A review on decontamination of arsenic-contained water by electrocoagulation: Reactor configurations and operating cost along with removal mechanisms

A review on decontamination of arsenic-contained water by electrocoagulation: Reactor configurations and operating cost along with removal mechanisms

Environmental Technology & Innovation 17 (2020) 100519 Contents lists available at ScienceDirect Environmental Technology & Innovation journal homep...

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Environmental Technology & Innovation 17 (2020) 100519

Contents lists available at ScienceDirect

Environmental Technology & Innovation journal homepage: www.elsevier.com/locate/eti

A review on decontamination of arsenic-contained water by electrocoagulation: Reactor configurations and operating cost along with removal mechanisms ∗

Mehmet Kobya a , , Reza Darvishi Cheshmeh Soltani b , Philip Isaac Omwene a , ∗ Alireza Khataee a,c , a

Department of Environmental Engineering, Gebze Technical University, 41400 Kocaeli, Turkey Department of Environmental Health Engineering, School of Health, Arak University of Medical Sciences, 38196-93345 Arak, Iran c Research Laboratory of Advanced Water and Wastewater Treatment Processes, Department of Applied Chemistry, Faculty of Chemistry, University of Tabriz, 51666-16471 Tabriz, Iran b

article

info

Article history: Received 6 August 2019 Received in revised form 23 October 2019 Accepted 24 October 2019 Available online xxxx Keywords: Arsenic pollution Arsenic removal Groundwater Electrocoagulation Operational parameters

a b s t r a c t Pollution of water resources by arsenic (As) that originates from both natural and anthropogenic sources is a serious matter causing health problems to millions of people worldwide due to the toxic effects of this ionic pollutant. To attain conformity with strict Maximum Contaminant Level (MCL) of As (10 µg/L), electrocoagulation (EC) is considered as an advantageous process for the removal of As because of high removal efficiency, simplicity, cost-effectiveness, feasibility of small scale operations and lower chemical requirement in comparison with other treatment processes. In this regard, this review discusses the applications and performance results of EC process for arsenic removal, taking into account the drawbacks and limitations of EC technologies. The mechanism and theoretical aspects of arsenic removal by EC was reviewed with details. The effects of operational parameters on the efficiency of EC process, including current density, charge loading and initial pH, as well as reactor configurations and operating cost of the process were reviewed. The amount of sludge produced during EC process, characterization and disposal methods were investigated and the simultaneous removal of As with other contaminants from water presented. Furthermore, examples of pilot and full-scale applications of EC for the removal of arsenic were provided. Concluding remarks and outlook of this field of study with respect to new areas of research are also discussed. © 2019 Published by Elsevier B.V.

Contents 1. 2. 3. 4.

Introduction............................................................................................................................................................................................... Arsenic removal technologies ................................................................................................................................................................. Mechanism and theoretical aspects of arsenic removal by EC........................................................................................................... Effect of operational parameters ............................................................................................................................................................ 4.1. Effect of electrode material and shape ..................................................................................................................................... 4.2. Effect of current density and charge loading ........................................................................................................................... 4.3. Effect of initial pH .......................................................................................................................................................................

∗ Corresponding authors. E-mail addresses: [email protected] (M. Kobya), [email protected] (A. Khataee). https://doi.org/10.1016/j.eti.2019.100519 2352-1864/© 2019 Published by Elsevier B.V.

2 3 9 10 10 13 16

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M. Kobya, R.D.C. Soltani, P.I. Omwene et al. / Environmental Technology & Innovation 17 (2020) 100519

5. 6. 7. 8. 9.

Sludge production during EC process, characterization and disposal................................................................................................ EC reactor configurations and operating cost ....................................................................................................................................... Simultaneous removal of As and other contaminants......................................................................................................................... Pilot and large scale EC plants for arsenic removal............................................................................................................................. Conclusion and outlook ........................................................................................................................................................................... Declaration of competing interest.......................................................................................................................................................... References .................................................................................................................................................................................................

16 17 20 21 22 23 23

1. Introduction Arsenic (As) is the twentieth most abundant element on earth’s crust. However, it is a toxic metalloid of natural geogenic origin and considered as a significant worldwide groundwater contaminant, profound along major rivers and large deltas of South and East Asia and South American countries like Argentina and Chile (Ravenscroft et al., 2009; Bundschuh et al., 2012; Buschmann and Berg, 2009; Hashim et al., 2019b). The population defined as being at risk of As exposure worldwide has increased due to the reduction of its MCL in drinking water from 50 µg/L to 10 µg/L (Organization, 2004). In view of natural As contamination for over 70 countries, groundwater concentration of As reported in the literature largely changes in the range of < 0.10–163,000 µg/L (Ravenscroft et al., 2009). Some countries affected by As contamination include: Argentina, Bangladesh, Chile, Hungary, Canada, Pakistan, China, Mexico, Taiwan, South Africa, USA, Vietnam and West Bengal (India). The South and Southeast Asian As Belt are considered the most arsenic polluted areas; these include India, Bangladesh, Nepal, Vietnam and China (Ravenscroft et al., 2009). The Pacific Plains of Chile in South America and Pampean Plains of Argentina are acknowledged to have areas of grave arsenic contamination (Bundschuh et al., 2012). The USA and Canada in North America experience very prevalent levels of arsenic contamination, however, the concentrations are distinctively lower in comparison with Asia (Sorg et al., 2014). Severe arsenic pollution in Europe has been noted in Hungary, in South and Central America, the volcanic mountains of the Pacific Rim are also characterized by extreme As contamination. Also, geothermal sources in Iceland and Hawaii are indicated to be a source of As contamination (Ravenscroft et al., 2009). Although dependable approximation of the number of people exposed to As via water is not available, rough approximations indicate that over 100 million people are at danger of exposure to As concentrations of higher than 10 µg/L, with the worst affected area currently being the Bengal delta (Ravenscroft et al., 2009; Buschmann and Berg, 2009). About 27% of groundwater samples in Bangladesh exceeded local As standard of 50 µg/L. Bangladesh is experiencing the highest ever reported mass poisoning of its population as a result of inorganic As contamination of groundwater. Studies suggest that 28.0 to 62.0% of the inhabitants of Bangladesh are susceptible to As-contaminated drinking water (Smith et al., 2000). Over 10 million habitants of Mekong and Red River (Bengal Delta and North Vietnam) are exposed to harmful effects of extreme concentrations of As in groundwater (Buschmann and Berg, 2009). At Ganges delta, about 69.0% of wells between 11.0 and 15.8 m depth contain As concentration of higher than 1000 µg/L. More than 5 million people in Northern China are exposed to high As concentrations in drinking groundwater ranging from 40 to 750 µg/L (Ravenscroft et al., 2009; Buschmann and Berg, 2009). Arsenic concentrations in groundwater of different parts of the world (worldwide occurrences) are displayed in Table 1. It is worth noting that most of these severely affected regions are rural areas of low-income communities. Oxidizing and reducing environmental conditions affect the global distribution of As contamination (Ravenscroft et al., 2009; Smedley and Kinniburgh, 2002). Anthropogenic activities such as geothermal discharges, mining and agricultural applications also lead to As pollution being dissipated into the environment (Ravenscroft et al., 2009; Alsina et al., 2014). Millions of humans worldwide are exposed to As beyond the threshold value of 10 µg/L via drinking. Arsenic is verified as a cause of various serious health effects. The International Agency for Research on Cancer (IARC) classified As in drinking water as carcinogenic, providing evidence for the lung, bladder, and skin cancers (I.W.G.o.t.E.o.C.R.t. Humans et al., 2004). Also, non-cancer health effects on human cardiovascular, neurologic, reproductive and pulmonary systems have been joined to chronic effects of As contamination (Rahman et al., 2009). Blackfoot disease hyperendemic regions of southern Taiwan have As levels in waters in the range of 470–897 µg/L (Tseng, 1977). In addition, a dramatic rise in mortality ascribed to internal cancers has been noted in Chile and Taiwan (Smith et al., 1998; Hopenhayn-Rich et al., 2000). Deaths of 5–10% of people in northern Chile were attributed to bladder and lung cancers, however, a reduction in deaths was observed after preventive strategies were applied. Other researchers have observed human health effects such as keratosis, depigmentation, hyperkeratosis and melanosis among people leaving in arsenic affected communities (Tseng, 1977; Smith et al., 1998; Hopenhayn-Rich et al., 2000). Thus, there is an urgent need for the development of modest, cost-effective and efficient As removal techniques. Furthermore, setting lower MCL value (<10 µg/L) necessitates As treatment techniques desirable for decontamination of water resources with low ambient As concentrations.

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Table 1 Arsenic concentrations in groundwater of different parts of the world (worldwide occurrences). Geographical region

Asia

Country

MPL (µg/L)

District

Resource and amount

C0 (µg/L)

Reference

Bangladesh

50

Ganges delta

69% of wells between 11 and 15.8 m depth

>1000

Chowdhury et al. (1999)

Pakistan

50

South-western Punjab

58% of samples

906

Nickson et al. (2005)

Shanxi

Sun (2004)

50

52% of 3079 investigated wells 11% of 5885 investigated wells 12% of 8200 investigated wells

>50

China

Inner Mongolia Jilin

India

10

>50 >50

West Bengal

52% of 101 934 samples collected from tube wells

>10

Rahman et al. (2001)

Central India

64 samples of well and tube-well water

15–825

Patel et al. (2005)

5% of 15 000 tube-wells

>50

Pokhrel et al. (2009)

48% of tube-wells

>50

Berg et al. (2001)

20% of tube-wells

>150

Nepal

50

Terai

Vietnam

10

Hanoi

Taiwan

10

NA

Groundwater of 200,000 population

10–1820

Sarkar and Paul (2016)

Thailand

10

NA

Groundwater of 15 000 population

1–>5000

Sarkar and Paul (2016)

USA

10

Verde Valley, Arizona Comarca Lagunera, New Mexico

456 selected samples

16

Camacho et al. (2011)

73 selected samples

41

Ellis Pool, Alberta

NA

230

Virden, Manitoba

NA

65–70

Zimapán Valley

NA

1097

Armienta et al. (1997)

Baja California Sur

41% of the monitored sites

>10 with maximum

Wurl et al. (2014)

La Pampa

73% of investigated groundwater samples

>50

Smedley et al. (2002)

Atacama Desert

Beneath 30–180 m of Miocene piedmont gravels NA

278

Leybourne and Cameron (2008)

60–80

Sancha (2006)

253

Abiye and Bhattacharya (2019)

America Canada

10

Mexico

50

Argentina

50

Chile

50

Northern zone Namaqualand South Africa

10

value of 450

Aquifers

Africa West of Johannesburg Karoo Ghana

50

Obuasi area in the Ashanti region Olgatanga area of the Upper East region

Wang and Mulligan (2006)

6150 500 Both shallow and deep groundwater

1–64

Smedley (1996)

1–141

(continued on next page)

2. Arsenic removal technologies As exists in natural water normally in either trivalent form (arsenite; As(III)) and/or pentavalent form (arsenate; As(V)). The latter is the most common form in surface water resources such as lakes and rivers (Tabaraki and Heidarizadi, 2018). Under anoxic conditions typically found in the groundwater, the predominant form is As(III) (Smedley and Kinniburgh, 2002; Nidheesh et al., 2018). At pH range of 2–12, predominant species of As(V) are H2 AsO4 − or HAsO4

4

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Table 1 (continued). Geographical region

Country

MPL (µg/L)

Hungary Europe

Oceania

Serbia

10

District

Resource and amount

C0 (µg/L)

Reference

Southern Hungary

Arsenic-contaminated wells 66 samples taken from 10 sampling sites

20 to 260

Petrusevski et al. (2007) Devic et al. (2014)

Natural geological sources Groundwater of 150,000 population

10–610

Banat

Croatia

Eastern Croatia

Greece

NA

Australia

10 Northern New South Wales

2−

25.7–137.8

1–1840

At a depth of 10–11 m

52–85

At a depth of 25 m

337

Habuda-Stanić et al. (2007) Sarkar and Paul (2016) Smith et al. (2003)

. On the other hand, As(III) mainly exist as H3 AsO3 at pH below 9 and H2 AsO3 − is dominant at pH values of 9–12. It is worth mentioning that As(III) is more soluble, which increases its mobility. As(III) has much more severe effects on human health and is the dominant arsenic species in groundwater (Vaclavikova et al., 2008; Mondal et al., 2013). The hylogenesis of As(V) in aqueous solution with respect to pH is H2 AsO4 − at 3.6 < pH < 7.2; HAsO4 2− at 7.2 < pH < 12.4; AsO4 3− at pH >12.4. Solution pH is an important operational parameter during the removal of As in water via various treatment techniques (Vaclavikova et al., 2008; Mondal et al., 2013). Several treatment technologies, including advanced and convectional treatment processes, have been suggested for arsenic removal from water both at laboratory scale and full scale of operation (Ng et al., 2004). Conventional and non-conventional treatment technologies for the remediation of arsenic-contained water, along with their features, are listed in Table 2. Some of the best available technologies for As removal as reported by United States Environmental Protection Agency (U.S.EPA) include activated alumina (95%), enhanced coagulation/filtration (95%), ion-exchange (95%), enhanced lime softening (90%), coagulationassisted microfiltration (90%), electrodialysis (85%), and oxidation/filtration (80%) (U.S.EPA, 2000). However, these methods are only effective for the treatment of pentavalent As which is mostly found in surface water, not groundwater. Preoxidation chlorine may be necessary for the efficient removal of As (III) (Hering et al., 1997). Fe based (ferric chloride and ferric sulfate) and Al based (aluminum sulfate) coagulants are the most commonly used chemicals for the removal of As through coagulation. Iron hydroxides show stability over a wider range of pH and have a high affinity for As compared to aluminum hydroxide (alum). These characteristics promote precipitation/co-precipitation of As ions in aquatic phase (Hering et al., 1997). The efficiency of treatment systems designed for As removal is highly influenced by the type and dosage of coagulant, pH of the solution and availability of other competing anions. The optimum pH range for both ferric and aluminum sulfate is between 5.0 and 8.0. Water quality, especially the iron content of water affects the required coagulant dosage. Nevertheless, it has been reported that FeCl3 dosage of > 20 mg/L or alum dosage of >40 mg/L is able to remove more than 90% of As(V) ions from aquatic solutions (McNeill and Edwards, 1995). The ion-exchange, reverse osmosis and electrodialysis processes have also been found effective for the removal of arsenic, but they are not cost-effective treatment processes, producing secondary wastes. The aforementioned conventional treatment processes for the removal of As from water have some advantages, but it is clear that they also have some drawbacks. The application of adsorption process only transfers the target pollutant from the liquid phase to the solid phase (Darvishi Cheshmeh Soltani et al., 2015; Hassani et al., 2015; Soltani et al., 2009). Over the past few years, EC treatment technology has attracted great attention among many researchers as an advanced treatment technology for water and wastewater treatment (Chen, 2004; Mollah et al., 2001; Hashim et al., 2020; Müller et al., 2019; Yavuz and Ögütveren, 2018; McBeath et al., 2018). This is primarily attributed to the compact treatment facility and high removal efficiency associated with EC process compared to conventional treatment technologies (Hashim et al., 2019a; Islam, 2019). Furthermore, EC process is recognized as the cheapest treatment technique because of its cost-effectiveness and suitability for small scale operations (Gomes et al., 2007; Kuan et al., 2009; Gilhotra et al., 2018; López-Guzmán et al., 2019a; Song et al., 2018; Mroczek et al., 2019a; Rosales et al., 2018; Graça et al., 2019). Comparatively, EC is considered as a unique and powerful treatment technique to solve arsenic problem (Ghosh et al., 2019). During EC process, the anode (as Fe or Al electrodes) is electrochemically dissolved, resulting in the formation of various metal hydroxides and coagulant species. As-generated metal hydroxides aid in the removal of both dissolved and suspended contaminants through the formation of aggregates and adsorption as described in the next section. Using EC process, the As removal efficiency of up to 99.9% from both ground and surface waters has been reported in the literature (Tables 3 and 4). The data provided in Tables 3 and 4 shows the high potential of the EC process for efficient removal of As with various species from aquatic phases. According to Table 3, As with wide concentration range of 36–100,000 µg/L has been effectively removed using Fe anode-equipped EC process in both surface and ground waters. This indicates high tolerance of the EC process against shock loads of As ions as the target pollutant. As can be observed in Table 4, similar results have been obtained in the case of the application of Al and hybrid electrodes. Moreover, it should be noted that flow regime does not meaningfully influence the efficiency of the EC process. The explanations about the findings, represented in Tables 3 and 4, are provided in following sections regarding reactor configurations and operational cost as well as the effect of operational parameters.

Table 2 Conventional and non-conventional treatment technologies for the remediation of arsenic-contained water. Type of agents

Advantages

Drawbacks

Case reports

Refs.

Coagulation

– Ferric chloride

– Simplicity of the process

– About 90% of As(V) ions have been removed from aquatic solutions using FeCl3 dosage of >20 mg/L or alum dosage of >40 mg/L.

Sarkar and Paul (2016), Hering et al. (1997) and McNeill and Edwards (1995)

– Ferric sulfate

– High stability of Fe-based coagulants over a wider range of pH having a high affinity for As compared to Al-based coagulants.

– Highly influenced by the type and dosage of coagulant, pH of the solution and availability of other competing anions, especially the iron content of the water sample. – Producing As-contained sludge with high handling cost

– Almost complete removal of 300 µg/L As(III) has been reported after 1 h oxidation in the case of the application of permanganate and sodium hypochlorite.

Sarkar and Paul (2016), Sorlini and Gialdini (2010) and Ungureanu et al. (2015)

– As(III) oxidation efficiency of 90% has been reported within 10 min by the UV/TiO2 process at the initial pH of 9.

Ungureanu et al. (2015), Molinari and Argurio (2017), Lan et al. (2016), Krishna et al. (2001), Lescano et al. (2012) and Rahimi and Ebrahimi (2019)

– Aluminum sulfate Chemical oxidation

– Hydrogen peroxide

– Not suitable and efficient for the removal of trivalent Arsenic – Pre-oxidation may be needed – Easy to be incorporated into existing treatment plant

– Permanganate

– High amounts of chemical agents required.

– Required pre-oxidation of As(III) to As(V) prior to coagulation, membrane filtration, and adsorption process. – Possible formation of hazardous byproducts due to the application of chlorine. – Dependence on solution pH

– Chlorine

– Sodium hypochlorite – Chlorine dioxide

– Chlorine dioxide was not able to effectively remove As

– Ferrate Advanced oxidation processes (AOPs)

– Photocatalysis (UV/TiO2 , UV/ZnO)

– High oxidation rate

– Only effective for the conversion of As (III) to As (V)

– UV/H2 O2

– No generation of toxic or hazardous intermediate byproducts during the conversion of the target pollutant

– Non-selectivity of the treatment process

M. Kobya, R.D.C. Soltani, P.I. Omwene et al. / Environmental Technology & Innovation 17 (2020) 100519

Treatment process

– UV/Ozone – Ozone/H2 O2 – ElectroFenton (continued on next page) 5

Table 2 (continued). 6

Type of agents

Advantages

Drawbacks

Case reports

Refs.

Adsorption

– Granular activated carbon (GAC)

– High efficient for the removal of As (V)

– Only transferring the target pollutant from one phase to another phase

– High adsorption capacity of 136 µg/g was attained when iron oxide-covered sand (IOCS) was used.

Mohan and Pittman (2007), Sigdel et al. (2016), Lunge et al. (2014) and Thirunavukkarasu et al. (2005)

– Activated alumina

– Need for the regeneration of spent adsorbent after four to five working run

– Significant adsorption capacity of 189 mg/g is reported when magnetic nanoparticles obtained from tea waste is used as adsorbent.

– Iron coated sand – Magnetite nanoparticles – Polymeric adsorbents – Biochar – Waste carbonaceous materials

– High cost of regeneration agents

Ion-exchange

Natural (zeolites or montmorillonite (MMT)) and synthetic resins (synthetic polymeric resins)

– High capacity for the sequestration

– High operation and maintenance costs

– Independence towards solution pH – High selectivity to remove arsenic – Possibility to achieve an arsenic concentration of below 2 µg/L in the effluent

– Sludge production because of the regeneration of applied resin – Difficulty to remove As(III)

– Maximum efficiency of 97.9% reported regarding the removal arsenate ions by a new type of exchange fiber in the pH range 3.5–7.0.

is of ion of

Ghosh et al. (2019), Ungureanu et al. (2015), Mohan and Pittman (2007) and Addy et al. (2011)

– Resin bed-life is limited

– Metal-coated polymers are effective than strong basic ion-exchange resins – High affinity of polymeric resins towards carbonate, bicarbonate and sulfate ions compared with As ions Membrane filtration

– Nano-filtration (NF)

– Well established process

– Removing other pollutants together with the target compound

– As(V) rejection by NF or RO was in the range of 85%–99% and As(III) rejection was between 61%–87%.

– Reverse osmosis (RO)

– Useful for household use

– High costs

– The removal efficiency of 90% is reported for the removal of As (V) ions by the RO made of cellulose acetate.

– Electrodialysis *Microfiltration (MF) and ultrafiltration (UF) are not effective to remove As species from the liquid phase

– Less efficiency for removal of As(III) – Fouling problems regarding utilization of membrane

– Production of toxic rejected water

Sarkar and Paul (2016), Ghosh et al. (2019), Ungureanu et al. (2015), Nguyen et al. (2009) and Schmidt et al. (2016)

M. Kobya, R.D.C. Soltani, P.I. Omwene et al. / Environmental Technology & Innovation 17 (2020) 100519

Treatment process

EC reactor and used water type

Initial pH

As species

Co (µg/L)

i, U or j (A, V, A/m2 )

tEC (min, h)

Re (%)

OC ($/m3 )

Ref.

CR for GW BR for GW BR for GW BR (pilot, Vr = 600 L) for GW BR for SW BR for SW BR for SW BR (pilot, Vr = 100 L) for GW BR for SW BR for SW BR for SW BR for SW BR for SW BECR and SW BR for GW BR for SW CR for SW CR for GW BR for GW BR for SW BR for SW CR for SW CR (pilot, Vr = 100 L) for GW

8.3 8.6 7.6 7.1

As(III) As(III) and As(V) As(III) and As(V) As(III) and As(V)

38.15 ± 2.01 36.03–1020.5 285 266 ± 42

1.9 A/m2 0.025–0.10 A 14.3 A m−2 U = 2.1 V

30 min (12 L/h) 2–16 min 6 min 400 C/L∗

96.0 97–99.9 96.9 98.2

0.0154 – 0.1010 0.84–1.04

Kuan et al. (2009) Kobya et al. (2017) Kobya et al. (2015) Amrose et al. (2014)

4.0 7.0 7.4 7.1

As(III) As(V) As(III) As(III) and As(V)

50,000 50–500 112.3 80–760

5.4 A/m2 45 A/m2 5.64 A m−2 0.2–1000 A/m2

45 min 5 min 5 min 85–456 C/L

99.9 95.0 93.9 99.9

– – 0.076 0.220

Can et al. (2014) Ucar et al. (2013) Kobya et al. (2013) Amrose et al. (2013)

7.0 7.0 7.0 6.5 6.5 7.3 8.1 8.0 7.2 7.2 7.2 7.0 4.0 7.0 7.1

As(V) As(V) As(V) As(III) As(III) As(III) As(V) As(III) As(V) As(V) As(V) As(V) As(III) As(V) + As(III) As(III)

100 500 500 150 150 1180 133 899–980 1000 133 131 100,000 15,000 100,000 40

– 20 A/m2 20 A/m2 2.5 A/m2 2.5 A/m2 3 A 45 A m−2 2.4 A/m2 20 V 15–45 A/m2 30 A/m2 50–150 A/m2 75 A/m2 120 A/m2 5 A

16 h 15 min 15 min 2.5 min 12.5 min 2 min 6.5 min 20 min 60 min 10.5–0.75 min 0.30 min 50 min 30 min 120 min 30 L min−1

95.0 98.6 98.6 94.1 93.5 98.6 92.5 99.0 75.0 99.0 92.4 94.0 99.3 98 99.0

0.120 – – 0.0054 0.023 – – – 1.00 – – – – – 0.002

Mólgora et al. (2013) Vasudevan et al. (2010b) Vasudevan et al. (2010b) Kobya et al. (2011a) Kobya et al. (2011b) Majumder and Gupta (2010) García-Lara and Montero-Ocampo (2010) Zhao et al. (2010) Kumar and Goel (2010) García-Lara et al. (2009) Martinez-Villafane et al. (2009) Balasubramanian et al. (2009) Thella et al. (2008) Hansen et al. (2007) Parga et al. (2005)

CR: continuous-flow mode EC reactor, BR: batch-flow mode EC reactor, GW: groundwater, SW: synthetic water sample, Co : initial As concentration, i: applied current, j = current density, q: charge loading (C/L), t EC : operating time, Re : removal efficiency, OC : operating cost.

M. Kobya, R.D.C. Soltani, P.I. Omwene et al. / Environmental Technology & Innovation 17 (2020) 100519

Table 3 Arsenic removal with EC process using Fe anodes.

7

8

EC reactor and water type

Electrode type

Initial pH

As species

Co (µg/L)

i, U or j (A, A/m2 )

tEC (min)

Re (%)

OC ($/m3 )

Ref.

BR BR BR BR BR BR BR BR BR BR BR BR CR CR BR BR BR BR CR BR BR BR BR BR BR

Fe plate Fe plate Fe rod Fe rod Fe rod Fe ball Fe ball Fe ball Fe ball Fe ball Fe scrap Al plate Al plate Al plate Al scrap Al ball Al ball Fe–Al hybrid Fe–Al hybrid Fe–Al hybrid Fe–Al hybrid Fe-Zn hybrid Fe–Al hybrid Fe–Al hybrid Fe–Al hybrid

7.1 6.4 7.0 6.5 – 7.2 7.6 8.5 7.5 7.5 7.42–8.01 7.0 6.8 5.0 7.42–8.01 7.03 7.5 7 .0 7.0 7.0 5.0–8.0 5.0–8.0 7.0 6.0–10.0 4.0

As(III) As(V) As(III) As(III) As(III) As(V) As(III) As(III) As(V) As(III) As(III) As(III) As(III) As(III) As(III) As(V) As(III) As(V) As(V) As(III) As(V) As(V) As(V) As(III) As(III)

2000 130 100,000 50 10,000 150 285 50 200 200 145–146.4 150 134 5.4 ± 16.4 145–146.4 613.4 200 150 150 500 2000–5000 2000 1500–500 10,000 13,400

15.3 A/m2 15 A m−2 220 A 282 A 52 A/m2 0.5 A 0.3 A 0.05 A 0.30 A 0.30 A 0.05 A 2.5 A/m2 60 A/m2 57.8 A/m2 0.10 A 0.29 A 0.15 A 2.5 A/m2 2.5 A/m2 47 A/m2 3 V 3 V 20 A/m2 48 A/m2 300 A/m2

60 min 40 min 15–90 min 0.50 L/min 10 min 1.2 min 20 min 4.94 min 12 min 14 min 8 min 4 min 0–1 L/min 4.3 L/h (90 min) 30 min 10.5 min 3 min 0.05 L/min (1 min) 0.05 L/min (3 min) 2 min 12 min 10 min 30 min 10 min 60 min

99.5 93.0 99.1 80–95 99.8 99.2 99.3 99.0 95.8 96.0 93.0 93.5 92.6 85.0 93.0 99.0 95.1 95.8 96.1 99.9 99.9 98.8 98.4 99.9 99.6

– – – – – 0.031 1.55 0.010 0.546 0.612 0.017 0.0073 – – 0.181 0.442 0.041 0.0023 0.0104 0.0782 – – – – –

Kumar et al. (2004) Maldonado-Reyes et al. (2007) Wan et al. (2011) Lakshmanan et al. (2010) Lakshmipathiraj et al. (2010) Sık et al. (2015) Kobya et al. (2015) Demirbas et al. (2019) Şık et al. (2017) Şık et al. (2017) Omwene et al. (2019) Kobya et al. (2011b) Alcacio et al. (2014) Mohora et al. (2012) Omwene et al. (2019) Demirbas et al. (2019) Kobya et al. (2018) Kobya et al. (2014) Kobya et al. (2014) Song et al. (2016) Ali et al. (2012) Ali et al. (2012) Vasudevan et al. (2010a) Daniel and Prabhakara Rao (2012) Gomes et al. (2007)

for for for for for for for for for for for for for for for for for for for for for for for for for

SW SW SW SW SW, SW GW GW GW GW GW SW GW GW GW GW GW SW SW SW SW SW SW SW SW

plate plate plate plate plate plate plate plate

+ As(V) + As(V) + As(V). + As(V) + As(V)

+ As(V)

+ As(V) + As(V)

+ As(V)

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Table 4 Removal of As by EC process using Fe (rod and scrap), Al (plate, ball and scrap) and Fe–Al hybrid anodes.

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3. Mechanism and theoretical aspects of arsenic removal by EC Application of direct current results in the dissolution of sacrificial anodes (Al or Fe electrodes) and consequently, in-situ production of coagulants in the solution for sequestering target compound (Xu et al., 2018). While metal cations are generated at the anode, the generation of hydrogen gas takes place at the cathode surface along with the simultaneous release of hydroxyl radical (OH− ) (Chen, 2004; Mollah et al., 2001; Binh et al., 2018). The anodic dissolution of Fe or Al electrode leads to the formation of ferric or aluminum ions depending on pH of the solution, which in turn form monomeric species and polymeric hydroxyl metallic complexes, causing coagulation (Chen, 2004; Daniel and Prabhakara Rao, 2012). Furthermore, iron/aluminum oxides such as hydrous ferric oxide (HFO), Al(OH)3 , amorphous Fe(OH)3 , lepidocrocite (γ -FeOOH) goethite (α -FeOOH) are known to strongly adsorb As ions in the bulk solution (Chen, 2004; Daniel and Prabhakara Rao, 2012). Thus, surface complex adsorption and/or compound formation are responsible for the removal of As by iron species. Precipitation at the surface of As–Fe phase can also take place on the surface of aggregated iron hydroxides. Other mechanisms that may take place are electrochemically facilitated coagulation of colloidal As oxyanions or occlusion that is the enmeshment of contaminants in the internal part of growing particles. Oxidation of As(III) to As(V) and subsequently, its surface complexation with iron hydroxides is suggested as one of the main mechanism of As(III) removal by the EC process equipped with Fe electrode (Kumar et al., 2004). The anodic (Eqs. (1)–(5)), cathodic (Eqs. (6) and (7)), co-precipitation (Eqs. (11)–(16)) and adsorption reactions (Eqs. (17) and (18)) are the main mechanisms for the removal of contaminants by the EC process which are presented as follows: (i) Main Anodic reactions with Fe or Al electrodes: First, oxidation of iron to ferrous ion occurs, and depending on the anode potential, its subsequent oxidation to ferric ion may occur as shown in the following equations (Kobya et al., 2011b; Kumar et al., 2004): 4Fe → 4Fe2+ + 8e−

(1)

4Fe → 4Fe3+ + 12e−

(2)

Fe2+ → Fe3+ + e−

(3)

2H2 O → O2(g) + 4H+ + 2e−

(4)

Al → Al3+ + 3e−

(5)

(ii) Main reaction for Fe or Al cathodes: Reactions at the surface of cathode which result in the formation of H2 and release of OH− in the bulk solution are exhibited in the following equations (Kobya et al., 2011b; Hansen et al., 2007). 2H2 O + 2e− → H2(g) + 2OH− (Fe electrode)

(6)

3H2 O + 3e → 3/2H2(g) + 3OH (Al electrode)

(7)





The presence of oxygen (air) in the bulk solution during the process, rapidly oxidize Fe2+ to Fe3+ ion as shown in the below equation: O2(g) + 4Fe2+ + 2H2 O → 4Fe3+ + 4OH−

(8)

Generally, the electrochemically released Fe3+ or Al3+ are slowly hydrolyzed to form Fe(OH)3(s) and Al(OH)3(s) , respectively. The rate of Fe2+ oxidation depends on the availability of dissolved oxygen. As the reaction time of EC process increases, the solution near the surface of cathode becomes alkaline and migration of OH− ion towards the anode occurs, thereby facilitating the formation of ferric and aluminum hydroxides (Chen, 2004; Mollah et al., 2001; Kobya et al., 2011b; Hansen et al., 2007): Fe3+ + 3OH− → Fe(OH)3(s)

(9)

Al3+ + 3OH− → Al(OH)3(s)

(10)

(iii) Co-precipitation or adsorption reactions: Complete reaction of OH− ions discharged from the cathode with the Al3+ or Fe3+ ions generated from the anode is beneficial to keep the solution pH constant during the EC process. Removal of As through co-precipitation (Eqs. (11)–(11)) and adsorption mechanism (Eqs. (17) and (18)) occurs by an alternative reaction in which As ions displace the OH− from their positions in the hydroxide (Chen, 2004; Daniel and Prabhakara Rao, 2012). These reactions are shown in the following equations: + − 2FeOOH(s) + H2 AsO− 4 → (FeO)2 HAsO4 H2 O + OH

(11)

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3FeOOH(s) + HAsO24− → (FeO)3 AsO4(s) + H2 O + 2OH−

(12)

mAl3+ + (3m − n)OH− + nHAsO24− → Alm (OH)(3m−n) (HAsO4 )n(s)

(13)

− ≡ Al − OH(s) + HAsO24− → ≡ Al − OAs(O)2 (OH)− (s) + OH

(14)

⏐ ⏐ ≡ Fe ⏐(OH)2(s) + HAsO24− → ≡ Fe ⏐O2 As(O)(OH)2(s) + 2OH−

(15)

+ ≡ Al − OH(s) + H3 AsO3 → ≡ Al − OH2 −As(O)2 (OH)− (s) + H

(16)

Fe (OH)3(s) + AsO34− → Fe (OH)3 ∗AsO34−

[

Al (OH)3(s) + AsO34− → Al (OH)3 ∗AsO34−

[

] (s)

] (s)

(17) (18)

where ‘‘≡’’ denotes the bonds of the cations with solid surface and the symbol ‘‘|’’ denotes a surface bidentate complex. 4. Effect of operational parameters The efficiency of As removal by the EC process is highly influenced by the following operational parameters: applied current density, type and shape of the electrode, initial concentration of pollutant, initial pH, reaction time, distance between electrodes, mode of electrode connection, and temperature. Electrode distances of between 0.50 and 3.0 cm are generally used in the EC processes (Bouguerra et al., 2015). Accordingly, a distance of 1 cm has been generally used in many studies (Kobya et al., 2011a, 2014; Omwene et al., 2018; Omwene and Kobya, 2018). As the temperature increases, the conductivity of water increases leading to an increase in removal efficiency. However, the temperature of the bulk solution in the EC system is normally dictated by the ambient conditions and ranges from 10 to 30 ◦ C. The effects of current density and charge loading are discussed in detail in the following sections. 4.1. Effect of electrode material and shape The shape and material of the applied electrode are very significant parameters, influencing the performance of the EC process (Mohora et al., 2018). Aluminum, iron and zinc are some of the metal electrodes that are commonly utilized for the removal of As ions from waters by the EC process (Chen, 2004; Kobya et al., 2016). Plate, mesh, rod and ball-shaped metal anodes of iron or aluminum have been widely used for the removal of As, with the removal efficiency of as high as 75–100% recorded by ball-shaped electrodes at different operational conditions (Sık et al., 2015). The cheaper cost and less toxicity of the residuals make Fe anodes more favorable in comparison with other electrodes for the EC process. In addition, Fe anodes can be used over a wide pH range due to the formation of Fe-hydroxyl species (Kobya et al., 2011a,b; Sık et al., 2015; Demirbas et al., 2019). The removal of As(III) and As(V) by the EC process equipped with Fe and Al plate electrodes is compared and exhibited in Tables 3 and 4. The operational costs of As removal by the EC process were 0.0054 and 1.04 $/m3 for synthetic and real wastewater samples, respectively. The removal efficiencies of As were obtained to be 75.0 and 99.9% regarding the treatment of synthetic and real wastewater samples by the EC process, respectively. Groundwater (from Guanajuato, Mexico) with As concentration of 134 µg/L was treated by uninterrupted filter press EC reactor using Al plate electrodes. At initial pH of 6.8, current density of 60 A/m2 , and flow rate of 0.91 1/cm, the As removal efficiency of 89.6% was obtained with calculated energy consumption of 0.89 kWh/m3 (Alcacio et al., 2014). At initial As concentration of 150 µg/L and applied current density of 2.5 A m−2 , the removal efficiency of 95.7% was reported in a batch flow mode EC reactor equipped with Al plate electrodes at initial pH of 7.0 and reaction time of 15 min, whereas the removal efficiency of 93.5% was reported when Fe plate electrode was used at initial pH of 6.5 and reaction time of 12 min (Kobya et al., 2011a). The operational cost of EC process equipped with Al plate anode was found to be cheaper than Fe plate anode by 0.0004 $/m3 . In the case of Fe plate electrodes, more than 97.0–99.9% removal efficiency of As was achieved at reaction time of 2–16 min and applied current of 0.025–0.10 A when As-contaminated spring and groundwater samples with initial concentration ranging from 36 to 1021 µg/L were treated by the EC process (Kobya et al., 2017). Recently, the application of composite electrodes (hybrid electrodes) composed of different materials has been considered to achieve higher removal efficiency of As from aqueous environments compared to pure electrodes (Table 4). Better floc formation was observed with Fe–Al hybrid electrodes over a wide pH range (Kobya et al., 2014). As-generated metal hydroxides and (oxy) hydroxides provide larger surface area for the adsorption of As, making the application of Al or Fe electrodes efficient as their hybrid form (Gomes et al., 2007; Kobya et al., 2014; Song et al., 2016; Vasudevan et al., 2010a; Daniel and Prabhakara Rao, 2012). The removal efficiency of 99% has been attained for the removal of As concentration in detail in the following sections ranging from 2000 to 5000 µg/L by the EC process as Al–Fe hybrid electrodes are implemented at initial pH range of 5–8, reaction time of 12 min and applied voltage of 3 V (Ali et al., 2012). Ali et al. in their studies,

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Fig. 1. Batch (a) and continuous (b) flow mode EC reactors for the removal of As using Al–Fe hybrid anodes.

observed and reported the removal of As by the adsorption or co-precipitation with hydrous aluminum oxide during the EC process (Ali et al., 2012). The As removal efficiencies were found to be 99.6, 97.5 and 99.6% when Fe–Al, Al–Al and Fe–Fe electrodes were employed in a batch flow mode reactor at current density of 300 A/m2 , reaction time of 60 min, initial pH of 4 and initial As concentration of 1230 µg/L (Gomes et al., 2007). The treatment of As-contaminated drinking water using both continuous and batch flow mode EC reactors was investigated by Kobya et al. (2014), using Al–Al, Al–Fe, and Fe–Fe hybrid electrodes (Fig. 1). Mohora and coworkers studied the application of a horizontal continuous-flow EC reactor for the removal of As from groundwater using iron electrodes. The optimum operational parameters were found to be charge loading of 54 C/L, flow rate of 12 L/h and current density of 1.98 A/m2 to remove 96.0% of As from the contaminated water resource. After the reaction time of 4 h, the obtained effluent met World Health Organization (WHO) guideline value (Mohora et al., 2018). In the case of batch flow mode EC system operated by monopolar electrode connection mode at initial solution pH of 7, current density of 2.50 A/m2 , and initial As concentration of 150 µg/L, the removal efficiency of 96.0% was obtained when Fe–Al hybrid plate pair was applied at reaction time of 1 min with 0.0023 $/m3 as calculated operational cost. For the continuous flow mode reactor with the same electrode combination, the required reaction time of EC process increased from 3.8 to 20 min as the flow rate increased from 0.050 to 0.20 L/min. Furthermore, the operational cost and the amount of sludge generated during the EC process varied from 0.0103 to 0.0684 $/m3 and 0.0095 to 0.025 kg/m3 as the flow rate increased from 0.05 – 0.20 L/min. In a recent study by Kobya and coworkers, a new compact set-up of EC process was designed with high practicability and adaptability, providing higher surface areas of Fe and Al ball anodes to achieve higher removal efficiency (Kobya et al., 2015; Sık et al., 2015; Demirbas et al., 2019; Şık et al., 2017; Kobya et al., 2018; Gören et al., 2018). The new fixed-bed air flow EC reactor with Fe ball anodes attained As(III) removal efficiency of 96.0% under optimum operational parameters, including initial As concentration of 200 µg/L, initial pH of 7.50, air flow rate of 6 L/min, Fe balls diameter of 7.50 mm, and anode height of 7.5 cm (Fig. 2). At applied current of 0.30 A, reaction time of 14 min, electrode consumption of 0.0752 kg Fe/m3 , energy consumption of 1.442 kWh/m3 , operational cost of 0.612 $/m3 , 2.55 µg removed As per mg Fe was recorded. Similarly, As(V) removal efficiency of 95.8% was reported considering electrode consumption, energy consumption, operational cost and removed As per mg Fe of 0.0628 kg Fe/m3 , 1.386 kWh/m3 , 0.546 $/m3 , and 3.05 µg, respectively (Şık et al., 2017). The arsenic removal efficiency of 99.3% was obtained for Fe ball anodes at initial pH of 7.6, air-flow rate of 6 L/min, applied current of 0.30 A, and anode surface area of 210 cm2 . At reaction time of 20 min, operational cost and charge density were 1.55 $/m3 and 360 C/L, respectively. Meanwhile, the Fe plate electrodes indicated As removal efficiency of 96.9% at reaction time of 6 min, operational cost of 0.101 $/m3 and charge density of 108 C/L (Table 4). Remarkably, Fe plate electrodes resulted in the reduced reaction time of EC process from 20 to 6 min, leading to the significant reduction in the operational cost. Also, as can be observed in Fig. 3, the reaction time needed to meet MCL decreased when applied current increased. Under optimum operational conditions, As concentration in the bulk solution decreased to < 10 µg/L when both ball and plate-shaped Fe anodes were implemented (Kobya et al., 2015). For low As(III) concentration of 50 µg/L, the removal efficiency of 99.0% was obtained using Fe ball anodes (diameter: 9.24 mm) at optimum operational conditions (applied current of 0.050 A, initial pH of 8.50, air flow rate of 9.98 L/min,

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Fig. 2. Appearances of a fixed-bed air flow EC reactor. (Reprinted from Demirbas et al. (2019) with permission from Elsevier).

operational cost of 0.010 $/m3 , and reaction time of 4.94 min (Demirbas et al., 2019). Under similar air flow rate, the removal efficiency of >99.0% was observed when Al ball anodes were used instead of Fe ball anodes (ball diameter: 7.50 mm, applied current: 0.29 A, initial pH: 7.03, air flow rate: 6.40 L/min, operational cost: 0.442 $/m3 , reaction time: 10.50 min (Gören et al., 2018). Utilization of waste Al and Fe scraps as sacrificial anodes for the treatment of As-contaminated groundwater is considered and the results are depicted on Fig. 4 (Omwene et al., 2019). The use of waste metal scraps as sacrificial anodes is beneficial both in terms of environmental aspects and cost-effectiveness point of view. For As removal efficiency of >93.0%, the applied current of 0.050 A and reaction time of 8 min is needed when Fe scraps were used, while in the case of Al scrap anodes, the applied current of 0.10 A and reaction time of 30 min were applied to achieve the same removal efficiency (Table 4). The energy consumptions, electrode consumptions and operational cost for Fe scraps were 0.070 kWh/m3 , 0.052 kg/m3 and 0.017 US $/m3 , respectively; while in the case of Al scraps, the aforementioned costs were calculated to be 0.876 kWh/m3 , 0.067 kg/m3 , and 0.181 US $/m3 , respectively. Consequently, Fe scraps showed superior As removal efficiency than that of Al scraps under all experimental conditions tested.

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Fig. 3. Effect of applied current on the removal of (a) As(III) and (b) As(V) by the EC process (Reprinted from Kobya et al. (2015) with permission from Elsevier).

4.2. Effect of current density and charge loading During an EC process, the production rate of coagulant is mainly controlled by the reaction time and applied current density based on Faraday’s law (Eq. (19)) (Amrose et al., 2014, 2013; Kumar et al., 2004). The amount of coagulant generated during the EC process is directly proportional to the current density and reaction time, which in turn affects the growth of flocs as represented in the following equation (Amrose et al., 2014): Celectrode =

i × tEC × Mw z×F ×v

=

q × Mw z×F

(19)

where i denotes current intensity (A), tEC is the reaction time (min), q denotes charge loading (C/L or F/m3 water), Mw is the atomic weight (Mw,Al = 26.98 g/mol, Mw,Fe = 55.85 g/mol), z refers to the number of electrons (zFe = 2 and zAl = 3), F denotes Faraday’s constant (96 487 Coulomb), v denotes volume of solution (L or m3 ) in the reactor. The amount of coagulant required to attain residual As the concentration of <10 µg/L depends on the charge density, which is directly influenced by anode consumption during the EC process (Faraday’s law). Charge loading in the EC process controls the rate of reaction in the EC reactor and is an important parameter to design EC systems (Chen, 2004; Amrose et al., 2013; Kumar et al., 2004). The charge loading for an EC system can be estimated using Eq. (20), representing the charges transferred in electrochemical reactions per volume of treated water as shown in the following equation: q(C/L) =

i × tEC

v

(20)

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Fig. 4. EC reactor using a fixed-bed air flow EC reactor using scrap anodes.

Excessive charge results in the generation of extra iron hydroxide flocs in the solution, making the concentration of residual ions higher. The excess iron hydroxide flocs may result in the operation shortfalls as they are hard to float and isolate due to their poor affinity to the air bubbles. Furthermore, high charge loading leads to the elevated operational cost owing to the high-energy consumption (Kobya et al., 2016). The effect of applied current (0.010–0.10 A) on the treatment of As-contaminated groundwater using Fe and Al scrap anodes is shown in Fig. 5. A significant decrease in the concentration of As was observed with increasing the reaction time under all applied current intensities. Substantial removal of As observed at the beginning of the EC process, however, in the following, the removal of As became gradual with increased reaction time. When Fe electrode was used, residual As concentration decreased to < 10 µg/L after a reaction time of 6–20 min. However, in the case of Al electrodes, higher reaction time (30–60 min) was required to achieve the same efficiency. The arsenic removal efficiency of 95.1% was obtained by the Fe electrode-equipped EC process

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Fig. 5. As removal at different applied currents with Fe and Al scrap anodes (Reprinted from Omwene et al. (2019) with permission from Elsevier).

within the reaction time of 6 min, while Al electrode-equipped EC process required a much longer time (30 min) at the same conditions. Similar results have been obtained by Rosales and coworkers in the case of the treatment of arseniccontaminated groundwater using a continuous electrocoagulation process equipped with a twelve-cell stack (Rosales et al., 2018). Based on response surface methodology, it is specified that current intensity of the EC process is more significant than other main operational parameters such as solution pH in terms of arsenic decontamination (Amrose et al., 2013). The results display a direct relation between applied current intensity and generation rate of coagulant, a result of the dissolution of anode electrode which creates hydroxide anions in the solution (Parga et al., 2005). Actually, current intensity is proposed as the most important operational parameter to control the coagulant generation rate and reaction rate of the process (Orescanin et al., 2013). Electrostatic attraction and surface complexation are two major mechanisms responsible for the removal of target ionic pollutants by the as-generated hydroxide anions. During surface complexation, the target ionic pollutant acts as a ligand to attach a hydrous iron moiety through adsorption and precipitation mechanisms (Orescanin et al., 2013). However, excessive current density leads to the cathode passivation due to the intensive formation of hydroxide anions. As-generated hydroxide anions create a thick sludge with high turbidity in the bulk solution, leading to the increased applied voltage and consequently, higher energy consumption (Amrose et al., 2013).

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4.3. Effect of initial pH The solution pH is another key parameter that controls the removal of As ions by the EC process. During the EC process, the solution pH increases by increasing the reaction time due to the production of OH− ions at the surface of cathode according to Eqs. (6)–(7). Moreover, during the adsorption or co-precipitation of As(V) and As(III) ions in the EC reactor, OH− positions in Fe-hydroxide or Al-hydroxide are substituted by As(V) and As(III). At pH range of 4.0–10.0, As(V) is the dominant species, yielding an overall negative charge. In contrast, As(III) species have no net charge. Consequently, As(V) is expected to have a higher removal efficiency than As(III) owing to adsorption of H2 AsO4 − , HAsO4 2− or AsO3 4− by Fe(OH)3 precipitates. The lepidocrocite (γ -FeOOH) produced during EC is neutral (pH 7.0). However, at pH values higher than 7, both arsenate and the lepidocrocite surface create a negative charge which may result in the lower adsorption of arsenate (Omwene et al., 2019). The lepidocrocite surface is positively charged at pH values lower than 7.0, resulting in the enhanced electrostatic adsorption of arsenate species. Nevertheless, pH values between 7.0 and 9.0 have been reported as the most suitable range for the removal of arsenate from most natural water resources. Oxidation of As(III) to As(V) is recommended to boost the removal rate of As by the EC process via coagulation. As(III) removal by Fe electrodes-equipped EC process is more efficient than that of Al electrode because of the stronger attachment of As(III) to electro-generated Fehydroxide surface. A study by Daniel and Prabhakara Rao (2012) indicated a significant increase in the removal efficiency of As with increasing pH values from 6.0 to 10.0 using Fe electrodes. This may be attributed to the fact that polymeric metal hydroxides are more easily formed at higher pH values, accelerating the coagulation of As ions in the bulk solution. Remarkably, an increase in the removal efficiency of As from 83.0 to 99.6% was observed when Fe anode and Zn cathode was used at initial pHs in the range of 2.0–7.0, applied voltage of 3 V, As concentrations of 2000 µg/L and reaction time of 10 min (Ali et al., 2012). 5. Sludge production during EC process, characterization and disposal Arsenic-laden sludge is produced during the removal of arsenic through EC process that should be safely handled and disposed. For example, a pilot EC reactor used for arsenic removal produced 10–20 g amount of sludge in the batch flow mode reactor of Electrochemical Arsenic Remediation (ECAR) with working volume of 100 L designed for decreasing initial arsenic concentration of 600–3000 µg/L to 10 µg/L in the Bangladesh groundwater (Addy et al., 2011). Arsenic-laden flocs in ECAR process are segregated from the bulk solution via gravitational sedimentation enhanced by the alum addition as coagulant (about 6–15 mg/L). Afterwards, leachate of the produced sludge were analyzed for Hg, P, Ag, As, Cd, Cr, Se and Ba metals. Accordingly, concentration of all metals were below maximum concentration limits. Therefore, waste sludge of the ECAR process was not categorized as hazardous solid waste. As-generated waste sludge by the ECAR process was insignificant and was not hazardous based on U.S.EPA regulations. Based on Toxicity Characteristic Leaching Procedure (TCLP) utilizing U.S.EPA method no. 1311, there was not detectable arsenic (1 µg/L) in the crushed concrete. In addition, arsenic-laden waste sludge was successfully stabilized in concrete. This approach may be proposed as an alternative to landfilling of the wastes. In another study, the ECAR reactor with working volume of 600 L in 3.5 monthly field trials produced about 245 mg waste sludge per liter of the reactor at optimum operational conditions. This ECAR reactor with coulombic dosage of 450 C/L decreased arsenic concentration of real groundwater from the West Bengal from 266 µg/L to lower than 5 µg/L and arsenic loading in the sludge ∼1.5% by weight (Amrose et al., 2014). Landfilling costs of the as-generated sludge was added 5% to the total cost of the production of arsenic-free water. As an alternative to landfilling, the concrete stabilization of the sludge was suggested. The arsenic-laden waste sludge was incorporated into the mixture of a Portland-cement concrete which resulted in replacing 6% concrete by weight and 40% by volume. As-generated concrete was not hazardous based on U.S.EPA TCLP standard. Arsenic-laden sludge stabilized by the concrete could effectively be packed into roadways with minimum detrimental effect to the ecosystem. During ECAR process, corrosion of iron as sacrificial anode generates ferric oxide (Fe(III) precipitates) in arsenic-polluted water stream (Gadgil et al., 2010). There are many complicated processes occurring at the same time during ECAR process such as Fe(II) oxidation, surface adsorption, electrochemical dissolution of the electrode, coagulation and hydrolysis of ferric oxide. On the other hand, co-occurring ions influence the removal of arsenic due to the competition with arsenic for occupying adsorptive sites placed on ferric oxide sludge, or by influencing the structure of ferric oxide sludge generated during oxidation of Fe(II) ions (Mollah et al., 2004). In addition, the oxidation of As(III) to As(V) occurs in the ECAR reactor by the electrolytic action at the surface of the electrode and highly reactive radical species generated through the oxidation of Fe(II) ions (Kumar et al., 2004). Arsenic creates some complexes with ferric oxide sludge, thereby forming a floc that can be segregated from water stream. Synthetic (total arsenic of 100–3000 µg/L) and real groundwater’s (total arsenic of 93–510 µg/L from Bangladesh and 80–760 µg/L from Cambodia) were studied to arsenic removal by the ECAR process (Gadgil et al., 2010). Extended X-ray absorption fine structure analysis of the sludge produced during the ECAR process was determined to be the structure of the as-generated iron precipitate with arsenic complexation. Shell-by-shell fits Fe K-Edge XAS spectra of arsenic-contained precipitates showed only edge linkages and did not support corner linkages caused by AsO4 3− , SiO3 2− and PO4 3− ions (Gadgil et al., 2010; van Genuchten et al., 2012). This compels ferric oxide sludge generated in the ECAR reactor to stabilize colloidal particles which need the long sedimentation time when no coagulants are added to the bulk solution. According to K-Edge spectra, arsenic is bound to ferric oxide sludge by strong binuclear complexes. Varying current density of the ECAR reactor in the range of 0.20–50 A/m2 caused no

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effect on the amount of as-generated ferric oxide sludge or arsenic complexes. Arienzo et al. (2002) characterized the sludge produced during EC using XRD analysis to find poorly ordered Fe mineral phases with a dominance of 5-line ferrihydrite, including no lines of ferric arsenate phases (current density of 4 A/m2 , process time of 3 min, and initial pH of 6.5) (Arienzo et al., 2002). The amount of ferric oxide sludge produced at above conditions was 162 mg/L and the percentage of arsenic recovery on the sludge was >96% (ferric oxide sludge-contained arsenic of 2613 µg/L). The results of FT-IR indicated the adsorption of both As(III) and As(V) ions onto ferric oxide sludge at neutral and acidic pH conditions. TEM observations exhibited the formation of spherical shape ferrihydrite, providing evidence for the generation of subrounded hematite and goethite with acicular shape. The removal of arsenic by the stainless steel electrodes-equipped electrocoagulation is optimized via response surface methodology (Gilhotra et al., 2018). This statistical approach has been widely implemented to optimize research projects with lower costs. At optimized conditions (current density = 20 A/m2 , initial pH = 5.2, reaction time = 20 min and electrode distance = 1.5 cm), the mass spectrometric analysis exhibited the generation of a complex between iron-contained sludge and arsenite ions during the EC process. SEM-EDX images of the settled sludge showed amorphous structures on the surface of flocs of iron oxide/hydroxide and confirmed the removal of arsenic from the water sample. The results of XRD analysis indicated the production of iron arsenate species (FeAsO4 and Fe3 AsO7 ) in the sludge. Gomes et al. (2007) showed that fine magnetite particles and amorphous iron oxyhydroxides produced during EC process efficiently remove both As(V) and As(III) from groundwater in the Mexico (removal efficiency of more than 99.0%) under optimum operational conditions (initial pH of 5.5–7.1, current density of 37–46 A/m2 , arsenic concentration of 25–50 µg/L and reaction time of 1 min). The results of XRD, SEM, and TEM indicated the nonstoichiometric production of goethite, fine magnetite particles and iron oxyhydroxides during the EC process with air injection using carbon and steel electrodes. Arsenic removal efficiencies of 99.6% by Fe–Fe electrode pair, 97.8% by Al–Al electrode pair and 78.9->99.6% by Al–Fe electrode pair (operational conditions: initial arsenic concentrations of 1.42–1230 mg/L current density of 300 A/m2 , initial pH of 2.4–6.0 and electrocoagulation time of 60–120 min) (Gomes et al., 2007). The analysis of the produced sludge confirmed the production of amorphous/poorly crystalline phases for mansfieldite: AlAsO4 ·2 (H2 O) and aluminum hydroxide/oxyhydroxides (magnetite: Fe3 O4 , iron oxyhydroxides: FeO(OH), diaspore: AlO(OH), and bayerite: Al(OH)3 ), and iron oxide: FeO). In another study, Kobya et al. (2014) obtained that Fe– Al–Al–Fe hybrid electrode pairs were the best electrode configuration for the removal of arsenic (95.8%) at reaction time of 1 min (initial arsenic concentration of 150 µg/L, current density of 2.5 A/m2 ). The total operational cost (including landfill disposal of arsenic-laden sludge (0.0034 kg/m3 ), energy (0.00358 kWh/m3 ) and electrode (0.00112 kg/m3 Fe and 0.00036 kg Al/m3 ) consumptions was calculated as 0.00202 e/m3 . The XRD pattern of the produced sludge displayed the presence of arsenolite (As2 O3 ), kankite (Fe3+ (AsO4 )3.5 (H2 O)), bayerite (Al(OH)3 ), parasymplesite (Fe2+ 3(AsO4 )2 . 8(H2 O)), and schneiderhohnite (Fe2+ Fe3+ 3As5 O13 ) in the sludge sample. According to XRD pattern, the amorphous nature of asgenerated sludge indicated rapid precipitation of electrogenerated metal hydroxide during the treatment of arsenic by the Fe and Al hybrid electrodes-equipped EC reactor. 6. EC reactor configurations and operating cost Physical factors such as mode of operation (batch or continuous flow modes), electrode connection modes, the geometry of reactor, distance between electrodes and surface area to volume ratio (S/V) of the electrodes are decisive parameters in the design of EC reactors. The results of As removal by batch and continuous flow mode EC reactors are given in Tables 2 and 3. Continuous flow mode EC reactors, like turbulent flow reactor, airlift reactor and modified flow continuous reactor have been proposed as more economical than batch flow reactor configurations (Song et al., 2017). For Fe electrode-equipped EC reactor, the airlift and modified flow EC reactors showed significant As(V) removal of >98.0% at current density of 1200 A/m2 within the reaction time of 9.40 min, far beyond the removal efficiency of 40.0% which was obtained when turbulent flow reactor was used (Hansen et al., 2007). Complete removal of As from groundwater was reported by Flores et al. when a pre-pilot-scale continuous filter press EC reactor using Al plate anodes was implemented at mean linear flow rate of 0.91 cm/s, initial pH of 7.50 and As concentration of 50 µg/L (Flores et al., 2013). A pilot study with cylindrical batch flow mode EC reactors (working volumes of 100 and 600 L) was utilized for the treatment of real As-contaminated groundwater in Cambodia, Bangladesh and West Bengal containing As concentration range of 80–760 µg/L. The results exhibited the removal efficiency of 98.70% at the current density of 0.2–50 A/m2 . Effluent from the EC reactor was channeled to a sedimentation tank, where the sedimentation of the electro-generated iron oxide precipitate was enhanced through the addition of alum (as 5.0–15.0 mg Al/L). The overall EC costs for 100 and 600 L reactors were calculated to be 0.22 and 0.83–1.04 US $/m3 , respectively (Amrose et al., 2014, 2013). In recent years, airfeed cylindrical batch EC reactors using Fe and Al ball anodes have verified to be highly effective in the removal of As from aquatic environments. Accordingly, As removal efficiencies between 95.1 and 99.2% were reported in such studies. As(III) removal efficiency of 96.1% was reported by the continuous flow mode EC reactor using hybrid Fe–Al–Al–Fe electrode at initial As the concentration of 150 µg/L, current density of 2.5 A/m2 , reaction time of 3 min, initial pH of 7.0 and flow rate of 0.050 L/min (Kobya et al., 2014). In another study, an As removal efficiency of >99.0% by a continuous EC reactor using Fe electrode was reported at flow rate of 30 L/min, initial As concentration of 25–50 µg/L, initial pH of 5.5–7.1, current density of 5 A/m2 and reaction time of 1 min (Parga et al., 2005). A batch flow mode EC reactor using Fe plate electrode has been reported (Wan et al., 2011). According to this study, As removal efficiency of 99.2–99.9% was obtained for the removal of As concentration range of 25–50 µg/L at initial pH of 6.8–7.1. At present, no pilot-scale

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Fig. 6. Various configurations of electrode connections in EC reactors.

study has been performed in continuous flow mode for the removal of As from the aquatic phase by the EC process. Regarding reactor geometry, no ideal reactor shape for the removal of As has been developed until now. Based on the literature, the higher retention time of up-flow long vertical-plate reactors generates more metal hydroxides, leading to the higher As removal efficiency by the EC process. The geometry of the EC reactor should not only be cost-effective and easily to operate, but should also boost the reactions in favor of the performance of EC process. Three electrode connection modes are possible in an EC reactor: bipolar-serial (BP-S), monopolar parallel (MP-P) and monopolar serial (MP-S). Various configurations of electrode connections are displayed in Fig. 6. The BP-S mode of connection is simple with lower maintenance requirements; however, it is flawed by the current loss in the EC process. Fe electrodes in MP-S mode improve chemical co-precipitation and decrease anodic potential, making it the most efficient connection with reference to the energy consumption, reactor efficiency and operational cost (Kobya et al., 2011b). Irrespective of the simple operation of the basic EC process, EC needs electrodes with high surface area to improve the dissociation rate of metals. Therefore, the utilization of monopolar or bipolar arrangements in parallel or series connections is proposed to overcome this limitation considering full-scale application of the treatment process. A smaller interelectrode spacing (0.50–3.00 cm) is favorable for the reduction of energy consumption and reaction time, ensuring adequate interelectrode mass transfer. The electrode surface area to volume of EC reactor (S/V) ratio is considered a significant factor for EC system scale-up. Moreover, the applied current density and subsequent bubble formation and distribution in the EC reactor are affected by the surface area of the applied electrode. At a constant S/V ratio, increasing charge loading results in the reduction of reaction time. Increasing charge loading can be done by either operating at a higher voltage or at higher current, which in turn increases the energy consumption. Kobya et al. (2016) observed an obvious increase in electrochemically dissolved iron and charge loading with increasing EC reaction time for high initial As concentration. In a batch flow mode Fe plate electrodes-equipped EC reactor in MP-S mode operated for the removal of As concentration of 500 µg/L, the As removal capacity improved from 6.17 µg/C (or 21.33 µg/mg Fe) to 16.38 µg C−1 (56.6 µg/mg Fe) along with decreasing S/V from 29.20 to 10 m2 /m3 (Fig. 7). The amount of As removed per quantity of electrochemically released aluminum or iron (qe , µg As per g Al or Fe) was estimated through the following equation: qe =

(Co − Ct ) × v Celectrode

(21)

where Co is initial concentration of As ions (µg/L), Ct is the concentration at specified reaction time (t), and v is the working volume of the EC reactor (L) Like any other treatment technology, to effectuate a feasible EC process, the operating cost is an important aspect to be considered. Electrode and electricity consumption during the removal of As are the principal costs ascribed to the EC process. The energy consumption encompasses power supply in EC reactor and pumping costs in the process. The cost of sludge handling generated in EC process such as Landfill deposition fees should also be added to the total operational costs. The operational cost (OC, US $/m3 ) can be calculated via the following equation (Omwene et al., 2019): OC = α × Cenergy + β × Celectrode

(22)

where α denotes cost per unit of electrical energy (US $/kWh), and β refers to the unit electrode cost (US $/kg). Reaction time and applied current highly influence the electrode and energy consumptions of an EC system. To attain the required removal efficiency, the operation of the EC process with longer reaction time at low applied current or a shorter reaction time at high applied current is required. Therefore, these two main parameters should be optimized in order to achieve the required treatment efficiency as well as lower electrode and energy consumption. The performance of an EC process, along with the energy consumption, is highly affected by the conductivity of the electrolyte and also electrode connection

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Fig. 7. (a) Effect of charge loading on As removal with regards to the reaction time of EC process at different S/V, (b) As removal capacity (RC) as a function of charge loading at different S/V (Reprinted from Kobya et al. (2016) with permission from Elsevier).

mode. In a comparative study on the removal of As by the EC process equipped with different anode types, it was reported that the operational cost for Fe plate electrodes were 0.0054–1.04 US $/m3 , whereas the operational cost for Fe ball anodes ranged from 0.010–1.55 US $/m3 . The respective operational costs for Fe and Al scrap anodes were, 0.017 and 0.181 US $/m3 , respectively. The calculated operational costs for both plate and ball-shaped Al anodes were 0.0073 and 0.442 US $/m3 , respectively. Finally, operational costs for Fe–Al hybrid plate electrodes was in the range of 0.0023–0.0782 US $/m3 (Tables 3 and 4) (Kobya et al., 2014). For the removal of As with initial concentration of 100 mg/L by the EC-microfiltration process (EC-MF), the estimated operational cost was 0.12 $/m3 under optimum operational conditions including, initial pH of 7.0, distance between electrodes of 0.4 cm, electrochemically released Fe3+ precipitate of 4 mg/L, and long reaction time of 16 h (Mólgora et al., 2013). Conversely, the operational costs of the chemical coagulation together with MF (CC-MF) was estimated to be 0.066 $/m3 under optimum conditions (Fe3+ dosage of 4 mg/L, initial pH of 7.0 and reaction time of 26 h), which resulted in the As removal efficiency of 97.0%. Notably, the EC-MF process designed for As removal was 1.8 times more expensive than that of CC-MF process (Mólgora et al., 2013). Using alternative methods, the operational costs to achieve arsenic concentration below the WHO guideline (I.W.G.o.t.E.o.C.R.t. Humans et al., 2004) were 3.40 US $/m3 for ion exchange, 3.20 $/m3 for activated alumina, 1.21 US $/m3 for coagulation–filtration, 3.72 US $/m3 for reverse osmosis, 0.054 US $/m3 for air oxidation–filtration, and 1.20 US $/m3 for granulated ferric hydroxide/oxide (Ahmed, 2004). Although comparison of the operational costs of different treatment processes may be unreasonable owing to the different conditions, nevertheless, the EC process has lower operational costs in comparison with abovementioned conventional treatment processes, making it a more suitable alternative for the cost-effective removal of As from aquatic environments such as groundwater. In recent years, an alternative technique named metal-air fuel cell EC (MAFCEC) has been proposed to address various drawbacks associated with the application of conventional EC process especially its high energy consumption. In this approach, microbial fuel cell (MFC) is combined with an EC process to remove target inorganic and organic pollutants (Kim et al., 2018, 2015; Tian et al., 2016). MFCs are emerging methods to generate electricity along with the treatment of contaminated solutions (Kim et al., 2015; Maitlo et al., 2019; Xue et al., 2013; Heijne et al., 2010).

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Fig. 8. The removal mechanism of As by metal-air fuel cell electrocoagulation (Reprinted from Maitlo et al. (2019) with permission from Elsevier).

The reduction/oxidation reactions occur at the surface of cathode electrode to generate hydrogen gas as presented in the following equations (Maitlo et al., 2019): + − M(0s) → M(naq ) + ne

(23)

2H2 O → 4H(+aq) + O2(g ) + 4e−

(24)

nH2 O + ne− →

n 2

H2(g) + nOH − (aq)

(25)

The MAFCEC is cost-effectively used to treat As-contaminated water resources. The operation of the MAFCEC process results in the treatment of As-polluted liquid streams as well as the generation of electrical energy (Maitlo et al., 2019, 2018). A schematic flow diagram of the MAFCEC process is presented in Fig. 8. The integration of EC process with an air–fuel cell is considered as a promising method for the removal of As without the need for an electricity supply. Therefore, this innovative technology is beneficial for the treatment of polluted water and wastewater in remote areas where supplying electricity power is not easy (Kim et al., 2017). In this regard, Maitlo et al. used a fuel cell equipped with an air-permeable carbon cathode and an iron anode to treat As-contaminated solution. According to their results, the As concentration decreased from 1 mg/L to less than 5 µg/L within 4 h at the initial pH of 5.0 (Maitlo et al., 2017). 7. Simultaneous removal of As and other contaminants The simultaneous presence of various inorganic contaminants is inevitable in different water resources. Thus, evaluation of the effect of co-existing compounds on the efficiency of the treatment process in the removal of target pollutants is essential from a practical point of view. Moreover, the removal of other pollutants from water is essential together with target pollutants assuming beneficial use of effluent is considered after the treatment process. For this purpose, many researchers have focused on the simultaneous removal of As and other organic and inorganic pollutants by the EC process (Song et al., 2018; Mroczek et al., 2019a; Thakur et al., 2019; Zheng et al., 2019). In this regard, Thakur and Mondal evaluated the efficiency of the EC reactor to individually and simultaneously remove As and fluoride from the aquatic phase. Their findings exhibited the suitability of the EC process for individual and simultaneous removal of both pollutants. Based on their results, the EC reactor was more efficient for As removal than it was for fluoride (Thakur and Mondal, 2017). Mroczek et al. investigated the simultaneous removal of As and silica from the liquid stream using EC process. Their results showed the effective removal of both As(III) and silica by the Fe electrodes-equipped EC process under acidic conditions (Mroczek et al., 2019b). In the case of simultaneous removal of As and fluoride, current density, initial pH of the bulk solution, and reaction time have been reported as critical parameters for the simultaneous removal

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Fig. 9. The ECAR reactors with 100 L volume. Source: Reprinted from Addy et al. (2011).

by the EC. As a result, acidic conditions were more suitable than basic conditions for the removal of fluoride, while the effective removal of As was achieved without pH adjustment (López-Guzmán et al., 2019b). It is reported that arsenate, along with other ionic compounds such as sulfate and phosphate, is adsorbed on aluminum or iron flocs, while fluoride ions may replace hydroxyl groups from electro-generated flocs (Rosales et al., 2018; Guzmán et al., 2016). 8. Pilot and large scale EC plants for arsenic removal Among all the efforts in developing EC process for the removal of arsenic from water, there are some studies on the advancement of pilot and full-scale EC reactors or processes for reclamation of naturally contaminated water samples. A brief summary of these studies in the literature is mentioned here. As an example, a 100-L batch flow mode ECAR prototype reactor was designed for the removal of arsenic from contaminated groundwater by Addy et al. (2011). The ECAR was designed to meet the needs of a suitable scheme that is locally affordable, financially viable, and offers long-term water access with sustainability. The ECAR was a cylindrical electrochemical reactor connected to a sedimentation tank for the addition of alum and separation of solid content of the reactor (Fig. 9). The ECAR was able to reduce arsenic levels below the MCL of WHO (10 µg/L). In short-duration field trials, real contaminated groundwater samples in Bangladesh (initial arsenic 93–510 µg/L) and Cambodia (initial arsenic 80–760 µg/L) were easily reached to below the MCL of WHO. Electrolysis and sedimentation times were around 2 h and 2–4 h, respectively. The total operating cost of this process was calculated to be 0.31 US $/m3 (electricity and iron electrode costs account for about 82.0% and 18.0%, respectively). In another study, a larger scale of the ECAR arsenic removal process for field trials was used. This process is comprised of three 750-L sintex tank reactors, one for the electrochemical process (600 L) and two for sedimentation (Amrose et al., 2014, 2013). A 600-L ECAR reactor was operated for 15 weeks for the treatment of real As-contaminated groundwater in West Bengal (Fig. 10). The electrolysis tank was comprised of 4 cores with parallel interdigitated mild-steel plates (5 anodes and 5 cathodes). As expressed by the authors, the design of the EC process permits for easy scale-up of the reactor by the addition of more cores to a larger EC reactor. Furthermore, this design facilitated current reversal of the EC process in order to minimize the passivation of electrodes and to reduce extensive rust build-up. During the electrolysis, water is recirculated by the tank with flow rate of about 3.4 L/h. After the electrolysis, water is pumped into the sedimentation tank and dissolved alum (6–15 mg/L) is added to accelerate the sedimentation. The coulombic dose, electrolysis time and total energy consumption were 450 C/L, 105 min and 2.31 kWh/m3 , respectively. The amount of produced dried sludge was 245 g/m3 . These trials decreased arsenic concentrations of ∼266 ± 42 µg/L to > 5 µg/L in real groundwater and total operational costs based on field trial results were in the range of 0.83–1.04 US $/m3 . Aluminum concentration of the treated water was below 0.10 mg/L. The ECAR process produced small amounts of sludge that can be easily stabilized by concrete. A pilot plant named Lamar Mobil Electrocoagulation pilot plant in field trials was used to remove arsenic most efficiently from groundwater (initial arsenic = 25–50 µg/L, conductivity = 600–4000 µS/cm, and pH = 5.5–7.1) containing naturally arsenic pollution in La Comarca Lagunera, Mexico (Parga et al., 2005). An EC reactor with air injection in the pilot plant was used. Groundwater containing 40 µg/L of arsenic with pH value of 7.0 was drawn directly from the well and transferred to the pilot plant comprised of EC reactor, pumps, sedimentation tank, and a rectifier. The rectifier timers were set to 5 min; however, the timer was set to 20 min near the end of the first pass because of the stable voltage. The residuals generated during the EC process were collected and processed by a filter press. The arsenic removal efficiency of this pilot plant (flow rate: 30 L/min) at reaction time 1 min, applied voltage of 20–30 V and applied current of 5 A was higher than 99.0% (residual arsenic: 2 µg/L, final pH: 8.5 and final conductivity: 500–2000 µS/cm). The energy cost of the process was

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Fig. 10. The ECAR process in a reactor with a volume of 600 L for field trials (Redraw from Amrose et al. (2014) with permission from Elsevier).

0.0020 US $/m3 . Results revealed that EC reactor injected by air could be a very promising technology to remove As(III) and As(V), and that this process produces magnetic particles of magnetite and amorphous iron oxyhydroxides. EC reactor with air injection was expressed to be having advantages such as high-arsenic removal efficiency, shortened retention time, minimum sludge and low operational cost. For the treatment of Zrenjanin and Temerin groundwater (Vojvodina, Serbia), an effective pilot-scale system was developed (Orescanin et al., 2013). The physicochemical parameters of the groundwater in Zrenjanin and Temerin were pH of 8.79 and 8.27, conductivity of 1.79 and 0.83 mS/cm, total dissolved solids of 1240 and 570 mg/L, color of 142 and 26 Pt-Co, total arsenic of 0.075 and 0.088 mg/L, total iron of 0.728 and 0.273 mg/L, ammonia of 0.85 and 1.54 mg/L, and COD of 13 and 9 mg/L, respectively. The treatment capacity of the pilot reactor was 1000 L/day. The volume of 70 L of raw water from the Temerin area was pumped into 100 L reactor and subjected to EC reactor equipped with iron electrode (electrodes thickness: 2 mm, dimensions: 20×50 cm, distance between the plates: 5 mm, current intensity: 65 A, applied voltage: 12 V), followed by the treatment with aluminum electrodes with the same dimensions to the iron electrode (current intensity: 65 A and applied voltage: 12 V) with reaction time of 10 min which was simultaneously treated with ozone (500 mg/h). The resulting effluent was pumped into the sedimentation tank to separate flocs formed during the previous stage within the retention time of 30 min. The sedimentation of sludge was speeded up by the electromagnetic treatment. Arsenic removal efficiency from groundwater’s of Zrenjanin and Temerin towns was 100% by the electrocoagulation with simultaneous ozonation. Pilot and large scale arsenic removal studies revealed that one of the important problems in arsenic removal with EC process is the operating cost, including emerging costs of energy consumption, stabilization and disposal of arsenic-laden waste sludge. The success of this treatment technology depends on its ability to attain the minimized operational costs and achieve a successful investment in the shortest time period. In addition, larger scale and longer-term field trial studies with different water samples containing arsenic pollution are needed. 9. Conclusion and outlook The EC process has been found to be an effective process for the treatment of As-contaminated aquatic phase. According to the reviewed literature, As removal from contaminated water samples by EC processes was is the following range 80.0– 99.99%. Different electrode types (like Fe, Al and Fe–Al hybrid) and shapes (like plate, rod and scrap) for As removal have

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been successfully used in the EC process. The removal of As by Fe electrodes-equipped EC process has been reported as more effective than other EC processes due to the enhanced adsorption and co-precipitation reactions. The removal efficiency of As with Fe plate electrodes-equipped EC process varied between 75.0 and 99.9% regarding operational costs between 0.0020 and 1.00 US $/m3 . The costs of As removal by hybrid Fe–Al plate (0.0023 US $/m3 ) were lower than Fe scrap (0.017 US $/m3 ) electrodes. Based on the presented data, charge loading can also be a key parameter in influencing the removal of As by the EC process. Increasing applied current density generally results in enhanced removal efficiency of As by the EC process, leading to the subsequent reduction in the reaction time (1–30 min) for pollutant removal. The high removal efficiency of As (99.9% for 80–760 µg/L) with operational costs of 0.22 $/m3 has been obtained when pilot-scale EC reactors were examined for the treatment (working volume of > 100 L). Overall, the EC process is an efficient technique for the removal of both As (III) and As (V) without the need to oxidize As (III) to As (V) unlike other As removal processes. As opposed to other conventional treatment methods, the EC process appears to be simple and easy to operate technique for the removal of As that does not require any chemical addition and process control. Moreover, this process does not produce any concentrated waste (like membrane processes) and secondary contaminants (such as ion exchange processes). The presence of other co-existing ions in the solution does not have a major adverse effect on the removal of target contaminant (As) by the EC process. Nevertheless, the EC process has some drawbacks such as electricity (energy) requirement and waste sludge management just like other processes designed for the removal of As. To minimize energy requirements for the EC process, clean energy sources such as solar energy can be utilized instead of expensive sources. This strategy also reduces the cost of water treatment process designed for decontamination of As. Clearly, EC is an effective process to attain the WHO guideline (10 µg/L) for safe drinking water without any long term adverse effects of As on human health. Overall, the application of EC process requires to be examined further to justify their ability as an efficient treatment process for removal of As-contaminated water streams. Furthermore, a comprehensive assessment of the process using different water matrices is required to determine the mechanism of the EC process. 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