Removal of silica from brackish water by electrocoagulation pretreatment to prevent fouling of reverse osmosis membranes

Removal of silica from brackish water by electrocoagulation pretreatment to prevent fouling of reverse osmosis membranes

Available online at Separation and Purification Technology 59 (2008) 318–325 Removal of silica from brackish water by electroc...

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Available online at

Separation and Purification Technology 59 (2008) 318–325

Removal of silica from brackish water by electrocoagulation pretreatment to prevent fouling of reverse osmosis membranes Walter Den ∗ , Chia-Jung Wang Department of Environmental Science and Engineering, Tunghai University, Taichung-Kan Road, Section 3, Number 181, Taichung 407, Taiwan, ROC Received 5 December 2006; received in revised form 25 June 2007; accepted 5 July 2007

Abstract Desalination of seawater and brackish water by reverse osmosis (RO) has become increasingly important for drinking water supply in a greater part of the world. The presence of high silica concentrations in some brackish water, however, limits the application of RO desalination due to the potential formation of silica scales that irreversibly deteriorate the membrane material and performance. This study investigates the feasibility of electrocoagulation as a pretreatment process to remove silica from the source brackish water. The effects of several electrical parameters, including electrode arrangement, current intensity and hydraulic retention time, were studied on the basis of silica removal efficiency. Bipolar configuration attained greater extent of silica removal as compared to monopolar configuration. Increases in charge loading generally improved the silica removal efficiency, but excessive hydraulic retention time (60 min) was detrimental to the system performance. In this study, with no modification of the source water, silica removal efficiency up to 80% was achieved under a current intensity of 0.5 A and a hydraulic retention time of 30 min. The subsequent nanofiltration studies demonstrated severe flux declines over the first 3 h, yielding only 70% of its initial flux for brackish water containing 100 mg/L silica, and progressively lower with higher silica concentrations. For the pretreated water by electrocogulation, the extent of flux decline was markedly improved, suggesting that the pretreatment was effective for the attenuation of the flux decline. Electron micrograph images of the membrane autopsy also confirmed the lack of scale formation for the pretreated water as compared to those without pretreatment. © 2007 Elsevier B.V. All rights reserved. Keywords: Electrocoagulation; Reverse osmosis; Membrane fouling; Silica; Brackish water; Desalination

1. Introduction Desalination of seawater and brackish water has become increasingly important as a source of water supply to contend with the population burden worldwide. Pressure-driven filtration processes such as reverse osmosis (RO) and to a lesser extent, nanofiltration (NF), has dominated the desalination market because of its technical maturation that promises high rejection rate of salts and inorganics with high production rate at reasonable costs. However, like other membrane processes, RO and NF are prone to rapid flux decline due to fouling problems, among which the most difficult ones to deal with are those caused by silica scaling [1,2]. The problem associated with silica scaling occurred not only when the source water contains high silica concentration, but also to the back-end of a multiple-stage

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1383-5866/$ – see front matter © 2007 Elsevier B.V. All rights reserved. doi:10.1016/j.seppur.2007.07.025

design of RO process even with low silica concentrations in the source water. Although the exact form of dissolved silica in water depends on its concentration in solution, silica scaling on membranes can conceivably occur either by monomeric silica (silicic acid, Si(OH)4(aq) ) deposition on membranes followed by polymerization, or by accumulation of silica colloids (SiO2(s) ) on the membrane surfaces. Approximating the solubility of amorphous silica in water to be 120 mg/L, the former scenario would likely occur for water with low silica content, whereas the latter scenario would likely occur for high silica content. Therefore, for desalination of brackish water with high silica content, a control strategy must be implemented to maintain the service life of the membranes. Prevention of silica scaling commonly follows two distinct approaches, namely the addition of an anti-scalant to inhibit scale formation, and the installment of a pretreatment process to reduce the silica concentration before membrane filtration. Scale inhibition relies on the promoting electrostatic and steric

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Table 1 Characteristics of the brackish water, RO reject water, and artificial water samples Property

Feed brackish water

First-stage RO rejecta

Artificial feed water

pH Conductivity (mS/cm) Zeta potential (mV) Total SiO2 concentration (mg/L) Dissolved SiO2 concentration (mg/L) Background matrix (mg/L)

7.9 2.3 −10.8 50–60 28 [Ca2+ ] = 28; [Mg2+ ] = 38; [Na+ ] = 130; [Cl− ] = 192; [SO2− ] = 161

6.5–7.5 5–5.3 −2.5 to −6.5 83–130 73–119 [Ca2+ ] = 164; [Mg2+ ] = 130; [Na+ ] = 560; [Cl− ] = 192; [SO2− ] = 724

7.0 5.0 −20 to −30 120 120 [Ca2+ ] = 200; [Mg2+ ] = 100


This is the reject at the first of two-stage RO, each operating at 75% recovery rate.

repulsion between the charged colloids and the membrane surface as well as between the colloids themselves. Commercially available anti-scalants and anti-foulants have claimed successful results in preventing silica scaling [3,4], but their proprietary status and specificity (e.g., choice of an agent is highly dependent of the membrane material and configuration) makes this approach somewhat precarious. Alternatively, several methods probing the reduction of silica content in the source water have been investigated, including chemical precipitation [5,6] and silica-gel seeding [7]. Chemical precipitation most commonly doses a combination of lime, soda ash, or caustic soda into solutions, sometimes with the addition of a precipitant aid (alum and ferric chloride) to promote coagulation and precipitation. The optimum dosage depends on the silica concentration and the background matrix of the solution, as well as the required extent of silica reduction. Pretreatment by silica-gel seeding intends to provide a seeding surface on which silica monomers deposit, thereby removing dissolved silica from solution. However, this method was ineffective when silica polymers formed in the solution under highly supersaturated conditions. In this study, a pretreatment method involving electrocoagulation process is proposed and investigated. Electrocoagulation has been investigated as a treatment technology for a variety of wastewaters containing finely dispersed particles [8–14] and metal ions [15,16]. In this process, the sacrificial anodic electrodes, commonly consisted of iron and aluminum, are used to continuously supply metallic ions as the source of coagulants. These electrochemically generated metallic ions can hydrolyze near the anode to form a series of metallic hydroxides capable of destabilizing dispersed particles. The simultaneous electrophoretic migration of the negatively charged ions, colloids, or particles toward the anodic surfaces forces chemical coagulation between particles and metallic hydroxides in the vicinity of the anode, forming flocs that either settles or redeposits onto the anode. Den and Huang [14] and Den et al. [17] have demonstrated that the electrocoagulation process was particularly effective for the removal of silica in its colloidal form because the colloids carry strong negative charges in neutral range. However, because silica in its reactive form is generally unionized at neutral pH range, they can only be subjected to coagulation with the metal hydroxides in the solution. In particular for the electrocoagulation with aluminum anodes, formation of stable hydroxyaluminosilicates (HAS) has been known as an

important process in the presence of silicic acids [18]: OH−


nAl(aq) 3+ ←→[Al(OH)3 ]n ←→4 (AlO)n (SiO)n/2 (OH)2n


Therefore, electrocoagulation with aluminum anodes takes advantage of fact that Al and Si(OH)4 are effective mutual scavengers in the process of removing dissolved silica from the solutions. This study intends to investigate the dissolved silica removal efficiency of the electrocoagulation process and its applicability in reducing silica scaling problem during membrane filtration. 2. Materials and methods 2.1. Water samples The water samples were original obtained from Penghu County (Taiwan), an agglomerate of isles off the west coast of Taiwan. Desalination plants using two-stage RO systems for seawater and brackish water desalination have been operative to meet the demand of drinking water supply for the islands. However, silica scaling has been identified as an operational problem that deteriorated the water production and shortened the membrane lives, particularly for the membranes on the backend of the two-stage system operating at 75% recovery rate. Therefore, samples of feed brackish water and the concentrates from the first-stage membranes were collected, and their properties are listed in Table 1. Artificial water samples simulating the feed brackish water were prepared for the laboratory experiments, including both electrocoagulation studies and membrane flux decline studies. Silica solutions were prepared by the addition of sodium metasilicate (Na2 SiO3 ·9H2 O, Hanawa, Japan). The pH and conductivity of the solution was then adjusted using HCl and NaCl, and calcium ions (100 mg/L) and magnesium ions (200 mg/L) were added to the solution such that the background matrices of the artificial brackish water were comparable with those of the feed brackish water. The surface potential of the solids in the experimental solutions was measured by a Malvern Zetasizer nanoZS instrument (Worcestershire, UK) featuring dynamic laser light scattering mechanism. The total silica and aluminum concentrations were determined using a Jobin-Yvon JY24 inductively coupled plasma-atomic emission spectrometry (ICP-AES) (Horiba Jobin-Yvon, NJ, USA) at emission


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Fig. 1. The overall experimental setup. The upper part represents the elctrocoagulation system, and the lower part is the cross-flow membrane filtration system.

wavelength of 212.412 nm and 309.271 nm, respectively. The dissolved reactive silica was determined using the molybdosilicate colorimetric method quantified by a Hach DR2400 UV–vis spectrophotometer (Hach Co., CO, USA). In this study, a term “dissolved silica” was generically used to describe reactive silica that included small molecular-weight polysilicic acids that sufficiently reacted with molybdic acid during the colorimetric analysis. 2.2. Electrocoagulation study A bench-scale acrylic continuous-flow reactor (Fig. 1) having interior dimensions of 26.5 cm × 14 cm × 17 cm (L × W × H) and channelized by the electrode plates was used to perform the electrocoagulation experiments. All electrodes were of identical size at 14 cm × 13.5 cm (W × H) with an electro-active area of 12 cm × 12 cm. The electrodes were alternately fitted into the reactor such that a completely submerged electrode was adjoined with a partially submerged electrode to form a series of vertical channels. Electrical current was provided by a manually controllable DC power supply (Model GPS, Goodwill Instruments, Taiwan). At the commencement of each experiment, the reactor was filled with brackish water such that the liquid level was approximately midway between two alternately positioned electrodes. The feed water was continuously pumped into the reactor using a microcomputer-controlled peristaltic pump (Masterflex L/S, Cole Parmer, IL) at various flow rates between 75 mL/min and 225 mL/min, corresponding to mean hydraulic retention times (HRTs) between 20 min and 60 min. A ball valve installed at the effluent port allowed the control of steady liquid level in the reactor.

The electrical connection followed either a parallel monopolar configuration or a serial bipolar configuration. The bipolar system (as shown in Fig. 1) consisted of four aluminum electrodes (an anode and three sacrificial electrodes) and a stainless steel cathode, whereas the monopolar system contained three aluminum anodes in alternate with two stainless steel cathodes. The electrocoagulation experiments used a total of five electrodes such that the inter-electrode spacing (4 cm) remained consistent between the monopolar and bipolar systems. Furthermore, the systems were operated with a constant-current mode ranging between 0.25 A and 1.0 A to give a range of current density between 8.7 A/m2 and 34.7 A/m2 for the contacting anode in the bipolar system, as well as a range between 2.9 A/m2 and 11.6 A/m2 for the three parallel anodes in the monopolar system. The responding potentials were monitored and recorded during each experimental run. Samples from the influent and effluent ports were taken and measured for dissolved silica concentration, pH and conductivity (inoLab WTW pH meter). The effects of current intensity, HRT, and charge loading were investigated based on the removal efficiency of dissolved silica. The charge loading is defined as the number of electrical charges (in Faraday) applied to the system per unit volume of treated water: charge loading(CF ) =

iτ FVr


where i is the applied current in ampere (A), τ the mean HRT in s, F the Faraday constant (96,500 C), and Vr is the working volume (m3 ) of the reactor. In addition, after the completion of each run, the sludge was carefully collected, and its dry weight was measured after oven drying at 105 ◦ C for 24 h. The current efficiency of the bipolar system was evaluated with a separate set of experiments using a fixed current intensity of 0.5 A (17.4 A/m2 ) and RO water whose conductivity was adjusted to 5 mS/cm. After each electrolysis experiment, the electrodes were removed and the solution/sludge was vigorously mixed for a few minutes. Two hundred milliliters of sample was then collected and immediately filtered through a 0.45-␮m membrane. The filtrate was then pretreated with nitric acid digestion (Standard Methods 3030E), and the retentate (solids) was pretreated with HNO3 –H2 O2 –HCl digestion (standard methods 3050B). The aluminum concentrations in the digested samples were then determined by ICP-AES. Similar procedure was also performed at the end of each actual electrocoagulation experiment to verify the current efficiency (εc ), which was defined as the ratio of the actual mass of aluminum ([Al]R ) to its theoretical mass ([Al]T ) liberated from the anodes: εc =

[Al]R [Al]T


2.3. Membrane fouling study Membrane fouling studies were investigated using both the untreated artificial brackish water samples containing different concentrations of dissolved silica, and those having been pretreated with electrocoagulation, as shown in the lower part of Fig. 1. For membrane filtration of the pretreated water, the elec-

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trocoagulation effluent was allowed to settle for 30 min, and only the supernatant was used for the filtration study. Membrane coupons were cut from a flat-sheet membrane roll and loaded into an Osmonics SEPA CFII crossflow membrane filtration unit (GE Osmonics, USA). The membrane was a commercial polyamide NF membrane (FCS TFC-2012, Rum-Tech Inc., Taiwan) having a molecular cutoff of 600–700 Da. Before use, the fresh NF membrane was conditioned with methanol and deionized water. For all filtration experiments, the SEPA test unit was operated at a constant pressure of 500 kPa and a constant volumetric flow of 0.8 L/min. The permeate flux was measured by transferring the permeate to an electronic microbalance, which recorded the accumulative mass of the permeate as a function of time. To observe the condition of the membrane surface after each fouling study, membrane autopsy tests were performed using a JEOL JSM-6700F scanning electron microscope (SEM) (JEOL Ltd., Tokyo, Japan) equipped with an Oxford INCA Energy 400 energy-dispersive spectrometer (EDS) (Oxford Instruments, Oxon, UK). 3. Results and discussion 3.1. Electrocoagulation studies The performance of the monopolar and the bipolar systems were evaluated with respect to the removal efficiency of dissolved silica. In the case of the parallel monopolar system, each individual channel (cell) resembles a complete electrical circuit, such that the assembly voltage is identical to the individual cell voltage, and the overall current is the sum of the individual cell currents. In contrast, with the serial bipolar system, electrical current flows through the outer electrodes physically connected to the power supply unit, transforming the interspersed sacrificial electrodes originally in neutral state into charged state via the conductive aqueous medium. In this case, the assembly voltage becomes the sum of the individual cell voltages. Fig. 2 shows the extents of dissolved silica removal by the monopolar system at current intensities of 0.5 A (equivalent to a current density of 5.8 A/m2 ) and 1.0 A (11.6 A/m2 ) and by the bipolar system at 0.25 A (8.7 A/m2 ), 0.5 A (17.6 A/m2 ) and 1.0 A (34.7 A/m2 ). After 60 min of electrocoagulation, the system performance with the bipolar configuration was significantly superior than that of the monopolar configuration with reference to the applied current intensity, having achieved dissolved silica removal efficiency of approximately 70% at 0.5 A and 80% at 1.0 A, as opposed to only 30% and 50% for


Fig. 2. The dissolved silica concentration profiles as a function of electrocoagulation time (solution conductivity = 5 mS/cm, pH 7.0) for monopolar and bipolar configurations at various applied current intensity (A) or current density (A/m2 ).

the monopolar system. It is also evidenced that the extent of removal was greater for the bipolar system when current density is the basis of comparison. Some studies have demonstrated that bipolar system were effective in removing fluoride ions from wastewater with aluminum electrodes [15,19] and phosphate ions with iron electrodes [20], while others have shown that monopolar systems were more effective in removing dye substance [21]. Although consensus could not be reached on whether a monopolar or a bipolar system is more effective than the other because the system performance is often dictated by the water/wastewater properties, a review of relevant literature strongly points to bipolar systems with aluminum electrodes are favorable for the removal of soluble or ionic substances from aqueous media. One of the primary reasons is that bipolar systems normally operate under Faradaic (current efficiency equal to unity) or super-Faradaic (current efficiency exceeds unity) conditions, insofar as sufficient ionic strength or conductivity is contained in the aqueous medium. Other studies have also indicated that the presence of high chloride concentration in solutions yielded Faradaic conditions due to the chloride-facilitated pitting corrosion mechanisms [15,22]. In the present study, Faradaic condition was realized (εc = 0.98–1.05) with the aluminum electrodes and aqueous conductivity equal to or greater than 5 mS/cm, thereby ensuring a surplus supply of coagulant. Despite its superior performance in the removal of dissolved silica, the bipolar system was also more energy intensive and produced greater quantity of sludge (measured in dry weight)

Table 2 Comparison of removal efficiency, energy consumption, and sludge production between monopolar and bipolar systems Current intensity (A)

0.5 1.0



SiO2 removal efficiency (%)

Energy consumption (kWh/m3 )

Sludge weight (kg/m3 )

SiO2 removal efficiency (%)

Energy consumption (kWh/m3 )

Sludge weight (kg/m3 )

28.3 48.2

0.11 0.22

0.32 0.54

69.1 80.8

1.3 4.9

0.73 1.85


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than its monopolar counterpart, as shown in Table 2. This is because the serial bipolar assembly essentially constitutes individual cells connected in series, as opposed to the assembly of monopolar configuration that forms parallel connection of cells. Consequently, a higher potential difference is required due to the higher resistance for cells connected in series in the bipolar system. Therefore, for practical evaluation of the electrocoagulation system configuration, factors other than removal efficiency must be taken into consideration. Nevertheless, these factors were not further discussed as the study focused on the technical feasibility of the electrocoagulation process. 3.2. Effect of current density and retention time Batch studies reported in the previous section clearly suggest that the bipolar system achieved greater dissolved silica removal efficiency, and thus the ensuing continuous-flow studies also followed the bipolar configuration. As shown in Fig. 3, the steady-state dissolved silica removal efficiency generally increased with the charge loading (Eq. (2)), which asymptotically approached a maximum removal efficiency of approximately 80%. Also, the removal profile of the actual brackish water generally fell into a similar trend, although the removal efficiencies corresponding to the same charge loading conditions were slightly lower than those for the synthetic brackish water. One should note that the current intensity exceeding 1.0 A (or current density of 34.7 A/m2 ) was not investigated because the energy consumption and the sludge production would expand beyond reasonable range while gaining limited increase in the removal efficiency. Also, the two data points corresponding to a charge loading of 4.1 F/m3 indicated distinctively different results (38% versus 82%), suggesting that there existed a window of parametric values within which charge loading could be used as an indicator for silica removal efficiency. Further analyses of the parametric component of charge loading indicated that the extent of dissolved silica removal continually increased with the current intensity for electrocoagulation

Fig. 3. The effect of charge loading on the removal efficiency of dissolved silica for (䊉) artificial brackish water (conductivity = 5 mS/cm, pH 7.0) and () actual brackish water (conductivity = 2.3 mS/cm, pH 7.9).

Fig. 4. The effect of (a) current intensity (at HRT = 30 min) and (b) HRT (at i = 0.5 A), on the removal efficiency of dissolved silica for artificial brackish water (conductivity = 5 mS/cm, pH 7.0).

of both the artificial and the actual brackish water (Fig. 4a). However, as shown in Fig. 4b, longer HRT did not necessarily result in improved removal efficiency. In fact, while extending HRT from 20 min to 30 min slightly improved the removal efficiency, further extending HRT to 60 min caused a significant deterioration in the system performance. This justifies the difference in the removal efficiency under the same value of charge loading (4.1 F/m3 ) by observing that the low data point corresponded to the condition with longer HRT (60 min) as compared to the high data point operated with short HRT (30 min). The abundance of aluminum hydroxides in the reactor is presumably counterbalanced by the amount of silicic acids entering the electrocoagulation reactor. Consequently, the excessively long HRTs (or low flowrates) would “stagnate” the continuous-flow process in which insufficient amount of silicic acids is available in the reactor, leading to accumulation of aged aluminum hydroxides that were adverse to the coagulation of silicic acids. Therefore, for the continuously flow aluminum bipolar system used in the study, the current intensity strongly affected the steady-state removal efficiency, whereas the HRT need to be limited within 30 min.

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During electrocoagulation of the actual brackish water, the effluent pH (6.9–7.3) and conductivity (5.0–5.2 mS/cm) did not change notably from its feed pH (6.8–7.5) and conductivity (5.0–5.3 mS/cm), indicating that the aqueous background matrix did not react with electrolyzed aluminum hydroxyl species. Exley and Birchall [18] and Strekopytov and Exley [23] have elucidated that silicic acid reacts with aluminum to form one of the two types of HAS depending on the ratio of Si(OH)4 to Al in the solution. The reaction, however, is thought to involve only the condensation of Si(OH)4 across hydroxyl groups on adjacent Al atoms which are part of an aluminum hydroxide (Al(OH)3(s) ) framework, but does not directly engage free Al ions [18]. The occurrence of this reaction obviously demands a neutral or acidic solutions (pH < 7) undersaturated with respect to Si(OH)4 such that neither polymeric nor deprotonated (SiOH)4 is significant. For the present study, Al(III) within the observed pH range of the brackish water was predominantly Al(OH)3 and Al(OH)4 − , and Si(IV) predominantly remained as undissociated Si(OH)4 . Additionally, the molar ratio between Si(OH)4 to Al was approximately 2 on the basis of feed dissolved silica concentration of 100 mg/L and an applied current intensity of 0.5 A. Therefore, formation of HAS between Al(III) and Si(OH)4 exemplified by the following known reactions may occur: 2Al(III) + 2Si(OH)4 + H2 O ↔ Al2 Si2 O5 (OH)4(s) + 6H+ (4) Al(III) + Si(OH)4 ↔ AlOSiO5 (OH)3 2+ + H+


where Al(III) indicates the Al atoms belonging to the framework of Al(OH)3 and Al(OH)4 − . The low solubility of these HAS suggests that precipitation will occur and eventually forms the sediment observed in the electrocoagulation reactor. Furthermore, the H+ released from reaction (4) and (5) recombine with the OH− generated from the cathodes, thereby keeping the pH of the solution essentially constant. The aforementioned mechanisms for the removal of dissolve silica by aluminum anodes was in stark contrast to the previous results [17] involving continuous-flow electrocoagulation with iron anodes for the separation of silica colloids, where at least two distinct mechanisms were identified, namely (i) a direct reaction (e.g. charge neutralization) between the ferrous ions liberated from the anodic sides that leads to the formation of gelatinous flocs that mostly redeposit onto the pits of the anodic surfaces through electrostatic attraction, and (ii) enmeshment of silica colloids by the hydrogenated iron hydroxyl species to form precipitating flocs. In the present study, no direct reaction occurred between silicic acids and aluminum ions, and thus the anodic aluminum surfaces remained free from any surface-floc deposition. 3.3. Membrane fouling studies The normalized flux data as a function of time for solutions containing dissolve silica concentrations of 100 mg/L, 200 mg/L, and 500 mg/L are plotted in Fig. 5. In all experimental conditions, permeate flux began to decline as soon as


Fig. 5. Flux decline profile for artificial water samples containing different dissolved silica concentrations (feed flux = 82 m3 /m2 -day, feed pressure = 500 kPa).

membrane filtration commenced. It is clearly seen that the extent of flux decline was the least severe for the solution containing 100 mg/L dissolve silica, and the flux started to stabilize after 3 h. In comparison, the solution containing 200 mg/L dissolved silica concentration experienced the most rapid initial flux decline, even more so than the solution containing 500 mg/L. The declining fluxes crossed each other after about 3.5 h, at which point the flux stabilized for the solution containing 200 mg/L dissolved silica concentration but that for the solution containing 500 mg/L continued to decline after 8 h when the experiments were terminated. The membrane autopsy tests performed with SEM examination are shown in Fig. 6. As compared with the fresh membrane, the membrane used for the filtration of 100 mg/L dissolved silica concentration mostly exhibited patches of semi-transparent deposits, likely a consequence of precipitation of monomeric silica that slowly polymerized on the membrane surface. In contrast, the membranes used for the filtration of 200 mg/L and 500 mg/L dissolved silica solutions showed large aggregates of opaque milky-to-white films that cracked during sample dehydration for the SEM analyses. These opaque films suggest that the fouling occurred by the rapid deposition of silica colloids, which presumably had been formed by polymerization in the supersaturated bulk solutions. As a result, the initial rate of flux decline was significantly greater for the membrane filtration of the supersaturated silica solutions (200 mg/L and 500 mg/L) as compared to that of the undersaturated solution (100 mg/L). In either case, formation of these silica films is difficult to remove by any of the cleaning procedures used in common practices, and thus pretreatment targeting the removal of dissolved silica becomes necessary. To verify whether pretreatment of silica-containing solution by electrocoagulation could realistically prevent or alleviate the fouling problems associated with silica precipitation, membrane filtration runs were performed for artificial solutions containing 200 mg/L of initial dissolved silica concentration pretreated with electrocoagulation process (i = 0.5 A, HRT = 30 min). The residual dissolved silica concentration in the pretreated efflu-


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Fig. 6. The SEM images of membrane autopsy for water samples containing (a) 100 mg/L, (b) 200 mg/L, (c) 500 mg/L of feed dissolved silica concentrations. Part (d) is the image for water sample (feed silica concentration = 200 mg/L) pretreated with electrocoagulation.

ent fluctuated between 39 mg/L and 66 mg/L, corresponding to approximately 70–80% removal efficiency that was consistent with the results reported earlier in Section 3.2. As shown in Fig. 7, the permeate flux for the pretreated water experienced a much lesser extent of initial decline as compared to the water sample without electrocoagulation pretreatment, and stabilized at a normalized flux of ∼0.8 during the 3 h run. Furthermore, the SEM membrane autopsy observation for the pretreated water sample (Fig. 6d) shows no visible film-like deposition during the membrane filtration study.

Residual turbidity ranging between 3 NTU and 7 NTU in the pretreatment effluents after 30 min of sedimentation indicated residual solids of HAS or aged aluminum hydroxides may still remain in the solution. To determine whether these residual flocs played any role in the permeate flux decline during the membrane study, a part of the supernatant of the pretreated solutions was pre-filtered with 0.45-␮m filter paper to remove the residual solids. The pre-filtered supernatant was then used for the membrane study under the identical conditions. As shown in Fig. 7, the permeate flux profiles for the water samples essentially overlapped, suggesting that the presence of residual flocs in the supernatant was inconsequential to extent of flux decline. 4. Conclusion

Fig. 7. Permeate flux decline data for artificial water samples (feed dissolved silica concentration of 200 mg/L) with electrocoagulation pretreatment only (), with pretreatment and pre-filtration (), and without pretreatment (䊉). The feed flux was 82 m3 /m2 -day and the feed pressure was 500 kPa.

The study investigated the removal of dissolved silica from the brackish water by electrocoagulation in an attempt to prevent silica fouling in the desalination process by reverse osmosis. The experimental results demonstrated that bipolar system with aluminum electrodes was more effective than its monopolar counterpart with respect to removal efficiency of dissolved silica from the brackish water. For initial concentrations of dissolved silica between 80 mg/L and 200 mg/L, the removal efficiency increased with charge loading up to 80%, with an optimum condition at current intensity of 0.5 A and HRT of 30 min. No discernable difference in the electrocoagulation performance was noted for the actual brackish water and the artificial brackish water, indicating that the presence of a more complex background matrix in the actual

W. Den, C.-J. Wang / Separation and Purification Technology 59 (2008) 318–325

brackish did not affect silica removal during electrocoagulation. Membrane filtration studies using a polyamide NF membrane clearly exhibited silica fouling phenomena for solutions both undersaturated (100 mg/L) and oversaturated (200 mg/L and 500 mg/L) with silica. The extent of permeate flux decline of the undersaturated solutions was expectedly less severe than those of the oversaturated solutions, which formed a layer of white opaque film as a consequence of rapid deposition of silica colloids from the bulk solutions. As a direct comparison, the flux decline was significantly reduced when the artificial brackish water was pretreated with electrocoagulation, suggesting that the pretreatment process presents a feasible alternative in dealing with brackish water containing high silica concentrations. Acknowledgements The research was financially supported through project grant NSC-93-2622-E-029-001-CC3 in collaboration with Te Cheng Sheng (TCS) Engineering Co. The authors also express their gratitude to Mr. Y.C. Lee Wang and Mr. Ke of TCS for the supply of brackish water samples and membranes. References [1] T. Koo, Y.J. Lee, R. Sheikholeslami, Silica fouling and cleaning of reverse osmosis membranes, Desalination 139 (2001) 43–56. [2] D. Lisitsin, D. Hasson, R. Semiat, Critical flux detection in a silica scaling RO system, Desalination 186 (2005) 311–318. [3] P.F. Weng, Silica scale inhibition and colloidal silica dispersion for reverse osmosis systems, Desalination 103 (1995) 59–67. [4] Z. Amjad, J.F. Zibrida, R.W. Zuhl, A new antifoulant for controlling silica fouling in reverse osmosis systems, in: Proceedings of the World Congress on Desalination and Water Reuse, Madrid, Spain, 1997. [5] R. Sheikholeslami, J. Bright, Silica and metals removal by pretreatment to prevent fouling of reverse osmosis membranes, Desalination 143 (2002) 255–267. [6] A.M. Al-Rehaili, Comparative chemical clarification for silica removal from RO groundwater feed, Desalination 159 (2003) 21–31. [7] I. Bremere, M. Kennedy, S. Mhyio, A. Jaljuli, G.-J. Witkamp, J. Schippers, Prevention of silica scale in membrane systems: removal of monomer and polymer silica, Desalination 132 (2000) 89–100.


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