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Electrochemical Treatment of Recirculating Aquaculture Wastewater Using a Ti/RuO2-IrO2 Anode for Synergetic Total Ammonia Nitrogen and Nitrite Removal and Disinfection

Y. Ruan, C. Lu, X. Guo, Y. Deng, S. Zhu

Published in Transactions of the ASABE 59(6): 1831-1840 (doi: 10.13031/trans.59.11630). Copyright 2016 American Society of Agricultural and Biological Engineers.

Submitted for review in November 2015 as manuscript number PAFS 11630; approved for publication by the Plant, Animal, & Facility Systems Community of ASABE in August 2016.

The authors are Yunjie Ruan, Postdoctoral Fellow, Chan Lu, Graduate Student, Xishan Guo, Associate Professor, Yale Deng, Graduate Student, and Songming Zhu,ASABE Member, Professor, College of Biosystems Engineering and Food Science, Zhejiang University, and Key Laboratory of Equipment and Informatization in Environment Controlled Agriculture, Ministry of Agriculture, Hangzhou, China. Corresponding author: Xishan Guo, College of Biosystems Engineering and Food Science, Zhejiang University, Hangzhou 310058, China; phone: +86-571-88982373; e-mail:

Abstract.  Wastewater treatment and biosecurity are essential for intensive recirculating aquaculture system (RAS) production. In this study, the viability of the electrochemical process using a Ti/RuO2-IrO2 anode for synergetic total ammonia nitrogen (TAN) and nitrite removal and disinfection of semi-commercial RAS wastewater was evaluated. During the electrochemical oxidation process, the effects of the applied current density, sodium chloride concentration, and initial pH on the removal of TAN and nitrite were investigated. Experiment results indicated that under the conditions of 1.7 g L-1 sodium chloride concentration and 60 min electrolysis time, TAN removal efficiencies reached 78% at a current density of 60 mA cm-2, while nitrite removal efficiencies reached more than 95% at a current density of 30 mA cm-2. TAN removal due to an indirect oxidation mechanism followed second-order kinetics, while nitrite removal was described by pseudo-first-order kinetics in low-salt water (1.7 g L-1). The kinetics for electro-oxidation of TAN and nitrite affected by the current density were expressed as kTAN = 1.0 × 10-5 J-0.0002 and kNO2- = 1.9 × 10-3 J-0.0041, respectively. For disinfection, the active chlorine in situ generated by electrochemical treatment caused and inactivation. The results showed that the two objective pathogens can be sterilized rapidly, which indicated that no extra treatment for disinfection was needed. Finally, a comparison was made of the energy consumption in this study with those in the literature. This study showed that the Ti/RuO2-IrO2 anodic oxidation process has a potential for sustainable RAS wastewater treatment.

Keywords.Disinfection, Electro-oxidation, Nitrite, Recirculating aquaculture system, TAN, Wastewater treatment.

Indoor recirculating aquaculture systems (RAS) provide opportunities to reduce water consumption and improve nutrient recycling, which make intensive fish production compatible with environmental sustainability (Martins et al., 2010). Thus, RAS have been regarded as a promising approach to meet the growing global demand for aquatic products due to the generally high stocking density (Badiola et al., 2012). However, the intensive culture implies increasing nutrient inputs from feeding, as well as added crowding stress, which might reduce the immunity levels of fish and potentially spread opportunistic pathogens (Naylor et al., 2000). In practice, RAS mainly employ biofilm-type technology, such as fixed-bed reactors and moving-bed reactors to remove inorganic contaminants from the protein-rich feed (Gutierrez-Wing and Malone, 2006). The functional microorganisms, including ammonia-oxidizing bacteria (AOB) or archaea (AOA) and nitrite-oxidizing bacteria (NOB), play key roles in transferring the toxicity of total ammonia nitrogen (TAN) and nitrite to nitrate for fish species (Grommen et al., 2002). It is not enough to transfer TAN and nitrite to nitrate in wastewater treatment using biofilm-type technologies because excessively high nitrate content seems not beneficial to fish species. For disease control, disinfection technologies that use chemical therapeutic agents (Noble and Summerfelt, 1996; Burridge et al., 2010), ozone (Summerfelt, 2003), UV radiation (Mamane et al., 2010; Sanchez-Roman et al., 2007), and a combination of ozone with UV irradiation (Sharrer and Summerfelt, 2007) are the most common methods. Among these methods, the amount of chemical therapeutic agents is not easy to control, and ozone or UV irradiation as an additional sterilization process consumes more energy.

Hence, although commercial RAS containing biofilters and UV/O3 sterilizers have proven feasible and have become established procedures for removing ammonia and pathogens, some improvements can still be made. Compared with other performance indexes in RAS, such as dissolved oxygen, temperature, salinity, and pH, which can be easily monitored and controlled, biofilters are more complicated, as the reactor performance relies on the interaction among the microbial communities, feed loading, and relevant environmental changes (Ruan et al., 2015). As a result, RAS lack rapid response ability when enhancement of the water quality is needed to meet emergencies, such as a disease outbreak due to water quality deterioration. In addition, the functional microorganisms (AOB and NOB) have different growth strategies (Gatti et al., 2015) and are easily affected by temperature, especially for fish cultures in cold water. NOB growing in a K-strategy are more sensitive than AOB, which have a negative effect on NOB in niche competition during biofilm formation (Nogueira and Melo, 2006). This might result in fluctuation of the nitrite concentration in marine water under various operating conditions. Some researchers have doubted whether continuous water disinfection in a mature RAS is necessary (Blancheton et al., 2013; De Schryver and Vadstein, 2014). These concerns are mainly that sterilization produces a selective pressure on specific pathogens, which may result in bacteria evolving to survive in the sterilized RAS. On the other hand, a biofilter enriched with functional bacteria may become a shelter in which pathogens survive, in which case higher sterilization intensity should be adopted to maintain biosecurity (Cogan, 2013). Improvements to solve these problems are needed for the sustainability and benefit of RAS.

A possible alternative for treatment of RAS wastewater is electrochemical oxidation technology, which has proven efficient and has some advantages, such as quick removal performance, synergetic disinfection, no sludge production, low sensitivity to temperature, easy operation, and a small footprint (Dash and Chaudhari, 2005; Mook et al., 2012). The performance of electrochemical oxidation treatment is mainly determined by the electrode material. Many different anode materials have been used for electro-oxidation. Graphite and stainless steel electrodes are inexpensive and easily available, but graphite anodes have relatively low removal efficiencies (Lin and Wu, 1996; Sun and Chou, 1999; Vijayaraghavan et al., 2008), and stainless steel anodes are unstable in marine water due to corrosion (Abuzaid et al., 1999). In contrast, dimensionally stable anodes (DSAs), as novel and efficient electrode materials, have good electrical conductivity and high corrosion resistance. For example, Xing and Lin (2011) used Ti/IrO2-SnO2-Sb2O5 anodes for electrochemical treatment of aquaculture water. Moreover, DSAs have relatively low chlorine evolution potential, which provides high oxidation performance and can also be used for disinfection of pathogens in water (Jeong et al., 2009). As another novel material, boron-doped diamond (BDD) electrodes have been used to treat aquaculture water (Diaz et al., 2011) due to their stable properties and wide potential window. However, BDD electrodes are far more expensive and generate less active chlorine for disinfection than DSAs.

In this study, another material, Ti/RuO2-IrO2, which has advantages of low chlorine evolution potential, high corrosion resistance, and good conductivity (He et al., 2013), was used as the DSA electrode, for the first time to the best of our knowledge, for TAN and nitrite removal and disinfection of RAS wastewater. Thus, this study aimed to evaluate the viability of electro-oxidation of TAN and nitrite by Ti/RuO2-IrO2 anode to enhance the quality of real RAS wastewater for reuse. The major factors influencing TAN and nitrite removal, such as applied current density, sodium chloride concentrations, and initial pH values, were investigated. The effect of applied current density on the kinetics of TAN and nitrite electro-oxidation are discussed. The disinfection efficiency of the electrochemical process and the production of residual chlorine were also studied in sustainable RAS operation.

Material and Methods

The wastewater was collected from a semi-commercial tilapia RAS that included four tanks (4 m3) and one moving-bed biofilm reactor (MBBR) biofilter (0.5 m3) in our laboratory. The RAS was relatively stable and in working order for more than one year (Zhu et al., 2015). Table 1 lists the physicochemical characteristics of the wastewater collected for the experiments. Different doses of NaCl were separately added into certain amounts of the experimental wastewater to simulate aquaculture wastewater with different salinities (from 1.7% to 25‰). In addition, moderate amounts of (NH4)2SO4, NaNO2 and NaNO3 were added to provide the desired initial concentrations of TAN, nitrite-N, and nitrate-N. Unless specified, the initial TAN, nitrite-N, and nitrate-N concentrations were determined as 10, 5, and 100 mg L-1, respectively. Initial concentrations of 10 mg L-1 TAN and 5 mg L-1 or less nitrite are very large concentration levels that can be found in an untreated, high-density fish hatchery. The pH was adjusted using NaOH and H2SO4 during experiments that assessed the effects of pH on TAN and nitrite removal.

Table 1. Initial physicochemical characterization of the recirculating aquaculture system wastewater used in this study.
pH6.8 to 7.6
Conductivity (mS cm-1)0.62 to 1.02
Dissolved organic carbon (mg L-1)28.75 to 51.50
Total ammonia nitrogen (mg L-1)0.08 to 1.15
Nitrite-N (mg L-1)0.15 to 0.54
Nitrate-N (mg L-1)30.61 to 47.35

Escherichia coli and Vibrio parahaemolyticus were used as potential pathogens in the experiments to determine the efficiency of electrochemical disinfection. E. coli was cultured in nutrient broth, and V. parahaemolyticus was cultured in 3% NaCl alkaline peptone water. All microorganisms were pre-cultured in each medium for 24 h at 37°? to reach the late logarithmic phase. Microbial cells were then collected by centrifuging at 5000 rpm for 10 min at 25°? and later resuspended in normal saline. For the disinfection experiments, RAS wastewater was spiked with the reagents mentioned above to simulate synchronous TAN and nitrite removal and disinfection and was sterilized before use to eliminate the influence of germs contained in the raw wastewater. Suspensions of E. coli and V. parahaemolyticus were added to the wastewater to create an initial population of approximately 105 to 106 CFU mL-1.

Experimental Setup

A schematic diagram of the experimental setup used in this study is shown in figure 1. The plexiglass electrochemical cell was comprised of a Ti/RuO2-IrO2 anode (Fenggang Co. Ltd., China) and a graphite cathode with an effective single-electrode surface area of 19.5 cm2 (3 cm × 6.5 cm) and inter-electrode gap of 1 cm. Wastewater volumes of 1 L and 2.19 L were stored in the electrochemical cell and the feed tank, respectively. A peristaltic pump (BT100-2J, Longer Co. Ltd., China) was used to recirculate the inflow from the tank through the cell at a flow rate of 400 mL min-1. During treatment, the reaction was controlled by a DC power supply (TPR3003T/3005T, Atten Technology Co. Ltd., China), while the range of the applied current density was between 15 and 60 mA cm-2.

Figure 1. Experimental setup: 1 = feed tank, 2 = pump, 3 = electrochemical cell, 4 = power supply, and 5 = stirrer.

Analytical Methods

During treatment, samples were taken regularly from the feed tank using a pipette for analysis of different parameters. The water samples were filtered through a 0.45 µm membrane filter before analysis. TAN and NO2--N concentrations were analyzed using a UV-Vis spectrophotometer (Cary 60, Agilent Technologies) according to standard methods (SEPA, 2002). The pH was measured with a pH meter (S40, Mettler Toledo), and conductivity was measured with a portable conductivity meter (DDB-303A, REX Instrument, China). Samples were filtered with a 0.22 µm micro-membrane for total dissolved organic carbon (DOC) analysis using a Multi N/C 2100 (Analytik Jena, Germany). Residual chlorine was analyzed using a pocket chlorimeter (RC-3F, KRK, Japan) according to the DPD (N,N-diethyl-p-phenylenediamine) method.

For the electrochemical disinfection experiments, 5 mL was sampled every 5 min from the feed tank and immediately quenched with 0.20 mL of Na2S2O3 (5 wt%) to eliminate the residual chlorine in solution. The samples were then diluted accordingly, and 1 mL of undiluted or diluted suspension was taken for counting using the pour plate method with violet red bile agar (VRBA) for E. coli and thiosulfate citrate bile sucrose agar (TCBS) for V. parahaemolyticus. Three replicates were carried out in each series for 24 h under 37°?. Except for the variable, the same initial conditions were reached for each series of experiments.

Results and Discussion

    Based on a previous study, the removal of TAN and nitrite-N can be explained as follows: free chlorine is generated in situ through electro-oxidation of the available chloride ions in wastewater (Gendel and Lahav, 2012). In addition, when the pH value is in the range of 6.8 to 7.6, most of the TAN and free chlorine is in the form of NH4+ and HClO, respectively. Thus, the reaction sequence during the electrochemical treatment can be represented by equations 1 through 7 (Diaz et al., 2011; Ding et al., 2015):

Anode:    2Cl- ? Cl2 + 2e-    (1)

    Cl2 + H2O ? HClO + H+ + Cl-    (2)

    HClO ? ClO- + H+    (3)

    2NH4+ + 3HClO ? N2 + 3H2O + 5H+ + 3Cl-    (4)

    HClO + NO2- ? NO3- + H+ + Cl-    (5)

Cathode:    2H2O + 2e- ? H2 + 2OH-    (6)

    NO3- + 6H2O + 8e- ? NH3 + 9OH-    (7)

According to the above analysis, the electro-oxidation process is affected by the electric charge. Figures 2a and 2b show the variations of TAN and NO2--N with different current densities during the electrochemical oxidation process for RAS wastewater that contained 1.7 g L-1 sodium chloride. From the curves shown in figures 2a and 2b, NO2--N was almost completely removed after 60 min of electrolysis when the applied current density was greater than 15 mA cm-2 after 60 min of electrolysis, while no similar behavior was found for TAN removal. This phenomenon was also reported in a previous study (Díaz et al., 2011). Figure 3 shows that higher TAN removal was achieved without adding nitrite, which indicated that nitrite could strongly affect TAN removal performance when co-existing in the system. A similar phenomenon was found in other work (Lin and Wu, 1996). In addition, equation 7, describing the reduction of NO3- in the cathodic area, could explain the slightly increase of TAN in the reaction system at a relatively low current density (figs. 2a and 2b).

The evolution of residual chlorine concentration during the electrolytic process at different applied current densities is shown in figure 2c. As expected, when the applied current density increased, a higher residual chlorine concentration was obtained. As the concentrations of chlorine produced in our system were much higher than the concentrations of residual chlorine that are lethal to a wide variety of fish (Tanaka et al., 2013), activated charcoal should be applied to remove the residual chlorine, as well as other chlorinated and bromated compounds, especially when the electrolytic process is used for seawater treatment (Gopal et al., 2007; Yeh et al., 2013). However, because synergetic disinfection is one of the main purposes of this study, optimized operation by mixing high residual chlorine water with other effluents might become a potential practice in RAS operation.

Effect of Cl- Concentration on Ammonia and Nitrite Removal

Because different kinds of fish have different water salinity requirements, diverse doses of sodium chloride were added to simulate RAS seawater and brackish water. In this study, the effect of Cl- on the removal of nitrogen compounds was evaluated. Figure 4 shows the influence of sodium chloride concentrations on TAN and nitrite removal and the residual chlorine production.

Four sodium chloride concentrations were tested (1.7, 8, 15, and 25 g L-1) with current density maintained at 30 mA cm-2. Figures 4a and 4b show that higher chloride concentration resulted in more TAN and nitrite removal at the same current density. The same results can be found in the literature (Ding et al., 2015; Liu et al., 2009). As shown in figure 4a, when the sodium chloride concentration was greater than 8 g L-1, TAN degradation followed a nearly linear trend, which was different from the trend observed with different current densities and a sodium chloride concentration of 1.7 g L-1 (fig. 3a).

Figure 2. Effects of current density on evolution of (a) [TAN]/[TAN0], (b) [NO2--N]/[NO2--N]0, and (c) production of residual chlorine ([NO2--N]0 ˜ 5 mg L-1, [TAN]0 ˜ 10 mg L-1, initial pH 7.3 to 7.6, and NaCl concentration 1.7 g L-1).
Figure 3. Effects of nitrite and nitrate existence on TAN removal (current density of 50 mA cm-2 and initial pH of 7.3 to 7.6, [TAN]0 ˜ 10 mg L-1): (1) [NO2--N]0 ˜ 5 mg L-1, [NO3--N]0 ˜ 100 mg L-1; (2) [NO2--N]0 ˜ 0.18 mg L-1, [NO3--N]0 ˜ 100 mg L-1; and (3) [NO2--N]0 ˜ 0.54 mg L-1, [NO3--N]0 ˜ 38.67 mg L-1.

The production of residual chlorine during the electrolytic process with different concentrations of sodium chloride is shown in figure 4c. Obviously, a higher sodium chloride concentration can produce more residual chloride at the same current density. When most of the TAN and nitrite were removed, the amount of residual chlorine increased dramatically, which was also observed by Kapalka et al. (2010) and Ding et al. (2015). Both TAN and nitrite reached more than 97% degradation after 40 min of electrolysis with a sodium chloride concentration of 25 g L-1. However, the residual chlorine concentration after electrolysis with a sodium chloride concentration of 25 g L-1 did not seem to be much higher than the residual chlorine concentration after electrolysis with a sodium chloride concentration of 15 g L-1, which means that higher concentrations of NaCl had higher removal efficiencies for TAN and nitrite without producing more residual chlorine.

Effect of Initial pH on Ammonia and Nitrite Removal

The effect of initial pH on TAN and nitrite removal was investigated with low-salinity water (8 g L-1 NaCl) at a current density of 30 mA cm-2. The selected pH values nearly covered the actual pH range of the RAS wastewater (i.e., 6.5, 7.5, and 8.5). The results agreed with previous findings that acidic conditions are relatively conductive to nitrite removal (Lin and Wu, 1996). The TAN removal rate was a little slower at pH 6.5 than at pH 7.5 or 8.5, and there was no obvious difference in TAN degradation rates at pH 7.5 and 8.5 (data not shown). Overall, the results indicated that an initial pH in the range of 6.5 to 8.5 slightly influenced the removal of TAN and nitrite.

At the same time, the pH variation was determined during the electrolysis process (data not shown). The pH in the reaction system tended to reach a moderate value and then dropped at 50 min of electrolysis time, when the TAN was almost removed. After that, the pH increased due to the ac-cumulation of OH- by the nearly pure electrolysis of Cl- after the complete degradation of ammonia (Xiao et al., 2009). The pH value was still in a reasonable range for most fish species.

Figure 4. Effects of NaCl concentration on evolution of (a) TAN/TAN0, (b) [NO2--N]/[NO2--N]0, and (c) residual chlorine production ([NO2--N]0 ˜ 5 mg L-1, [TAN]0 ˜ 10 mg L-1, current density of 30 mA cm-2, and pH of 6.8 to 7.0).

Ammonia and Nitrite Removal Kinetics

In the studies by Anglada et al. (2009) and Diaz et al. (2011), TAN oxidation in an electrochemical process followed second-order kinetics. In this study, the model (eqs. 8 through 10) was also used to predict the data obtained with different current densities at a low concentration of sodium chloride (1.7 g L-1). The model assumes that TAN removal occurs mainly through indirect oxidation mechanisms by means of electro-generated active chlorine. In the pH range of our reaction system, the major component of active chlorine is HOCl, so the model can be given as:


where k is the second-order rate constant (L mg-1 min-1), and [HOCl] is the concentration of hypochlorous acid (mg L-1). Assuming that the rates of chlorine loss due to cathodic reduction of active chlorine, anodic oxidation to ClO3-, and homogeneous reaction with TAN are much lower than the chlorine production rate, the variation of chlorine concentrations with time (until saturation is reached) can be described by equation 9 (Anglada et al., 2009; Diaz et al., 2011):


where F is the Faraday constant (96485 C mol-1); ? is the current efficiency for chlorine evolution, which depends on the current density, mass transport rate coefficient, and chloride concentration; A is the effective area of the electrode (m2); J is the current density (A m-2); and V is the volume of solution (m3). The model assumes that chloride concentration remains constant throughout the oxidation process and that free chlorine evolution is the main anodic reaction. Consequently, the substitution of the integrated form of equation 9 into equation 8 gives:


Therefore, TAN removal due to an indirect electro-oxidation mechanism follows second-order kinetics. Values of k' (min-2) are derived for different applied current densities from the slopes of the logarithms of [TAN]t/[TAN]0 versus the square of the electrolysis time. When the current density was 30, 40, 50, and 60 mA cm-2, k' was 8.8 × 10-5, 2 × 10-4, 3 × 10-4, and 4 × 10-4, respectively. In all cases, the correlation coefficient was greater than 0.98. A linear equation was obtained to describe the dependency of the kinetic parameter k' on the applied current density:

    k' = 1.0 × 10-5 J-0.0002 (min-2)    (11)

    (R2 = 0.999)

The simulated values from the model fitted well with the experimental data, as 90% of the simulated data fell within the experimental data range of 10%.

Kinetics of Nitrite Oxidation

In previous studies describing the kinetics of nitrite electro-oxidation, Sun and Chou (1999) used the Butler-Volmer equation to describe effect of nitrite ion concentrations on the kinetics of nitrite oxidation, and Lin and Wu (1997) and Diaz et al. (2011) described the behavior of nitrite by a zero-order expression. In contrast to the nitrite removal kinetics mentioned above, pseudo-first-order kinetics were assumed in this study, and the rate of nitrite disappearance can be expressed as:


After integration, equation 12 gives:


Values of k (min-1) are derived for different applied current densities from the slopes of the logarithms of [NO2-]t/[NO2-]0 versus the electrolysis time. When the current density was 15, 30, 40, 50, and 60 mA cm-2, k was 0.0232, 0.0594, 0.0692, 0.0943, and 0.1121, respectively. Under various conditions, the correlation coefficients were all greater than 0.97. The kinetic constant of nitrite removal (k) increased linearly with the applied current density, and the fitting relationship is given as equation 14:

    k = 1.9 × 10-3 J-0.0041 (min-1)    (14)

    (R2 = 0.989)

The simulated values from the model are nearly consistent with the experimental data, as about 80% of the simulated data fell within the experimental data range of 20%.

Inactivation of Microorganisms

Electrochemical treatment of water has shown potential for disinfection of different types of water, such as drinking water, industrial wastewater, and domestic wastewater (Martínez-Huitle and Brillas, 2008; Anglada et al., 2009; Pikaar et al., 2011). One advantage of electrochemical disinfection is that the damage produced to the bacterial cells is more severe than that produced by pure chemical disinfection with chlorine (Wang et al., 2010).

In this study, electrochemical disinfection experiments were performed by treatment of the spiked RAS wastewater with E. coli (as an indicator bacterium in freshwater) and Vibrio parahaemolyticus (as an indicator bacterium in seawater). Figures 5a and 5b show the effects of current density on the inactivation of E. coli and V. parahaemolyticus during the electrolysis process for aquaculture water containing 1.7 and 25 g L-1 sodium chloride, respectively. The results showed that higher current density enabled better efficiency in sterilization of E. coli at the same sodium chloride concentration. The average germicidal rate at 15 mA cm-2 was 1.74 × 108 CFU min-1, which was higher than that of 8.56 × 107 CFU min-1 using a ß-PbO2 anode doped with fluorine at 7 mA cm-2 and 5 min, as found by Cong et al. (2008). The disinfection rate was also higher than when using Pt or BDD as the anode material (Jeong et al., 2007, 2009; López-Gálvez et al., 2012). Inactivation of V. parahaemolyticus followed the same trend, with an average germicidal rate of 6.51 × 107 CFU min-1 at 15 mA cm-2 after 10 min. Additionally, it has been proven that electrochemical disinfection is effective for many other pathogens (Tanaka et al., 2013; Feng et al., 2004; Liang et al., 2005).

Figure 5. Effects of current density on inactivation of E. coli and V. parahaemolyticus: (a) [NaCl] = 1.7 g L-1 and (b) [NaCl] = 25 g L-1.

Under the electrolysis conditions, enough chlorine was generated in situ from the chloride ions in the solution to contribute most of the disinfection. Contrary to TAN and nitrite removal, less time was required for inactivation, which meant no extra energy was needed for disinfection. Therefore, electrochemical disinfection can be a promising supplement to UV for decreasing the equipment investment and operating cost, which is mostly used for disinfection in conventional RAS. However, the potential disinfection byproducts, such as trihalomethanes, bromate, and haloacetic acids, have toxic effects on fish. These compounds can damage aquatic animals by oxidizing cell membranes, especially in their gills, lateral line, and skin, and can even cause death (Katayose et al., 2007; Yeh et al., 2013). Therefore, further treatment, such as activated carbon, is needed after electrochemical disinfection to remove residual chlorine and disinfection byproducts, which will be further studied in our future work.

Energy Consumption Analysis

Figure 6. Removal efficiency and energy consumption versus current density during electrochemical oxidation for (a) TAN and (b) NO2--N and (c) energy consumption at different NaCl concentrations.

As electricity is almost the only consumable in the electrochemical oxidation process, energy consumption has to be considered for the technical feasibility of the process, and it is dependent on the percentage removal of TAN and nitrite-N. The energy consumption can be calculated according to equation 15:


where W is the energy consumption (kWh g-1 TAN or kWh g-1 NO2--N), U is the electrolysis voltage between the anode and the cathode (V), I is the applied current (A), t is the electrolysis time for total removal of TAN or NO2--N (min), and ?m is the removal amount of TAN or NO2--N (g). The energy consumption and removal efficiency of TAN and NO2--N at different current densities in wastewater containing 1.7 g L-1 NaCl are shown in figures 6a and 6b. At current density of 50 mA cm-2, TAN could reach 68.9% removal with the lowest energy consumption of 0.51 kWh g-1 TAN, while NO2--N could reach 99.8% removal with an energy consumption of 0.63 kWh g-1 NO2--N.

Figure 6c shows the energy consumption for total TAN and NO2--N removal at different NaCl concentrations. As shown in figure 6c, there is no great difference in energy consumption between 8 and 25 g L-1 NaCl at a current density of 30 mA cm-2 (i.e., 0.11 kWh g-1 TAN and 0.11 kWh g-1 NO2--N), whereas the energy consumption was much higher at 1.7 g L-1 NaCl (i.e., 0.64 kWh g-1 TAN and 0.34 kWh g-1 NO2--N).

The energy consumption of electrochemical oxidation for TAN removal has been extensively studied in the literature. A comparison of reported energy consumption values with our study at a NaCl concentration of 25 g L-1 is listed in table 2. Based on the analysis, the energy consumption calculated in this work was lower than that obtained in most other studies. A lower energy cost than our study was found by Díaz et al. (2011) using BDD anodes. However, comprehensive consideration is needed due to the high cost of BDD materials and lower disinfection ability. In addition, the energy consumption for nitrite-N removal was 0.11 kWh g-1 NO2--N at a NaCl concentration of 25 g L-1 (fig. 6c), which was lower than the range of 1.2 to 4 kWh g-1 NO2--N obtained by Lin and Wu (1996), who used graphite anodes for nitrite removal at current densities of 44.2 to 110.6 mA cm-2 for simulated aquaculture wastewater. Table 2 also shows that the TAN removal efficiency of the Ti/RuO2-IrO2 anode is higher than that of the other reported electrodes used in aquaculture systems.

Furthermore, many other factors should be taken into consideration for reducing the energy consumption. Nitrite in low-salinity water would decrease the activity of heme oxygenase, which makes nitrite more toxic than TAN for fish species. However, in high-salinity water, more TAN would be in the form of NH3 due to the relatively high pH. At the same time, a high concentration of Cl- suppresses the negative effects of nitrite on fish, which makes TAN more toxic than nitrite for aqueous animals. Therefore, based on the analysis mentioned above, it is necessary to choose a suitable current density based on the aquaculture conditions, and to control energy consumption based on the stress level of the target pollutants.


Based on this study of anodic oxidation of TAN and nitrite and inactivation of E. coli and Vibrio parahaemolyticus during an electrochemical process, the following conclusions can be drawn:

Table 2. Efficiency and energy consumption for TAN removal in different electrochemical oxidation systems.
Experimental ConditionsRemoval Efficiency
(min/[mA cm-2 mg L-1])
(kWh g-1 TAN)
Anode MaterialEffective
Anode Area
(mA cm-2)
BDD14021150.041Díaz et al.
Ti/IrO2-SnO2-Sb2O51001.82.51.550.68Xing and Lin
Ti/RuO2-IrO218907.515.46-0.13He et al.
Saline industrial
BDD701-230-0.43Anglada et al.
Ti/RuO2-IrO219.53.19300.20.11This study


This study was financially supported by the National Science and Technology Support Project of China (2014BAD 08B09), the National High Technology Research and Development Program of China (2013AA103007), the Special Fund for Agro-Scientific Research in the Public Interest (201303091), and the Natural Science Fund of China (31402348). We also thank the reviewers for their insightful comments and suggestions.



Abuzaid, S., Al-Hamouz, Z., Bukhari, A. A., & Essa, M. H. (1999). Electrochemical treatment of nitrite using stainless steel electrodes. Water Air. Soil Pollut., 109(1), 429-442.

Anglada, A., Ibañez, R., Urtiaga, A. M., & Ortiz, I. (2010). Electrochemical oxidation of saline industrial wastewaters using boron-doped diamond anodes. Catal. Today, 151(1-2), 178-184.

Anglada, A., Urtiaga, A. M., & Ortiz, I. (2009). Pilot-scale performance of the electro-oxidation of landfill leachate at boron-doped diamond anodes. Environ. Sci. Tech., 43(6), 2035-2040.

Badiola, M., Mendiola, D., & Bostock, J. (2012). Recirculating aquaculture systems (RAS) analysis: Main issues on management and future challenges. Aquacult. Eng., 51, 26-35.

Blancheton, J. P., Attramadal, K. J., Michaud, L., d’Orbcastel, E. R., & Vadstein, O. (2013). Insight into bacterial population in aquaculture systems and its implication. Aquacult. Eng., 53, 30-39.

Burridge, L., Weis, J. S., Cabello, F., Pizarro, J., & Bostick, K. (2010). Chemical use in salmon aquaculture: A review of current practices and possible environmental effects. Aquaculture, 306(1-4), 7-23.

Cogan, N. G. (2013). Concepts in disinfection of bacterial populations. Math. Biosci., 245(2), 111-125.

Cong, Y., Wu, Z., & Li, Y. (2008). Electrochemical inactivation of coliforms by in situ generated hydroxyl radicals. Korean J. Chem. Eng., 25(4), 727-731.

Dash, B. P., & Chaudhari, S. (2005). Electrochemical denitrificaton of simulated ground water. Water Res., 39(17), 4065-4072.

De Schryver, P., & Vadstein, O. (2014). Ecological theory as a foundation to control pathogenic invasion in aquaculture. ISME J., 8(12), 2360-2368.

Diaz, V., Ibañez, R., Gómez, P., Urtiaga, A. M., & Ortiz, I. (2011). Kinetics of electro-oxidation of ammonia-N, nitrites, and COD from a recirculating aquaculture saline water system using BDD anodes. Water Res., 45(1), 125-134.

Ding, J., Zhao, Q., Zhang, Y., Wei, L., Li, W., & Wang, K. (2015). The eAND process: Enabling simultaneous nitrogen removal and disinfection for WWTP effluent. Water Res., 74, 122-131.

Feng, C., Suzuki, K., Zhao, S., Sugiura, N., Shimada, S., & Maekawa, T. (2004). Water disinfection by electrochemical treatment. Bioresour. Tech., 94(1), 21-25.

Gatti, M. N., Gimenez, J. B., Carretero, L., Ruano, M. V., Borras, L., Serralta, J., & Seco, A. (2015). Enrichment of AOB and NOB population by applying a BABE reactor in an activated sludge pilot plant. Water Environ. Res., 87(4), 369-377.

Gendel, Y., & Lahav, O. (2012). Revealing the mechanism of indirect ammonia electrooxidation. Electrochim. Acta, 63, 209-219.

Gopal, K., Tripathy, S. S., Bersillon, J. L., & Dubey, S. P. (2007). Chlorination byproducts, their toxicodynamics and removal from drinking water. J. Hazard. Mater., 140(1), 1-6.

Grommen, R., Van Hauteghem, I., Van Wambeke, M., & Verstraete, W. (2002). An improved nitrifying enrichment to remove ammonium and nitrite from freshwater aquaria systems. Aquaculture, 211(1), 115-124.

Gutierrez-Wing, M. T., & Malone, R. F. (2006). Biological filters in aquaculture: Trends and research directions for freshwater and marine applications. Aquacult. Eng., 34(3), 163-171.

He, X., Chai, Z., Li, F., Zhang, C., Li, D., Li, J., & Hu, J. (2013). Advanced treatment of biologically pretreated coking wastewater by electrochemical oxidation using Ti/RuO2-IrO2 electrodes. J. Chem. Tech. Biotech., 88(8), 1568-1575.

Jeong, J., Kim, C., & Yoon, J. (2009). The effect of electrode material on the generation of oxidants and microbial inactivation in the electrochemical disinfection processes. Water Res., 43(4), 895-901.

Jeong, J., Kim, J. Y., Cho, M., Choi, W., & Yoon, J. (2007). Inactivation of Escherichia coli in the electrochemical disinfection process using a Pt anode. Chemosphere, 67(4), 652-659.

Kapalka, A., Katsaounis, A., Michels, N.-L., Leonidova, A., Souentie, S., Comninellis, C., & Udert, K. M. (2010). Ammonia oxidation to nitrogen mediated by electrogenerated active chlorine on Ti/PtOx-IrO2. Electrochem. Commun., 12(9), 1203-1205.

Katayose, M., Yoshida, K., Achiwa, N., & Eguchi, M. (2007). Safety of electrolyzed seawater for use in aquaculture. Aquaculture, 264(1), 119-129.

Liang, W., Qu, J., Chen, L., Liu, H., & Lei, P. (2005). inactivation of Microcystis aeruginosa by continuous electrochemical cycling process in tube using Ti/RuO2 electrodes. Environ. Sci. Tech., 39(12), 4633-4639.

Lin, S. H., & Wu, C. L. (1996). Electrochemical removal of nitrite and ammonia for aquaculture. Water Res., 30(3), 715-721.

Lin, S. H., & Wu, C. L. (1997). Electrochemical nitrite and ammonia oxidation in sea water. J. Environ. Sci. Health. A, 32(8), 2125-2138.

Liu, Y., Li, L., & Goel, R. (2009). Kinetic study of electrolytic ammonia removal using Ti/IrO2 as anode under different experimental conditions. J. Hazard. Mater., 167(1), 959-965.

López-Gálvez, F., Posada-Izquierdo, G. D., Selma, M. V., Pérez-Rodríguez, F., Gobet, J., Gil, M. I., & Allende, A. (2012). Electrochemical disinfection: An efficient treatment to inactivate Escherichia coli O157:H7 in process wash water containing organic matter. Food Microbiol., 30(1), 146-156.

Mamane, H., Colorni, A., Bar, I., Ori, I., & Mozes, N. (2010). The use of an open-channel, low-pressure UV reactor for water treatment in low-head recirculating aquaculture systems (LH-RAS). Aquacult. Eng., 42(3), 103-111.

Martínez-Huitle, C. A., & Brillas, E. (2008). Electrochemical alternatives for drinking water disinfection. Angewandte Chemie Intl. Ed., 47(11), 1998-2005.

Martins, C. I. M., Eding, E. H., Verdegem, M. C. J., Heinsbroek, L. T. N., Schneider, O., Blancheton, J. P., ... Verreth, J. A. (2010). New developments in recirculating aquaculture systems in Europe: A perspective on environmental sustainability. Aquacult. Eng., 43(3), 83-93.

Mook, W. T., Chakrabarti, M. H., Aroua, M. K., Khan, G. M. A., Ali, B. S., Islam, M. S., & Abu Hassan, M. A. (2012). Removal of total ammonia nitrogen (TAN), nitrate, and total organic carbon (TOC) from aquaculture wastewater using electrochemical technology: A review. Desalination, 285, 1-13.

Naylor, R. L., Goldburg, R. J., Primavera, J. H., Kautsky, N., Beveridge, M. C., Clay, J., ... Troell, M. (2000). Effect of aquaculture on world fish supplies. Nature, 405(6790), 1017-1024.

Noble, A. C., & Summerfelt, S. T. (1996). Diseases encountered in rainbow trout cultured in recirculating systems. Annu. Rev. Fish Diseases, 6, 65-92.

Nogueira, R., & Melo, L. F. (2006). Competition between Nitrospira spp. and Nitrobacter spp. in nitrite-oxidizing bioreactors. Biotech. Bioeng., 95(1), 169-175.

Pikaar, I., Rozendal, R. A., Yuan, Z., Keller, J., & Rabaey, K. (2011). Electrochemical sulfide oxidation from domestic wastewater using mixed metal-coated titanium electrodes. Water Res., 45(17), 5381-5388.

Ruan, Y.-J., Guo, X.-S., Ye, Z.-Y., Liu, Y., & Zhu, S.-M. (2015). Bacterial community analysis of different sections of a biofilter in a full-scale marine recirculating aquaculture system. North American J. Aquacult., 77(3), 318-326.

Sanchez-Roman, R. M., Soares, A. A., de Matos, A. T., Sediyama, G. C., DeSouza, O., & Mounteer, A. H. (2007). Domestic wastewater disinfection using solar radiation for agricultural reuse. Trans. ASABE, 50(1), 65-71.

SEPA. (2002). Water and wastewater monitoring methods (4th ed.). Beijing, China: Chinese Environmental Science Publishing House.

Sharrer, M. J., & Summerfelt, S. T. (2007). Ozonation followed by ultraviolet irradiation provides effective bacteria inactivation in a freshwater recirculating system. Aquacult. Eng., 37(2), 180-191.

Summerfelt, S. T. (2003). Ozonation and UV irradiation: An introduction and examples of current applications. Aquacult. Eng., 28(1), 21-36.

Sun, C.-C., & Chou, T.-C. (1999). Kinetics of anodic oxidation of nitrite ion using in situ electrogenerated HCLO in a NaCl aqueous solution. Ind. Eng. Chem. Res., 38(12), 4545-4551.

Tanaka, T., Shimoda, M., Shionoiri, N., Hosokawa, M., Taguchi, T., Wake, H., & Matsunaga, T. (2013). Electrochemical disinfection of fish pathogens in seawater without the production of a lethal concentration of chlorine using a flow reactor. J. Biosci. Bioeng., 116(4), 480-484.

Vijayaraghavan, K., Ahmad, D., & Bin Fadzin, T. S. (2008). In situ hypochlorous acid generation for the treatment of brackish shrimp aquaculture wastewater. Aquacult. Res., 39(5), 449-456.

Wang, Y., Claeys, L., van der Ha, D., Verstraete, W., & Boon, N. (2010). Effects of chemically and electrochemically dosed chlorine on Escherichia coli and Legionella beliardensis assessed by flow cytometry. Appl. Microbiol. Biotech., 87(1), 331-341.

Xiao, S., Qu, J., Zhao, X., Liu, H., & Wan, D. (2009). Electrochemical process combined with UV light irradiation for synergistic degradation of ammonia in chloride-containing solutions. Water Res., 43(5), 1432-1440.

Xing, Y., & Lin, J. (2011). Application of electrochemical treatment for the effluent from marine recirculating aquaculture systems. Procedia Environ. Sci., 10(1), 2329-2335.

Yeh, S.-P., Hsia, L.-F., & Liu, C.-H. (2013). Usage of electrolytic water system in the giant freshwater prawn, Macrobrachium rosenbergii (de Man) larval hatchery system. Aquacult. Res., 44(5), 713-727.

Zhu, S.-M., Deng, Y.-L., Ruan, Y.-J., Guo, X.-S., Shi, M.-M., & Shen, J.-Z. (2015). Biological denitrification using poly(butylene succinate) as carbon source and biofilm carrier for recirculating aquaculture system effluent treatment. Bioresour. Tech., 192, 603-610.