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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
To appear in:
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 [email protected]
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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[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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