1. Introduction
The sludge obtained from primary clarifiers and the excess sludge produced in biological wastewater treatment facilities require specialized additional treatment within wastewater treatment systems [1]. Typically, this waste sludge undergoes stabilization in an anaerobic digester, followed by a dewatering process. Dewatering systems are generally employed intermittently, resulting in the generation of reject water in a somewhat discontinuous fashion [2, 3]. The wastewater generated from sludge digestion and subsequent dewatering process (also known as reject water) contains high levels of ammonia nitrogen (up to 1000 mg/L), along with substantial concentrations of phosphate and COD (chemical oxygen demand) [4]. There is no separate treatment process for this reject water, and often it is returned to the inlet of the wastewater treatment plant where it causes significant increase in the influent pollutant load [5]. This can even lead to occasional overloading situations when the treatment plant operates near its designed loading capacity. Hence, side stream treatment of this excess nitrogen is necessary to keep the treatment situation under control and also to provide better treatment efficiency by reducing incoming pollutant load to the biological treatment facility.
There are many different technologies explored for the treatment of ammonia-containing wastewater, including ammonia stripping (with air/steam) [5, 6], Magnesium-Ammonia-Phosphate precipitation (MAP/CAFR process) [7], Anammox, membrane technologies [8], sequential batch reactor [9], biofilm airlift suspension reactor [6] and SHARON® process [6]. Each of these methods have their own merits and demerits and has been applied suitably to treat specific nitrogen-containing wastewater based on their qualities. Among these technologies, membrane-based technologies such as membrane bioreactor (MBR) have attained widespread acceptability based on their extensive advantages [10]. MBR allows optimum conditions for the growth of nitrifying bacteria, which are responsible for converting ammonia to nitrate [11, 12]. In addition, MBRs have lower energy requirement, smaller footprints, and reduced sludge production compared with traditional bioreactors [13]. However, despite having so many advantages, one of the major challenges associated with MBR is membrane fouling [14]. Membrane fouling refers to the accumulation of unwanted materials on the surface or within the pores of a membrane, hindering its performance in the separation process. One of the options that can help to overcome membrane fouling is the pretreatment of wastewater before utilizing MBR [15, 16].
There are a number of pretreatment methods available in the literature, among which advanced oxidation process (AOP) seems to be more advantageous. Different AOPs namely Fenton’s oxidation [17], electrooxidation [18], electro-Fenton [19], ozonation [20], and photocatalysis [21] have been previously used for wastewater pretreatment to improve treatment efficiency and reduce membrane fouling. Electrooxidation (EO) is one such AOP which involves the use of an electric current to promote oxidation reactions, leading to the breakdown or removal of organic and inorganic contaminants from wastewater [22, 23]. This process offers several advantages over other AOPs including potential for selective pollutant removal, the ability to treat a wide range of organic/inorganic pollutants, and the avoidance of costly chemical additives [24]. In addition, it is often used as part of an integrated water treatment strategy to achieve comprehensive and efficient removal of pollutants, showing its potential to be used as a pretreatment method before applying MBR [25, 26]. However, to the best of our knowledge, electrooxidation process has rarely been used previously for the treatment of reject water and condensate generated during sludge dewatering, neither individually nor in combination with MBR. Moreover, as the success of electrooxidation depends on factors such as the choice of electrode material, cell configuration, applied current density, pH and the nature of wastewater [22], it is pertinent to explore them for a given wastewater in order to achieve the best performance.
Hence, this current study focuses on exploring electrooxidation as a pretreatment for reject wastewater from sludge dewatering process prior to its treatment by MBR. Key process parameters, namely total current, current density, pH and Cl− ion concentration was optimized for the electrooxidation process. The pretreated wastewater obtained from electrooxidation process was further subjected to biological treatment using an MBR, and the effect of wastewater pretreatment on membrane fouling was studied in detail. The effect of hydraulic retention time (HRT) and electrooxidation pretreatment duration on wastewater treatment performance by MBR was also studied.
2. Materials and Methods
2.1. Electrooxidation Experiment
Electrooxidation experiment was performed in a 2000 mL glass beaker with 1800 mL working volume. The anode was made of Ti/IrO2 with a working geometric area of 50 cm2 (5 cm length × 10 cm height × 0.1 cm width). The cathode used was made of Ti/Pt with same size as the anode. The gap between the electrodes was set to be 0.5 cm. Mixing inside the electrochemical reactor was achieved with the help of a magnetic stirrer.
The wastewater was obtained from Namwon sewage treatment plant, South Korea and stored at 4°C prior to its use in the experiments. The characterization of the collected wastewater was carried out and the results obtained are presented in Table 1. The electrooxidation experiment was carried out under batch mode at pH of 7.0 and agitation of 40 RPM. In the first experiment, total applied current was varied from 0.58 A/L to 1.15, 1.173, 2.88. 3.45, 4.03 and 5.18 A/L by changing the current density from 10 mA/cm2 to 20, 30, 50, 60, 70 and 90 mA/cm2 and keeping the reaction time fixed at 2 h. The effect of current density at a fixed total current of 4.03 A/L was studied by changing it from 10 mA/cm2 to 20 30, 50, 60, 70 ad 90 mA/cm2. In order to keep the total current fixed at changing current densities, the reaction time also varied as 14, 7, 4.67, 2.8, 2.33, 2 and 1.56 h, respectively. The effect of pH was evaluated for a total of four pH values, viz. 3, 5, 7 and 9. Finally, the increment in Cl− ion by adding 500, 1000, 1500, 2500 and 4000 mg/L of Cl− on the treatment efficiency was studied. The experimental conditions for these two experiments to study the effect of pH and Cl− concentration were same, i.e., 4.03 total current, 50 mA/cm2 current density and 2.8 h of reaction time.
2.2. Membrane Bioreactor Experiment
2.2.1. Bioreactor setup
The experimental setup for biological experiment consisted of four chambers with sequentially anoxic and aerobic in nature, i.e., first and third chambers were anoxic and second and fourth chambers were aerobic (Fig. 1). The membrane bioreactor was placed in the fourth chamber. The volume of four compartments were 7.5, 25, 5.5 and 4 L, respectively. Mechanical stirrers were provided in the first and third chambers for proper mixing of the contents inside the compartments. The second and fourth tanks had a diffuser placed at the bottom of the tanks though which the tank is aerated at 1.5 L/min aeration rate.
The membrane module was of hollow fiber type and made of polyvinylidene fluoride (PVDF). The fiber had an outer diameter of 1.2 mm and average pore size of 0.1 μm. The membrane had an active surface area of 0.26 m2. The operating pressure maintained was 0.01~0.08 kPa. The membrane operating cycle had 5 min filtration time and 30 sec of backwashing. The filtration and back-washing flux were 10 and 30 LMH, respectively.
2.2.2. Bioreactor operation
The bioreactor was initially operated with a 4 d HRT using electrooxidation treated (6 h treatment) reject wastewater from sludge dewatering. A control experiment with the same HRT was also run using untreated wastewater as the control experiment to evaluate the extent of improvement in MBR performance due to electrooxidation treatment. In the following experiment, both the electrooxidation pretreatment duration and MBR’s HRT were increased to understand its effect on reactor performance. In this case the wastewater was subjected to a 12 h electrooxidation pretreatment prior to its treatment in MBR. The MBR was operated at an extended HRT of 10 d using 12 h electrooxidation treated wastewater. For all the different experiments, sampling was performed every day and samples were analyzed for different nitrogen species, total organic carbon (TOC), polysaccharide and protein. The transmembrane pressure (TMP) was also measured regularly.
2.3. Analytical Methods
COD was measured using a spectrophotometer (DR3000, USA) from Hach according to closed reflux standard method. TOC was measured using Elemental’s TOC Analyzer (vario TOD cube, USA). Conductivity and pH were measured using a conductivity meter (iSETK, ecomet C65) and pH meter (OHAUS, ST 300, USA), respectively. MLSS was measured according to the standard method (APHA, 1998). Viscosity was measured using a viscometer from CAS (CL-1, Korea). Protein was measured using the Pierce Modified Lowry Protein Assay kit (Thermo Fisher Scientific Co., USA), and polysaccharide was measured using the phenol sulfuric acid method using a UV-vis spectrophotometer (UV 2600, Japan) from Shimadzu. TMP was measured using method as described in Chung et al. [26].
3. Results and Discussion
3.1. Electrooxidation Treatment
3.1.1. Effect of total current
The effect of total applied current on different process parameters in the electrooxidation treatment of reject water and condensate generated during sludge dewatering is shown in Fig. 2. The increase in total current showed positive impact on nitrogen removal, with maximum TN and NH4-N removal efficiency of 22.7% and 18.5%, respectively, obtained at 4.03 A/L (Fig. 2a). The NH4-N, NO3-N and NO2-N values at this total applied current were 1279.9, 83.7 and 6.1 mg/L, respectively. The TOC and solids removal efficiencies also increased with increasing total applied current (Fig. 2a). For the first four total current values, the TOC and TSS removal efficiencies remained low (~30%), which gradually increased to 39.9% TOC removal and 82.8% TSS removal at the maximum total current used (5.28 A/L) in this study (Fig. 2b). The VSS to TSS ratio was consistently good (0.9–1.0) for all the applied current values. This represents a considerable improvement from the original VSS/TSS ratio of 0.48 of the raw wastewater due to electrooxidation treatment. The colour removal followed similar pattern with an increase in removal efficiencies at high total current values (Fig. 2b). A maximum of 65.2% colour removal could be achieved at 5.18 A/L total current. The polysaccharide and protein concentration in the wastewater were also monitored to understand the change in EPS composition due to electrooxidation treatment (Fig. 2c). The polysaccharide (PS) and protein (PN) removal efficiency improved with an increase in total applied current, with maximum respective removals of 94% and 62% obtained at 5.18 A/L total current.
The increase in total applied current is achieved by increasing the current density. The increase in current density usually increases the generation of hydroxyl radicals as well as reactive chlorine species which are responsible for pollutant removal from wastewater [27]. Multiple previous studies have observed that at elevated current density and consequently high total applied current, both the COD and ammonia nitrogen removal efficiency improved [28–30]. Although, the maximum treatment efficiency were achieved at 5.18 A/L for majority of parameters, removal efficiencies obtained at 4.03 A/L are often close enough or sometimes even better. Hence, spending extra energy for electrooxidation treatment to slightly improve treatment efficiency does not any make economic sense. Considering all these different factors, 4.03 A/L was selected as the optimum total applied current and used for further experiments wherever applicable.
3.1.2. Effect of current density
The effect of different current densities on electrooxidation treatment is shown in Fig. 3. It could be observed from the results that the nitrogen removal efficiencies fluctuated for different current density values, and no clear pattern could be discerned. The best nitrogen removal efficiency, i.e., 7.1% TN and 8.6% NH4-N were at 20 mA/cm2 (Fig. 3a). The concentrations of different nitrogen constituents in wastewater also did not show any significant change for samples treated at different current densities including the influent wastewater. The TOC removal efficiency remained almost similar (36.4–40.6%) for all seven different current densities used in this study (Fig. 3b). Solids removal improved with increase in current density, and a maximum of 43.3% TSS removal efficiency could be obtained at 60 mA/cm2. Further increase in current density to 70 and 90 mA/cm2 did not show any betterment in treatment efficiency (Fig. 3b). The VSS to TSS ratio improved significantly from 0.76 at 10 mA/cm2 to 0.95 at 50 mA/cm2 and further increased to 1.0 at higher current densities. The most prominent impact of variation in current density could be observed for colour removal, with gradual improvement with an increase in current density (Fig. 3b). The maximum colour removal efficiency (94.6%) was obtained for 70 and 90 mA/cm2, at which current density the colour intensity reduced to 5 CU from its original value of 92 CU for raw wastewater. The polysaccharide removal efficiency was consistently high (>90%) for all the different current densities (Fig. 3c). In case of protein removal, the increase in current density showed positive impact with a maximum 62% removal at 90 mA/cm2.
Although, this particular experiment studied the effect of current density on pollutant removal from wastewater, in order to keep the total applied current fixed, the reaction time was varied accordingly. For example, the reaction times were 14 h and 7 h at low current densities of 10 and 20 mA/cm2, whereas these were ~2 h at high current densities (70 and 90 mA/cm2). As a result, the impact of current density on TOC and NH4-N removal does not come out prominently from the results; rather, this change in reaction time showed more significance. This is the reason behind not following the general pattern of increased pollutant removal with increase in current density. Regardless of that, the colour removal was excellent at high current density of 90 mA/cm2 indicating high rate of hydroxyl ion generation at such current density could be responsible for better removal [28]. The membrane fouling causing agents such as TSS, and the protein and polysaccharide concentration in wastewater, were much low in effluent treated at 60 mA/cm2 or above. Hence, this current density value was taken to be optimum.
3.1.3. Effect of pH
Fig. 4 illustrates the effect of pH on different parameters of wastewater during electrooxidation treatment. The nitrogen removal was best at a high pH value of 9, where the TN and NH4-N removal efficiency reached 19.9% and 18.2%, respectively (Fig. 4a). The TOC removal showed inversely proportional relation with increase in pH as the removal efficiency gradually depleted with pH increase. A maximum TOC removal efficiency of 44.9% was obtained at pH 3 (Fig. 4b). However, TSS removal efficiency showed a completely opposite pattern, improving with increasing pH value. For instance, the TSS removal efficiency increased from 66.7% at pH 3 to 78.1% at pH 9. The VSS to TSS ratio was best (0.98) at pH 5, closely followed by pH 7 (0.95). The colour removal efficiency increased from 65.2% at pH 3 to 88.4% at pH 7; however, further increase in pH resulted in drastic reduction in colour removal efficiency to 32.6% (Fig. 4b). The polysaccharide and protein concentrations in the treated samples and their respective removal efficiencies varied based on pH value (Fig. 4c). The best pH for maximum polysaccharide and protein removal were different, i.e., pH 7 and 3, respectively, at which 90% and 67% removal efficiency could be achieved.
It is clear from the results that pH is a crucial factor affecting the performance of electrooxidation treatment. The difference in the optimum pH for TOC and nitrogen removal has to do with their respective removal mechanisms during electrooxidation process. Organic compounds are generally removed by hydroxyl radical generated at the anode surface, whereas nitrogen compounds are degraded with active chlorine/hypochlorite species produced during the process [27, 31]. At low pH conditions, various hydroxyl ion scavenging compounds such as carbonate and bicarbonate are depleted, resulting in enhanced oxidation of organic compounds present in wastewater. Moreover, acidic pH prevents (or at least restricts) the generation of undesirable O2 evolution as it requires higher energy input at such pH condition [27, 32]. This reduces unnecessary energy consumption for O2 producing reaction and enhances organic degradation efficiency. However, reactive chlorine species such as hypochlorous acid (HOCl) and hypochlorite ions (OCl−) are produced from chlorine at high pH (pH > 7), which reacts with nitrogen compounds and are responsible for its removal [33]. Similar observations for alkaline and acidic pH to be optimal for organic carbon and nitrogenous compounds, respectively, have been previously reported by other authors as well [31, 34]. Hence, with such conflicting results for different pollutants, selecting the optimum pH value for electrooxidation treatment becomes difficult. For this purpose, only the critical pollutants that cannot be removed during subsequent MBR treatment and can cause membrane fouling are taken into consideration. Based on these factors, pH 7 was selected to be optimum and used in further experiments.
3.1.4. Effect of Cl− ion concentration
The effect of Cl− ion on electrooxidation treatment of reject wastewater from sludge dewatering is shown in Fig. 5. The addition of Cl− ion showed a positive impact on nitrogen removal, as both TN removal and NH4-N removal efficiency improved with increase in Cl− ion. The maximum TN removal and NH4-N removal efficiencies were 22.4% and 21.6%, respectively at 4000 mg/L of Cl− ion (Fig. 5a). The NH4-N, NO3-N and NO2-N concentrations in the treated wastewater were 1206.2, 21.3 and 6.8 mg/L, respectively, at this high chlorine concentration. The TOC and TSS removal efficiencies also improved with an increase in Cl− ion concentration, with maximum values reaching 49.9% and 56.7%, respectively, at 4000 mg/L of Cl− concentration (Fig. 5b). The VSS to TSS ratio showed significant improvement due to an increase in Cl− concentration, increasing from 0.36 at 500 mg/L of Cl− to >0.92 at 1500 mg/L and above Cl− concentration. The colour removal also followed a similar pattern to other pollutants, with a maximum colour removal efficiency of 96.7% could be obtained at the highest Cl− concentration of 4000 mg/L (Fig. 5b). In addition, both polysaccharide and protein removal values increased with increase in Cl− addition (Fig. 5c). The maximum polysaccharide and protein removal efficiencies were 95% and 87%, respectively, at 4000 mg/L of Cl− concentration.
The addition of chlorine ions helps increase the generation of reactive chlorine species, which are responsible for pollutant degradation during electrooxidation process. Agustina et al. [29] have observed that due to presence of high Cl− concentration (1500–3550 mg/L), the chlorine reactive species were the dominant oxidative species during the anodic oxidation of landfill leachate. Nidheesh et al. [35] also made a similar observation, noting that the increase in COD and colour removal at high Cl− concentration (5 g/L) during electrooxidation treatment of mixed industrial wastewater was due to an increase in active chlorine species. Chlorine ions are added generally in the form of NaCl, which also acts as electrolyte and can enhance the electrical conductivity of the solution to be treated in an electrochemical system [27]. Compared to other electrolytes such as sodium sulfate and perchloric acid, NaCl has been proven to be better option for pollutant degradation [36, 37]. However, at very high Cl− concentration, the anodic reaction to form chlorine became kinetic-depended, i.e., on current density rather than chloride concentration. In this current study, no such negative effect on pollutant removal due to increase in Cl− concentration was observed, possibly because of relative range of Cl− concentrations used. However, considering the negative impact of Cl− concentration on biological treatment, which could affect the subsequent MBR operation [38], a moderate level of 1500 mg/L was chosen as optimum.
3.2. Membrane Bioreactor Performance
3.2.1. Effect of electrooxidation pretreatment on MBR performance
The MBR performance with electrooxidation-treated and untreated wastewater is depicted in Fig. 6. The results clearly demonstrate the better performance of the electrooxidation-treated wastewater in comparison to the the untreated wastewater. The most important parameter in the case of membrane performance with respect to membrane fouling is TMP. Monitoring and controlling transmembrane pressure is important for optimizing the performance and lifespan of any membrane-based system [39]. The increase in TMP indicates that there is formation of foulant over the membrane surface which is hindering the free flow of water through it. From the results, it can be seen that the TMP gradually increases from 2 kPa on the first day of MBR operation to 13 kPa on 8th day for EO treated wastewater. It remained almost steady after that and reached 15 kPa with another 5 days of continuous operation. Whereas, in case of untreated raw wastewater, the TMP at the start of MBR operation is 13 kPa, indicating membrane fouling from the beginning of the reactor operation itself. The TMP value reached 15 kPa by day 5 and remained constant for the rest of the MBR operation with untreated wastewater. The reason behind such stark difference between TMP can be explained by the TSS and EPS (measured as polysaccharide and protein) concentration of the EO-treated and untreated wastewater fed to the two different MBRs. The primary membrane fouling agents, namely suspended solids and EPS, exhibited a considerable reduction in wastewater treated with EO. Conversely, these substances were notably more abundant in untreated wastewater, contributing to a rapid membrane fouling rate in the control MBR [40, 41].
The performance in terms of nitrogen and TOC removal was comparable in both the bioreactors; however, as the influent concentrations of these pollutants were less in EO-treated wastewater than in untreated wastewater, the final effluent concentration achieved after MBR treatment is better for EO-pretreated wastewater. The nitrification and denitrification rates also improved due to EO treatment as evidenced by enhanced NO2-N and NO3-N concentrations during MBR operation using treated wastewater (Fig. 6). In addition to removing pollutants from wastewater, EO treatment can also breakdown complex organics and nitrogenous compounds leading to their better degradation during subsequent biological treatment in the MBR [42]. This could be attributed to the better performance of the MBR with EO-treated wastewater than the control MBR.
3.2.2. Effect of increase in HRT and EO treatment duration on MBR performance
Additional experiments were later performed to understand the effect of prolonging the HRT of MBR to 10 days on treatment efficiency using a wastewater treated with electrooxidation for a longer duration (12 h) than the prior experiment. The results obtained from that experiment are depicted in Fig. 7. The nitrogen removal efficiency, in terms of both ammonia and total nitrogen, improved with 10 d HRT and a 12 h EO pretreatment in comparison to the previous experiments. Even at the initial stage of MBR operation, the NH4-N removal was at 25%, which gradually rose to reach 51% by the end of reactor operation on 18th day (Fig. 7a). The final NH4-N concentration value obtained was 826.1 mg/L, which is ~42% lower than the value obtained at 4 d HRT. The TOC removal efficiency also consistently increased to 84% along with a low effluent TOC concentration of 56 mg/L (Fig. 7b). This effluent TOC value is 89.4% less than the final TOC obtained at 4 d HRT using 6 h EO treated wastewater. The TMP value was 0 during initial 8 days of MBR operation and never went >1 for rest of the experimental period (Fig. 7b). This increase in HRT increased the reaction time available for the microorganisms present in the biofilm to treat wastewater, which inadvertently improved the treatment efficiency [43]. The increase in EO treatment period also had significant impact on MBR performance and membrane longevity by removing foulants from wastewater. This is also supported by the low polysaccharide (99.3–114.5 mg/L) and protein (108.6–121.4 mg/L) concentration in the wastewater (Fig. 7b).
The goal for this novel combined treatment system was not to produce effluents that can confirm to the release standards, but rather reduce the pollutant load in the wastewater so that it can be successfully recycled back to the conventional wastewater treatment plant without causing additional problems. The findings from this study strongly suggest that this combined treatment system involving electrooxidation treatment and MBR can potentially achieve that set target. The NH4-N and TN concentration could be brought down from their initial value of 1465 mg/L and 1440 mg/L, respectively, in raw wastewater to 826.1 mg/L and 830.9 mg/L following an EO treatment of 12 h followed by MBR treatment at 10 d HRT. Similarly, the organic content of the raw wastewater reduced from its original value of 868 mg/L to 56 mg/L following combined EO and MBR treatment. Additional advantages in the form colour reduction, suspended solids removal and improvement in biodegradability due to the EO treatment proved to be beneficial for the ultimate treatment of this reject water and condensate generated during sludge dewatering. However, in order to make this technology applicable in actual industry, pilot-plant studies need to be conducted where some of other parameters including water flowrate, volume of the treatment reactors etc. need to be taken care of. Future research direction should also be focused towards kinetics and modelling of the system through a data-driven process from which automation of the entire system can be achieved.
4. Conclusions
The study demonstrated a novel method for the treatment of high nitrogen containing reject water and condensate generated during the sludge dewatering process though integrated electrooxidation and membrane bioreactor systems. The effect of different process parameters, namely, total applied current, current density, pH and chorine ion concentration on pollutant removal from wastewater by electrooxidation process, was studied. The results revealed that these parameters have significant impact on pretreatment performance, and a very high treatment efficiency can be achieved through their optimization. The main target pollutants for EO treatment were suspended solids and EPS that cause membrane fouling, could be removed from wastewater at a very high degree (57% and 86%, respectively). The biodegradability of the wastewater could also be improved due to EO pretreatment as the VSS to TSS ratio increased from 0.49 to >0.9 for most of the different experimental conditions. Another wastewater parameter which is difficult to be treated in MBR, i.e., colour removal (>90%), could be achieved using EO treatment. During MBR operation, initially a comparative experiment was conducted with EO pretreated and raw wastewater. Although, the performance of pollutants removal was comparable for both the wastewater, MBR treating EO pretreated wastewater showed better performance regarding membrane fouling. Further improvement in pollutant removal performance was obtained by increasing the EO treatment time to 12 h and extending the MBR operation by prolonging the HRT to 10 days. Finally, a low effluent concentration for NH4-N and TOC, i.e., 826.1 mg/L and 56 mg/L, respectively were obtained using MBR operated at 10 d HRT for the 12 h EO treated wastewater.
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