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Wellner, Daniel, Couperthwaite, Sara, & Millar, Graeme(2018)The influence of coal seam water composition upon electrocoagulationperformance prior to desalination.Journal of Environmental Chemical Engineering, 6(2), pp. 1943-1956.

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https://doi.org/10.1016/j.jece.2018.02.042

Accepted Manuscript

Title: The Influence of Coal Seam Water Composition uponElectrocoagulation Performance Prior to Desalination

Authors: Daniel B. Wellner, Sara J. Couperthwaite, Graeme J.Millar

PII: S2213-3437(18)30114-3DOI: https://doi.org/10.1016/j.jece.2018.02.042Reference: JECE 2234

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Received date: 22-11-2017Revised date: 4-1-2018Accepted date: 24-2-2018

Please cite this article as: Daniel B.Wellner, Sara J.Couperthwaite, GraemeJ.Millar, The Influence of Coal Seam Water Composition upon ElectrocoagulationPerformance Prior to Desalination, Journal of Environmental ChemicalEngineering https://doi.org/10.1016/j.jece.2018.02.042

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The Influence of Coal Seam Water Composition upon Electrocoagulation Performance

Prior to Desalination

1Daniel B. Wellner, Sara J. Couperthwaite and *Graeme J. Millar

Institute for Future Environments, 1School of Chemistry, Physics & Mechanical Engineering,

Science and Engineering Faculty, Queensland University of Technology (QUT), Brisbane,

Queensland 4000, Australia.

*Corresponding author:

Graeme J. Millar | Professor

Science and Engineering Faculty | Queensland University of Technology

P Block, 7th Floor, Room 706, Gardens Point Campus, Brisbane, Qld 4000, Australia

ph (+61) 7 3138 2377 | email graeme.millar@qut.edu.au

EC was investigated to ascertain its applicability to remove dissolved species from a variety of

associated water samples typical of coal seam gas (CSG) operations. The hypothesis was that

the CSG water composition may impact EC performance for the removal of problematic

species such as alkaline earth ions and dissolved silicates. Bench top studies of a range of CSG

associated water samples revealed that the greater total salinity (conductivity from 5290 to

15680 μS/cm) the less alkaline earth ions were removed. However, dissolved silicate

remediation maintained high efficiency (89.5 to 98.0 %) regardless of water salt content.

Residual aluminium was present in treated water when aluminium electrodes were employed

(4.6 to 39.0 mg/L) and correlated with increasing solution pH. In contrast, steel electrodes

did not result in notable residual iron. Whether steel or aluminium electrodes were optimal

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depended upon the CSG water salinity. Aluminium based flocs were discovered to settle

significantly slower than iron based flocs, with salinity influencing aluminium flocs properties

more than iron flocs. Differences in the presence of amorphous species and crystalline

gibbsite may in part explain the floc settling behaviour. In either case, dewatering of flocs

represents a technical challenge. The major cost in terms of economics was electrode

consumption whether iron or aluminium electrodes were used. The system with lowest

operating cost was always iron (A$2.50 to 2.68 per kL compared to A$2.70 to 4.32, per kL for

aluminium) regardless of water salinity.

Key Words: Electrocoagulation (EC); Coal Seam Gas (CSG); Coal Bed Methane (CBM);

Production Water; Associated Water

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1. Introduction

The development of unconventional gas resources such as coal seam gas (coal bed methane)

in recent years has provided an opportunity for economic growth [1]. Despite the advantages

that the availability of natural gas gives to mankind, there are also problems to be solved in

relation to the by-product saline water which is produced when the gas is extracted [2].

Invariably, the dominant species dissolved in the coal seam (CS) water are sodium chloride

and sodium bicarbonate [3, 4]. Also present are alkaline earth ions, silicates, sulphate,

potassium, iron, boron, aluminium and a variety of minor species [4]. A complicating factor

is the fact that the coal seam water composition is highly variable and related to the gas

exploration region and location of the wells [5, 6].

A primary goal for the CSG industry is to regard the produced water as a resource and not a

waste material; hence, beneficial reuse options such as crop irrigation, dust suppression, coal

washing and livestock watering should all be considered [3]. In some instances, coal seam

water can be used without further treatment; however, in many cases the water contains

concentrations of dissolved ions in exceedance of reuse guidelines and/or exhibits a sodium

absorption ratio (SAR) value which is not conducive to discharging the water to soil [7].

Therefore, usually a desalination stage needs to be implemented in order to recover purified

water for beneficial reuse and to concentrate the dissolved species into a brine solution which

can be disposed of [8]. In terms of coal seam water treatment, reverse osmosis (RO), ion

exchange, and membrane distillation (MD) have been proposed as desalination methods to

recover purified water. RO is presently the preferred technology in Australia [9], with several

facilities utilizing this process [8, 10]. Ion exchange has been deployed in the Powder River

Basin region of the USA, with a cation resin sufficient to remove sodium ions and decrease

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bicarbonate concentrations due to the acidic process conditions [11]. Bench trials of ion

exchange for desalination of coal seam water have also proven the effectiveness of this

technique [12-14]. MD has also been reported for the recovery of purified water from coal

seam water by Duong et al. [15]. The combination of MD and membrane electrolysis allowed

not only recovery of water but also sodium hydroxide.

Membrane based technologies are known to be susceptible to fouling or scaling by species

such as organics, salts of alkaline earth ions and silicates [16]. As such, membrane systems

usually employ extensive pre-treatment processes to protect the membranes and maintain

high water recovery rates [3]. For example, Lipnizki et al. [17] recommended the use of ion

exchange resins prior to a RO stage in order to remove alkaline earth ions and thus reduce

the potential for scaling. Duong et al. [18] investigated the mitigation of scale formation in a

MD unit used to treat RO brine derived from CS water. Application of cleaning chemicals

removed the majority of scalants but residual silicates inhibited the membrane performance.

Lowering of the brine temperature limited the extent of scaling but decreased the flux rates

obtained. Chun et al. [19] demonstrated the application of forward osmosis to minimise

problems with downstream reverse osmosis systems.

Protection of the central desalination unit operation by suitable pre-treatment technologies

is therefore of importance. Ideally, a method which can remove multiple contaminants from

solution would be practically preferred. Electrocoagulation (EC) has been the subject of

numerous studies due to its demonstrated ability to remove a variety of species from a range

of wastewater types including heavy metals, alkaline earth ions, silicates, boron, dyes and

suspended solids [20-22]. The application of EC to treat produced water from the oil and gas

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industry has gained interest in recent years. Lobo et al. [23] described the use of an EC unit

equipped with aluminium electrodes with or without the addition of biochar to remove

turbidity and suspended solids from hydraulic fracturing flowback water from the Denver-

Julesburg basin in USA. Alternating current was proposed to be more effective than direct

current and overall the reduction in turbidity and suspended solids was typically >98 %.

However, chemical oxygen demand and total dissolved solids were relatively difficult to

reduce in value. Sari and Chellam [24] reported the use of EC which used aluminium

electrodes for the control of boron in hydraulic fracturing wastewater. These authors found

that boron was probably removed via a ligand exchange mechanism and important

parameters were the quantity of aluminium dissolved into solution, current density, and

solution pH. Esmaeilirad et al. [25] examined the impact of softening in conjunction with EC

to treat hydraulic fracturing flowback water. Softening before the EC stage was deemed

better in terms of the overall degree of alkaline earth ion removal.

A recent study has shown that EC has potential to significantly reduce the concentrations of

calcium, magnesium, strontium, barium, silicate, dissolved organic carbon (DOC) and species

causing turbidity from coal seam water produced by an operating coal seam gas field [26]. In

the best case, all four alkaline earth species present could be almost completely removed

from the coal seam water sample. Notably, dissolved silicate species which are known to be

particularly problematic [27] and difficult to remove from solution by alternate methods such

as sorption on alumina [28, 29] were also diminished by >90 %. The formation of

aluminosilicate materials was proposed as one reason for the high efficiency silica removal.

One significant aspect of previous studies was the observation of aluminium anode

consumption in excess of Faradaic amounts. Super-faradaic dissolution of aluminium in an

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EC process has been described in detail by Mechelhoff et al. [30] and related to the presence

of chloride ions in solution. Canizares et al. [31] also reported substantial chemical dissolution

of aluminium electrodes in salt solutions particularly at alkaline pH.

From the literature review, it was apparent that information did not currently exist in relation

to the applicability of electrocoagulation to treat a range of coal seam water types which vary

not only in total concentration but also composition. In addition, iron (steel) electrodes have

not been studied previously for CSG associated water treatment despite their extensive use

in EC systems [32]. Therefore, this study focussed upon extending previous studies using EC

to treat CSG associated water with a range of salinity values and hardness. The hypothesis

was that the variability in salt content, solution pH, alkaline earth ions, and dissolved silica

species may impact electrocoagulation performance. In particular research questions

addressed included: (1) what is the impact of CS water concentration upon electrode

consumption and electricity requirements? (2) are the removal efficiencies of dissolved

species influenced by variability of CS water composition? (3) do the characteristics of flocs

produced change as a function of the type of CS water treated? (4) what is the preferred

electrode material to employ? To answer the aforementioned research questions a

continuous bench top EC unit equipped with either aluminium or steel (iron) electrodes was

used to treat a range of simulated CS water types which allowed precise control of solution

parameters.

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2. Materials and Methods

2.1 Electrocoagulation Cell

A continuous electrocoagulation unit was constructed which comprised of 13 electrode plates

configured in a bipolar arrangement with the direct current supplied via external electrodes

[33]. Vertical flow of the solution through the EC cell from bottom to top was via a 6 mm

diameter inlet. A diffuser plate was located above the inlet to ensure even dispersion of the

solution across the electrode surfaces. The plates used were 10 cm length, 15 cm height and

0.3 cm width, spaced 3 mm apart. Aluminium plates were laser cut from 5005 grade

aluminium sheets which contained up to 0.7 % iron content. The steel electrodes were

similarly constructed from mild steel (< 0.25 % carbon). When used in a bipolar configuration

the 13 electrodes resulted in 1800 cm2 of active anodic surface area. For these experiments

the residence time of the fluid in the EC cell was fixed at 30 s which equated to a flow rate of

1.08 L/min. The unit was usually operated in constant current mode and the maximum

voltage which could be applied was 37.9 V. A polarity reversal rate of 30 seconds was chosen

in all instances based upon previously published studies [26, 34]. Fresh aluminium and iron

plates were used for each test in order to facilitate comparison of results. Aluminium

electrodes were cleaned using acetone and light scrubbing to remove the oil layer on the

plate. Aluminium electrodes were subsequently placed in an oven for 1 h to dry and then

weighed before use. After use, a light rinse with deionised water was applied to remove any

material adhered to the plate, then the electrode was dried for 1 h at 100 oC before being

weighed to determine mass loss. Mild steel plates were also washed using deionised water

accompanied by light scrubbing with a scouring pad. After cleaning the surface of any residual

oils or slight surface corrosion, they were immersed in deionised water until they were dried

quickly to remove the bulk of the fluid and placed in an oven at 100 oC for 10 min. The plates

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were weighed and immersed in deionised water in the electrocoagulation cell until use. The

plates were treated in the same fashion after use to obtain a total mass loss per plate. The

process effluent was collected in 20 L Nalgene containers. To determine the settling rate of

the process waters, the bulk process water was agitated and then placed into a 2 L measuring

cylinder.

2.2 Chemicals

Simulated analogues of CS water were prepared by the addition of requisite salts to deionised

water. All reagents used were purchased from Chem-Supply unless otherwise stated. Solid

analytical reagent grade sodium chloride, barium chloride dihydrate, boric acid, magnesium

chloride, strontium chloride (Sigma-Aldrich), calcium carbonate, potassium carbonate,

sodium bicarbonate, sodium carbonate, calcium chloride, sodium hydroxide, ammonium

chloride and anhydrous potassium silicate (Alfa Aesar) were used in conjunction with solution

based hydrochloric acid (32 wt/V %), sodium hydroxide (40 wt/V%), and absolute ethanol

diluted as required. Sodium metasilicate pentahydrate and iron(III) chloride were also used

when necessary. The precise quantities of reagents required to obtain the desired CS water

composition were calculated using an Excel model.

2.3 Analysis

2.3.1 Inductively Coupled Plasma- Optical Emission Spectroscopy (ICP-OES)

Dissolved species such as sodium, potassium, calcium, magnesium, barium, strontium, iron,

aluminium, boron, and silica were analysed using an inductively coupled plasma optical

emission spectrometer (PerkinElmer Optima 8300 DV). Aqueous samples were first filtered

through a 0.45 µm syringe filter to remove any residual solids, then diluted as required using

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a Hamilton auto diluter by addition of nitric acid (2.5 %). In cases where undiluted samples

were required the sample filtered and spiked with purified nitric acid (200 µL, 70 %) to adjust

the pH to ca. 2 prior to analysis. Calibration standards from Australian Chemical Reagents

were used to calibrate the ICP-OES unit. Elemental analysis of floc materials by ICP-OES

involved digestion of ca. 1 g of the sample in 5.75 mL of aqua regia, prepared by the addition

of 4.0 mL HCl (32 wt/V %) and 1.75 mL HNO3 (70 wt/V%), diluted to 50 mL using deionized

water. A sample of 5.75 mL aqua regia was diluted to 50 mL in order to be used as a blank for

each sample set collected. Undissolved solids from the digestion were collected by filtration

through a glass fibre filter paper. The filter paper was dried for at least one hour in an oven

at 105 °C and weighed before and after filtration to obtain the mass of undissolved solids.

2.3.2 X-Ray Diffraction (XRD)

Specimens were prepared for powder X-ray diffraction by the addition of accurately weighed

(ca. 0.1 g) corundum (Al2O3, Bai) internal standard to the samples. The total mass of sample

and standard was then recorded in order to calculate the known wt% of corundum. The

corundum and sample mixtures were then micronised in a McCrone mill for 6 minutes with

ethanol (ca. 10 mL). The resultant slurry was dried overnight at 40 °C, then the dried

homogenous powders front-pressed into sample holders. Powder X-ray diffraction patterns

were collected with a PANalytical X’Pert Pro MPD in Bragg-Brentano geometry with a cobalt

source operating at 40 kV and 40 mA. The incident optics included 0.04 radian Soller slits, 15

mm mask, 0.5° fixed divergence slit and a 2° anti-scatter slit. The receiving optics before the

X’Celerator detector included a 0.04 radian Soller slit, 5.0 mm anti-scatter slit and an iron Kβ

filter. The samples were spun during analysis. The scan range was 4 – 90 °2θ at a step size of

0.0167° with a total scan time of 30 minutes per sample. Phase identification used various

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databases (American Mineralogist Crystal Structure Database, PDF4+, Crystallography Open

Database) in both Jade (V4.1.0, Materials Data Inc.) and X’Pert Highscore Plus (V4,

PANalytical) software. Quantitative phase analysis was performed using the Rietveld method

as implemented in TOPAS (V5, Bruker). An instrument function previously determined from

LaB6 (SRM 660a) was used to model the peak shapes. This instrument function also

incorporated a custom Co emission profile which included a Kβ component for the specific

diffractometer employed. Refined parameters during quantitative phase analysis were 15

term Chebyshev background and specimen displacement, and for each phase: scale factor,

unit cell parameters, and Lorentzian crystallite size and strain terms as appropriate. All phase

concentration estimates (wt %) reported were absolute and corresponded to the original

sample. The degree of crystallinity method was also used for some samples that had a

significant highly disordered component, or unidentified peaks that could not be successfully

attributed to a known crystalline phase.

2.3.3 Scanning Electron Microscopy (SEM) and Energy Dispersive Spectroscopy (EDS)

Sections of the used and as received electrodes were analysed using a JEOL-7001 scanning

electron microscope, at a working distance of 10 mm and an operating voltage of 15.0 kV.

The electrodes were analysed using ATLAS EDS imaging software to assist in determining the

fate of removed contaminants. Electrodes were cut to size (20 mm x 20 mm) using a

discatom-600 with an aluminium cutting wheel. The cutting rate was set to 0.5 mm/s and the

cooling streams were active. After cutting, the samples were rinsed with deionized water and

dried in an oven at 60°C for two hours. The samples were then mounted on the sample stages

and blown with compressed dry nitrogen before mounting on the microscope stage.

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2.3.4 pH and Conductivity

pH and conductivity measurements were taken using TPS Aqua probes, which were calibrated

daily using appropriate standards. The probes were cleaned after every use by rinsing with

deionised water and the pH probes were stored in saturated potassium chloride adjusted to

ca. pH 2. The probes were cleaned weekly according to the method recommended by the

manufacturer.

2.4 Mass Balance Studies

Mass balance of aluminium/iron during electrocoagulation tests was achieved by weighing

each electrode plate prior to and immediately after EC use. The percentage of aluminium or

iron in the produced flocs was obtained by acid digestion.

2.5 Power Consumption and Faradaic Yields

Power consumption was calculated using the formula shown in Eq. 1.

Eq. 1: 𝑷 = 𝑰𝑽𝑸⁄

Where P = average power consumption (W), I is the average current (A), V is the average

voltage (V) and Q is the flow rate.

Faradaic yields were based upon the following expression:

Eq. 2: 𝒎 = 𝑰𝒕𝑴𝒛𝑭⁄

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Where m is the mass of electrode material released to solution (g), M is the atomic mass of

either aluminium or iron as appropriate, z is the number of electrons transferred during

anodic dissolution (=3) and F is Faradays constant (96,486 C/mol).

Calculation of the economics of electrode consumption was based upon assigning a price for

aluminium sheet of A$7.50 per kg and a price for mild steel sheet of A$2.50 per kg (based

upon local market assumptions in Australia). A recent study by Demirci et al. [35] used a price

for aluminium sheet the Turkish marketplace of 3.06 euros/kg (A$4.40/kg based upon current

exchange rate as of January 2017). Similarly, Touahria et al. [36] assumed a price of US$3.08

(A$4.42), hence our assignation appears reasonable accounting for higher cost basis in

Australia. Kobya et al. [37] assumed an iron electrode cost of 0.85 euros/kg (A$1.22/kg) based

upon prices in 2009, and the ratio we have used of 3:1 Al:Fe cost is in line with the reported

literature. An electricity price of A$0.2 per kWh was used to calculate the power costs during

EC operation.

2.4 Coal Seam Water Compositions

Three samples of CS water were simulated to represent a cross-section of samples typical of

the Surat basin in Queensland. The three waters were chosen as representations of a low,

medium, and high TDS solution that would typically be found from coal seam waters [6]. The

compositions of each water can be found in Table 1.

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3. Results and Discussion

3.1 Impact of Total Dissolved Solids Content upon Contaminant Removal

Electrocoagulation tests were conducted in order to compare not only the ability of EC to

treat the three CS water compositions of interest but also to examine the differences between

aluminium and iron electrodes [Figure 1]. In general, the concentration of sodium ions was

relatively constant (not shown for sake of brevity) as expected due to the fact that

electrocoagulation does not provide a means to remove large concentrations of singly

charged cations from solution [38]. Application of aluminium electrodes appeared to be more

effective than iron electrodes in terms of removal of alkaline earth ions and dissolved silicate

species for low and medium TDS CS water samples [Table 2]. Heffron et al. [39] noted that

one difference between the use of aluminium and iron electrodes was the surface charge of

the flocs formed. In solutions near neutral in pH, aluminium hydroxide species were anionic

in character whereas iron hydroxide species were cationic. Thus uptake of cations may be

preferred when using aluminium electrodes. Interestingly, for higher TDS CS water [Table 2]

the removal performance was generally better for iron based electrodes compared to

aluminium electrodes. Thus, selection of the most appropriate type of electrode in the EC

cell may depend upon the composition and salinity of the CS water to be treated.

Of the alkaline earth ions present in the CS water samples, typically barium ions were

removed to the greatest extent whether with aluminium or iron electrodes, albeit at the

highest CS water concentration magnesium ions were most easily removed when using

aluminium electrodes. The general trend was that as the CS water total concentration

increased the ability of the electrocoagulation process to remove alkaline earth ions

decreased when employing aluminium electrodes. For example, for the low TDS solution the

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degree of barium reduction was 100 % with aluminium electrodes but only 23.2 % when the

highest TDS CS water sample was treated [Table 2]. With iron electrodes the extent of alkaline

earth removal was diminished when changing from a low to medium TDS CS water sample,

but upon treating a high TDS CS water sample the degree of alkaline earth ion removal

exhibited a slight increase (albeit not to the same magnitude as with the low TDS sample).

The reduction of barium ion content of brackish water using electrocoagulation has not been

reported in many previous studies.

De Oliveira Da Mota et al. [40] used an electrocoagulation/electroflotation cell to treat

solutions which simulated wastewater from an operation which had soil contaminated with

drilling fluids. The removal of barium ions was demonstrated to be favourable over a wide

range of pH values (4 to 10) with typical removal rates in excess of 85 %. Notably, increasing

current density was found to favour the remediation of pollutants in this study, whereas

greater ionic strength of the wastewater diminished the removal rates slightly. For instance,

reduction in lead ion content of the solution was lowered by ca. 1 % when the ionic strength

was elevated from 0.0032 to 0.08 M. The rationale for this latter behaviour was proposed to

relate to increasing competition between ions. In the present study, the decrease in removal

efficiency for alkaline earth ions was more substantial. Esmaeilirad et al. [25] reported the

remediation of hydraulic fracturing flowback water which comprised of a range of alkaline

earth ions and concentrations using electrocoagulation. However, it was difficult to ascertain

the effectiveness of the EC unit for alkaline earths as a softening stage was employed either

before the EC or after the EC system.

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(a) Low TDS CS Water

(b) Medium TDS CS Water

(c) High TDS CS Water

Figure 1: Electrocoagulation of Various CS water compositions as indicated: Flow rate = 1.08

L/min; Polarity reversal period 30 s; hydraulic retention time 30 s; treatment time = 40 min.

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Murthy and Parmar [41] studied electrocoagulation of solutions of strontium chloride in

water containing dissolved sodium chloride. Key observations included strontium removal by

application of greater current density or sodium chloride concentration; optimal strontium

removal in the pH range 5 to 7 and decreasing strontium remediation as the initial amount of

strontium in solution increased from 25 to 100 mg/L. These latter authors found that stainless

steel electrodes consistently outperformed aluminium electrodes. Oncel et al. [42] studied

the ability of electrocoagulation to remove a wide range of contaminants from coal mine

drainage wastewater (CMDW), including calcium, magnesium and strontium. A critical factor

was the current density, which when raised to 500 A/m2 was sufficient to reduce

concentrations of all the dissolved species by > 99.9 %. Notably, the optimal pH was acidic

which inferred that precipitation of the alkaline earth species was not the primary mechanism

of removal. Unfortunately, detailed investigation of the fate of the calcium, magnesium, and

strontium was not provided. Zhao et al. [43] also demonstrated both increasing current

density and initial solution pH promoted the reduction in water hardness by application of

iron electrocoagulation. These authors were of the opinion that the hardness was removed

by precipitation of calcium carbonate and calcium sulphate and also enhanced sweep

flocculation at alkaline conditions. The extent of precipitation as a mode of hardness

reduction in coal seam water was uncertain as the solution was buffered and did not show a

large change in solution pH, albeit it trended to higher alkaline values. Of course, bulk solution

pH was only an average value of the various pH regions present in an electrocoagulation

system [44].

Kamaraj and Vasudevan [45] indicated that a solution pH in the range 7 to 8 was optimal for

strontium ion removal from solution using EC and indicated that strontium ions could be

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reduced in abundance by up to 97 %. Our study with CS water was in agreement with Murthy

and Parmar [41] in that higher initial strontium ion concentrations resulted in decreased

removal using electrocoagulation. However, aluminium electrodes were more favoured with

CS water for strontium control and elevation of sodium chloride concentrations in fact

inhibited strontium removal from solution. This investigation has revealed that complex,

multi-component solutions behave differently from simplified test solutions. Malakootian

and Yousef [46] used an EC unit equipped with aluminium electrodes to reduce hardness

levels of water and observed a trend where hardness was reduced as the pH was increased

from acidic to alkaline conditions. The different forms and efficiencies of the aluminium

coagulants produced under the various pH conditions was suggested to be responsible for the

alkaline earth ion degree of removal from solution. Esmaeilirad et al. [25] also examined the

solubilities of various carbonate, sulphate and hydroxide salts of alkaline earth ions in relation

to flowback water treatment by EC. Barium sulphate was the least soluble species evaluated

and thus suggested to be able to precipitate from solution under the applied experimental

conditions. However, the presence of this latter salt was not considered to be feasible to

explain the current results. The concentration of sulphur containing species (presumably

sulphate anions) in the CS water sample was typically ca. 3.5 mg/L and during the EC tests the

level of removal was less than 5 %.

Regardless of the CS water composition or the identity of the electrodes used in the EC unit, dissolved

silicate species were consistently removed with high efficiency (89.5 to 98 %). Silica is of particular

interest in terms of pre-treatment due to the high propensity for scale formation on process

equipment and membranes [28]. The removal of silicate species when aluminium electrodes were

used is proposed to be at least in part to the formation of aluminosilicates [Equations 3 & 4]. Den

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and Wang [47] applied electrocoagulation using aluminium electrodes to brackish water and

deduced that increasing the charge across the cell typically resulted in greater removal efficiencies.

However, they noted that higher hydraulic retention times were potentially detrimental to EC

performance.

Equation 3: 𝟐 𝐀𝐥(𝐈𝐈𝐈) + 𝟐 𝐒𝐢(𝐎𝐇)𝟒 + 𝐇𝟐𝐎 ↔ 𝐀𝐥𝟐𝐒𝐢𝟐𝐎𝟓(𝐎𝐇)𝟒 (𝐬) + 𝟔 𝐇+

Equation 4: 𝐀𝐥(𝐈𝐈𝐈) + 𝐒𝐢(𝐎𝐇)𝟒 + 𝐇𝟐𝐎 ↔ 𝐀𝐥𝐎𝐒𝐢𝐎𝟓(𝐎𝐇)𝟑𝟐+ + 𝐇+

Application of coagulants to treat water and wastewater samples can lead to concerns

regarding the presence of residual aluminium or iron species in solution post-treatment. He

et al. [48] studied the use of aluminium chloride and polyaluminium chloride (PACl)

coagulants for the removal of fluoride species from drinking water and noted the production

of up to 16 mg/L residual aluminium in solution. The quantity of residual aluminium present

was dependent upon factors such as the identity of the coagulant, with aluminium chloride

producing larger amounts of residual aluminium than PACl, and initial concentration of

fluoride ions in solution. Chen et al. [49] tested a range of iron based coagulants for the

clarification of solutions comprising of humic acid and kaolin particles. Residual iron levels of

up to 7.5 mg/L were observed for pH values less than 7.

Consequently, we examined whether the application of EC to treat CS water resulted in the

presence of significant quantities of residual metals in the effluent. In all cases, when iron

electrodes were used we did not detect any notable increase in the iron content of the treated

CS water sample. As outlined by Jimenez et al. [50] at alkaline pH values iron species convert

from monomeric hydroxoiron ions to iron hydroxide precipitates, thus the presence of

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residual dissolved iron species in high pH CS water solutions was not expected or indeed

recorded. However, when aluminium electrodes were employed in the EC unit it was found

that significant amounts of dissolved aluminium were present [Figure 2]. The recorded

residual values were 39.0, 32.1, and 4.6 mg/L Al for the low, medium, and high TDS CS water

samples, respectively. Inspection of the effluent solution pH values displayed in Figure 3

suggested that there was a correlation between the highest solution pH recorded for each CS

water sample and the amount of residual aluminium detected; namely, the higher the

solution pH the greater the quantity of residual aluminium present.

Figure 2: Aluminium ions released into solution during electrocoagulation with aluminium

electrodes of various CS water compositions: Flow rate = 1.08 L/min; Polarity reversal period

30 s; Contact time 30 s; Current fixed at ca. 5 A; Treatment time = 40 min.

This latter observation was consistent with the known chemistry of aluminium in aqueous

solution whereupon aluminium dissolution is promoted at alkaline pH values due to the

formation of soluble complexes such as Al(OH)4- [31]. At a pH of ca. 9 the percentage of

aluminium hydroxide precipitates was expected to begin to reduce in concentration and as

the pH is increased further the presence of soluble monomeric hydroxoaluminium species

was accelerated [50].

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The question arose as to means to reduce the presence of aluminium species in the treated

water after electrocoagulation. Sinha and Mahur [51] addressed the issue of residual

aluminium present in solution following electrocoagulation of fluoride solutions. Activated

silica sol was added to the effluent from the EC treatment of solutions containing both sodium

fluoride and sodium chloride. Aluminium ions were initially in the range 11.6 to 15.3 mg/L

due to the solution pH obtaining a value of 9 or greater wherein the formed Al(OH)4 species

become soluble. The presence of activated silica sol consistently reduced the aluminium

concentration to less than 0.1 mg/L, and thus may be one possible means to control the high

levels of residual aluminium ions when EC treated CS water. Sinha et al. [52] also assessed

the effectiveness of added bentonite clay to control residual aluminium levels to below 0.2

mg/L and discovered that 2 g/L doses were sufficient to attain the latter goal. Heffron et al.

[39] evaluated the ability of electrocoagulation using either aluminium or iron electrodes to

remove a range of heavy metal ions from drinking water. Residual iron and aluminium

concentrations in the treated water were both in excess of drinking water regulations.

Interestingly, filtration of the EC effluent was effective in removing aluminium species when

the pH was 6.5 but significantly less effective at pH 8.5. The alkaline pH values obtained when

using EC to treat CS water of 9 to 10 would be expected to make a filtration option even less

desirable as the solubility of aluminium hydroxide species increased. Baciu et al. [53] reported

a means of minimizing residual aluminium formation when treating groundwater with an EC

unit equipped with aluminium electrodes. These authors proposed a new methodology

which they termed “adaptive electrocoagulation” which comprised of an initial treatment

step wherein the current density (100 A/m2) applied was relatively high over a small time

period (5 min). A second step was introduced wherein a low current density (25 A/m2) was

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applied for 10 min. The underlying idea was to simulate the fast mixing and maturation stages

in conventional coagulation processes. As a result residual aluminium ion concentration was

reduced to only 0.01 mg/L (compared to values of 12 to 32 mg/L Al observed during

conventional electrocoagulation). Another strategy which could be employed is to simply add

acid to the treated water and lower the pH to less than 9 in order to remove the presence of

soluble Al(OH)4- species.

3.2 Impact of Total Dissolved Solids Content upon Electrocoagulation Performance

Parameters

3.2.1 Aluminium Electrodes

It was pertinent to examine whether solution pH variation in the CS water EC tests was a

possible explanation for the observed behaviour in Figure 1. Figure 3 displays the pH of the

EC treated effluent solution as a function of experimental time. As a general comment, the

effluent pH was not observed to be stable and indeed oscillated in value. The extent of the

pH oscillation appeared to be dependent upon the CS water concentration as the oscillation

amplitude was greater as the solution concentration increased. The period of the pH

oscillation correlated with the polarity reversal time employed and was found to repeat

approximately every 60 s (i.e. twice the polarity reversal period). Oscillatory pH behaviour

within an electrocoagulation unit was reported by Moreno et al. [54] when using a static

system. It is generally agreed that addition of sodium chloride aids in removal of the

passivation layer which can accumulate on the electrode surfaces during EC operation. Thus

it would appear that the cleaner the electrode surface the greater the variation in effluent

solution pH. It has to be determined whether the pH fluctuation is a positive or negative

aspect to the EC treatment method. If the removal of the dissolved contaminant was by

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precipitation for example, then reduction in solution pH may redissolve the precipitate which

could be detrimental. It is also widely believed that Al13 polymer species

((AlO4Al12(OH)24(H2O)12)7+) are optimal in terms of coagulation performance. As indicated by

Pi et al. [55] the precursor to Al13 species is thought to be Al(OH)4− which is typically formed

at pH values above 9. Hence, the decrease of pH to acidic values during the oscillatory pH

behaviour would be expected to inhibit Al13 production and thus reduce coagulation

performance.

It was noted that the voltage required to maintain a constant current value of ca. 5 A,

generally decreased with increasing electrocoagulation time before ultimately stabilizing as

steady state conditions were achieved. Mouedhen et al. [56] showed similar trends when

they applied electrocoagulation with aluminium electrodes to solutions of sodium sulphate

and sodium chloride. For aluminium electrodes, the voltage initially reduced in value during

the initial stages of the CS water treatment before plateauing. As the current was reasonably

constant a reduction in voltage indicated that the resistance in the EC system had decreased.

During the initial EC treatment period pitting of the electrode surface occurs [56] and this

process may have influenced the behaviour noted in Figure 3. As the salinity of the CS water

increased the voltage required exhibited a substantial decrease from an average of 35.26 (low

TDS CS water) to 17.14 V (high TDS CS water) [Table 3].

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Low TDS CS Water

Medium TDS CS Water

High TDS CS Water

Figure 3: Variation of effluent solution pH and voltage during electrocoagulation with

aluminium electrodes of various CS water compositions: flow rate = 1.08 L/min; polarity

reversal period 30 s; hydraulic retention time 30 s; current 5 A

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Power and electrode consumption per L of water treated are important criteria in terms of

the economics of electrocoagulation. Factors which impact these latter parameters include

solution pH and conductivity [35]. For aluminium electrodes the mass loss ranged from 0.504

to 0.325 kg/kL of CS water treated, with less mass lost as the solution concentration increased.

Demirci et al. [35] noted that approximately 0.5 kg of aluminium electrode was consumed per

kL of textile wastewater remediated with electrocoagulation, which was similar to the values

found when pre-treating CS water in this study. Similarly, Kobya and Demirbas [57] found

that aluminium electrode consumption was 0.215 to 0.247 kg/kL when they applied

electrocoagulation to treat can manufacturing wastewater. Whereas, El-Ashtoukhy et al. [58]

reported that the electrode consumption when treating petrochemical wastewater using EC

could vary by as much as ca. 400 % depending upon operating conditions. These studies along

with our investigation illustrate the dependence of electrode consumption during EC to the

precise operating conditions and solution composition. The need to carefully evaluate a

range of water compositions in relation to the problem to be solved when using EC is

emphasised. Hence, why we have evaluated multiple coal seam water samples encompassing

the generic range of water types found in Queensland.

The theoretical aluminium or iron loss was calculated by application of the Faraday expression

[Equation 2]. It was deduced that with aluminium electrodes the disparity between actual

and theoretical mass electrode mass loss was greater as the solution dissolved solids content

decreased. In all cases the quantity of material removed from the electrode during the

electrocoagulation tests was in excess of the amount predicted from the Faraday model which

was consistent with previous studies [30, 31]. Notably, the aluminium removed from the

electrodes was greater for the low TDS CS water (0.544 g/min) than the high TDS CS water

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(0.350 g/min). Mouedhen et al. [56] investigated the influence of sodium chloride

concentration upon aluminium electrode performance during electrocoagulation of mixed

sodium sulphate/sodium chloride solutions. Increasing the sodium chloride concentration

was found to reduce the anodic potential and promote the formation of pits on the electrode

surface due to reaction of the passive oxide layer with chloride ions. However, the size and

distribution of the pits was dependent upon NaCl concentration. We note that with the high

TDS CS water the process was practically an electrochemical process only with minimal

evidence for alternate means for removing electrode material such as chemical dissolution.

This observation was somewhat surprising based upon the reported impact of chloride ions

from salt solutions increasing the consumption of aluminium electrodes during

electrocoagulation [30]. In addition, the presence of sodium bicarbonate/carbonate in

solution was expected to passivate the electrodes, thus inhibiting dissolution of aluminium

[59].

Although the low TDS CS water purification was the most successful in terms of the degree of

removal of dissolved contaminants [Table 2] it came at a cost. Namely, the power

consumption (2.72 kWh/kL) was significantly higher than observed for the high TDS CS water

(1.32 kWh/kL) as an example. In addition, as previously noted the consumption of electrode

material was also higher (0.504 kg compared to 0.325 kg/kL, respectively). Therefore, the

combined cost of electrode and electricity consumption was estimated as A$4.32, 3.08 & 2.70

per kL for the low, medium & high TDS CS water samples, respectively. Of note was the fact

that electrode consumption was the major cost relating to CS water treatment when using

aluminium electrodes. The value for electrode consumption was also seen as an

underestimate as practically achieving 100 % electrode use is not possible due to electrode

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wear leading to problems with structural integrity. As outlined by Tsioptsias et al. [60] in a

pilot plant or full scale EC system the electrodes will not be consumed fully and thus the

replacement cost for electrodes will be higher than the consumption values estimated here.

Labour and maintenance costs need to be considered and in addition, sludge handling and

dewatering costs are potentially substantial. In line with the study of Cesar Lopes Geraldino

et al. [61] it was observed that the electrodes wore more rapidly not only for electrodes at

the extremity of EC unit but also in the mid-section of each individual electrode relative to

the top and bottom sections of the vertical plates.

3.2.2 Iron Electrodes

Similar to the case where aluminium electrodes were studied, plots of pH variation in the

effluent stream and EC voltage are shown in Figure 4. The effluent pH again oscillated for

each CS water sample tested and as for the aluminium plates the general trend was for the

magnitude of the oscillation to be greater as the salinity of the CS water increased. However,

the pH and voltage profiles were more complex when using iron electrodes compared to the

aluminium plates. For the low TDS CS water the voltage decreased over the first 450 s of EC

treatment before plateauing and then gradually rising in value especially after 1000 s of

treatment time. A decrease in voltage suggested that the resistance of the EC unit may have

decreased which is usually assigned to de-passivation of the electrode surface [62]. The

relatively low ratio of chloride to bicarbonate/carbonate ions for the low TDS sample may

have promoted passivation of the electrode surface by various carbonate containing species

such as green rust during the latter stages of the treatment time [63]. In contrast, for the

medium and high TDS CS water the voltage profiles were similar with the voltage increasing

overall with EC treatment time and ultimately exhibiting a reasonably stable plateau in

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voltage value. Therefore, it could be inferred that in these instances the electrodes probably

passivated to some degree which would have increased cell resistance.

Low TDS CS Water

Medium TDS CS Water

High TDS CS Water

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Figure 4: Variation of effluent pH and voltage during electrocoagulation with iron electrodes

of various CS water compositions: flow rate = 1.08 L/min; polarity reversal period 30 s;

hydraulic retention time 30 s; current 5 A

It was apparent that the higher the concentration of the coal seam water sample the lower

the specific power consumption was in terms of kWh/kL of water treated [Table 4]. Brahmi

et al. [64] found that increasing the conductivity of mining wastewater with sodium chloride

addition decreased the energy consumption in the EC process using aluminium electrodes.

However, they also added the caveat that high concentrations of sodium chloride in solution

could accelerate electrode use due to chemical dissolution.

With iron electrodes the amount of iron removed from the electrodes was relatively constant

and not as impacted by the TDS of the CS water sample compared to the situation with

aluminium electrodes. An unequivocal reason for this observed behaviour with the iron

electrodes in unclear. However, there is no doubt that the interaction of mixtures of chloride

and bicarbonate/carbonate ions is relatively complex based upon previously published

studies of electrochemical corrosion of iron [65-67]. Power consumption with iron electrodes

was comparable to that measured for the aluminium electrodes, but electrode consumption

was substantially greater [c.f. Table 3]. The combined cost of electrode and electricity

consumption when using iron electrodes was estimated as A$2.68, 2.53 & 2.50 per kL for the

low, medium & high TDS CS water samples, respectively. For each water type, use of iron

electrodes was less expensive compared to electrocoagulation using aluminium electrodes.

3.3 Examination of Floc Behaviour as a Function of CS Water Composition

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The mass of the dried flocs produced in each test was measured and it was found that the

mass of floc produced using aluminium electrodes [Table 3] was significantly greater than that

when iron electrodes were used [Table 4]. The behaviour and nature of the flocs formed

during an electrocoagulation process is critically important to the practical use of EC, yet floc

settling properties, mass produced and water content are not routinely reported. For

industrial applicability, it is desirable for flocs created in a coagulation process to settle within

60 min [68]. Consequently, the flocs formed during the EC experiments to treat the three CS

water samples were allowed to settle in a 2 L measuring cylinder following agitation [Figure

5].

Aluminium Electrodes Iron Electrodes

Figure 5: Settled volume as a function of time of flocs formed by electrocoagulation of low,

medium and high TDS CS water samples

As a general observation, the flocs formed from iron electrode use were discovered to settle

notably faster than those created when aluminium electrodes were used. In addition, the

flocs were visually like cotton wool and gel-like in character. Mahesh et al. [69] studied the

floc settling behaviour after treatment of pulp and paper mill effluent with EC equipped with

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iron electrodes. These authors also noted that the flocs were like “fluffy bubble-like jelly” and

that upon settling the sludge volume occupied between 23.5 and 60 % of the column

depending upon EC treatment time; which were in line with our observation that the settled

floc volume was significant.

Hu et al. [70] studied the influence of solution pH upon the growth characteristics of flocs

produced when using EC equipped with aluminium electrodes to treat water containing

humic acid. It was reported that the flocs formed during EC treatment were comparatively

fragile and porous in character relative to flocs produced using chemical coagulation. It was

also noted that EC produced flocs over a wider pH range than with conventional coagulation

using aluminium sulphate. At the high pH values noted in this study, the presence of Al(OH)3

and larger, polymeric aluminium hydroxide species was expected, in contrast to monomeric

and oligomeric aluminium species formed at lower pH values [71]. Gomes et al. [72] studied

flocs formed when using either aluminium, iron or aluminium-iron electrodes during

electrocoagulation of truck wash water. X-ray diffraction suggested that the flocs were mainly

amorphous in character and infrared spectroscopy indicated the formation of iron/aluminium

hydroxides and oxyhydroxides. Kim et al. [73] further added that the structure of the flocs

produced using EC depended upon the precise composition of the water treated. Lee and

Gagnon [74] investigated the evolution and character of flocs produced using iron electrodes

in an EC unit using water containing only dissolved sodium and calcium chlorides.

Transmission Electron Microscopy (TEM) revealed the flocs to be not only amorphous but also

fractal in nature. Floc growth reached a steady state condition usually within ca. 10 min of

the commencement of the EC test, which corresponded to compaction of initial porous flocs.

Interestingly, the flocs formed were relatively insensitive to the electrolyte strength which

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was in harmony with the settling data shown in Figure 6 wherein the iron flocs settled at

approximately the same rate.

3.4 Characterization of Electrodes and Flocs

3.4.1 Electrodes

Figure 6 presents electron microscopy images of the electrode surfaces after treatment of

CSG associated water comprised of varying levels of salinity. The pitting corrosion of the

electrode surface of aluminium exhibited different surface structure depending upon the

salinity of the CSG associated water treated. It was observed that the pits became finer in

appearance as the salinity increased which was in accord with the results of Mouedhen et al.

[56] when these authors examined micrographs of aluminium plates after EC treatment of

sodium sulphate/sodium chloride solutions.

High TDS CS Water

Aluminium

Iron

Medium TDS CS Water

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Aluminium

Iron

Low TDS CS Water

Aluminium

Iron

Figure 6: SEM images of aluminium and iron electrodes after treating CSG water

The corresponding SEM images of resin mounted iron electrodes presented information

about the depth and distribution of the pits. The higher concentration CSG associated water

appeared to have created more pits than the lower salinity samples albeit it was not

unequivocal if the pits were deeper into the electrode.

3.4.2 Floc Characteristics

Quantitative XRD analysis of the flocs formed during the electrocoagulation of CS water

revealed that in all cases the flocs were predominantly composed of amorphous material

[Table 5]. Gamage and Chellam [75] concluded from XRD analysis of aluminium based flocs

resultant from EC treatment of surface water that the majority of the material was

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amorphous Al(OH)3 in harmony with our characterization data. The presence of AlO(OH) was

tentatively suggested to also be present but definitive identification of this latter species was

not possible. With the aluminium electrodes there was evidence for significant quantities of

gibbsite (Al(OH)3) present and smaller amounts of calcite (CaCO3) and quartz (SiO2) [Table 5].

Notably the medium TDS CS water treatment provided flocs with the highest percentage of

crystalline material but no further correlation between CS water concentration and

composition, with degree and nature of crystalline products could be discerned. Under

alkaline conditions the formation of Al(OH)3 has been described by Malakootian and Yousefi

[46] to occur as a result of Equation 5.

Equation 5: 𝟐 𝐀𝐥 + 𝟔 𝐇𝟐𝐎 ↔ 𝟐 𝐀𝐥(𝐎𝐇)𝟑 + 𝟑 𝐇𝟐

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Digestion of the flocs was also performed in order to provide further detail about their

composition [Table 6]. As expected the major element present was aluminium from electrode

dissolution. The presence of sodium was consistent with residual sodium chloride trapped in

the floc material and the noted increase in the presence of sodium ions with increasing CSG

associated water salinity was in harmony with this latter conclusion. A similar rationale was

appropriate for the presence of potassium in the floc. Trends observed for the aluminium

based flocs were an enhancement in calcium, magnesium, and boron relative content as the

CSG associated water concentration increased. Species such as strontium, barium and iron

all peaked with medium TDS CS water. In contrast, the presence of silica in the sample was

practically constant regardless of the CSG associated water salinity.

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For the EC tests with iron electrodes [Table 7] in general the presence of amorphous material

was notably higher than the comparable analysis of aluminium based flocs [Table 5].

Maldonado-Reyes et al. [76] also observed that flocs produced from EC with iron electrode

treatment of arsenic containing solutions resulted in highly amorphous hydrous ferric oxides.

Crystalline goethite (FeO(OH)) and magnetite (Fe3O4) were particularly evident in the flocs

produced when treating the low TDS CS water sample, of lesser importance for the medium

TDS CS water and non-existent for the high TDS CS water [Table 7]. However, for the flocs

obtained from EC treatment of the high TDS CS water the presence of calcite and quartz was

recorded. Parga et al. [77] conducted XRD analysis of flocs generated when using an EC unit

equipped with iron electrodes to treat solutions comprising of dissolved strontium hydroxide

and sodium chloride.

The presence of magnetite was discerned and the formation of this material was ascribed to

the process shown in Equation 6.

Equation 6: 𝟐 𝐅𝐞(𝐎𝐇)𝟑 + 𝐅𝐞(𝐎𝐇)𝟐 ↔ 𝐅𝐞𝟐𝐎𝟑 + 𝟒 𝐇𝟐𝐎

Moreno et al. [78] noted the formation of magnetite and goethite during electrocoagulation

treatment of a variety of aqueous solutions of dissolved salts when using iron electrodes. The

composition of the solution controlled the amount and identity of the iron containing

minerals, which was in harmony with our data [Table 7]. Dubrawski et al. [79] corroborated

and extended the findings of Moreno et al. [78] by completing a detailed evaluation of the

transformation mechanisms of materials generated using electrocoagulation with iron

electrodes. In the presence of dissolved oxygen FeO(OH) was often identified as the final

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product of the EC process, with lepidocrocite and goethite prominence dependent upon the

aging time of the material. Parga et al. [77] indicated that iron oxyhydroxide species were

formed by means of Equation 7.

Equation 7: 𝐅𝐞(𝐎𝐇)𝟑 ↔ 𝐅𝐞𝐎(𝐎𝐇) + 𝐇𝟐𝐎

When carbonate species were in solution the formation of double-layered green rust was also

found to be possible. Although we did not unequivocally characterize the presence of this

latter species the visual appearance of dark green material in the floc supported the presence

of green rust which may have been amorphous. It is also noted that Dubrawski et al. [79]

emphasised that in aerobic solutions containing dissolved carbonate the green rust was not

stable and transformed to other species upon aging, with the precise range of products

depending upon test conditions.

To probe the identity of the amorphous material in the floc samples, the material collected

was digested and subsequently analysed to obtain information regarding the proportion of

elements present [Table 8]. As expected, the dominant species was iron with typically < 2 %

of the floc sample comprised of the listed species.

For the iron based flocs the trends noted included growth in the presence of sodium,

magnesium, calcium, strontium, barium, and boron as the concentration of the CS water

samples increased [Table 8]. Whereas, the quantity of silica detected decreased with

increasing CS water salinity.

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Conclusions

Electrocoagulation has been proven to be able to remove a variety of contaminants from

simulated coal seam water samples such as calcium, barium, strontium, magnesium, boron,

and silica. In general, aluminium electrodes appeared to be more efficient than iron

electrodes, albeit with aluminium electrodes the presence of residual aluminium was found

which may be problematic for downstream technologies such as reverse osmosis. Dissolved

silicates were particularly well removed with either electrode type and thus

electrocoagulation may of interest to coal seam gas producers.

The ability of electrocoagulation to remove contaminants from CS water was however

influenced by the composition and salinity of the solution. As a rule, the greater the

concentration of the CS water the lower removal of contaminants recorded; apart from

dissolved silicates which appeared highly removed in all instances. This latter observation

indicated the removal mechanism for dissolved silicates was not the same as for alkaline earth

ions and boron.

The settling properties and water content of the produced flocs was an important parameter

which respect to implementation of an EC system. The fact that iron based flocs settled to

significantly smaller volumes than aluminium based material may be important in terms of

clarifier operation. The flocs formed by EC treatment of CS water were voluminous and

delicate with presumably a relatively high water content.

A major question to answer is the choice of iron or aluminium electrodes for

electrocoagulation of coal seam gas associated water. Based simply on economics, the

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combined cost of electrode materials and electricity was estimated to be less for iron

electrodes than aluminium electrodes across the entire range of water salinities tested.

However, the process selection will ultimately depend upon the overall reduction in costs for

the entire water treatment process. It may be the case that high removals of all contaminants

which can cause scaling of downstream equipment is necessary.

Future work should extend these studies to the study of the mechanism for the removal of

the outlined contaminants by investigating simpler mixtures of alkaline earth ions, alkalinity

and dissolved silicates; thus determining the key factors which control the degree of pollutant

reduction. Furthermore, addition of materials causing solution turbidity, dissolved organic

species and/or algae which may be present in actual CS water samples from the field is of

interest to determine if they can not only be removed by EC but also if they impact the

removal of other species present.

Acknowledgements

We thank Dr. Chris East for help with Scanning Electron Microscopy characterization of

electrodes and Dr. Henry Spratt for collection and interpretation of XRD patterns.

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[61] H. Cesar Lopes Geraldino, J. Izabelle Simionato, T. Karoliny Formicoli de Souza Freitas, J. Carla Garcia, N. Evelázio de Souza, Evaluation of the electrode wear and the residual concentration of iron in a system of electrocoagulation, Desalination and Water Treatment, 57 (2016) 13377-13387. [62] Z.H. Yang, H.Y. Xu, G.M. Zeng, Y.L. Luo, X. Yang, J. Huang, L.K. Wang, P.P. Song, The behavior of dissolution/passivation and the transformation of passive films during electrocoagulation: Influences of initial pH, Cr(VI) concentration, and alternating pulsed current, Electrochimica Acta, 153 (2015) 149-158. [63] J. Gomes, D. Cocke, H. Moreno, D. Tran, A. Mahmud, K. Das, M. Guttula, Green rust: Its electrochemical generation, characterization, and implications, in: TMS Annual Meeting, EPD Congress, 2009, pp. 211-218. [64] K. Brahmi, W. Bouguerra, K. Missaoui, Z. Tlili, E. Elaloui, M. Loungou, B. Hamrouni, Highly Cost-Effective and Reuse-Oriented Treatment of Cadmium-Polluted Mining Wastewater by Electrocoagulation Process, Journal of Environmental Engineering (United States), 142 (2016). [65] F.F. Eliyan, A. Alfantazi, Corrosion mechanisms of iron and pipeline steels in bicarbonate and carbonate solutions - A review, in: EUROCORR 2013 - European Corrosion Congress, 2013. [66] F.F. Eliyan, A. Alfantazi, Mechanisms of corrosion and electrochemical significance of metallurgy and environment with corrosion of iron and steel in bicarbonate and carbonate solutions - A review, Corrosion, 70 (2014) 880-898. [67] C.T. Lee, M.S. Odziemkowski, D.W. Shoesmith, An in situ Raman-electrochemical investigation of carbon steel corrosion in Na<inf>2</inf>CO<inf>3</inf> NaHCO<inf>3</inf>, Na<inf>2</inf>SO <inf>4</inf>, and NaCl solutions, Journal of the Electrochemical Society, 153 (2006) B33-B41. [68] T.G. Milligan, An examination of the settling behaviour of a flocculated suspension, Netherlands Journal of Sea Research, 33 (1995) 163-171. [69] S. Mahesh, K.K. Garg, V.C. Srivastava, I.M. Mishra, B. Prasad, I.D. Mall, Continuous electrocoagulation treatment of pulp and paper mill wastewater: Operating cost and sludge study, RSC Advances, 6 (2016) 16223-16233. [70] C. Hu, J. Sun, S. Wang, R. Liu, H. Liu, J. Qu, Enhanced efficiency in HA removal by electrocoagulation through optimizing flocs properties: Role of current density and pH, Separation and Purification Technology, 175 (2017) 248-254. [71] C. Hu, S. Wang, J. Sun, H. Liu, J. Qu, An effective method for improving electrocoagulation process: Optimization of Al13 polymer formation, Colloids and Surfaces A: Physicochemical and Engineering Aspects, 489 (2016) 234-240. [72] A.J. Gomes, K.K. Das, S.A. Jame, D.L. Cocke, Treatment of truck wash water using electrocoagulation, Desalination and Water Treatment, 57 (2016) 25991-26002. [73] D.G. Kim, R.J.S. Palacios, S.O. Ko, Characterization of sludge generated by electrocoagulation for the removal of heavy metals, Desalination and Water Treatment, 52 (2014) 909-919. [74] S.Y. Lee, G.A. Gagnon, Growth and structure of flocs following electrocoagulation, Separation and Purification Technology, 163 (2016) 162-168. [75] N.P. Gamage, S. Chellam, Aluminum electrocoagulation pretreatment reduces fouling during surface water microfiltration, Journal of Membrane Science, 379 (2011) 97-105. [76] A. Maldonado-Reyes, C. Montero-Ocampo, J. Medina-Garcia, S. Bolado-Rodríguez, J. Alvárez-Benedí, A. Herrera-Vazquez, V.M. Castaño, Electro Coagulation Removal of As from Water: the Role of Phases Formation, Water, Air, and Soil Pollution, 226 (2015) 1-6. [77] J.R. Parga, G. González, H. Moreno, J.L. Valenzuela, Thermodynamic studies of the strontium adsorption on iron species generated by electrocoagulation, Desalination and Water Treatment, 37 (2012) 244-252. [78] H.A.C. Moreno, D.L. Cocke, J.A.G. Gomes, P. Morkovsky, J.R. Parga, E. Peterson, C. Garcia, Electrochemical generation of green rust using electrocoagulation, in: ECS Transactions, 2007, pp. 67-78.

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[79] K.L. Dubrawski, C.M. Van Genuchten, C. Delaire, S.E. Amrose, A.J. Gadgil, M. Mohseni, Production and transformation of mixed-valent nanoparticles generated by Fe(0) electrocoagulation, Environmental Science and Technology, 49 (2015) 2171-2179.

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Table 1: Composition of simulated CS water solutions used for electrocoagulation tests

Value

Species Low TDS CS

Water

Medium TDS

CS Water

High TDS CS

Water units

Sodium 1060 2260 3698 mg/L

Potassium 6.23 10.90 21.67 mg/L

Calcium 1.933 5.50 78.16 mg/L

Magnesium 2.447 10.40 33.14 mg/L

Barium 0.658 2.30 7.97 mg/L

Strontium 1.123 4.39 17.49 mg/L

Iron 0 5.37 0 mg/L

Dissolved silica 15.61 11.31 12.73 mg/L

Boron 0.482 11.03 17.09 mg/L

Chloride 1040.06 3240.67 5910.34 mg/L

Alkalinity as CaCO3 980.62 524.52 558.72 mg/L

Solution pH 8.32 8.61 8.21

Solution

Conductivity 5290 9550 15680 µS/cm

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Table 2: Summary of Removal of Dissolved Species by Electrocoagulation using Aluminium

and Iron Electrodes

Aluminium Electrodes Iron Electrodes

Low TDS CS Water Low TDS CS Water

Species Initial

Conc.

(mg/L)

Final Conc.

(mg/L)

Removal

Efficiency

(%)

Initial

Conc.

(mg/L)

Final Conc.

(mg/L)

Removal

Efficiency

(%)

Mg 2.45 0.60 75.5 1.41 0.21 85.1

Ca 1.94 0.47 75.8 1.47 0.67 54.4

Sr 1.12 0.01 99.1 1.32 0.25 81.1

Ba 0.66 0.00 100.0 0.74 0.024 96.8

B 0.48 0.39 18.8 0.72 0.58 19.4

Si 15.61 0.37 97.6 18.13 0.918 94.9

Medium TDS CS Water Medium TDS CS Water

Mg 10.40 1.70 83.6 10.32 5.72 44.6

Ca 5.50 1.44 73.9 7.631 5.73 24.9

Sr 4.39 0.59 86.6 4.65 3.37 27.5

Ba 2.30 0.14 93.8 2.517 0.67 73.4

B 11.03 9.55 13.5 11.7 10.87 7.1

Si 11.31 0.32 97.2 11.98 1.25 89.5

High TDS CS Water High TDS CS Water

Mg 33.14 18.01 45.7 27.99 11.96 57.27

Ca 78.16 64.75 17.2 36.26 22.73 37.31

Sr 17.49 16.03 8.3 13.65 9.92 27.33

Ba 7.97 6.12 23.2 7.75 2.09 73.09

B 17.09 13.95 18.4 14.95 13.46 9.77

Si 12.73 0.25 98.0 12.56 0.63 95.00

Table 3: Summary of voltage, current and power consumption when treating CS water by

electrocoagulation using aluminium electrodes

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Low TDS CS

Water

Medium TDS

CS Water

High TDS CS

Water

Average EC Voltage 35.26 19.97 17.14

Number of Electrode Plates 13 13 13

Average EC Voltage per Plate 2.94 1.66 1.43

Electrode Surface Area (cm2) 1800 1800 1800

Average Current (A) 5.00 4.56 4.99

Average Current Density (mA/cm2) 2.78 2.53 2.77

Treatment Time (min) 37.72 32.00 32.92

Total Volume of Water Treated (L) 40.73 34.56 35.55

Specific Power Consumption (kWh/kL) 2.72 1.40 1.32

Total Electrode Mass Loss (g) 20.51 12.88 11.54

Aluminium Theoretical Loss (g) 12.67 9.78 11.02

Aluminium Loss per Minute (g/min) 0.544 0.402 0.350

Mass loss per kL of water treated (kg/kL) 0.504 0.373 0.325

Mass of Dried Flocculent from 2L Sample

of Treated Solution (g)

3.49 3.52 3.62

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Table 4: Summary of voltage, current and power consumption when treating CS water by

electrocoagulation using iron electrodes

Low TDS CS

Water

Medium TDS

CS Water

High TDS CS

Water

Average EC Voltage 32.07 22.00 18.84

Number of Electrode Plates 13 13 13

Average EC Voltage per plate 2.67 1.83 1.57

Total Active Electrode Surfaces 12 12 12

Electrode Surface Area (cm2) 1800 1800 1800

Average Current (A) 5.04 5.12 5.29

Average Current density (mA/cm2) 2.80 2.84 2.94

Treatment Time (min) 30.60 32.67 32.20

Total Volume of Water Treated (L) 33.05 35.28 34.78

Specific Power Consumption (kWh/kL) 2.49 1.74 1.54

Total Electrode Mass Loss (g) 28.85 30.85 30.49

Iron Theoretical Loss (g) 21.40 23.24 23.69

Iron Loss per Minute (g/min) 0.943 0.944 0.947

Mass loss per kL of water treated (kg/kL) 0.873 0.874 0.877

Mass of Dried Flocculent from 2L Sample

of Treated Solution (g)

1.81 1.80 1.89

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Table 5: Quantitative XRD analysis of flocs produced after EC of CS water with aluminium

electrodes

Low TDS CS Water

Medium TDS CS

Water High TDS CS Water

Quartz, SiO2 0.6 0.1 0.5

Calcite, CaCO3 1.5 0.2

Gibbsite, Al(OH)3 17.5 45.5 15.8

Halite, NaCl 0.1

Non-diffracting

("Amorphous") 81.9 52.9 83.4

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Table 6: Digestion analysis of flocs produced after EC of CS water with aluminium electrodes

Solution Concentration (mg/L)

Element Low TDS CS Water Medium TDS CS

Water

High TDS CS Water

Na 313.8 419.9 566.2

K 2.6 2.8 3.7

Mg 89.4 175.7 276.9

Ca 70.4 295.5 301.6

Sr 18.8 68.0 38.3

Ba 9.6 31.9 24.5

Fe 29.3 107.1 12.0

Al 4460 4368 4092

B 1.4 28.8 58.5

S 1.3 2.4 15.8

Si 174.2 170.8 131.2

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Table 7: Quantitative XRD analysis of flocs produced after EC of CS water with iron electrodes

Low TDS CS Water

Medium TDS CS

Water High TDS CS Water

Quartz, SiO2 0.4

Magnetite, Fe3O4 5.5

Calcite, CaCO3 8.2

Goethite, FeO(OH) 17.8 3

Halite, NaCl 0.3

Non-

diffracting/unidentified

("Amorphous")

76.7 97 91.1

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Table 8: Digestion analysis of flocs produced after EC of CS water with iron electrodes

Solution Concentration (mg/L)

Element Low TDS CS Water Medium TDS CS Water High TDS CS Water

Na 110.1 189.7 301.9

K 0.0 0.0 0.8

Mg 15.4 83.1 186.1

Ca 45.3 71.8 778.7

Sr 16.9 29.5 40.2

Ba 13.6 40.0 69.0

Fe 10260 9699 8228

Al 10.4 8.4 4.4

B 0.0 9.0 13.0

S 0.0 0.0 0.1

Si 329.0 235.2 178.3

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