The eAND process: Enabling simultaneous nitrogen-removal and disinfection for WWTP effluent

w a t e r r e s e a r c h 7 4 ( 2 0 1 5 ) 1 2 2 e1 3 1

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The eAND process: Enabling simultaneous nitrogen-removal and disinfection for WWTP effluent Jing Ding, Qingliang Zhao*, Yunshu Zhang, Liangliang Wei, Wei Li, Kun Wang State Key Laboratory of Urban Water Resources and Environment (SKLUWRE), School of Municipal and Environmental Engineering, Harbin Institute of Technology, Harbin 150090, China

article info

abstract

Article history:

To mitigate potential eutrophication risk caused by nitrogen species in the effluent of

Received 14 November 2014

wastewater treatment plant (WWTP), nitrogenous compounds failed to be removed dur-

Received in revised form

ing biological wastewater treatment should be further eliminated. In this paper, an

30 January 2015

electrochemical process for ammonia-oxidation, nitrate-reduction and disinfection

Accepted 3 February 2015

(eAND process) of WWTP effluent was developed and its performance for tertiary treat-

Available online 12 February 2015

ment of synthetic wastewater and actual effluent was evaluated. Results indicated ammonia and nitrate removal efficiencies in actual effluent reached 96% and 36% at

Keywords:

1.23 Ah l
1, while coliforms were totally inactivated at 0.072 Ah l
1 under the optimal

Ammonia

operation conditions. Ammonia removal due to the anodic indirect oxidation followed a

Nitrate

pseudo first kinetic, while the modified model expressed as exponential decay fitted well

Disinfection

to the experimental data with the presence of nitrate. The coliforms inactivation was

Electrochemical

attributed to the in situ generated active chlorine, indicating no extra addition of disin-

WWTP effluent

fectant. Nitrate reduction in cathodic area fitted to pseudo first order kinetic with kinetic constants of 0.13e0.54 l A
1 h
1. These results clearly showed the potential of this eAND process to serve as a tertiary treatment of WWTP effluent for simultaneous removal of ammonia, nitrate and disinfection. © 2015 Elsevier Ltd. All rights reserved.

1.

Introduction

Aqueous nitrogen species left in domestic wastewater effluent are major nutrient sources for environmental nature water (Pernet-Coudrier et al., 2012). As one of the most effective and economical processes in WWTP, nitrification/denitrification is greatly influenced by the ambient temperature. In many regions of the world, especially in the mid- and high latitudes,

* Corresponding author. Tel./fax: þ86 451 8628 3017. E-mail address: [email protected] (Q. Zhao). http://dx.doi.org/10.1016/j.watres.2015.02.005 0043-1354/© 2015 Elsevier Ltd. All rights reserved.

environmental temperature varied greatly throughout the year. It leads to that many conventional WWTPs have good performance of nitrogen removal in summer, whereas very poor in winter, as the activity of nitrifying bacteria reduced and the nitrification rates dropped significantly at temperature below 15  C (Sudarno et al., 2011). Thus, on the global scale, nutrient species existing in the WWTP effluent should be further removed to mitigate eutrophication risk when the performances of nitrification and denitrification are poor.

w a t e r r e s e a r c h 7 4 ( 2 0 1 5 ) 1 2 2 e1 3 1

Therefore, other methods with good performance and high efficiency for nitrogen removal for tertiary treatment of WWTP effluent are urgently needed. The secondary effluent is characterized by high total nitrogen (TN) and low organic matter, further removal of TN is difficult to be achieved by the conventional nitrification and denitrification processes (Zhao et al., 2013). TN in secondary effluent contains ammonia ðNH4 þ Þ, nitrate ðNO3
Þ, nitrite ðNO2
Þ and organic nitrogen (org-N), with the higher percentage of NO3
and NH4 þ (Simsek et al., 2012). There have already several potential methods such as membrane filtration, precipitation, flocculation and wetland ecosystem for tertiary removal of nitrogen from the effluent, however, some drawbacks like secondary pollution of concentrate, extra dosage of coagulant and negative influence by low ambient temperature are still concerned (Xiao et al., 2009; He and Xue, 2010). Regarding conventional disinfection treatments, chlorination and UV technologies are the most widely used methods (Cano et al., 2012). Electrochemical technologies have attracted lots of attention due to their advantages of high efficiency, no sludge production, independence of temperature, small area occupied and relatively low cost (Szpyrkowicz et al., 1995), which have been attempted to remove such pollutants as organic matter (Zhu et al., 2009), NH4 þ (Kim et al., 2006), turbidity (Cotillas et al., 2013), phosphorous (Mahvi et al., 2011), NO3
(Mook et al., 2012), sulfide (Pikaar et al., 2011) and coliforms (Cotillas et al., 2013) under different operational conditions. Among these electrochemical technologies, NH4 þ electrooxidation process has been widely studied and currently used in a diversity of wastewater treatments, such as landfill leachate (Deng and Englehardt, 2007), domestic effluent (Li and Liu, 2009), reverse osmosis concentrates (Bagastyo et al., 2011), coking wastewater (Zhu et al., 2009), tannery wastewater (Szpyrkowicz et al., 2005), and aquaculture wastewater (Dı´az et al., 2011). The NH4 þ present in wastewater can be oxidized to nitrogen through the in situ generated free chlorine (Cl2, HClO, ClO
), and the removal rate of NH4 þ is dependent on chloride concentration and electric charge (Kapałka et al., 2010a). Moreover, the formation of free chlorine in the process is also efficient in coliforms inactivation (Cano et al., 2012). The inactivation rate of Escherichia coli accelerates with increasing current density by a faster generation of electrochemical oxidants (Schmalz et al., 2009), which includes the chlorine and reactive oxygen species ($OH, O$, HO2$, O3 or H2O2). The type of electrochemical oxidants depends on anode material largely because of the different reaction pathways. Free chlorine was generated as predominant oxidants on conventional dimensionally stable anodes (DSA), while the efficiency of $OH production was in the order of boron-doped diamond (BDD)>>Ti/RuO2 (DSA) z Pt (Jeong et al., 2009). The anodic oxidation had also been tried to remove pathogenic microorganism as tertiary wastewater treatment (Frontistis et al., 2011). The electrochemical technologies have also been investigated to reduce NO3
on the cathode. During the electroreduction of NO3
, electrons on the cathode are provided by current and balanced by anodic water electrolysis. The operational conditions influence this process significantly, including current input, pH, chloride and sulfate anions

123

contained (Szpyrkowicz et al., 2006; Mattarozzia et al., 2013; Dortsiou et al., 2013). Nitrogen gas is the desired products of NO3
reduction, but NH4 þ and NO2
are usually formed, and the formation of NH4 þ is considered as the major limitation to efficient electrochemical denitrification (Mook et al., 2012). To overcome this issue, chloride is usually added and the oxidation of the denitrified products to nitrogen has been proven in various electrolytic cells (Li et al., 2009a, 2010; Reyter et al., 2010; Reyter et al., 2011; Fan et al., 2013). However, these works just attempted to combine the cathodic NO3
reduction and anodic NH4 þ oxidation for the improvement of selective removal of NO3
to nitrogen instead of simultaneous removal NH4 þ and NO3
from wastewater (Reyter et al., 2010). To the authors' knowledge, there are no investigations into electrochemical treatment of WWTP effluent containing both NH4 þ and NO3
. The objective of this study was to develop an eAND process as the tertiary treatment of WWTP effluent, focusing on simultaneous cathodic reduction of NO3
, anodic oxidation of NH4 þ , and coliforms inactivation. The scopes of this work are: (1) to investigate the major factors influencing NH4 þ , NO3
and coliforms removal, such as current density, chloride concentration, and initial pH; (2) to reveal nitrogen removal mechanism and determine the kinetic model of NH4 þ oxidation and NO3
reduction; (3) to evaluate the feasibility of this eAND process for simultaneous removal of nitrogen species and inactivation of coliforms from actual WWTP effluent; and (4) to explore the formation of chlorinated by-products during the process.

2.

Materials and methods

2.1.

Experimental setup

The eAND reactor made of Pyrex glass was consisted of DSA anode and graphite felt cathode with an effective electrode surface area of 36 cm2 (60 mm  60 mm). DSA anode with IrO2 and RuO2 cover (purchased from Baoji, China) was used, exhibiting high electrocatalytic activity, high corrosion resistance, and excellent mechanical stability (Fan et al., 2013). The graphite felt cathode with carbon content higher than 99.9% was purchased from Beijing, China, which had been commercially used in fuel cell field with the advantages of wide potential range and good stability. The reactor comprised anodic and cathodic chambers, which were pressed up onto either side of a cation exchange membrane (CEM, Ultrex CMI-7000, America). The electrolysis experiments were conducted in batch mode using a potentiostat (Zhaoxin Corp. JPS-3005, China). The synthetic wastewater was prepared with NaCl, NH4Cl and NaNO3, and the secondary effluents were collected from a local WWTP in Harbin, northeast China. The main characteristics of the wastewater used in this study were shown in Table 1.

2.2.

Operation and sampling

A volume of 0.4 l synthetic wastewater or actual WWTP effluent was stored in a reservoir (V ¼ 1 l) and continuously recirculated first through cathodic, and then anodic chamber

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w a t e r r e s e a r c h 7 4 ( 2 0 1 5 ) 1 2 2 e1 3 1

Table 1 e The main characteristics of the applied wastewater. Parameters pH
1 NHþ 4 eN (mg l )
1 eN (mg l ) NO
3 Cl
(mg l
1) SO4 2
(mg l
1) Coliforms (CFU ml
1) TN (mg l
1)

Synthetic wastewater

Actual WWTP effluent

6,7,9 20 20 100,250,350,500 250 2500~3000 40

7.2 8.98 18.88 95.9 80.2 1500~2000 30.12

with peristaltic pumps (LongerPump BT100-1F, China) at a rate of 100 ml min
1 (Fig. 1). Electrolysis process was carried out at constant ambient temperature (T, 20 ± 0.5  C) and different current density (2.5, 5.0, 7.5, 10.0, 12.5 mA cm
2). The water pH was not adjusted. Samples were taken from the reservoir after fixed time interval (5, 10, 20, 30, 45, 60, 90 min), analyzed in 24 h or stored in 4  C refrigerator prior to analysis. The concentrations of initial pollutants and products were represented against the electric charge (Q, Ah l
1), which was calculated according to electrolysis time and current density as follows: Q¼

J$A$t V

(1)

where J is the applied current density (A cm
2), A is the valid electrode area (cm2), t is the electrolysis time (h), and V is the volume of electrolyte (l).

2.3.

Analytical methods

NO3
, NO2
and chloride inorganic anions (Cl
, ClO3
, ClO4
) were measured by ion chromatography (Dionex ICS3000, America) with a Dionex Ion Pac AS11-HC 4  250 mm anion exchange column, using KOH as eluent at 1 ml min
1. The measurement of TN was performed with carbon analyzer (Shimadzu TOC-VCPN, Japan), and NH4 þ was quantified by Nessler reagent spectrophotometry (APHA, 1998). The org-N in the effluent was calculated by subtracting dissolved inorganic nitrogen ðNH4 þ ; NO3
and NO2
Þ from TN. Total active chorine (free chlorine and inorganic chloramines, e.g. monochloramine and dichloramine) was measured with the N,Ndiethylp-phenylenediamine (DPD) standard method (APHA, 1998).

The organic chlorinated products were determined by Gas Chromatography (GC, Agient 7890, America). Trihalomethanes (THMs) including chloroform (CHCl3), bromodichloromethane (CHCl2Br), chlorodibromomethane (CHClBr2) and bromoform (CHBr3), were analyzed by headspace sampler with a temperature program from 40  C (5 min) to 100  C at 8  C min
1 and from 100  C to 200  C (5 min) at 6  C min
1. The temperature of headspace oven and injector was set at 60 and 250  C. The column used was DB1301 (60 m  0.25 mm  1.0 mm). Other disinfection by-products (DBPs) included carbon containing DBPs (C-DBPs) and nitrogen containing DBPs (NDBPs). Most researches focused on haloacetonitriles (HANs, NDBPs), haloacetamide (HAcAm, N-DBPs) and haloketons (HKs, C-DBPs), which would cause chronic cytotoxicity and genotoxicity (Muellner et al., 2007). In our study, the typical DBPs of dichloroacetonitrile (DCAN), dichloracetamide (DcAcAm), and trichloroacetone (TCP) were detected, which was liquideliquid extracted by methyl tertiary butyl ether (MTBE). 10 ml of sample and 2 ml of MTBE were pipetted into 40 ml screw cap vials, which then were shaken vigorously for 2 min and stood for 3 min to favor phase separation. MTBE extract was collected and analyzed by GC equipped with an electron capture detector (m-ECD) at 250  C. The column used was Agilent 19091J-413, 30 m  0.32 mm  0.25 mm. The temperature program stayed at 55  C for 5 min, and increased to 200  C (3 min) at 40  C min
1. The coliforms were screened from the actual WWTP effluent, and then added to the synthetic wastewater. The germs were grown and selected using a coliform-selective nutrient agar (Schmalz et al., 2009). The final suspension with lysogeny broth (LB) substrate containing 109 CFU ml
1 coliforms was stored at 4  C. Prior to use, the bacteria concentrate revived at 37  C for 2 h, and then was diluted to get an initial concentration of approximate 103 CFU ml
1. The population of samples was measured by the spread plate method through plating 1 ml of undiluted or diluted solution on agar, and then enumerating at 37  C for 20 h incubation. Three replicates of suspensions were carried out and showed good reproducibility with a standard deviation of 15%.

3.

Results and discussion

3.1. Major factors influencing nitrogen removal and disinfection The influence of operating condition (current density) and wastewater quality (pH, chloride and sulfate ions) on the process performance for synthetic wastewater was firstly evaluated.

3.1.1. Effect of the applied current density on NH4 þ , NO3
and coliforms removal

Fig. 1 e Schematic diagram of the experimental set-up.

Fig. 2 shows the variations of NH4 þ , NO3
and coliforms during the eAND process of synthetic wastewater with the applied current density. The increase of current density from 5 to 7.5 mA cm
2 enhanced NH4 þ removal (Fig. 2a), however, the NH4 þ removal efficiency correlated only with the charge when the current densities were above 7.5 mA cm
2, at which more than 90% of NH4 þ was removed at an electric charge of

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w a t e r r e s e a r c h 7 4 ( 2 0 1 5 ) 1 2 2 e1 3 1

HOCl 4 ClO
þ Hþ

(4)


2NH4 þ þ 3 HOCl/N2 þ 3 H2 O þ 5 Hþ þ 3Cl

(5)

þ

Different from that of NH4 oxidation, it appeared that 10 mA cm
2 was the most efficient current density for NO3
reduction with the highest efficiency of 49% at an electric charge of 1.2 Ah l
1 (Fig. 2b). As pointed out by Reyter et al. (2008), the cathodic potential was the crucial parameter for NO3
reduction. The increase of current density could enhance the potential on cathode and provide more electrons, however, it might result in hydrogen evolution whilst enhancing NO3
reduction. Thus, NO3
reduction should be performed under the threshold of applied current density in order to limit the occurrence of side reactions. In contrary to NH4 þ removal, less electric charge was required to inactivate coliforms present in wastewater (Fig. 2c). The required charge for complete removal of coliforms at 10 mA cm
2 was only 0.066 Ah l
1. The complete inactivation of coliforms could be achieved at a current density greater than 7.5 mA cm
2, owing to the in situ generation of enough chlorine. The presence of NH4 þ could increase the disinfection capability due to the generation of chloramines (Eqs. (6)e(8)) (Cotillas et al., 2013). NH4 þ þ HOCl4NH2 Cl þ H2 O þ Hþ

(6)

NH2Cl þ HOCl4NHCl2 þ H2O

(7)

NHCl2 þ HOCl4NCl3 þ H2O

(8)

Based on the above-mentioned results, it might be concluded that a working current density of 10 mA cm
2 for eAND process was suitable for the simultaneous removal of NH4 þ , NO3
and coliforms from water.

3.1.2.

Fig. 2 e Effect of current density on the removal of NH4 þ (a), NO3
(b) and inactivation of coliforms (c) (Initial NH4 þ eN and NO3
eN: 20 mg l¡1; initial coliforms: 2500 ~ 3000 CFU ml¡1).

1.2 Ah l
1. As revealed by Gendel and Lahav (2012), free chlorine was in situ generated through electrooxidation of the available chloride ion in wastewater. Consequently, the NH4 þ in wastewater reacted with the generated chlorine to form gaseous nitrogen through the anodic indirect oxidation mechanism as shown in Eqs. (2)e(5). Most NH4 þ removal at a current density greater than 7.5 mA cm
2 indicated the abundance of free chlorine in the electrochemical reactor. 2Cl
/Cl2 þ 2 e


(2)

Cl2 þ H2O/HOCl þ Cl
þ Hþ

(3)

Effect of initial pH on nitrogen removal

The effect of pH on NH4 þ and NO3
removal at an electric charge of 0.9 Ah l
1 is shown in Fig. 3. Moderate acidic condition (pH 6) benefited the removal of NH4 þ , with the highest removal efficiency of about 60%. One reason mentioned by Chen et al. (2007) and Kapałka et al. (2010b) was high pH promoted an undesired disproportion reaction to chlorate (Eq. (9)), which was a stable anion and caused the loss of chloride ion in water. In addition, Anglada et al. (2009) found the pH value affected NH4 þ removal efficiency because it determined the primary free chloro species. At pH < 7.5, the primary free chloro specie was HClO (Eq. (3)), and ClO
at pH > 7.5 (Eq. (4)). Thus, NH4 þ removal under acidic condition was considered to be a better choice since HClO was a stronger oxidant than ClO
in principle.








3ClO 4ClO3 þ 2Cl


(9)

NO3 eN removal increased slightly when the pH was increased since pH affected mainly indirect oxidation processes in such a system. Dortsiou et al. (2013) reported that NO3
reduction with tin cathode did not depend on the pH at pH > 4, as the proton donor was water molecule. NO3
eN reduction rate was proportional to the hydronium cation (proton donor) concentration in the pH range of 0~4. Although

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Fig. 3 e Effect of pH on NH4 þ and NO3
removal (Q: 0.9 Ah l¡1; current density: 5 mA cm¡2; chloride ion: 250 mg l¡1; initial NH4 þ eN and NO3
eN: 20 mg l¡1).

the pH had very different effect on the removals of NH4 þ and NO3
, an initial pH range of 6~9 of wastewater would not significantly affect the removal of TN, indicating no need of pH adjustment when such eAND process was applied for tertiary treatment of WWTP effluent.

3.1.3.

Effect of chloride and sulfate ions on nitrogen removal

Chloride and sulfate are the most common co-existent ions in actual wastewater, so it is necessary to evaluate their effects on the removal behavior of nitrogen species by using synthetic wastewater. Fig. 4 shows the influence of chloride concentration on NH4 þ removal and free chlorine production. With the presence of NO3
in water, the higher chloride concentration, the more NH4 þ removal under the same electric charge (Fig. 4a). Higher chloride concentration (250 mg l
1, Cl
=NHþ 4 eN molar ratio  5) brought about more than 90% removal of NH4 þ , while less efficient with lower chloride concentration (<100 mg l
1, Cl
=NH4 þ eN molar ratio < 2) because of mass transfer control. Further increase of chloride concentration (350 mg l
1) did not cause significant NH4 þ removal, ascribing to charge control at Cl
=NH4 þ eN  5. Similar to the intensively studied “breakpoint chlorination”, the electrolysis treatment accelerated the NH4 þ oxidation process because of the recycle use of the chloride (Eq. (5)). Thus, one advantage of this eAND process was the lower Cl
=NH4 þ eN requirement than that in conventional chlorination (Gendel and Lahav, 2012) when chloride was present in the wastewater. In the absence of NO3
in wastewater, more NH4 þ could be removed at the same chloride concentration of 250 mg l
1 (Fig. 4a). The reason was that NO3
reduction in cathodic area contributed more NH4 þ to the reaction system (Eq. (10)) when NO3
was present, which was also considered as the primary restriction to the efficacy of electrochemical denitrification (Mook et al., 2012). NO3
þ 6H2 O þ 8e/NH3 þ 9OH


(10)


The production of free chlorine (Cl2, HOCl and OCl ) in anodic area at different concentration of chloride (Fig. 4b)

Fig. 4 e Effect of chloride concentration on NH4 þ reduction (a) and free chlorine production (b) (Initial NH4 þ eN and NO3
eN: 20 mg l¡1; current density: 10 mA cm¡2; pH: 7).

indicated that the higher chloride concentration produced more free chlorine. When a higher electric charge was applied (Q > 0.9 Ah l
1), the amount of free chlorine increased linearly with electric charge (when Cl
>250 mg l
1) after removal of most of the NH4 þ from wastewater, which was also observed by Kapałka et al. (2010a) who reported active chlorine and chlorate sharply increased once NH4 þ was removed at 7 Ah l
1. This verified again the indirect oxidation mechanism of NH4 þ in the anodic compartment, the produced chlorine reacted with NH4 þ quickly instead of being further electrooxidized to chlorate (Eqs. (5) and (9)). Free chlorine (>1 mg l
1 when Cl
> 250 mg l
1 and Q > 0.9 Ah l
1) would continuously produce by using the electric charge, preferentially being consumed for NH4 þ oxidization and then enriching in water after depletion of NH4 þ (Eqs. (2)e(4)). Fig. 5 shows the influence of chloride concentration on NO3
electro-reduction and removal constant (kNO3
) with different electric charge. For comparison, the influence of sulfate is also included. Both chloride (100~500 mg l
1) and sulfate (250 mg l
1) ions significantly affected NO3
reduction (Fig. 5a) and removal constant (Fig. 5b), maximizing 0.54 l A
1 h
1 at 250 mg l
1 of chloride. Previous studies also reported either a positive or negative effect of chloride on
rez et al., NO3
removal (Li et al., 2009b; Lacasa et al., 2012; Pe

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3.2.

NH4 þ and NO3
removal mechanism and kinetics

In the eAND process, NO3
was mainly converted to N2, NH4 þ and NO2
by direct electron reduction on the cathode, and thereafter NH4 þ existed in wastewater was further oxidized to N2 through anodic indirect oxidation mechanism. Meanwhile, coliforms were also inactivated by in situ generated active chlorine in the anodic chamber. The aqueous nitrogen removal and disinfection were both achieved by this eAND process. The membrane separated anodic and cathodic chamber alleviated the mutual influence between anodic oxidation and cathodic reduction, whilst enhanced the transfer of cation (e.g. Hþ) from anodic to cathodic chambers.

3.2.1. NO3


Electro-oxidation kinetics of NH4 þ in the absence of

As described by Chen et al. (2007), NH4 þ oxidation in electrochemical process followed the pseudo first kinetics. To investigate into the NH4 þ oxidation behavior in our eAND system, a total of 13 experiments were carried out in the absence of NO3
with the experimental design and results being summarized in Table 2. Based on the 1st Opt Software, it could be determined that the decrease of NH4 þ concentration against Q (Ah l
1) followed the model below:
kNH

CNH4 þ ¼ CðNH4 þ Þ e

4

þQ

(13)

0




Fig. 5 e Effect of chloride and sulfate anions on NO3 reduction (a) and NO3
removal constants (b) during electrolysis (Initial NO3
and NH4 þ : 20 mg l¡1 ; current density: 10 mA cm¡2; pH: 7).

kNH4 þ ¼ 0:934 

2012a). The negative influence of higher chloride concentration was attributed to the fast oxidation of the intermediate products (NO2
and NH4 þ ) to NO3
again (Eqs. (11)e(12)) besides chloride's oxidation of NH4 þ to nitrogen (Eq. (5)). The decrease of NH4 þ would in turn promote the NO3
reduction. The fluctuant change of NO3
reduction constant could be explained by the double-face influence of chloride.


NH4 þ þ 4 HOCl/NO3
þ H2 O þ 6 Hþ þ 4Cl


NO2
þ HOCl/NO3
þ Hþ þ Cl

(11) (12)

Sulfate ion exerted a greater negative impact than chloride, a lower NO3
removal constant of 0.16 l A
1 h
1 was observed at a sulfate concentration of 250 mg l
1 (Fig. 5b). These results were consistent with previous reports (Vanlangendonck et al., 2005; Szpyrkowicz et al., 2006) that the oxygen-contained sulfate would retard NO3
reduction due to its competition of adsorption sites with the NO3
on cathode surface. Lacasa et al. (2012) also indicated higher NO3
removal with chloride as supporting media than with sulfate. To sum up, a chloride concentration of 250 mg l
1 was the optimal operational condition to realize simultaneous removal of NH4 þ and NO3
in wastewater.

CNH4 þ 14

!
1:277

 

CCl
35:5

0:705 R2 ¼ 0:985

(14)

where CCl
is chloride concentration (mg l
1), CðNH4 þ Þ0 and CNH4 þ are the initial and residual NH4 þ concentration (mg l
1), and kNH4 þ is NH4 þ oxidation kinetic constant (l A
1 h
1). The above result illustrated that kNH4 þ in the absence of NO3
was proportional to chloride concentration, also evidenced by more free chlorine production (Fig. 4). Conversely, higher initial NH4 þ concentration would result in more residual NH4 þ in wastewater. These results were accordant with the study by Vanlangendonck et al. (2005) who reported kNH4 þ primarily depended on such parameters as current density, chloride and NH4 þ concentrations.

3.2.2.

Electro-reduction kinetics of NO3


The pseudo first order kinetic model previously proposed by Dash and Chaudhari (2005) and Katsounaros et al. (2006) for NO3
reduction versus time also fitted for this study against the change of electric charge (Eq. (16)), with the kNO3
being proportional to the initial NO3
concentration: kNO3


NO3
!N2 ; NH4 þ ; NO2
; NO

(15)

CNO3
¼ CðNO3
Þ e
kNO3
Q

(16)

0

where CðNO3
Þ0 and CNO3
are the initial and residual NO3
concentrations (mg l
1), kNO3
is the observed first order NO3
removal constant (l A-1h
1).

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w a t e r r e s e a r c h 7 4 ( 2 0 1 5 ) 1 2 2 e1 3 1

Table 2 e NH4 þ oxidation constants obtained from experiments in the absence of NO3
. No. 1 2 3 4 5 6 7 8 9 10 11 12 13

Cl
(mg l
1)

NH4 þ eN (mg l
1)

Cl
=NH4 þ eN molar ratio

NH4 þ oxidation constant (kNH4 þ , l A
1 h
1)

R2

200 40 90 200 300 400 200 800 200 400 1000 1500 2000

10 20 20 20 20 20 40 40 100 100 100 100 100

8 0.8 1.8 4 6 8 2 8 0.8 1.6 4 6 8

4.498 0.515 1.365 2.083 2.854 3.691 0.773 2.426 0.148 0.273 0.542 0.803 1.006

0.983 0.970 0.938 0.992 0.972 0.920 0.992 0.985 0.996 0.990 0.975 0.988 0.995

Note: current density ¼ 10 mA cm
2.

3.2.3.

Modification of NH4 þ electro-oxidation kinetics

As discussed previously (3.1.3), part of NO3
was converted to NH4 þ during NO3
reduction in the eAND process (Eq. (10)). With the presence of NO3
in raw wastewater, the model for NH4 þ oxidation failed to be simply described as pseudo first order, thus a modification of NH4 þ electro-oxidation kinetics was needed. NO3
conversion into NH4 þ and gaseous nitrogen in the eAND process can be described by Eq. (17) as follows: kNO

kNH


/NH þ 4

þ

NO3
!NH4 þ !N2 3

4

(17)

As proposed by Fan et al. (2013), the kinetics differential equation is: dCNH4 þ ¼
kNO3
/NH4 þ CNO3

kNH4 þ CNH4 þ dQ

(18)

In consideration of Eq. (16), it should be: dCNH4 þ ¼
kNO3
/NH4 þ CðNO3
Þ e
kNO3
Q
kNH4 þ CNH4 þ 0 dQ

(19)

The Laplace conversion of the above differential equations yields (Katsounaros et al., 2006): CNH4 þ ¼

kNO3
/NH4 þ CðNO3
Þ  0

kNH4 þ
kNO3


e
kNO3
Q
e


kNH

4

þQ



(20)

Considering that the total NH4 þ removal is the result of the initial and produced NH4 þ oxidation with the presence of NO3
, the residual NH4 þ concentration in the eAND process will be a combination of Eqs. (20) and (13) as below:

CNH4 þ ¼

 kNO3
/NH4 þ CðNO3
Þ 
k Q 0 e
kNO3
Q
e NH4 þ kNH4 þ
kNO3

kNH

þ CðNH4 þ Þ e

4

þQ

(21)

0

Since Eq. (21) is similar to the exponential decay, the following simplified model for NH4 þ oxidation (Eq. (22)) can be used to fit the experimental data at different chloride concentrations.
kNH

CNH4 þ ¼ A1 e
kNO3
Q þ A2 e

4

þQ

(22)

The regression of the experimental results according to Eq. (21) gives kinetic constants shown in Table 3, while a parity graph of NH4 þ concentration with different chloride concentration is illustrated in Fig. 6. The simulated values from the model fitted well with the experimental data since 82% of the simulated data fell within the experimental data range of 15%. Thus it might be concluded that the modified kinetic model are beneficial for estimation of NH4 þ removal under other operational condition to meet more stringent NH4 þ nitrogen discharge standards.

3.3. Fesibility of the eAND process for actual effluent treatment The effectiveness of the eAND process had been proven for NH4 þ oxidation and NO3
reduction in synthetic wastewater as discussed earlier, however, its feasibility for treatment of actual WWTP effluent remained to be confirmed. Under the

Table 3 e Kinetic constants of NO3
reduction and NH4 þ oxidation.
1
1

2
(mg l
1) kNH4 þ a(Simulated, l A-1h
1) kNO3
(l A
1h
1) kNO3
/NH4 þ (l A-1h
1) No. NHþ 4 eN (mg l ) NO3 eN (mg l ) Cl or SO4

1 2 3 4 5 6 a

20 20 20 20 20 20

0 20 20 20 20 20

250 100 250 350 500 250(SO4 2
)

kNH4 þ was simulated and calculated from Eq. (13).

2.34 1.23 2.34 2.97 3.82 e

e 0.13 0.54 0.33 0.17 0.16

e 0.11 0.46 0.28 0.14 e

w a t e r r e s e a r c h 7 4 ( 2 0 1 5 ) 1 2 2 e1 3 1

129

discharge standard (GB18918-2002) in China (TN < 15 mg l
1;
1 at T > 12  C and <8 mg l
1 at T < 12  C). These NHþ 4 -N<5 mg l results demonstrated the feasibility and capability of this eAND process as an alternative for further removal of nitrogen species from WWTP effluent. NO3
reduction in actual effluent followed a pseudo first kinetic with a removal constant of 0.37 l A-1h
1, lower than that with synthetic wastewater (0.54 l A-1h
1). This was also due to the presence of other oxygen-contained anions in the actual WWTP effluent (e.g. SO4 2
ions: 80.2 mg l
1).

3.3.2. Coliforms inactivation and distribution of chlorine species Fig. 6 e Parity graph of NH4 þ concentration at different chloride concentration with or without NO3
.

optimal conditions determined with synthetic wastewater (10 mA cm
2 current density, 250 mg l
1 chloride concentration, pH 6 ~ 9), the collected actual WWTP effluent was fed into the eAND system (ref. to Fig. 1) to examine the performance of nitrogen removal and coliforms inactivation as well formation of DBPs. The deficiency of chloride ions in this actual effluent was made up by adding NaCl.

3.3.1. Removal of nitrogen species from actual WWTP effluent The distribution of various nitrogen species after tertiary treatment of an actual WWTP effluent at different electric charge (Fig. 7) indicated that NH4 þ , NO3
and Org-N could be removed whilst a trace amount of NO2
formed. At an electric charge of 1.23 A h l
1, the NH4 þ and NO3
could be removed 96% and 36%, respectively. About 70% Org-N was also removed in the eAND system, corresponding to the reports in previous studies (Kim et al., 2013; Ding et al., 2014). Consequently, a final effluent with 14.18 mg l
1 TN and 0.30 mg l
1 NH4 þ eN was obtained, meeting the strict Class IA criteria of the

Fig. 7 e Distribution of various nitrogen species (NH4 þ , NO2
, NO3
and Org-N) after the tertiary treatment of an actual WWTP effluent at different electric charge (chloride ion: 250 mg l¡1; current density: 10 mA cm¡2).

When the actual WWTP effluent was treated, the eAND process also exhibited as good disinfection performance as for the synthetic wastewater. Complete inactivation of coliforms could be achieved at an electric charge of 0.072 A h l
1, ascribing to the in situ generated germicide from the available chloride ions. Haaken et al. (2012) conducted an individual study of electrochemical disinfection and also found coliforms in secondary effluent could be effectively inactivated at an electric charge of 0.10e0.15 A h l
1 in an electrochemical system with BDD as anode. In this scenario, the eAND process provided another potential alternative for disinfection while realizing simultaneous removal of NH4 þ and TN. In addition, this eAND process was free of chemical dosing in comparison with the conventional chlorination disinfection. To clarify the chlorine species during the eAND treatment of the actual WWTP effluent, both inorganic and organic chlorinated compounds were examined. The distribution of chlorine species under different electric charge (Fig. 8) indicated that chloride ions were more or less converted to chlorine gas (calculated based on the mass balance of the chlorine species), free chlorine, chloramines and chlorate. After most NH4 þ was removed at an electric charge of 1.23 Ah l
1, the proportion of chlorine gas increased significantly (at 1.65 Ah l
1). The free chlorine remained relatively constant until a sharp increase to 3.6 mg l
1 at the electric charge of 1.65 Ah l
1. The amount of in situ generated chlorine was in the range of 2e5 mg l
1 during the chlorination for a full-scale WWTP in NE

Fig. 8 e Distribution of chlorine species (chloride ion, chlorate, chlorine gas, free chlorine and chloramine) under different electric charge (chloride ion: 250 mg l¡1; current density: 10 mA cm¡2).

130

w a t e r r e s e a r c h 7 4 ( 2 0 1 5 ) 1 2 2 e1 3 1

Spain (Matamoros et al., 2007). As an intermediate product, chloramines performed the similar behavior like free chlorine,
rez et al. (2012b). With the which was also mentioned by Pe increase of free chlorine, chloramines would be further decomposed (Gendel and Lahav, 2012), however, the total active chlorine (the sum of free chlorine and chloramines) remained almost unchanged regardless of the electric charge. Chlorate of 1.8 mg l
1 was first observed at 1.23 A h l
1 and no perchlorate was detected, indicating the better performance of DSA anode than BDD in formation of chlorate and perchlorate, which were critical to human beings due to their toxic effects (hemolytic anemia and thyroid dysfunction) (Polcaro et al., 2008; Jung et al., 2010). As for the organic chlorinated compounds, only trace amount of DBPs were detected at an electric charge of 1.23 Ah l
1 (THMs 12 mg l
1, DCAN 33 mg l
1, TCP 13 mg l
1, and DcAcAm 9 mg l
1), with THMs being much lower than the threshold of 80 mg l
1 defined in international drinking water regulations and within the range of 2e30 mg l
1 for the chlorinated reclaimed water (Matamoros et al., 2007). These low concentrations of organic chlorinated compounds demonstrated the safe application of this eAND process to the disinfection of WWTP effluent.

4.

Conclusions

Based on the study of anodic oxidation of NH4 þ , cathodic reduction of NO3
and inactivation of coliforms in the eAND process treating synthetic wastewater and actual WWTP effluent, the conclusion can be drawn as follows: (1) The eAND process was feasible for tertiary treatment of actual WWTP effluent in achieving simultaneous nitrogen-removal and disinfection. Under the optimal operational conditions of 10 mA cm
2 and 250 mg l
1 chloride, 96% NH4 þ and 36% NO3
were removed at an electric charge of 1.23 Ah l
1 while realizing complete coliforms inactivation. (2) The presence of free chlorine indicated that indirect oxidation was the predominant mechanism for NH4 þ oxidation and coliforms inactivation in anodic area, and NO3
removal was achieved by electro-reduction in cathodic area. (3) NO3
or NH4 þ removal followed pseudo first kinetics model. With the presence of both NH4 þ and NO3
, the modified model of Exponential Decay for NH4 þ oxidation was satisfactory to simulate the experimental results. (4) Disinfection of WWTP effluent during the eAND process was free of chemical dosing and less formation of DBPs, indicating the features of simplicity and safety of this eAND process.

Acknowledgments The authors gratefully acknowledge funding from Project 51121062 (National Creative Research Groups) supported by

National Nature Science Foundation of China, and from the National Critical Scientific and Technological Project of Water Pollution Control and Management (2012ZX07201003-002).

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