for Al scrap anode. At

the end of each experimental run, the Al or Fe scrap anodes were first rinsed with diluted HCl

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followed by water, dried and re-weighed to reduce the effect of electrode passivation.

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2.4. Analytical procedures

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The samples for analysis were filtered through 45 μm membrane filter, reacted with KI and ascorbic acid for at least 1 h prior to determination of total As concentration by ICP-OES

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(Optima 7000 DV, Perkin-Elmer, USA). The detection limit of this study was 0.10 μg/L and

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analysis of the duplicates was within 2% of errors. Also, conductivity and pH of groundwater samples before and after the EC treatment process were determined by a pH meter (Mettler

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Toledo Seven Compact) and by a conductivity meter (Mettler Toledo Seven Go), respectively. The generated EC sludge samples at the end of the EC process were oven dried (105 °C ± 5) for analyses by Fourier-transform infrared spectroscopy (FTIR; Bio Rad FTS

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175 C spectropho-tometer), scanning electron microscopy (SEM; Philips XL30S-FEG) and energy-dispersive x-ray spectroscopy (EDAX; Rigaku 2000 D/max with CuK-radiation,  =

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0.154 nm at 40 kV and 40 mA).

3. Results and discussion

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3.1. Theoretical background of Electrocoagulation process for arsenic removal In EC process, sacrificial electro-dissolution of anode generates the coagulants (Eqs. (1) and (3-5)). While the anode dissolves to generate metal ions, the cathode liberates hydrogen gas (Eqs. (2) and (6)). The released H2 (g) aids in flotation of the flocculated matter in solution (Mollah et al., 2001).

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Reactions for Al electrodes in the EC process: Anode: Al  Al3  3e -

(1)

Cathode: 3H O  3e-  3/2H 2(g)  3OH

(2)

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Reactions for Fe electrode in the EC process: Anode: Fe  Fe2  2e -

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(3)

Fe2  Fe3  e-

(4)

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Fe  Fe3  e Cathode: 2H O  2e-  H 2(g)  2OH

(6)

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2

(5)

(7)

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O 2(g)  4Fe2  2H 2 O  4Fe3  4OH 

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Depending on the available oxygen, Fe2+ may be oxidized as shown in Eqn (7).

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Also, the oxidation of ferrous ions occurs due to alkalization from water reduction (Eqn. (8)): (8)

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Fe2  3OH  FeOOH(s)  H 2O  e 

Depending on the solution pH, amorphous hydroxide complexes are formed from the reaction

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of OH  with the electro-generated Fe3+, Fe2+ and Al3+ ions. Consequently, pollutants removal is attributed to formation of Fe(OH)3(s), Al(OH)3(s), monomeric and polymeric metal species

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due to electro-oxidation, precipitation, adsorption, co-precipitation and coagulation (Omwene and Kobya, 2018; Mollah et al., 2001). Arsenate (As(III) as H3AsO3) is generally the main arsenic species in oxygen-rich surface

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waters and arsenite (As(V) as HAsO42–, H2AsO4– and AsO43-) usually predominates in groundwater. As(III) species in the solution has neutral character in the pH < 9, whereas As(V) exists in the forms of H2AsO4– at pH 2-7, HAsO42–at pH 7-11, and AsO43-at pH 12-14 (Ravenscroft et al., 2009). Also, As(III) oxidation to As(V) occurs during the EC process either through electrolytic oxidation at the electrode or via highly reactive radical species produced by the oxidation of Fe(II) by dissolved oxygen (Amrose et al., 2014). The main effective mechanisms in arsenic removal by Fe and Al anodes in EC process are ligand 6

exchange, co-precipitation and adsorption. Mainly, the presence of polymeric aluminium hydroxides such as bayerite (Al-hydroxide), diaspore and mansfieldite in the pH range 4.710.5 would provide significantly larger surface areas for arsenic species adsorption due to their amorphous nature (Parga et al., 2005).

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Al(OH)3  HAsO24  Al(OH)3 * HAsO24 (s)

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(9)

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Al2 O3  HAsO24  Al2 O3 * HAsO24 (s)

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(10)

During the reactions As(V) substitutes for hydroxyl group of iron hydroxides/oxyhydroxides

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such as amorphous Fe(OH)3, hydrous ferric oxide (HFO), lepidocrocite (γ-FeOOH), goethite (α-FeOOH) and magnetite (Kobya et al., 2011; Song et al., 2017; Flores et al., 2013; Gomes et al., 2007; Hu et al., 2003). As(V) adsorption on HFO, lepidocrocite and goethite below pH

5-6 is more favourable than As(III), whereas above pH 7-8, As(III) has a higher affinity (Dixit

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and Hering, 2003). Generally, arsenate anion bound to HFO in chemical coagulation and EC

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processes can form common naturally occurring arsenate minerals FeAsO4·2H2O (scorodite) and Fe3(AsO4)2·8H2O (symplesite) as the dominant solid phase (Gomes et al., 2007; Parga et

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al., 2005). It has been reported that arsenic removal occurs by ligand exchange, arsenate

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displaces a hydroxyl group of FeOOH giving rise to insoluble surface complex reactions.

(Kobya et al., 2015):

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As(V) generally forms bidentate-binuclear bridging complexes as seen in Eqs. 11 and 12

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2FeOOH(s)  H AsO4  (FeO)2 HAsO4(s)  H 2O  OH 2

(12)

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3FeOOH(s)  HAsO24  (FeO)3 AsO4(s)  H 2O  2OH

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Literature studies have acknowledged that solution pH greatly affects pollutant removal mechanisms by EC process (Kobya et al., 2014; Mohora et al., 2018). However, studies have also indicated final arsenic removal efficiency to be independent of pH range 5-8 (Kumar et

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al., 2004; Wan et al., 2011). Consequently, we did not adjust the pH of the groundwater samples since it was within the stated ranges. Besides, we chiefly focused on EC treatment as an alternative for affected remote communities. Modifications in pH would result into increase in operational costs related to chemicals and hiring more skilled personnel. Moreover, the final pH of the treated groundwater ranged from 7.5-8.6, an acceptable WHO drinking water pH range.

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3.2. Effects of applied current and charge loading Applied current (i) in EC treatment is directly related to the rate of coagulant dosage, size of generated flocs and bubble production rate. These in turn influence the effectiveness of pollutant removal in EC treatment process (Amrose et al., 2014; Martinez-Villafane et al., 2009). The quantity of dissolved metal electrode is directly proportional to the charge passed through the solution (Faraday’s law). It is therefore expected that at high applied current,

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anodic dissolution (Faraday’s law, Eqn. 13) increases resulting into increase of hydroxide cationic complexes and their subsequent roles in As removal (Kobya et al., 2015; Vasudevan

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electrode



it

EC

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et al., 2010).

w

zF v

(13)

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where Celectrode (kg Al or Fe electrode per m3 treated groundwater sample) refers to theoretical amount of Al or Fe ions produced by current i (A) for operating time of tEC (s), z refers to the

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number of electrons involved (zFe = 2 and zAl = 3). Mw is the atomic weight (Mw,Al = 26.98

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g/mol, Mw,Fe = 55.85 g/mol), v is the volume (m3) of solution used (400 mL) and F is the

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Faraday’s constant (96485 C/mol).

For Al electrode, the Celectrode to achieve WHO guideline (for ≤ 10 µg As /L) at applied current

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(i) of 0.010 A, 0.025 A, 0.050 and 0.10 A ranged from 0.018-0.026, 0.057-0.062, 0.076-0.179 and 0.154-0.066 kg Al per m3, respectively. Similarly, for Fe electrode, the Celectrode to achieve WHO guideline at applied currents (i) of 0.010 A, 0.025 A, 0.050 and 0.10 A were 0.019,

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0.03, 0.052, 0.092 kg Fe per m3, respectively. Fig. 2 and Fig. 3 illustrate the effect of applied current on treatment of As contaminated ground water with Fe and Al scrap anodes. As can be

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seen, a tremendous drop in As concentration occurred as EC time increased for all values of applied current. The arsenic reduction rate was very sharp in initial stages of the process, but a gradual reduction is observed towards the end of EC process. EC time (tEC) to obtained

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residual As concentration of < 10 µg/L ranged from 6-20 min for Fe electrode and 30-60 min for Al electrodes, for variation in applied current from 0.010-0.10 A, respectively. At higher applied current, particularly 0.1 A, operating time of 6 min for Fe and 30 min for Al electrode was required to achieve As of ≤ 10 μg/L, giving removal efficiencies of > 95.05% for Fe and > 94.52% for Al electrodes. The reduction of required operating time with increase in applied current is associated with the higher metal dissolution rate. It is reported that more 8

Fe2+ or Al3+ ions are obtained at high applied current values according to Eqn. 13, and these consequently increased the removal of As (Gomes, et al., 2007). Initial As concentration for SW-1 and SW-2 were in the same range (145-147 μg/L), and this couldn’t provide a basis for assessment of the effect of initial As concentration for the present study. Nevertheless, literature studies have reported significant decrease in removal efficiencies with rising As concentrations, and As removal trends were the same as those obtained in the present study,

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with sharp removals at the beginning of the EC process (Şık et al., 2017; Wan et al., 2011). From energy consumption point of view, the operational cost per m3 of treated water increases

with charge loading. It is therefore important to optimize the charge loading for EC processes

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depending on the systems requirements and raw water characteristics. The charge loading

(delivered coagulant dosage) in EC process is related to applied current and EC time. The charge loading was calculated from Eq.14 (Chen et al., 2000; Vik et al., 1984):

(14)

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q (C)  i  t EC

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where q is the charge loading (Coulomb), i is applied current (A), and tEC is EC time (s). Very high charge loading leads into excess production of metal hydroxide flocs, which cause

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the concentration of Fe2+ or Al3+ ions to be higher than the required dosage hence affecting the residual metal concentration in the EC treated solution (Mohora et al., 2018). Excess metal

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ions in solution may also cause adverse impact on human health if the EC treated water is consumed. The variation of residual As concentration with charge loading is presented in Fig.

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4. In general, residual As concentration decreased as the charge loading (q) increased. The charge loading required to achieve Csafe (MCL As < 10µg/L) was 12, 15, 24 and 36 Coulombs

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at applied current of 0.010, 0.025, 0.050, and 0.10 A, respectively, for Fe scrap anode. Whereas, the respective charge loading to achieve Csafe for Al scrap electrodes was 36, 90, 240 and 180 Coulombs. Residual arsenic concentration decreased with increase in charge loading at constant current value. In other words, the arsenic removal efficiency increased

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with charge loading. The required EC times to obtain sufficient As removal by Fe scrap anodes are lower than the Al scrap anodes and therefore the calculated charge loading for Fe anodes is lower than Al anodes. Similar results were obtained in the literature (Omwene and Kobya, 2018; Kobya et al., 2016; Heidmann and Calmano, 2008). The As removal capacity per quantity of aluminum or iron dissolved electrochemically (qe, µg As/g Al or Fe) was computed from Eqn.15, and the results are presented in Fig. 5 and Fig. 6.

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qe 

(Ci  Ct )  v Celec, exp

(15)

where Ci is initial As concentrations (µg/L), Ct is the As concentrations at any EC time (µg/L), and v is volume of solution in the reactor (L).

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For applied current of 0.010 - 0.10 A, As removal capacities by Al electrode ranged from 37.89 – 0.62 mg As/g Al for SW-1, and 22.89-0.54 mg As/g Al for SW-2 (Fig. 6). On the

other hand, the removal capacities for Fe electrode ranged from 12.87-0.31 mg As/g Fe for

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SW-1 and 21.31 - 0.36 mg As/g Fe for SW-2 (Fig. 5). Generally, both electrodes showed a decreasing trend in As removal capacity with increase in applied current and EC time (except

at EC time < 5 min, lag phase). The As removal capacity in this study was calculated as the amount of As removed per amount of electrochemically dissolved aluminium or iron (qe, mg

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As/g Al or Fe), as seen in Eqn.15. The initial As removal rate is high, then gradually reduces

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to an almost constant rate, this trend could be explained by a reduction in the amount of As

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available for ligand exchange, co-precipitation and reductıon in adsorption capacity of metal

systems as it tends to equilibrium.

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hydroxides at reduced residual As levels. Also, another reason could be variations in the

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As removal capacities for applied current values of 0.010, 0.025, 0.050 and 0.100 A at EC time of 30 min were 4.99, 1.42, 0.75 and 0.31 mg As/g Fe for SW-1 and 4.04, 1.53, 0.72 and 0.36 mg As/g Fe for SW-2, respectively. In this case, iron dosage dissolving from Fe scrap

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anodes for 0.010, 0.025, 0.050 and 0.100 A were 11.36, 40.63, 77.62 and 183.90 mg Fe for

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SW-1 and 14.17, 13.02, 81.00 and 163.06 mg Fe for SW-2, respectively. Al dosage dissolving from aluminum scraps anodes for 0.010, 0.025, 0.050 and 0.100 A were 7.35, 22.58, 40.27 and 92.28 mg for SW-1 and 10.57, 24.59, 73.82 and 86.57 mg for SW-2, respectively at EC time of 60 min. Arsenic removal capacities at applied currents of 0.010,

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0.025, 0.050 and 0.10 A at EC time of 60 min were calculated as 7.41, 2.08, 1.43 and 0.62 mg As/g Al for SW-1 and 5.26, 2.27, 0.79 and 0.67 mg As/g Al for SW-2, respectively. Detailed arsenic removal capacities for MCL arsenic value (Csafe <10 μg/L) are given in Table 2 and Table 3 for Al and Fe scrap electrodes. For all the ground water samples, arsenic removal efficiency for both scrap anodes was over 99%, and as the applied current increases, the Fe dosage in the EC reactor increases. At the beginning of the EC process, the amount of metal 10

dissolved from the anodes is low, and as the EC time increases, the dissolved scrap anode increases. On the other hand, the amount of arsenic removed from groundwater decreases as the EC time increases. As a result, the amount of arsenic removed per g aluminum and iron dosage in the packed-bed EC reactor decreased with increase in EC time (except for the lag phase) (Figs. 5 and 6).

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Compared with at other literature results on As removal, studies on treatment of Songhua

River water contaminated with As by synthesized amino-functionalized coffee cellulose adsorbent, obtained high adsorption capacity of 46.1 and 13.2  mg/g for As(V), and As(III)

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respectively, at temperature of 313 oK (Hao et al.,2018). Zhu et al., 2018 conducted arsenic

removal in batch adsorption experiments by bismuth-impregnated aluminum oxide, they reported the quantity of adsorbed As to increase with increasing equilibrium As concentration,

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up to a maximum adsorption capacity of 26.8 mg/g, and achieved 91.6% As(III) removal

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efficiency. Nano-iron oxide coated single-wall carbon nanotubes were applied for As removal from water and obtained maximum As (V) adsorption capacity of the composite as 42.30

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mg/g at pH 8 and 49.65 mg/g at pH 4 (Ma et al., 2018). In another study by Vieira et al.,

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2017, on arsenic removal from water using iron-coated seaweeds, 100% of arsenite removal from a mining-influenced water was achieved with maximum adsorption capacities of 4.2

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mg/g and 7.30 mg/g obtained at pH 7 and 20 °C for arsenite and arsenate, respectively. Hu et al., 2015 reported hickory chips biochar impregnated with Fe(III) to have a relatively high

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As(V) removal efficiency, with an As(V) sorption capacity of 2.16 mg/g at pH of around 6. In comparison with the present study, electrochemically generated Al or Fe-hydroxides illustrated in Eqs. 8, 9, 10 and 11 have demonstrated a satisfactory As removal capacity and

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also exhibit a shorter removal time compared to other techniques. It is noteworthy that the optimum conditions for Fe electrode for both ground water samples to achieve residual As of ≤ 10 μg/L were determined to be EC time of 8 min and applied

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current of 0.050 A, whereas for Al scrap electrode, to achieve the same removal conditions, optimums were EC time of 30 min and applied current of 0. 10 A. 3.3. Effects of packed-bed density The impact of packed-bed density on arsenic removal was investigated by altering the mass of aluminium and iron scrap anodes in the packed-bed reactor at applied current of 0.10 A with 11

SW-1 (initial arsenic concentration of 145 μg/L). Packed anode density(mb) of 0.10, 0.20 and 0.40 kg Fe/m3 were investigated for Fe anode, whereas for Al anode, packed anode densities of 0.020, 0.040 and 0.080 kg Al/m3 were investigated. Optimum EC times and residual As concentrations (to achieve permissible As of ≤ 10µg/L) for Fe scrap anodes at 0.10, 0.20 and 0.40 kg Fe/m3 were 15 min and 8.2 μg/L, 12 min and 6.10 μg/L, and 6 min and 7.95 μg/L, respectively. On the other hand, the respective values for Al anodes were 40 min and 11.40 μg/L at 0.020 kg/m3, 35 min and 8.80 μg/L at 0.040 kg/m3 and 25 min and 8.97 μg/L at 0.080

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kg/m3. It was observed that the EC times required for arsenic removal decreased with increase

of aluminium and iron scraps dosage in the packed-bed EC reactor. Generally, the number of

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microscopic galvanic cells in the packed-bed EC reactor is positively proportional to the dosage of Al and Fe scraps. Therefore, it is reasonable that the arsenic removal rates were enhanced by increase of aluminium and iron scrap dosage in the reactor. According to these results, the Fe scrap anodes still exhibited better performance than the Al scrap electrodes at

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reduced bed density. In addition, the decline in removal performances with decrease in packed

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bed density could be explained by the fact that loose metal scraps cause a reduction in contact of anode material, which raises the external resistance from the solution, hence lowering EC

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performance (Ye et al., 2016). For Fe scrap anodes of 0.10, 0.20 and 0.40 kg/m3, energy

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consumptions at optimum EC times were 0.344, 0.275 and 0.138 kWh/m3, respectively, whereas the respective electrode consumptions were 0.020, 0.048, 0.092 kg/m3. Similarly, for

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Al scrap anodes of 0.020, 0.040 and 0.080 kg/m3, energy consumptions were calculated as 1.187, 1.037, and 0.741 kWh/m3, respectively, whereas the electrode consumptions were

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calculated as 0.041, 0.082, 0.154 kg/m3, respectively. The relatively higher energy and electrode consumptions at lower packed bed density could be due to the increase in EC time to achieve the required As removal efficiency. It is observed that the consumption of Al

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anodes is higher than that of Fe anodes due to the higher optimum EC time. Higher arsenic removal efficiencies were achieved at lower EC times with Fe scrap anodes (such as 6 min at

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0.40 kg/m3) compared to Al scrap anodes.

At present, very limited literature exists on As removal by Fe and Al scrap anodes in EC process from real groundwater. In this regard, we discuss related studies for comparison purposes. Arsenic removal by electrocoagulation using Al plate anodes was performed in a 1.4 L batch reactor for real groundwater (initial As concentration of 512 μg/L) from Kaudikasa village in India, and obtained 98.41% removal efficiency at optimum conditions 12

(current density of 10 A/m2, initial pH of 7, EC time of 95 min). The estimated operating cost of the study was 0.357 US $/m3 (Thakur et al., 2017). Alcacio et al. 2014 investigated As removal from a deep well (initial As concentration of 134 μg/L and initial pH of 6.8) in Central Mexico by electrocoagulation using sacrificial Al anode and the respective As removal efficiency and energy consumption were reported as 89.6% and 0.89 kWh/m3 at the optimums (mean linear flow rate of 0.91 cm/s and current density of 60 A/m2). A Similar study by Sandova et al.,2018 involving simultaneous removal of fluoride and arsenic from

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contaminated groundwater containing initial arsenic and fluoride concentrations of 50.4 µg /L

and 5.5 mg/L using sacrificial aluminum anode in a continuous filter-press reactor, achieved

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WHO standards (fluoride < 1.5 mg/L, arsenic < 10 µg/L) at mean linear flow rate of 0.23 cm/s

and current densities of 60  and 70 A/m2. Another research by Mohora et al., 2018 on As removal from raw groundwater without pH modifications, obtained 96% (effluent concentration of 1.52 µg/L) As removal within 4 h of experimental runs at optimums; charge

U

loading of 54 C/L, current density of 1.98 A/m2 and flow rate of 12 L/h. It is challenging to

N

compare the results of various As removal studies by EC due to the relatively different operational conditions of EC reactors such as flow rates, applied current, electrode types,

A

operation pH and initial As concentration, among others. However, all the results clearly point

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out that EC is a very efficient technology for aqueous As removal.

3.4. Operating cost, energy and electrode consumptions

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For a treatment process to be embraced by any community or funder, its economic viability is also an important factor to be taken into consideration, besides its technical efficiency.

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Electricity and electrode consumptions are the major costs of concern in electrocoagulation treatment process. In this regard, optimizing the electricity consumption is of paramount importance. The energy (Cenergy) and electrode (Celectrode) consumptions for this study were

A

calculated using the Eqs. 16 and 17. Cenergy (kWh/m 3 ) 

Celectrode (kg/m 3 ) 

i  tEC  U v

i  t EC  M Al or Fe z F v

(16)

(17)

13

where i is applied current (A), tEC is EC time (sec for electrode consumption and hour for energy consumption), U is cell voltage (V), v is volume (m3) of the wastewater in the EC reactor. Details on operating cost, energy and electrode consumptions for Fe and Al electrodes are given in Table 2 and Table 3, respectively. Although As removal efficiency was enhanced by increase of applied current and EC time, the cell voltage raised at higher current values, which

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increased overall energy consumption. For SW-1 groundwater, Cenergy to obtain As ≤ 10 μg/L for applied current of 0.010, 0.025, 0.05 and 0.1 A for Al and Fe electrodes were 0.009 and

0.042 kWh/m3, 0.016 and 0.185 kWh/m3, 0.070 and 0.040 kWh/m3, and 0.138 and 1.187

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kWh/m3, respectively. Similarly, for SW-2, Cenergy to obtain As ≤ 10 μg/L at applied current

of 0.010, 0.025, 0.050 and 0.100 A for Al and Fe electrodes were respectively, 0.04 and 0.01 kWh/m3, 0.132 and 0.017 kWh/m3, 0.328 and 0.070 kWh/m3, and 0.876 and 0.131 kWh/m3.

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Meanwhile, Celectrode for SW-1 groundwater, to obtain residual As ≤ 10μg/L at 0.010, 0.025,

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0.050 and 0.10 A for Al and Fe electrodes were 0.018 and 0.019 kg/m3, 0.057 and 0.034 kg/m3, 0.076 and 0.052 kg/m3, and 0.154 and 0.092 kg/m3, respectively. For the SW-2, the

A

electrode consumptions per m3 treated groundwater to obtain As ≤ 10 μg/L at 0.010, 0.025,

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0.050 and 0.10 A for Al and Fe electrodes were 0.0264 and 0.01892 kg/m3, 0.0615 and 0.0315 kg/m3, 0.1790 and 0.0518 kg/m3, and 0.0663 and 0.0920 kg/m3, respectively. Generally,

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Cenergy and Celectrode for were lower for Fe electrode compared to Al electrode. The operating cost (OC, US $/m3) was calculated using Eqn. (18):



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

OC $/m 3  α  Cenergy  β  Celectrode

(18)

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Where α electrical energy is cost per unit and β is unit electrode cost. According to March 2018 survey in the Turkish market, the price of electricity was 0.120 US $/kWh and Al and Fe scraps costs were 1.15 US $/kg and 0.20 US $/kg, respectively. To achieve an effluent of As

A

< 10 µg/L, the operating costs for both the groundwater samples ranged from 0.005 to 0.02 US $/m3 for variation in applied current of 0.010 to 0.10 A, for Fe electrode. On the other hand, the operating costs for Al electrodes ranged from 0.026 to 0.32 US $/m3 for the same variations in applied current (Table 2).

14

If the EC process is to be developed into a large-scale treatment plant to meet the water demands of affected communities, we recommend that renewable energy sources like solar, wind and small-scale hydropower plants should be used depending on the feasibility of the site. In this way, the operating costs shall be tremendously cut down. In comparison with a survey by Sorg at al., 2015, in which the cost of arsenic removal from groundwater in USA using

different

removal

technologies

such

as

adsorptive

media,

iron

removal,

coagulation/filtration, ion exchange and reverse osmosis to reduce arsenic to less than 10 μg/L

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was investigated, the operation costs obtained in the present study are lower and more

promising. In the above-mentioned study, Sorg at al., 2015 reported operation and

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maintenance cost at system flow rate of 379 L/min for adsorptive media, ion exchange, and

combined iron removal with coagulation/filtration to be respectively, 6.66, 1.85 and 1.06 US $/m3. It is obvious that the inconsistencies in operating costs arise from the differences in experimental conditions. Thus, the use of scrap electrodes proves to be a cost-effective

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3.5. Characterization of Treatment Sludge

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technology for arsenic removal from groundwater.

The sludge produced by EC process is another important factor considered in application of EC treatment. The properties of generated EC sludge are dictated by the characteristics of

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water or wastewater, applied current and hydraulic residence time. Sludge from Electrocoagulation is better to handle due to its dewatering potential and floc size. However,

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arsenic-sludge may pose a threat to groundwater if disposed to open environment. Safe disposal of As- containing sludge necessitates long term solutions like solidifications of As-

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sludge and applying it for purposes such as making bricks and concretes among others (Mandal et al.,2016 ). In this study, the electro-generated sludge was oven dried (105 ± 5 oC) before FTIR and SEM analyses. FTIR analysis was used to assess the chemical bonds of the obtained flocs. Fig. 8 presents the FTIR spectra of the generated sludge with both Fe and Al

A

scrap electrode. The FTIR analysis for generated sludge was performed in 3600–650 cm-1 range. The FTIR spectrum for both sludge samples was characterized by a broad peak at about 3436 cm-1, this is attributed to O-H stretching vibrations (Gross et al.,2007). For Fe scrap electrode, the peak at about 794 cm-1 is attributed to the bonding between Fe(III) and As(V) before hydrolysis (Jia et al., 2007; Myneni et al., 1998; Sun et al., 1996). Moreover, studies by Tokoro et al., 2010 on arsenic removal from arsenic contaminated waters showed 15

the formation of FeAsO4 at near neutral pH conditions. For Al electrodes, the speaks at about 1435 cm-1 and 851 cm-1 could be due to Al-O-Si and As-O bond, respectively, confirming the As removal by adsorption (Drouiche, et al, 2009; Ghosh et al., 2008; Guzman et al., 2016). The SEM images in Fig. 9(a) and Fig. 9 (b) were used to analyze the surface topography of the generated EC sludges (SEM, Philips XL30S-FEG) for Fe and Al scrap electrodes, respectively. These images are characterized by an abundant amorphous structure with an

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aggregate size of 10-50 μm. The floc sizes obtained from Al electrode are larger than the flocs obtained from Fe electrode.

EDAX analysis was used to determine the elements present in the electro-coagulated sludge.

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The results obtained for Fe and Al electrodes are presented in Fig.10 and Fig.11 respectively. For Fe electrode, half of the sludge consisted of iron by weight (Table 4). Neglecting oxygen, which is naturally found in the environment, the second highest element in the sludge is As.

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Whereas for Al electrode sludge, As and Al were the most predominant elements, neglecting oxygen and aurum, which is used for coating during the EDAX analysis. It is supported by

N

EDAX analysis that As is removed by floc formation. Besides, the amount of As in the sludge

A

from Fe electrode is slightly high.

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4. Conclusion

Electrocoagulation for treatment of As contaminated ground waters was performed in an EC

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reactor operated in a batch mode with Fe and Al scrap anodes. The influence of applied current, electrolysis time and packed bed density to obtain the required WHO drinking water

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standard of 10 µg/L were investigated. Both Al and Fe scrap anodes reduced the residual As level to the WHO standard. However, Fe scrap anodes performed much better than Al electrodes. Optimum conditions for Fe electrode to achieve residual As level of 10 μg/L

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(93.5% removal) were calculated as EC time of 8 min and applied current of 0.050 A. Similarly, for Al scrap electrode, the optimums were EC time of 30 min and applied current of 0.10 A. The operating cost, energy (Cenergy) and electrode (Celectrode) consumptions for Fe

A

scrap electrodes at their optimums were 0.017 US $/m3, 0.070 kWh/m3 and 0.052 kg/m3, respectively. Whereas for Al scrap electrode, operating cost, energy (Cenergy) and electrode (Celectrode) consumptions were 0.181 US $/m3, 0.876 kWh/m3 and 0.0663 kg/m3, respectively at its optimums. The results also showed that As removal slightly decreased with decrease in anode bed density. The SEM images of the sludge samples showed irregular shaped particles (amorphous structure).

16

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lab-scale column. J. Environ. Manag. 211, 73-82.

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aluminum oxide for arsenic removal: Adsorption mechanism study and application to a

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Figure 1. Experimental set-up using packed-bed electro-coagulator 160 SW-1

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80 60 40

Csafe

20 0

0

5

10

15 20 EC Time (min)

Current (i, A) 0.010 0.025 0.050 0.100

SW-2 Residual arsenic concentration (g/L)

100

160 140

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120

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Residual arsenic concentration (g/L)

140

Current (i, A) 0.010 0.025 0.050 0.100

120 100 80 60 40

Csafe

20 0

25

30

35

0

5

10

15 20 EC Time (min)

25

30

35

A

Figure 2. Effect of applied current on As removal with Fe scrap anodes.

23

160

160

SW-1

Current (i, A) 0.010 0.025 0.050 0.100

120

SW-2

Current (i, A) 0.010 0.025 0.050 0.100

140 Residual arsenic concentration (g/L)

100 80 60 40 Csafe 20

120 100 80 60 40

Csafe

20 0

0 0

5

10

15

20

25

30 35 40 EC Time (min)

45

50

55

60

65

0

5

10

15

20

25 30 35 40 EC Time (min)

100 80 60 40 Csafe 20

60

65

120 100

Current (i, A) 0.010 0.025 0.050 0.100

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120

SW-2

140

80 60 40

A

Residual arsenic concentration ( g/L)

160

Current (i, A) 0.010 0.025 0.050 0.100

55

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SW-1

140

Residual arsenic concentration (g/L)

160

50

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Figure 3. Effect of applied current on As removal with Al scrap anodes.

45

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Residual arsenic concentration (g/L)

140

Csafe

20

0

0 5

10

15 20 25 30 Charge loading (Coulombs)

35

40

0

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0

5

10

15 20 25 30 Charge loading (Coulombs)

35

40

10 8

22

Current (i, A) 0.010 0.025 0.050 0.100

PT

12

CC E

Arsenic removal capacity (mg As/g Fe)

SW-1

6 4 2 0

A

0

5

10

15 20 EC Time (min)

SW-2

20 Arsenic removal capacity (mg As/g Fe)

14

ED

Figure 4. Effect of charge loading on As removal from groundwater samples with Fe scraps.

Current (i, A) 0.010 0.025 0.050 0.100

18 16 14 12 10 8 6 4 2 0

25

30

35

0

5

10

15 20 EC Time (min)

25

30

35

Figure 5. Arsenic removal capacity for Fe scrap anodes.

24

40

24

30 25 20 15 10 5

5

10

15

20

25 30 35 40 EC Time (min)

45

50

55

60

Current (i, A) 0.010 0.025 0.050 0.100

20 18 16 14 12 10 8 6 4 2

0 0

SW-2

22

0

65

0

5

10

15

20

25

30 35 40 EC Time (min)

45

160

160

65

3

60 40 Csafe

0

120 100 80

U

80

20

60

0.020 0.040 0.080

60

N

100

140

40 20

A

120

55

mb (kg Al/m )

0.10 0.20 0.40

Residual arsenic concentration (g/L)

Residual arsenic concentration (g/L)

3

mb (kg Fe/m )

140

50

SC R

Figure 6. Arsenic removal capacity for Al scrap anodes.

IP T

Current (i, A) 0.010 0.025 0.050 0.100

Arsenic removal capacity (mg As/g Al)

Arsenic removal capacity (mg As/g Al)

SW-1 35

Csafe

0

5

10

15 20 EC Time (min)

25

30

35

0

5

10

15

20

25 30 35 EC Time (min)

40

45

50

55

M

0

104.5 104.0

Fe electrode Al electrode

103.5 103.0

PT

102.5 102.0 101.5 101.0 100.5

CC E

Transmittance (A)

ED

Figure 7. Effects of packed anode bed density for SW-1 (Current = 0.10 A).

100.0

99.5 99.0 98.5 98.0

A

97.5 97.0 96.5 96.0

3500

3000

2500

2000

1500

1000

-1

Wave number (cm )

Figure 8. FTIR analysis of EC generated sludge with Fe scrap and Al scrap electrodes.

25

(b)

IP T

(a)

Figure 9. SEM micrographs of Fe (a) and Al (b) flocs obtained from the EC process.

9000

7000

Fe K

6000

0 0.0

0.5

1.0

1.5

2.0

2.5

3.0

N

M

1000

A

Ca K

2000

Si K

Ca L

3000

Ca K

Fe L

Au M

4000

3.5

4.0

4.5

5.0

5.5

6.0

6.5

7.0

Au L

5000

U

As L

Counts

8000

Fe K

10000

SC R

OK

11000

7.5

8.0

8.5

9.0

9.5

10.0

Energy (keV)

OK

14000

PT

13000

Al K

12000 11000 10000 9000

CC E 8000

As L

7000 6000

Ca L

5000 4000

Au M

Counts

ED

Figure 10. EDAX analysis of sludge generated by Fe electrode

Au L

Fe K

Fe K

Ca K

1000

Ca K

2000

Si K

Fe L

A

3000

0 0.0

0.5

1.0

1.5

2.0

2.5

3.0

3.5

4.0

4.5

5.0

5.5

6.0

6.5

7.0

7.5

8.0

8.5

9.0

9.5

10.0

Energy (keV)

Figure 11. EDAX analysis of sludge generated by Al electrode.

26

SW-1

SW-2

pH

7.42

8.01

Conductivity (µS/cm)

1004

904

TDS (mg/L)

431

387

Alkalinity (mg/L HCO3-)

440

444

DO (mg/L)

7.8

6.9

TOC (mg/L)

3.2

4.8

PO4-P (mg/L)

1.52

1.16

Chloride (mg/L)

9

11

Sulphate (mg/L)

30

Al (mg/L)

0.026

Fe (mg/L)

0.012

Ca (mg/L)

79.6

Mg (mg/L)

74.74

As (µg/L)

145

Si (mg/L)

28.64

SC R

Parameter

IP T

Table 1. Chemical analyses of arsenic contaminated groundwater samples.

23

0.037 0.024

U

81.2

146.4 30.05

A

CC E

PT

ED

M

A

N

70.06

27

Table 2. Parameters to achieve the required MCL As concentration (10 μg/L ) by Fe scraps. i

Sample (A)

q

(μg/L) (min) (C)

U

Cenergy

Celectrode OC

qe

(V)

(kWh/m3) (kg/m3) (US $ /m3) (mg/g) (mg/C)

0.010 7.36

20

12

1.05

0.009

0.019

0.005

7.270

0.005

0.025 6.09

10

15

1.58

0.016

0.034

0.009

4.100

0.004

0.050 4.27

8

24

4.2

0.070

0.052

0.017

2.630

0.002

0.100 7.95

6

36

5.5

0.138

0.092

0.020

1.490

0.002

0.010 7.68

20

12

1.16

0.010

0.019

0.005

5.870

0.005

0.025 9.01

10

15

1.65

0.017

0.032

0.008

4.370

0.004

0.050 4.26

8

24

4.2

0.070

0.052

0.019

2.630

0.002

0.100 7.67

6

36

5.25

0.131

0.092

0.020

1.700

0.002

A

CC E

PT

ED

M

A

N

U

SW-2

tEC

IP T

SW-1

Cfinal

SC R

Water

28

Table 3. Parameters to achieve the required MCL As concentration (10 μg/L ) by Al scraps. Cfinal

tEC

q

Sample (A)

(μg/L) (min) (C)

U

Cenergy

Celectroe

OC

qe

(V)

(kWh/m3) (kg/m3) (US $ /m3) (mg/g) (mg/C)

60

36

1.67 0.042

0.018

0.026

7.410

0.0015

SW-1 0.025 8.63

50

75

3.56 0.185

0.057

0.087

2.420

0.0007

0.050 8.66

45

270

4.27 0.040

0.076

0.135

1.820

0.0002

0.100 6.54

40

240

7.12 1.187

0.154

0.319

0.900

0.0002

0.010 7.33

60

36

1.59 0.040

0.026

0.035

5.260

0.0016

SW-2 0.025 6.90

60

90

2.11 0.132

0.062

0.087

2.270

0.0006

0.050 7.81

40

240

3.94 0.328

0.179

0.100 7.25

30

180

7.01 0.876

0.066

IP T

0.010 8.95

SC R

Water i

0.770

0.0002

0.181

2.100

0.0003

A

CC E

PT

ED

M

A

N

U

0.245

29

Net Intensity Error 0.00

AsL

15.18

7.45

478.65

0.01

AsL

16.8

6.01

631.02

0.01

SiK

0.15

0.20

14.22

0.27

SiK

0.32

0.3

25.22

0.12

AuM

6.95

1.30

272.07

0.01

AuM

17.45 2.37

549.04

0.01

CaK

3.23

2.96

204.62

0.03

CaK

3.25

2.17

156.29

0.04

FeK

50.66

33.34

1173.24 0.01

FeK

1.69

0.18

31.81

0.16

AlK

19.1

18.98 1562.53 0.00

A

CC E

PT

ED

M

A

N

U

SC R

IP T

Table 4. Distribution of elements detected in EDAX analysis Element Fe electrode Element Al electrode Weight Atomic Net Net Weigh Atomi Net (%) (%) Intensity Intensity t (%) c (%) Intensity Error OK 23.83 54.76 1196.19 0.00 OK 41.4 69.36 1479.43

30