Degradation of nitrogen-containing refractory organic wastewater using a novel alternating-anode electrochemical system

Science of the Total Environment 697 (2019) 134161

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Degradation of nitrogen-containing refractory organic wastewater using a novel alternating-anode electrochemical system Yang Deng a, Nan Chen a,⁎, Chuanping Feng a,⁎, Haishuang Wang a, Yuhan Zheng a, Fangxin Chen a, Wang Lu a, Peijing Kuang a, Hanguang Feng a, Yu Gao b, Weiwu Hu c a b c

School of Water Resources and Environment, MOE Key Laboratory of Groundwater Circulation and Environmental Evolution, China University of Geosciences (Beijing), Beijing 100083, PR China College of Chemical and Environmental Engineering, Shandong University of Science and Technology, Qingdao 266590, PR China China University of Geosciences (Beijing), Journal Center, Beijing 100083, China

H I G H L I G H T S

G R A P H I C A L

A B S T R A C T

• AAES can simultaneously remove COD and TN in a single electrolytic cell. • AAES can promote magnetic force formation for electro-coagulation precipitates. • AAES can accelerate TN removal by decreasing the conversion of NO− x -N to NH+ 4 -N. • Obtained precipitate under Fe and DSA anode was Fe3O4 and γ-Fe2O3, respectively. • The available chlorine played a greater role in promoting the removal of COD.

a r t i c l e

i n f o

Article history: Received 12 May 2019 Received in revised form 11 August 2019 Accepted 27 August 2019 Available online 30 August 2019 Editor: Ching-Hua Huang Keywords: Alternating-anode electrochemical system Biologically treated landfill leachate Total nitrogen Magnetic precipitates formation Precipitates separation

a b s t r a c t This study presented a novel alternating-anode electrochemical system (AAES) based on single electrolytic cell for the treatment of nitrogen-containing refractory organic wastewater (NOW). The core of AAES lies in the alternating working of iron anode and DSA anode to integrate different electrochemical processes. The biologically treated landfill leachate (BTLL) was selected as a practical NOW for assessing the performance of AAES. The results indicated that after 140 min of electrolytic reaction, the removal efficiency of chemical oxygen demand and total nitrogen (TN) using AAES was found to be 76.9 and 98.9%, respectively. The main component of dissolved organic matter (DOM) in BTLL included humic-like substances, which could be degraded into smallmolecule DOM, such as fulvic-like substances and protein-like substances, by available chlorine and hydroxyl + radicals present in AAES. Cathode reduction (NO− x -N → NH4 –N and N2) under iron anode and indirect oxidation + (NH4 –N → N2) under DSA anode were the main pathways to remove TN from NOW. Owing to the redox conditions created by the alternating anodes, the main stable crystalline forms of precipitates obtained from AAES were Fe3O4 and γ-Fe2O3, which could be separated by using the external magnetic field. The findings of this study may provide a feasible solution for the advanced electrochemical treatment of NOW in a single electrolytic cell as well as rapid separation of precipitates. © 2019 Elsevier B.V. All rights reserved.

1. Introduction ⁎ Corresponding authors. E-mail addresses: [email protected] (N. Chen), [email protected] (C. Feng).

https://doi.org/10.1016/j.scitotenv.2019.134161 0048-9697/© 2019 Elsevier B.V. All rights reserved.

Nitrogen-containing refractory organic wastewater (NOW) is a common type of wastewater, which is widely generated from domestic

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waste discharges (Liang et al., 2018), aquaculture wastewater (He et al., 2019), toilet wastewater (Jasper et al., 2017), and landfill leachate (Fernandes et al., 2015). The direct discharge of NOW may reduce the oxidation–reduction potential of the receiving environment (Pan et al., 2019), induce eutrophication (He et al., 2019), trigger regional ecological environment problems, and endanger human health (Wang et al., 2019). At present, there are various methods including physicochemical processes, chemical processes, and biological processes for the removal of NOW, each with its own advantages, disadvantages, and applications scope. Moreover, it has also been confirmed that the microbial technology is feasible for the removal of nitrogen and organic matters (Im et al., 2001). However, the performance of microbial technology is vulnerable to influent water quality (e.g. C/N ratio and acidity–alkalinity) and seasonal temperature (Xiaomeng Zhang et al., 2018), making it difficult to obtain stable effluent quality. Furthermore, it is difficult to establish large-scale centralized biological treatment systems in developing regions (Dodane et al., 2012). Recently, electrochemical processes (e.g. electrochemical oxidation (Pérez et al., 2012), electrochemical reduction (Kuang et al., 2016), electro-coagulation (Ilhan et al., 2008), electro-Fenton (Fernandes et al., 2017), and electro-flotation (Alam and Shang, 2017)) have been proven to be suitable approach for NOW treatment owing to their high treatment efficiency, stable operation, and strong adaptability toward environment (Rocha et al., 2017). However, two difficulties are encountered in the current electrochemical treatment process of NOW, as follows: First, although the available chlorine produced by electrochemical oxidation is conducive to the removal of chemical oxygen demand (COD) and ammonium-N, it inhibits the cathode reduction of nitrite-N and nitrate-N (Li et al., 2010), indicating that simultaneous removal of total nitrogen (TN) and COD from NOW through a single electrolytic cell is difficult. In the previous studies, method involving connection of multiple electrolytic cells have been used as the main method for simultaneously removing TN and COD (Feng et al., 2003; Fernandes et al., 2017; Deng et al., 2018; Deng et al., 2019b). Moreover, simultaneously connecting two different anodes in single electrolytic cell can increase the removal efficiency toward COD and ammonium-N (Ding et al., 2018). Noteworthy, from the perspective of reaction engineering, the summation of mass transfer efficiency among individual reactors of combined electrochemical system based on connecting multiple electrolytic cells is often lower than that of an integrated single electrolytic cell (Scott Fogler, 1987). Moreover, one reactor is easier to operate than multiple reactors, and the removal efficiency is also significantly higher (Ding et al., 2018). Furthermore, transportation and maintenance of the single reactor is convenient. Therefore, the research and development of electrochemical reactor based on the simultaneous removal of TN and COD in a single electrolytic cell containing two different anodes is crucial for promoting the practical application of electrochemical technology in the field of NOW treatment. However, few studies have reported the simultaneous removal of TN and COD from NOW by constructing a single electrolytic cell reactor. Second, although the flocs produced by the electrocoagulation process are conducive to the separation of macro-molecular hydrophobic organic matters and rapid removal of COD from the liquid phase, the gravity sedimentation of electro-coagulated precipitates is slow. This indicates that long-term solid–liquid separation has a significant impact on the overall efficiency of wastewater treatment (Zhang et al., 2012; Lv et al., 2019). In the past, the sedimentation of electro-coagulated precipitates was accelerated by adding oxidants (e.g. ferrate (Ciabatti et al., 2010) and hypochlorite (Deng et al., 2019b)), coagulants (e.g. polyaluminum chloride (Deng et al., 2011), and polyferrous sulphate (Verma et al., 2012)). Unfortunately, the use of additional chemicals inevitably leads to difficult maintenance, high operating costs, as well as the risk of secondary contamination. In this study, a new alternating-anode electrochemical system (AAES) was proposed to simultaneously remove TN and COD from NOW in a single electrolytic cell. Moreover, the magnetic performance

of electro-coagulated precipitates was obtained by AAES operation process, and these precipitates could be easily and rapidly separated by using external magnetic field. The biologically treated landfill leachate (BTLL) was selected as a practical NOW to evaluate the performance of AAES, and the feasibility of AAES to remove TN and COD was demonstrated. The magnetic performance of electro-coagulated precipitates was explored. Furthermore, the formation of oxidants in AAES system and the changes of organic matters were also discussed in detail to support the mechanism and practical application of AAES. 2. Materials and methods 2.1. Samples and materials The BTLL was collected from Asuwei Refuse Sanitary Landfill (Beijing, China) and it was stored in a polypropylene bucket and kept at 4 °C prior to utilization. The characteristics of BTLL used in this study are listed in Table 1. The reagents used in this study were analytical grade or higher, and were purchased from Sinopharm Chemical Reagent Co., Ltd. (Shanghai, China). Solutions were prepared using ultrapure water. Iron anode, Ti/RuO2 anode, and iron net cathode, all with the surface area of 51.6 cm2 (12 cm × 4.3 cm), were purchased from a local factory in Beijing, China. The selection of mesh cathode was mainly due to its advantages including improvement in electrolyte flow, current distribution, and mass transfer in the reactor (Vilar and Coeuret, 1995). 2.2. Reactor setup of alternating-anode electrochemical system The reactor of AAES was designed in a single cycle electrolytic cell, and two anodes (iron and Ti/RuO2 anodes) were placed in the reactor as shown in Fig. 1. An iron mesh cathode was placed in the middle of the two anodes and the distance between each plate was 1.5 cm. A double-throw switch was used to control the switching between the two anodes. Direct current (DC) regulated power supply (KXN-645D, Zhaoxin Company, China) with a voltage range of 0–64 V and a current range of 0–5 A was used for the electrochemical reaction. During electrolysis, peristaltic pump (LongerPump, China) was used to circulate the electrolyte in the reactor in order to ensure the uniformity of electrolyte. 2.3. Experimental procedure BTLL (30 mL) were placed in the AAES reactor and then they were circulated using peristaltic pumps (LongerPump, China) at a flow rate of 40 mL min−1. The first working anode was iron anode, and then the system was switched to the DSA anode. The switching period between the anodes was set as 20 min. Electrolytic process was carried out at different current densities (25, 50, and 75 mA cm−2). Aliquotes of samples were taken out from the reactor at a fixed time interval (0, 20, 40, 60, 80, 100, 120, and 140 min) and filtered through 0.45 μm membrane filter. Each test was conducted in duplicate and their standard deviations were reported in order to minimize the experimental errors. Table 1 The Characteristics of BTLL. Property

BTLL

pH ORP (mV) COD (mg/L) TOC (mg/L) Cl− (mg/L) TN (mg/L) Ammonium-N (mg/L) Nitrite-N (mg/L) Nitrate-N (mg/L)

8.1 ± 0.6 −63.4 ± 5.2 2004.2 ± 58.3 678.8 ± 6.6 3012.2 ± 6.5 944.3 ± 6.3 159.7 ± 8.5 621.1 ± 9.7 59.6 ± 1.2

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organic matter and desorbed from the floc surface into the solution. This is discussed in detail in Section 3.4. When the chloride ions exist in the solution, the formation of available chlorine is an important support for the electrochemical oxidation process because indirect oxidation based on available chlorine can remove pollutants (e.g. COD and ammonium-N) more effectively than direct oxidation that occurs mainly on the surface of anode (Fernandes et al., 2015). In general, free chlorine (Cl2) is generated by the anode oxidation of chloride ion on the surface of anode (Eq. (1)) (Scialdone et al., 2009), and then it reacts with water and promote the formation of available chlorine (Eq. (2)) (Alfassi et al., 1987) as follows: −

Fig. 1. Schematic diagram of alternating-anode electrochemical system (AAES).

2Cl →Cl2 þ 2e−

ð1Þ −

Cl2 þ H2 O→2Hþ þ ClO þ Cl

−

ð2Þ

2.4. Analytical methods The analytical methods used in this study are detailed in Supporting Information. 3. Results and discussion 3.1. Chemical oxygen demand removal The change of COD concentration during electrolysis is shown in Fig. 2a. A significant decrease of COD was observed due to the flocculation of iron flocs dissolved from the iron anode. At all experimental current densities, about 50% of total COD was removed within the first 20 min of electrolytic process. In this process, the removal rate of COD increased with increasing current density, because higher current density promoted the formation of more iron flocs (Top et al., 2011). Noteworthy, the increase of COD concentration was observed using DSA anode during electrolysis (for example, at electrolysis time of 20–40 min and current density of 75 mA cm−2). This phenomenon was almost consistent with the change of TOC concentration (Fig. 2b). It was thus interpreted that the macro-molecule organic matter adsorbed on the surface of iron flocs was oxidized to smaller-molecule

The oxidation of chloride ion strongly depends on the material of anode (Zhang et al., 2017). Fig. 2c shows the linear sweep voltammetry (LSV) curve of iron anode, exhibiting that the anode current remains almost same when the anodic potential exceeds 0.5 V. This indicates that the electron transfer was mainly used for the Fe dissolution process. In general, potential of 0.5 V was less than the chlorine emission potential (ca. 1.359 V (Neodo et al., 2012)); therefore, it was difficult for iron anode to produce high concentration available chlorine. The Tafer curve of iron anode showed that the dissolution potential of Fe2+ in BTLL was higher than that in sodium chloride (NaCl) solution. In fact, this process can be explained in terms of the strong electron-holding capacity and electron shuttle capacity of dissolved organic matter (DOM) in BTLL (Yang et al., 2016). The electron shuttle properties of DOM may make it an important electron transfer medium, capable of participating in the electron transfer process on the anode plate before it gets mineralized. Therefore, DOM adsorbed on the anode surface can be regarded as an electrode parallel to the anode, thereby increasing the corrosion potential of the Fe anode. Moreover, the starting potential of LSV obtained in BTLL (−0.9 V) was lower than that in NaCl solution (−0.2 V), and higher current value was observed in BTLL. This phenomenon also indicated the organic matter adsorbed on the surface of iron

Fig. 2. (a) the change of COD during electrolysis in AAES; (b) the change of TOC during electrolysis in AAES; (c) the linear sweep voltammetry and Tafer curves of iron anode; (d) the change of available chlorine during electrolysis in AAES; (e) the change of hydroxyl radical during electrolysis in AAES (Current density: 70 mA/cm2); (f) the change of dissolved oxygen during electrolysis in AAES (Current density: 70 mA/cm2).

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anode could participate in electron transfer, and this might also be an important reason for the formation of available chlorine under the condition of iron anode. However, the generated available chlorine was difficult to detect under the condition of iron anode because it can be drastically removed by Fe2+ (Eq. (3)) (Folkes et al., 1995) and ammonium-N (Eq. (9)) (Li et al., 2010). −

2þ

ClO þ Fe

3þ

þ H2 O→Fe

–

−

þ HO  þHO þ Cl

ð3Þ

In this study, a significant negative correlation was observed between the concentration of available chlorine and ammonium-N during electrolysis, which is discussed in detail in Section 3.2. This result is consistent with the results of previous study (Jasper et al., 2017), which reported that the accumulation of available chlorine was observed when ammonium-N was almost consumed. It has been confirmed that available chlorine can rapidly oxidize ammonium-N, some reductive ions (e.g. Fe2+ and nitrite-N), and organic nitrogen (Deborde and von Gunten, 2008). Moreover, the removal of ammonium-N competed with the removal of organic matter (Fernandes et al., 2017). Therefore, at the current density of 70 mA cm−2 during the 100–120 min of electrolysis, the concentration of TOC showed a significant downward tendency because of the complete removal of ammonium-N. BTLL was the effluent from biological aerated filter of landfill leachate treatment, thus it contained certain concentration of dissolved oxygen. The change in concentration of dissolved oxygen during electrolysis is shown in Fig. 2f. Under the condition of iron anode, dissolved oxygen was hardly detected in solution. On the one hand, it was easy for Fe2+ released from iron anode to consume dissolved oxygen in solution. On the other hand, consumption of dissolved oxygen promoted the formation of hydrogen peroxide (H2O2) near the cathode during electrolysis (Eq. (4)) (Rocha et al., 2017). The generation of H2O2 was conducive to the initiation of Fenton reaction. The presence of hydroxyl radicals was detected when the iron electrode was used as anode (Fig. 2e), mainly due to the electro-Fenton process (Eq. (5)) (Özcan and Özcan, 2018). Under the condition of DSA anode, based on Comninellis theory (Comninellis, 1994; Ukundimana et al., 2018), the electrolysis of water on the surface of metal oxide (MOx) anode was bound to go through the following steps: first, water was electrolyzed by anode catalysis to produce adsorbed hydroxyl radicals (Eq. (6)). Second, the hydroxyl radicals reacted with each other to form molecular oxygen to complete the electrolysis of water (Eq. (7)). In this study, the accumulation of dissolved oxygen in solution also confirmed the interaction among adsorbed hydroxyl radicals on the surface of DSA anode. However, only low concentration of hydroxyl radicals could be detected under the condition involving DSA anode. It was mainly attributed to the fact that the chloride ion and available chlorine could also react with hydroxyl radicals (Eqs. (8)–(11)), resulting in inhibition of their accumulation (Lutze et al., 2015; Ji et al., 2017). Concentration of dissolved oxygen is an important factor affecting the concentration of hydroxyl radicals during electro-Fenton process (Özcan and Özcan, 2018). However, high concentration of hydroxyl radicals was detected in this study even though dissolved oxygen concentration was very low under iron anode condition at 80–100 min and 120–140 min of electrolysis, respectively. This indicated that the generation of hydroxyl radical had other process besides electro-Fenton process. In this study, formation of these hydroxyl radicals was mainly attributed to the reaction of available chlorine with Fe2+ (Eqs. (3) and (12)). This result is consistent with the result of study by Folkes et al. (1995). The formation of oxidants (available chlorine and hydroxyl radicals) promoted the mineralization and decomposition of organic matter and the removal of COD. The role of available chlorine and hydroxyl radicals in promoting the evolution and degradation of organic matter is discussed in Section 3.4. O2 þ 2Hþ þ 2e− →H2 O2

ð4Þ

Fe2þ þ H2 O2 →Fe3þ þ HO– þ HO

ð5Þ

MOx þ H2 O→MOx ðOHÞads þ Hþ þ e−

ð6Þ

2MOx ðOHÞads →2MOx þ O2 þ 2Hþ þ 2e−

ð7Þ

−

Cl þ HO→Cl  þOH–

ð8Þ

−

ClO þ HO→ClO þ OH– ClO2 Cl2

−

−

ð9Þ

þ HO  →ClO2 þ OH–

ð10Þ

−

þ HO  →HClO þ Cl

ð11Þ −

HClO þ Fe2þ →Fe3þ þ HO  þCl

ð12Þ

3.2. Removal of total nitrogen The concentrations of TN, nitrate-N, nitrite-N, and ammonium-N during electrolysis in AAES are shown in Fig. 3. Under the condition of iron anode, cathode reduction was the main reaction responsible for removal of nitrite-N (Deng et al., 2018). In general, the main products of cathode reduction of nitrite-N were nitrogen gas (Eq. (13)) and ammonium-N (Eq. (14)) (Takaoka et al., 2005; Garcia-Segura et al., 2018). Moreover, Under the condition of DSA anode, the available chlorine was produced during anode oxidation process of chloride ion, which led to rapid oxidation of ammonium-N into nitrogen gas (Li et al., 2010). Therefore, the most intuitive reactions were NO− x -N + → NH+ 4 -N (iron anode) and NH4 -N → N2 (DSA anode), respectively, thus the subsequent anodic alternation process could achieve the purpose of removing TN. 2NO2 − þ 8Hþ þ 6e− →N2 þ 4H2 O E0 ¼ 1:15 V vs SHE

ð13Þ

NO2 − þ 8Hþ þ 6e− →NH4 þ þ 2H2 O

ð14Þ

E0 ¼ 0:90 V vs SHE

In this study, we also observed some interesting phenomenon. At the current density of 25, 50, and 75 mA cm−2, the decrease in concentration of nitrite-N was 29.5 ± 0.7 (from 620.2 ± 0.4 to 590.7 ± 0.3), 146.9 ± 1.1 (from 620.9 ± 0.7 to 474.0 ± 0.4), and 302.8 ± 0.9 (from 623.8 ± 0.4 to 321.0 ± 0.5) mg/L, respectively; and the increase in concentration of ammonium-N was 78.2 ± 2.1 (from 152.0 ± 0.9 to 230.2 ± 1.2), 123.6 ± 3.1 (from 158.2 ± 0.5 to 281.8 ± 2.6), and 163.8 ± 2.7 (from 169.0 ± 1.4 to 332.8 ± 1.3) mg/L, respectively. At the current density of 25 mA cm−2, the increase of ammonium-N was higher than the decrease of nitrite-N, indicating the occurrence of other pathways in the production of ammonium-N in addition to nitrite-N reduction. The most direct process was the decomposition of organic nitrogen. It was thus confirmed that ammonium-N was the main product of the reaction of hydroxyl radicals with organic nitrogen (Maletzky and Bauer, 1998; X. Wang et al., 2015). With the increase of current density, the removal of nitrite-N increased sharply, while the accumulation of ammonium-N was obviously less than the removal of nitrite-N. First, when the pH was higher than 9.0, the cathode reduction of nitrite-N could also lead to the formation of small amount of nitrogen (Dash and Chaudhari, 2005). Second, Fe2+ released from the iron anode could reduce nitrite-N into ammonium-N and nitrogen gas (Eqs. (15) and (16)). However, the ΔG obtained by using Eq. (16) (162 kJ mol−1) was lower than that obtained by using Eq. (15) (245 kJ mol−1) (Melany, 2014; Stüeken et al., 2016), which indicated that nitrogen was the main product in the reduction of nitrite-N by Fe2+. It was observed that the strong reactivity of the available chlorine and nitrite-N led to the transformation of nitrite-N to nitrate-N when the DSA anode was used as working anode (Pressley et al., 1972). In fact, the reaction rate between the available chlorine and nitrite-N

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Fig. 3. The changes of TN, nitrate-N, nitrite-N and ammonium-N during electrolysis in AAES.

was faster than that of ammonium-N, because of the slight increase in ammonium-N at current density of 25 mA cm−2 during 20–40 min of electrolysis, while the conversion of nitrite-N into nitrate-N was obviously observed. 6Fe2þ þ NO2 − þ 16H2 O→6FeðOHÞ3 þ NH4 þ þ 10Hþ

ð15Þ

6Fe2þ þ 2NO2 − þ 14H2 O→6FeðOHÞ3 þ N2 þ 10Hþ

ð16Þ

Furthermore, besides the two above mentioned reasons for decreased ammonium-N accumulation, the generation of nitrogen gas promoted by the suitable oxidation–reduction potential (ORP) was also an important reason for the higher TN removal efficiency of AAES than that of traditional multi-cell reactor. It was confirmed that a strong reduction environment (ORP b -463 mV) could be formed by single iron anode as working anode for nitrate-N removal. Under this condition, the + conversion of NO− x -N to NH4 -N did not lead to the formation of intermediate products (Dash and Chaudhari, 2005). Therefore, excessively + low ORP led to a very high conversion efficiency of NO− x -N to NH4 -N, which was not conducive to TN removal. In order to achieve high TN removal efficiency, it was necessary to convert NO− x -N to N2 as much as possible during cathode reduction process. In this study, owing to the consumption of excess electrons from solution by the oxidation of DSA anode during anode alternation process, ORP of the solution was effectively controlled between −100 to 200 mV during 0–100 min of electrolysis. It was advantageous to promote the conversion of NO− x -N into N2 in this ORP range (Dash and Chaudhari, 2005). Therefore, AAES exhibited advantages in increasing removal efficiency toward TN.

This study further evaluated the removal of TN and COD by AAES from the perspective of energy consumption. In general, specific energy consumption (Esp) can be used to characterize the energy consumption (Fernandes et al., 2016). The formula used for calculation is as follows (Eq. (17)): Esp ¼

UIΔt 3:6VΔPollutants

ð17Þ

where U is electrode potential (V), I is applied current (A), Δt is electrolysis time (s), V is the volume of electrolytic cell (L), and the ΔPollutants is the removal concentration of contaminants (which was TN or COD in this study) (mg L−1). The Esps of TN and COD after 120 min of electrolysis are listed in Table 2. The Esp of COD removal by AAES was slightly higher than that by double-cell electrochemical system, which was mainly due to the release of organic matter from precipitates during the process of anode exchange. This result is consistent with the slight increase of TOC reported in Section 3.1. On the other hand, it was exciting that AAES could significantly reduce the Esp of TN removal and enhance TN removal efficiency at the same current density. In general, cathode reduction of NO− x -N was negatively affected by several factors such as the inhibition of available chlorine (Li et al., 2010) and hydrogen evolution side reaction (Kuang et al., 2016)), and reduced the efficiency of NO− x -N reduction. Moreover, the removal of TN was usually a speed-limiting step for electrochemical treatment of NOW, which indicated that TN removal would consume more energy. On the premise of achieving the same treatment effect, due to the high efficiency of TN removal, the energy consumption could be reduced, which made AAES have a significant advantage over the two-cell reactor in total energy consumption.

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Table 2 the Esps of TN and COD removal by AAES and double electrolyzer electrochemical reactor.

AAES

Double electrolyzer

Current density (mA/cm2)

Target wastewater

TN removal efficiency (%)

TN removal Esp Wh/(g-TN)

COD removal efficiency (%)

COD removal Esp Wh/(g-COD)

Reference

25 50 75 30 50 70

BTLL BTLL BTLL BTLL BTLL BTLL

60.1 95.3 98.8 59.7 88.2 96.1

72.2 179.3 387.1 233.3 432.1 778.2

56.8 59.9 71.0 73.2 77.7 82.1

36.8 137.9 243.0 35.6 93.2 172.8

This study This study This study Deng et al., 2018 Deng et al., 2018 Deng et al., 2018

3.3. The magnetic formation of precipitates Fig. 4a and b show the change of X-ray diffraction (XRD) pattern and ORP-pH diagram of precipitates obtained at different electrolysis time in AAES. The ORP-pH diagram demonstrates that when Fe2+ was released from the iron anode, it was converted to Fe(OH)+ by the redox environment of AAES. In general, the green iron compound (green rust) can be rapidly produced by the evolution of Fe(OH)+ under the condition of pH N 8.0 (Ochi et al., 1981). Green rust is a common metastable intermediate product of iron in anaerobic environment (Wang et al., 2013). In this study, the formation of green rust showing XRD characteristic of amorphous iron compounds was observed during 0–20 min and 40–60 min of electrolytic reaction. The formation of amorphous iron oxides can be attributed to the incomplete electron exchange of Fe2+–Fe3 + in the presence of organic matters (Bhattacharyya et al., 2018). The evolution of fixed crystalline iron compounds and amorphous iron compounds occurred during the anodic alternation process. The crystallinity index (CI) is defined as the volume fraction of crystallinity of one phase in a given sample (Reyes-Gasga et al., 2013). At present, the CI is usually evaluated by XRD, and the equation is presented as follows (Eq. (18)) (Marco et al., 2018): CI ¼ 1–AAM =ATOT

ð18Þ

where AAM is the area of the amorphous iron compounds peak and ATOT is the total area of all peaks. With the oxidation of green rust, increasing CI was observed during electrolysis (Fig. 4c). This result is consistent with the result of a previous study (Z. Wang et al., 2015), which reported that the oxidation of green rust could achieve the formation of iron oxides with stable crystalline state. In this study, the oxidation of green rust promoted the formation of Fe3O4 (strong magnetism) and γ-Fe2O3 (strong magnetism). The produced Fe3O4 can be interpreted as octahedral position of iron compounds occupied both by divalent ions Fe2+ and partially by Fe3+ (Kalska-Szostko et al., 2014). Generation of Fe3O4 was exciting because Fe3O4 is currently considered to be almost cytotoxic (Lindley, 1996; Kalska-Szostko et al., 2014). Moreover, the strong magnetism of Fe3O4 was conducive to the separation of the precipitates from the solution. The formation of Fe3O4 was significantly influenced by the pH. When the pH of the solution exceeded 10.0, the oxidation of Fe(OH)2 directly promoted the formation of Fe3O4 (Tamaura, 1981). Furthermore, it has also been confirmed that the accumulation of Fe3O4 was inhibited under the condition of pH N 13.0 (Z. Wang et al., 2015). In this study, the pH was below 13.0 during electrolysis (Fig. 4c), which was suitable for the formation of Fe3O4. Besides Fe3O4, magnetic γ-Fe2O3 was also detected in the XRD spectrum during electrolysis, which was attributed to the oxidation of Fe3O4 under DSA anode conditions (Ochi et al., 1981). Noteworthy, Fe3O4 and γ-Fe2O3 showed similar crystal surface shapes

Fig. 4. (a) XRD analysis of crystal form evolution of precipitate during electrolysis; (b) ORP-pH diagram in AAES during electrolysis; (c) the crystallinity index of precipitates during electrolysis process in AAES; (d) the Raman spectrums of the precipitates from the condition of DSA anode (120 min) and iron anode (140 min); (e) the VSM analysis of the precipitates from the condition of DSA anode (120 min) and iron anode (140 min). (Current density: 75 mA/cm2).

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in the XRD spectrum. However, the color of the precipitates obtained under iron anode was significantly different from that obtained under DSA anode, thus it was necessary to further distinguish Fe3O4 from γFe2O3. By comparing the Raman spectra of the precipitates from DSA anode (120 min) and iron anode (140 min) (Fig. 4d), it was confirmed that when the DSA electrode was used as working anode, the mixture of Fe3O4 and γ-Fe2O3 constituted the main crystal forms, while Fe3O4 was the main crystal form obtained from the iron electrode as a working anode. Vibrating sample magnetometer (VSM) analysis revealed that the magnetic property of precipitates produced by the iron anode was significantly higher than that produced by the DSA anode (Fig. 4e). On the one hand, Fe3O4 has stronger magnetism than γ-Fe2O3 (Zhang et al., 2016). When iron electrode was used as the anode, reduction based on Fe2+ promoted the transformation of γ-Fe2O3 to Fe3O4, and led to the enhancement of magnetism (Drits, 1995; Wang et al., 2005). On the other hand, although no significant α-Fe2O3 (nonmagnetic) formation was observed during the reaction, further oxidation of γ-Fe2O3 might result in the formation of α-Fe2O3, thereby leading to the loss of magnetism (Drits, 1995). Therefore, selection of the iron electrode as the final use of the anode is more favorable for the formation of magnetic precipitates that are easily separated by using the external magnetic field. 3.4. The transformation in dissolved organic matter Three-dimensional fluorescence spectroscopy combined with parallel factor analysis was applied to discriminate the components of BTLL (Supporting Information Fig. S1). As a result, four peaks of humic-like substances (Component 1 (Ex/Em = 350/450 nm), Component 2 (Ex/ Em = 325/400 nm), Component 3 (Ex/Em = 300/450 nm), and Component 4 (Ex/Em = 425/475 nm)); and one peak of fulvic-like substances (Component 5 (Ex/Em = 250/425 nm)) were detected (He et al., 2011; He et al., 2013). In general, humic-like substances and fulvic-like substances are collectively referred to as humus-like substances. These substances are similar to the organic matter ultimately produced during the biogeochemical cycles (Xi et al., 2016). The biodegradable organic matter of landfill leachate can be utilized by microorganisms in microbial reactors and the remaining organics in BTLL are mainly bio-refractory organics with high stability and aromaticity. This indicates that these organic matter substances may exist for a long time, and they are also associated with increasing cytotoxicity to the environment (SaldañaRobles et al., 2017; Yu et al., 2018). Fluorescence spectra of DOM during electrolysis are shown in Fig. 5. Significant blue shift of the peak related to the humic-like substances was observed from Ex of 425 nm to Ex of 325 nm in 0–20 min. The surface of humic-like substances and fulvic-like substances was negatively charged (Ulu et al., 2014) and those substances were easily adsorbed by positively charged iron hydroxides (e.g. Fe(OH)+ and Fe(OH)+ 2 ). Moreover, it was also confirmed that humic-like substances generally have higher molecular weights than fulvic-like substances, and their precipitation is more preferable because of their higher ability to react with iron hydroxides and co-precipitates (Dia et al., 2018). During 20–40 min of electrolysis, the DOM concentration in region V and III slightly increased, similar to the concentration of TOC (Fig. 2b), indicating that some humic-like substances were desorbed from the precipitates. Moreover, some humic-like substances were converted into fulvic-like substances in solution or on the surface of precipitates. The correlation analysis between various types of DOM and oxidants during electrolysis is shown in Fig. 6c. The strong statistical correlation among various concentrations of DOM and oxidants indicated that the principal component analysis (PCA) was suitable for assessing the relationship between oxidants formation and DOM removal (Zhao et al., 2018). Therefore, PCA was performed to reduce the data dimension and describe the degradation process of DOM more clearly. As a result, the generation of available chlorine promoted the formation of region

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(I + II + III) organic matter. Usually, the available chlorine is mainly produced by DSA anode and it can oxidize humic-like substances into small-molecule organic matter, which can be easily desorbed from the surface of precipitates. Therefore, increase in TOC concentrations was detected when using DSA anode. Accumulation of hydroxyl radicals mainly occurred when using iron anode, and the conversion of organic matter from region V to region IV was promoted by hydroxyl radicals. The organic matter of region IV showed a negative correlation with available chlorine, indicating that the available chlorine oxidation might lead to the transformation of region IV organic matter to region (I + II + III) organic matter. Moreover, the region (I + II + III) organic matter showed a negative correlation with hydroxyl radicals, which indicated that hydroxyl radicals promoted further mineralization of region (I + II + III) organic matter. In order to further elucidate the role of two oxidants, namely available chlorine and hydroxyl radicals, in the degradation process of BTLL, canonical correspondence analysis (CCA) or redundancy analysis (RDA) was used to reveal the contribution of two oxidants in this electrochemical system. First, detrended correspondence analysis (DCA) was performed to determine the selection of CCA or RDA. In this study, length of the first DCA ordination axis was 4, which indicated that CCA was suitable for evaluating the relationship between the above mentioned random variables (Xiaohui Liu et al., 2018). Contribution efficiency of hydroxyl radicals and available chlorine toward COD removal and organic evolution was calculated by CCA analysis. After 999 times of Monte–Carlo model tests, it was found that hydroxyl radicals were not significant (P = 0.507), while available chlorine was significant (P = 0.094), clearly demonstrating that available chlorine played a greater role in promoting the evolution of BTLL organic matter and COD removal compared to hydroxyl radicals. This result is consistent with the result of a previous study (Fernandes et al., 2016), which reported that the available chlorine played a dominant role in promoting removal of organic matter from landfill leachate when the chloride ion was present in solution. However, significant negative correlation was observed between hydroxyl radicals and Region V organic compounds observed in CCA diagram; therefore, it could not be said that the oxidation of hydroxyl radicals was negligible. The possible reason was that the concentration of hydroxyl radicals produced was low compared to that of available chlorine. On the one hand, the production of hydroxyl radicals was affected by the presence of chloride ion. On the other hand, the generation of hydroxyl radicals was limited by the low concentration of dissolved oxygen. The material composition, retention time, peak area, and peak percentage of initial BTLL and the effluent were qualitatively analyzed by gas chromatography–mass spectrometry (GC–MS) (Supporting Information Figs. S2 and S3). As a result, 40 types of organic compounds were detected, and their molecular structures were correlated with the humic acid-like substances. The result was similar to the research of Hui et al. (2019), which reported that the organic components in initial BTLL included alkanes, olefins, heterocyclic compounds, aromatic hydrocarbons, phenols, esters, ketones, acids, and aldehydes. Moreover, the formation of carboxylic acid and alkane during electrolysis was observed in previous studies (Xiao et al., 2013; Deng et al., 2019b; Kateb et al., 2019), and this phenomenon accompanied the increasing biodegradability. In this study, the main organic components obtained from effluent were alcohols (e.g. 1-nonanol and 1-hexanol), aldehyde and ketones (e.g. hexanal, 3-heptanone, and benzaldehyde), carboxylic acids and halogenated hydrocarbons (e.g. 3-chloropropionic acid, octyl, and trichloromethane). The formation of alcohols, aldehydes, and acids can be attributed to the oxidation of macro-molecule humic-like substances by available chlorine and hydroxyl radicals, which can also promote the breakage of macro-molecule (Deborde and von Gunten, 2008). In fact, the oxidation of organic matter was carried out in the following order: alcohols (C–OH, sp3), aldehydes (H–C=O, sp2), acids (–COOH, sp2), and CO2 (O=C=O, sp) (Yan Zhang et al., 2018), and it was also the main reason for the simultaneous existence of alcohols, aldehydes,

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Fig. 5. The fluorescence spectra of dissolved organic matters during electrolysis in AAES (Current density: 75 mA/cm2).

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Fig. 6. (a) The concentration of dissolved organic matter in different fluorescence regions; (b) the percentage of dissolved organic matter in different fluorescence regions; (c) correlation analysis of oxidants and organic matter evolution; (d) principal component analysis of oxidants and organic matter evolution; (e) canonical correspondence analysis of oxidants and organic matter evolution; (f) the GC–MS spectrum of BTLL obtained at 0 and 140 min electrolysis. (Current density: 75 mA/cm2).

carboxylic acids, and hydrocarbons in this study. In our previous studies (Deng et al., 2018; Deng et al., 2019a, 2019b), the final effluent products of BTLL degradation by electrochemical oxidation were small molecules of carboxylic acid, aromatic acid, and hydrocarbon, and the effluent cytotoxicity was significantly reduced by electrolytic process. If available chlorine was eliminated, significant cell growth would be detected during microbial culture using this organic matter in effluent (Deng et al., 2019b). Therefore, the organic matter can be utilized by microorganisms as carbon source, and it can be further degraded by natural attenuation in the environment (Ying Liu et al., 2018). 3.5. Mechanism analysis When the iron electrode was used as working anode, production of available chlorine was low because the dissolution of Fe2+ was the major electrode reaction of iron anode with the anode potential exceeding 0.5 V. In this process, nitrate-N and nitrite-N of BTLL were mainly reduced to ammonium-N and nitrogen gas. Noteworthy, compared to the traditional two-cell reactor (electro-coagulation + electrochemical oxidation), AAES could create a suitable redox environment (−100 mV

b ORP b 200 mV) to increase the efficiency of nitrate-N reduction to N2, thus improving the removal efficiency of TN (process (1) in Fig. 7). Electro-coagulation of iron flocs produced by dissolving of the iron anode promoted the separation of macro-molecular humic acid-like substances from solution, which resulted in achievement of rapid decrease of TOC (process (2) in Fig. 7). Furthermore, the hydroxyl radicals produced by the reaction of available chlorine with Fe2+ and by electroFenton process also promoted the removal of COD (process (3) in Fig. 7). When the DSA electrode was used as working anode, high concentration of available chlorine was generated, which also promoted the removal of ammonium-N (process (5) in Fig. 7). Therefore, the removal of TN could be achieved by alternating the two anodes in AAES. CCA showed that available chlorine and hydroxyl radicals were the main oxidants produced in AAES, and the available chlorine played a greater role than hydroxyl radicals in promoting the evolution of DOM and the removal of COD in AAES. Furthermore, the crystallinity of iron compounds was increased by the oxidation of amorphous iron compounds produced from iron anode due to the appropriate pH and redox environment in AAES. The

Fig. 7. The mechanism diagram of AAES for BTLL treatment.

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main iron compounds with stable crystallinity were Fe3O4 and γ-Fe2O3, which were magnetic in nature and could be separated rapidly by applying external magnetic field. Therefore, AAES provided a new approach for rapid separation of electro-coagulation precipitates. Noteworthy, the selection of iron electrode as final working anode was necessary because the residual available chlorine could also be removed by iron anode (Eq. (3)), thereby reducing the ecological risk of discharge. Furthermore, Fe3O4 was the main magnetic constituent under iron anode (process (4) in Fig. 7), which showed stronger magnetism than the main magnetic constituent γ-Fe2O3 obtained under DSA anode (process (7) in Fig. 7). This indicated that the precipitates produced by iron anode exhibited better magnetic separation performance than DSA anode. 3.6. Economic analysis and precipitates disposal suggestion In this study, the treatment cost of AAES at 50 mA cm−2 was $0.0135 per L. Other technologies for treatment of landfill leachate, such as coagulation/flocculation and microwave processes (C-F-MW) and solar photo-Fenton process have been reported with the corresponding treatment costs of $0.0123 (Tripathy and Kumar, 2019) and $0.0147 (Silva et al., 2016) per L, respectively. In fact, It has been confirmed that the main cost of electrochemical technology is power consumption (Dash and Chaudhari, 2005; Deng et al., 2019b). Electric energy can be compensated by solar energy and biomass energy which are easily available in natural environment (Yuan et al., 2017; Antolini, 2019), and provide a feasible alternative for further reducing the treatment cost of AAES. Moreover, the residual COD in the effluent of AAES was easy to be used by microorganisms, thus it could be further removed through phytoremediation (Sánchez et al., 2018) or small-scale constructed wetlands (Yuan Zhang et al., 2018). These methods are conducive to the improvement of the overall cost-effectiveness of the treatment process. Furthermore, undeniably, precipitates as solid waste by-products are produced by AAES during electrolytic process. Fortunately, compared to the reduction iron compounds (with strong cytotoxicity and difficult to deal with (Xiaobo Liu et al., 2018; Vaz et al., 2018)) obtained by iron anode electro-coagulation process, the precipitates obtained by AAES mainly include non-cytotoxic Fe3O4 and organic matter degraded by electrochemical oxidation. At present, it has been confirmed that the complexation ability, hydrophilicity, and acute cytotoxicity of organic matter after electrochemical oxidation treatment can be reduced (Deng et al., 2019a). This indicated that the precipitates obtained by AAES have the potential to be used as soil amendments for carbonlosing soil remediation, and it is of great significance to improve soil carbon content. 4. Conclusions The alternating-anode electrochemical system (AAES) constructed based on single electrolytic cell was demonstrated to be feasible for simultaneous removal of COD and TN from NOW (in this study, BTLL was selected as a typical NOW). The main conclusions of this study are as follows: + (1) Cathode reduction (NO− x -N → NH4 -N and N2) under iron anode + and indirect oxidation (NH4 -N → N2) under DSA anode were the main pathways to remove TN from BTLL. Owing to the appropriate oxidation– reduction condition (ORP: from −100 to 200 mV) created by alternating anode, the conversion of NO− x -N to N2 was increased which promoted the TN removal. (2) The main dissolved organic matter (DOM) in BTLL included humic-like substances. During the electrolytic process, humic-like substances were degraded into small-molecule DOM, such as fulvic-like substances and protein-like substances by available chlorine and hydroxyl radicals. The available chlorine played a greater role than

hydroxyl radicals in promoting the evolution of DOM and the removal of COD. (3) The main stable crystalline forms of precipitates obtained from AAES were Fe3O4 and γ-Fe2O3, which can be easily attracted by magnetic field and favorable for rapid solid–liquid separation. The findings of this study provide an important theoretical and experimental basis for the electrochemical advanced treatment of NOW in a single electrolytic cell as well as rapid separation of precipitates. Acknowledgments The authors acknowledge financial support from the National Natural Science Foundation of China (NSFC) (No. 51578519), the Major Science and Technology Program for Water Pollution Control and Treatment (No. 2017ZX07202002), and the Fundamental Research Funds for the Central Universities (No. 2652018185). Appendix A. Supplementary data Supplementary data to this article can be found online at https://doi. org/10.1016/j.scitotenv.2019.134161. References Alam, R., Shang, J.Q., 2017. Removal of bitumen from mature oil sands tailings slurries by electro-flotation. Journal of Water Process Engineering 15, 116–123. Alfassi, Z.B., Huie, R.E., Mosseri, S., Neta, P., 1987. Kinetics of one-electron oxidation by the cyanate radical. J. Phys. Chem. 91, 3888–3891. Antolini, E., 2019. Photoelectrocatalytic fuel cells and photoelectrode microbial fuel cells for wastewater treatment and power generation. Journal of Environmental Chemical Engineering 7, 103241. https://doi.org/10.1016/j.jece.2019.103241. Bhattacharyya, A., Schmidt, M.P., Stavitski, E., Martínez, C.E., 2018. Iron speciation in peats: chemical and spectroscopic evidence for the co-occurrence of ferric and ferrous iron in organic complexes and mineral precipitates. Org. Geochem. 115, 124–137. Ciabatti, I., Tognotti, F., Lombardi, L., 2010. Treatment and reuse of dyeing effluents by potassium ferrate. Desalination 250, 222–228. Comninellis, C., 1994. Electrocatalysis in the electrochemical conversion/combustion of organic pollutants for waste water treatment. Electrochim. Acta 39, 1857–1862. Dash, B.P., Chaudhari, S., 2005. Electrochemical denitrificaton of simulated ground water. Water Res. 39, 4065–4072. Deborde, M., von Gunten, U., 2008. Reactions of chlorine with inorganic and organic compounds during water treatment—kinetics and mechanisms: a critical review. Water Res. 42, 13–51. Deng, S., Zhou, Q., Yu, G., Huang, J., Fan, Q., 2011. Removal of perfluorooctanoate from surface water by polyaluminium chloride coagulation. Water Res. 45, 1774–1780. Deng, Y., Feng, C., Chen, N., Hu, W., Kuang, P., Liu, H., Hu, Z., Li, R., 2018. Research on the treatment of biologically treated landfill leachate by joint electrochemical system. Waste Manag. 82, 177–187. Deng, Y., Chen, N., Feng, C., Chen, F., Wang, H., Feng, Z., Zheng, Y., Kuang, P., Hu, W., 2019a. Research on complexation ability, aromaticity, mobility and cytotoxicity of humiclike substances during degradation process by electrochemical oxidation. Environ. Pollut. 251, 811–820. Deng, Y., Chen, N., Feng, C., Chen, F., Wang, H., Kuang, P., Feng, Z., Liu, T., Gao, Y., Hu, W., 2019b. Treatment of organic wastewater containing nitrogen and chlorine by combinatorial electrochemical system: taking biologically treated landfill leachate treatment as an example. Chem. Eng. J. 364, 349–360. Dia, O., Drogui, P., Buelna, G., Dubé, R., 2018. Hybrid process, electrocoagulationbiofiltration for landfill leachate treatment. Waste Manag. 75, 391–399. Ding, J., Wang, K., Wang, S., Zhao, Q., Wei, L., Huang, H., Yuan, Y., Dionysiou, D.D., 2018. Electrochemical treatment of bio-treated landfill leachate: influence of electrode arrangement, potential, and characteristics. Chem. Eng. J. 344, 34–41. Dodane, P.-H., Mbéguéré, M., Sow, O., Strande, L., 2012. Capital and operating costs of fullscale fecal sludge management and wastewater treatment systems in Dakar, Senegal. Environmental Science & Technology 46, 3705–3711. Drits, V.A., 1995. An improved model for structural transformations of heat-treated aluminous dioctahedral 2:1 layer silicates. Clay Clay Miner. 43, 718–731. Feng, C., Sugiura, N., Shimada, S., Maekawa, T., 2003. Development of a high performance electrochemical wastewater treatment system. J. Hazard. Mater. 103, 65–78. Fernandes, A., Pacheco, M.J., Ciríaco, L., Lopes, A., 2015. Review on the electrochemical processes for the treatment of sanitary landfill leachates: present and future. Appl. Catal. B Environ. 176–177, 183–200. Fernandes, A., Santos, D., Pacheco, M.J., Ciríaco, L., Lopes, A., 2016. Electrochemical oxidation of humic acid and sanitary landfill leachate: influence of anode material, chloride concentration and current density. Sci. Total Environ. 541, 282–291. Fernandes, A., Labiadh, L., Ciríaco, L., Pacheco, M.J., Gadri, A., Ammar, S., Lopes, A., 2017. Electro-Fenton oxidation of reverse osmosis concentrate from sanitary landfill leachate: evaluation of operational parameters. Chemosphere 184, 1223–1229.

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