Simultaneous removal of ammonia nitrogen and recovery of phosphate from swine wastewater by struvite electrochemical precipitation and recycling technology

Accepted Manuscript Simultaneous removal of ammonia nitrogen and recovery of phosphate from swine wastewater by struvite electrochemical precipitation and recycling technology Haiming Huang, Peng Zhang, Zhao Zhang, Jiahui Liu, Jing Xiao, Faming Gao PII:

S0959-6526(16)30234-7

DOI:

10.1016/j.jclepro.2016.04.002

Reference:

JCLP 7000

To appear in:

Journal of Cleaner Production

Received Date: 24 February 2016 Revised Date:

1 April 2016

Accepted Date: 2 April 2016

Please cite this article as: Huang H, Zhang P, Zhang Z, Liu J, Xiao J, Gao F, Simultaneous removal of ammonia nitrogen and recovery of phosphate from swine wastewater by struvite electrochemical precipitation and recycling technology, Journal of Cleaner Production (2016), doi: 10.1016/ j.jclepro.2016.04.002. This is a PDF file of an unedited manuscript that has been accepted for publication. As a service to our customers we are providing this early version of the manuscript. The manuscript will undergo copyediting, typesetting, and review of the resulting proof before it is published in its final form. Please note that during the production process errors may be discovered which could affect the content, and all legal disclaimers that apply to the journal pertain.

ACCEPTED MANUSCRIPT Simultaneous removal of ammonia nitrogen and recovery of phosphate from swine wastewater by struvite electrochemical precipitation and recycling technology

a

RI PT

Haiming Huang a *, Peng Zhang a, Zhao Zhang b, Jiahui Liu a, Jing Xiao a, Faming Gao a

Hebei Key Laboratory of Applied Chemistry, School of Environmental and Chemical Engineering, Yanshan University, Qinhuangdao 066004, PR China b

Liaoning Cleaner production Instruct center, Shenyang 110033, PR China

SC

Abstract: The removal of phosphorus and nitrogen from wastewater has been a matter of great concern for several decades. In this study, a coupled electrochemical process, involving the

M AN U

electrochemical recovery of phosphate as struvite, electrochemical decomposition of the struvite recovered, and removal of the ammonia nitrogen by recycling the struvite electrolysis product, was tested to simultaneously recover phosphate and remove the ammonia nitrogen from swine wastewater. The results demonstrated that when a magnesium alloy was electrolyzed as the magnesium source of struvite crystallization, the recovery efficiency of the phosphate was 99%, at

TE D

a current density of 2 mA/cm2 for 45 min. When the struvite recovered was electrolyzed using seawater as a supporting solution, the ammonium in the struvite could be completely removed. The characterization analysis revealed that the active component of the electrolysis product was

EP

present as dissolved phosphate and magnesium and insoluble cattiite. An efficiency of >90% of ammonia nitrogen removal could be achieved by recycling the electrolysis product of struvite. A

AC C

pilot-scale test performed for 30 recycle cycles demonstrated that approximately 93% of the recovery/removal efficiencies of the phosphate and ammonia nitrogen from swine wastewater could be achieved stably by the process proposed, which implied the presence of an important application value in the water eutrophication control and recovery of the phosphorus resources. Keywords: struvite, swine wastewater, electrolysis, phosphate, recycling

*Corresponding Author: Phone: +86 335 8387 741; Fax: +86 335 8061 569; E-mail: [email protected]

1

ACCEPTED MANUSCRIPT 1. Introduction In recent years, the pig industry in the world has witnessed rapid growth in the global demand for red meat. According to the statistical data of the Food and Agricultural Organization (FAO), the amount of swine produced in the world reached 963,044,187 heads in 2012 (Lim and

RI PT

Kim, 2015). On the one hand, since large quantities of swine wastewater containing high concentrations of total orthophosphate (PT), total ammonia nitrogen (TAN), and organic substances was generated from the feeding process of pigs (Ye et al., 2010), the serious effect of wastewater on the ecological environment is of great concern. On the other hand, the phosphorus

SC

(P) element in swine wastewater is a valuable resource. P is an important non-renewable resource in the nature, which is essential to maintain the development of societies (Zhang et al., 2013).

M AN U

Unfortunately, the exploitable reserves of phosphate rock are progressively decreasing since large amounts of P are utilized as the fertilizers in agriculture and the raw materials in some industries. It has been reported that the available accessible reserves of clean phosphate rock would be exhausted in the next 50 years unless some effective measures are taken (Gilbert, 2009). Therefore, based on the consideration of the sustainable development of P resource, recovery of P from

2012; Zhang et al., 2013).

TE D

wastewaters has received great attention worldwide (Cardoso et al., 2015; Ichihashi and Hirooka,

Currently, several techniques are available such as the osmotic membrane bioreactor (Qiu et

EP

al., 2014), selectrodialysis/crystallization (Tran et al., 2014), amorphous calcium silicate hydrates adsorption (Okano et al., 2013), biosorption (Rathod et al., 2014) and struvite crystallization

AC C

(Romero-Güiza et al., 2015) for the recovery of phosphate. Among these processes, struvite precipitation has been proven to be a promising technique for the removals of PT and TAN. Struvite (MgNH4PO4·6H2O) is an insoluble double salt that can be potentially used as a slow-release fertilizer as it contains both P and N (Hao et al., 2013). Precipitating struvite in the form of stable white orthorhombic crystals should be of a suitable pH, with the required concentrations of Mg2+, NH4+, and PO43–. Although swine wastewater is rich in PT and TAN, its magnesium content is scarce. Hence, a large amount of magnesium needs to be added to swine wastewater for the crystallization of struvite. Soluble salts such as MgCl2 and MgSO4 are the most frequently used magnesium sources for such processes. Besides, other low-grade magnesium sources such as seawater (Crutchik and Garrido, 2011), bittern (Lee et al., 2003), and seawater 2

ACCEPTED MANUSCRIPT nanofiltration concentrate (Lahav et al., 2013) have also been used. Nonetheless, when these magnesium sources are used to prepare struvite, a large amount of NaOH is required for pH adjustment, which may significantly increase salt concentration in the wastewater, which in turn is highly inhibitory toward the microbial activity in the following biological treatment process (An

RI PT

and Gu, 1993, Demirer et al., 2008; Long et al., 2007). Although the use of MgO/Mg(OH)2 as a magnesium source can resolve the problem of increased salinity, it significantly decreases the purity of recovered struvite due to the excessive dosing of alkali compounds. To resolve this issue, Ben Moussa et al. (2006) and Wang et al. (2010) used electrolytic cell with inert anodes to provide

SC

hydroxide anions for struvite crystallization to obtain pure struvite. In addition, Kruk et al. (2014) and Hug and Udert (2013) used magnesium sacrificial anode to simultaneously provide

M AN U

magnesium ions and hydroxide anions for the formation of high-purity struvite. Kruk et al. (2014) reported that struvite with purity of >90% can be obtained by struvite precipitation using magnesium sacrificial anode and by controlling the pH at the range of 7.5–9.3. Although phosphate could be completely recovered from swine wastewater by struvite precipitation with magnesium sacrificial anode, only a very small proportion of TAN was removed

TE D

as the amount of TAN in swine wastewater was much higher than that of phosphate (Liu et al., 2011). To achieve greater TAN removal, it is necessary to supplement additional phosphate salt to swine wastewater; however, this process is not economical for TAN removal. Recycling of struvite

EP

is a highly feasible method to reduce the dosage of phosphate salt and the cost of struvite precipitation. At present, direct pyrolysis (Huang et al., 2011a; Sugiyama et al., 2005) and NaOH

AC C

pyrogenation (He et al., 2007; Türker et al., 2007; Zhang et al., 2009) have been used most often for struvite decomposition. However, both the methods readily result in the generation of byproducts of magnesium pyrophosphate (Mg2P2O7), which has no contribution to TAN removal. Moreover, the amount of Mg2P2O7 in the decomposition product increased with the increase in the number of times of recycling (Türker et al., 2007). Mg2P2O7 is favorably produced from the further dehydration of magnesium hydrogen phosphate (MgHPO4) at temperatures of >80°C (Sugiyama et al., 2005; Türker et al., 2007). Therefore, given that struvite was decomposed in the water solution at room temperature, Mg2P2O7 should not be formed. Electrolysis oxidation is a simple and cost-effective technique for the removal of ammonium, which has been widely used for the treatment of various types of ammonium-containing wastewaters (Pérez et al., 2012; 3

ACCEPTED MANUSCRIPT Vanlangendonck et al., 2005; Xie et al., 2006). Therefore, removal of ammonium in struvite by the electrolysis process is completely feasible. The main objective of the present study was to achieve the simultaneous recovery of PT and removal of TAN from swine wastewater by a coupled electrochemical process. The specific

RI PT

investigations involve the following aspects: 1) effectiveness of electrochemically dissolved magnesium as a precipitant for the recovery of PT from swine wastewater and the purity of the struvite recovered; 2) optimal conditions for the electrochemical decomposition of the struvite recovered and the decomposition mechanism involved; 3) recycling process of the product of

SC

decomposition for TAN removal; and 4) evaluation of the stability and feasibility of the combined process proposed.

M AN U

2. Materials and methods 2.1. Raw wastewater

The swine wastewater used in the lab-scale experiments was collected from a pig farm located in a Beijing suburb and stored in a refrigerator at 5 oC. Prior to use, the wastewater was pretreated by filtering it through a 0.45-µm filter membrane to remove the suspended solids. The

TE D

characteristics of the pretreated wastewater are given in Table 1. In this study, the raw seawater collected from a bathing beach in the Qinhuangdao city was the supporting solution used in the experiments of the electrochemical decomposition of struvite. It was filtered through a filter paper

Table 1 here Table 2 here

AC C

EP

prior to use, and the main features of the filtrate are shown Table 2.

2.2. Analytical methods

The pretreated swine wastewater was analyzed according to the standard methods (APHA,

1998) to identify the concentration of the relevant PT, TAN, COD, alkalinity, and other cations. The TAN and PT constituents were analyzed by Nessler’s reagent spectrophotometry and Mo–Sb anti-spectrophotometry (752 N-spectrophotometer, JINGKE, China), respectively. The cation concentrations were determined using an Atomic Absorption Spectrophotometer (AA-6800; Shimadzu, Japan). The solution pH was recorded with a pH meter (pHS-3C, JINGKE, China). The morphology and composition of the solid components were investigated using an SEM-energy dispersive spectrometer (SEM-EDS; SUPRA 55 SAPPHIRE; Germany) and an XRD (DMAX-RB; 4

ACCEPTED MANUSCRIPT Rigaku, Japan), respectively. In this study, all the tests were performed in triplicate, and the average value was recorded. The purity of the struvite precipitates was determined according to the following method. At first, 0.5 g of the struvite precipitates was dissolved in 0.5% nitric acid solution and this solution was then diluted with ultrapure water to 100 ml. Finally, the solution pH

RI PT

was adjusted to around 4 by employing NaOH solution (1 M) prior to determining the ammonium concentration. The quantity of struvite in the precipitates could be calculated based on the result of the ammonium determination. Therefore, the struvite purity (Ps) can be calculated according to the equation given,

Qs ×100% Qp

(1)

SC

Ps =

M AN U

where Qs and Qp represent the struvite mass in the precipitates and the total mass of the precipitates, respectively.

2.3. The combined electrochemical process design

The combined electrochemical process for the simultaneous recovery of PT and removal of TAN chiefly involved three coupled reactors: the electrochemical precipitation reactor for the PT

TE D

recovery (labeled R1), electrolysis reactor for the struvite decomposition (R2), and the struvite precipitation reactor for the TAN removal (R3). The schematic illustration of the combined process is shown in Fig. 1. As evident, the PT in the swine wastewater was recovered first by the R1 reactor,

EP

which was a 1.5-L jar provided with an electrode set, a magnetic stirrer, and a pH probe (pHS-3C, JINGKE, China). The anode used in the R1 reactor for the electrochemical reaction was composed of a magnesium plate, while the cathode was a stainless steel mesh plate. The pair of electrodes

AC C

was installed parallel to each other at a 4-cm distance with an active surface area of 96 cm2 (12 cm × 8 cm). After PT recovery, the supernatant at the completion of the reaction was fed to the R3 reactor for the further removal of the TAN. The R3 reactor included a jar with an effective 1 L reaction zone. A mechanical agitator was mounted on top of the reactor to provide the stirring power during solution mixing. As explained in Fig. 1, the struvite recovered from the R1 reactor was transferred to the R2 reactor for struvite decomposition. The main structure of the R2 reactor (with the 100-ml reaction zone) was similar to that of the R1 reactor, except for the electrodes, which were made of graphite. These two graphite electrodes were 1.5-mm thick rectangular plates with an active 12 cm2 surface area (6 cm × 2 cm) and arranged parallel to each other at 2 cm. In 5

ACCEPTED MANUSCRIPT this study, the electricity supply for the R1 and R2 reactors was provided by a digital direct current power supply (12 V 30 A, Model: KXN-3050, ZHAOXIN, China). Fig. 1 here 2.4. PT recovery

RI PT

To determine the optimal electrochemical reaction conditions for the PT recovery, the experiments were conducted using different current densities (1–5 mA/cm2). The specific experimental procedures involved are as follows: At first 1-L swine wastewater was fed to the R1 reactor, followed by feeding the direct current power and then stirring the solution for 70 min at

SC

150 rpm. The pH of the reaction solution was continuously monitored by a pH meter throughout the experiments, and at different time intervals, 2 ml samples were drawn from the reaction

M AN U

solution and filtered through a 0.2-µm membrane for component analysis. Besides, the struvite precipitates obtained during the different reaction times were collected and washed thrice with ultrapure water and then oven dried at 35 ± 1oC for 24 h. 2.5. Struvite decomposition

The experimental procedures of struvite decomposition are described below. At first, 2 g of

TE D

the struvite obtained under the optimal conditions of PT recovery was added to the R2 reactor. Next, 80 ml of the filtered seawater used as the supporting solution of electrolysis was fed to the decomposition reactor, and the initial pH of the mixing solution was adjusted to the desired value (6–9) using 0.1 M HCl/NaOH solution. Third, the electrochemical decomposition of the struvite

EP

solid-liquid system was run at the desired current density of 42–168 mA/cm2 for 180 min. In the

AC C

experiments, the magnetic stirrer just below the R2 reactor was started to maintain the solid-liquid homogenization of the reaction system, and the stirring rate was controlled at approximately 500 rpm. Then, during the 180 min procedure, 1 ml of the mixing solution was taken from the reactor every ten minutes and directly dissolved into 4 ml of 0.5% nitric acid solution to determine the decomposition rate of ammonium in struvite. Finally, to elucidate the electrochemical decomposition mechanism of struvite, 1 ml of the mixing solution was drawn out at different time intervals and filtered through a 0.2 µm filter membrane, and filtration was done to determine the PT and TAN concentrations in the solution. Also, the solid which remained in the solution at different reaction stages was collected and washed thrice with ultrapure water, and oven dried at 35 ± 1 oC for 24 h. The dried solid was characterized by the X-ray diffraction analyzer (XRD) and 6

ACCEPTED MANUSCRIPT scanning electron microscope (SEM). 2.6. TAN removal To study the performance of TAN removal by recycling the struvite decomposition product, a series of experiments were performed at a molar ratio of P in the decomposition product to the

RI PT

TAN in the wastewater of 1:1 and at set points of pH 8, 8.5, 9, 9.5, and 10 for 120 min. Briefly, 1 L of the supernatant discharged from the R1 reactor was added to the R3 reactor, followed by the addition of the struvite decomposition product produced from the R2 reactor. During the stirring reaction of the mixture, the pH of the solution was constantly controlled at the desired value by 2

SC

M HCl/NaOH solution. At different time intervals throughout the 120 min procedure, 5 ml samples of wastewater were removed from the reaction system and filtered through a 0.45-µm

M AN U

filter membrane to determine the composition. Besides, for investigation of the multiple recycle effect of the process proposed, the decomposition product was repeatedly used for 8 recycle cycles as per the methods of the electrochemical decomposition and the recycling of the product mentioned above. 3. Results and discussion

TE D

3.1. Recovery of phosphate by electrochemical precipitation of struvite

Fig. 2 shows the changes in the solution pH, PT-recovery efficiency, and struvite purity with electrolysis time for the different current densities. As evident in Fig. 2a, the solution pH was

EP

obviously affected by the current density and electrolysis time. With respect to the recovery of PT, it is observed from Fig. 2b that the PT recovery efficiency increased with the rise in the current

AC C

density. Within the range of the current density tested, the PT-recovery efficiencies rapidly reached >85% for a 30 min period, followed by a slow increase until a plateau was reached. In the experiments, the struvite purity obtained at different current densities and electrolysis times was analyzed (Fig. 2c). It is found that the current density and electrolysis time significantly influence the struvite purity. As the current density and electrolysis time increased, the struvite purity rapidly decreased. When the current density was 2 mA/cm2 and the electrolysis time was 45 min, the purity of the recovered struvite reached 95.7%, and the contents of Mg, N, P, K, and Ca in it were 98.4, 55.2, 127.2, 1.8, and 3.1 mg/g, respectively. Besides, SEM and XRD analysis were performed to identify the features of the struvite recovered (Fig. 3). The SEM micrograph demonstrated the presence of regular, needle-shaped crystalline products 20–100 µm in diameter. 7

ACCEPTED MANUSCRIPT Based on the XRD patterns these crystals were identified as struvite. Although Fig. 2 implies that the struvite crystallization was significantly influenced by the current density and electrolysis time, the struvite formation was actually controlled by the practical ionic activity product of struvite in solution. Only when it exceeded the thermodynamic

RI PT

solubility product of struvite, did the crystallization occur (Wang et al., 2006). Nonetheless, the existing species and the activity of the constituted struvite ions (Mg2+, NH4+, and PO43–) were further affected by the solution pH (Kemacheevakul et al., 2011). Therefore, in the electrochemical precipitation, the PT recovery by struvite crystallization was essentially achieved

SC

both by the adjustment of solution pH and the release of the Mg2+ ions. In this study, while running the direct current power supply, the magnesium metal plate at the anode rapidly lost

M AN U

electrons and released the Mg2+ ions into the solution for struvite crystallization. Meanwhile, the H+ near the steel mesh which was connected with the cathode of the electric source obtained electrons to produce hydrogen gas, which caused the pH of the solution to increase. This could have caused the struvite formation. Besides, the corrosion of the surface of the magnesium plate (Song et al., 2003; Cao et al., 2013; Thomas et al., 2015) could also contribute to the struvite

TE D

crystallization process. In the published literature, Cao et al., (2013) reported that the amount of Mg2+ dissolved in solution at the applied anodic current is greater than expected from Faraday Law because some Mg2+ was produced by the Mg metal corrosion.

EP

The effective crystallization of struvite is commonly believed to occur within the optimal pH range of 8.5–9.0 (Çelen et al., 2007; Di Iaconi et al., 2010; Li et al., 1999). In this study, after 5 min of electrolysis, the solution pHs at all the current densities tested were within the optimal

AC C

range, where the ratio of the HPO42– species in the solution increased and reached 99% at pH 9 (Saidou et al., 2009). This process could promote the reaction in which the HPO42– reacted with the Mg2+ and NH4+ to form struvite, resulting in an increase in the PT recovery efficiency. In the optimal pH range, although other phosphate-based compounds like potassium magnesium phosphate (MgKPO4·6H2O), magnesium phosphate [Mg3(PO4)2], and calcium phosphate [Ca3(PO4)2] may be formed (Huang et al., 2014; Wilsenach et al., 2007; Wu and Zhou, 2012),

their quantities were very small. However, as the solution pH increased to >9, some Mg2+ in the solution was found to react with the OH– to form amorphous Mg(OH)2 (Wu and Zhou, 2012), and the quantities of the phosphate-based compounds mentioned above increased progressively (Hao 8

ACCEPTED MANUSCRIPT et al., 2013; Huang et al., 2014). This event could be responsible for the rapid decrease in the struvite purity with pH of >9. Fig. 2 here Fig. 3 here

RI PT

3.2. Electrochemical decomposition of the struvite formed 3.2.1. Electrochemical decomposition efficiency

To determine the effect of the initial pH of the solution on the electrochemical decomposition of struvite, batch experiments were performed first at the current density of 84

SC

mA/cm2. The experimental results are shown in Fig. 4 (a, b). The results revealed that the initial pH in the solid-liquid system clearly influenced the decomposition of ammonium in struvite. As

M AN U

observed in Fig. 5a, the solution pH during the decomposition gradually decreased with the increase in the initial pH of the solution from 6 to 7; however, it progressively increased with the increase in the initial pH of the solution in the pH range of 7–9. The changing profile of the decomposition ratio of ammonium in the struvite was similar to that of the solution pH during decomposition (Fig. 4b). Nevertheless, the maximum decomposition ratio of ammonium in

TE D

struvite took place in the initial pH of 8. When the reaction time was 150 min and the current density was 84 mA/cm2, the ratio of ammonium decomposition in the struvite was >99%. This finding was consistent with the decomposition ratio of ammonium reported by Liu et al., (2011),

EP

who investigated the electrolytic dissolution of struvite and showed that, when the electric voltage, NaCl concentration, reaction time, and initial quantity of struvite were 7V, 0.06%, 1.5 h, and 1.25

AC C

g/L, respectively, the ammonium in struvite was completely removed from the solution. Further experiments were performed to investigate the effect of the current density during

the initial solution pH of 8. The results in Fig. 4c indicate that the rise in the current density and reaction time could result in the rapid drop in the solution pH during decomposition. For example, with the increase in the reaction time between 10 min and 180 min, the solution pH during the decomposition process at the current density of 42 mA/cm2 decreased from 6.78 to 5.75, whereas at 168 mA/cm2 it decreased from 6.30 to 4.89. As observed in Fig. 4d, at the given reaction time, the decomposition ratio of ammonium in the struvite rapidly increased with an increase in the current density from 42 to 126 mA/cm2, followed by a slow increase in the current density range of 126–168 mA/cm2. At a given current density, the decomposition ratio of ammonium in the 9

ACCEPTED MANUSCRIPT struvite rapidly rose during the first reaction time of 100 min, followed by a gradual rise to plateau with a further increase in the reaction time. In the literature, several other methods dealing with the decomposition of struvite have been reported. Zhang et al., (2009) pyrolyzed the struvite in a NaOH:NH4+ molar ratio of 2:1 and at 110 oC for 3 h, achieving >90% of the ammonium removal

RI PT

from the struvite. Huang et al., (2014) used sodium hypochlorite as an oxidant to decompose the struvite and found that all the ammonium in the struvite was almost completely removed as nitrogen gas. Fig.4. here

SC

3.2.2. Electrochemical decomposition mechanism

The experiments to investigate the electrochemical decomposition mechanism of struvite

M AN U

were conducted at the initial solution pH of 8 and current density of 84 mA/cm2 for 180 min. In these experiments, the changes in the TAN and PT concentrations and pH in the solution with reaction time were observed. Besides, it was assumed that the phosphate resulting from the struvite decomposition became completely dissolved into the solution. Consequently, the PT concentrations assumed could be approximately calculated based on the decomposition rate of the

TE D

ammonium in struvite (Figs. 5a and b). Fig. 5a shows that the TAN concentration in the solution progressively decreased to <5 mg/L during the electrolysis, whereas the curve profile of the solution pH appeared to be a periodic change, which rapidly decreased during the first reaction

EP

time of 50 min, followed by a plateau between 50 min and 110 min. After a rapid decrease between 110–130 min, finally a plateau of approximately 5.4 was achieved after 130 min. As

AC C

observed in Fig. 5b, the actual concentration of PT in the solution was much lower than the calculated concentration of PT, suggesting that most of the struvite may be transformed into other insoluble phosphate-based compounds during electrolysis. Based on the results mentioned above and the well-known electrochemical degradation

mechanism of ammonium (Chen, 2004; Mook et al., 2012), the main reactions involved in the reaction stage of struvite electrochemical decomposition can be expressed as follows:

2Cl− − 2e− → Cl2

(2)

Cl 2 + H 2 O → HOCl + H + + Cl −

(3)

2NH 4 + + 3HOCl → N 2 + 5H + + 3Cl − + 3H 2 O

(4) 10

ACCEPTED MANUSCRIPT MgNH 4 PO 4 ⋅ 6H 2 O + H + → Mg 2+ + NH 4 + + HPO 4 2− + 6H 2 O

(5)

Mg 2+ + HPO 4 2− + 3H 2 O → MgHPO 4 ⋅ 3H 2 O

(6)

3Mg 2+ + 2H 2 PO 4 − + 22H 2 O → Mg 3 (PO 4 ) 2 ⋅ 22H 2 O + 4H +

(7)

RI PT

In order to elucidate the electrochemical decomposition process of struvite, the electrolysis reaction process was divided into three stages (Fig. 6). At the first reaction stage, chlorine was formed at the anode via the electrolysis of NaCl in the seawater (Eq. (2)), followed by its conversion to hypochlorous acid and hypochlorite (Eq. (3)) (Liu et al., 2011). The hypochlorous

SC

acid could react with the ammonium in the solution to form nitrogen gas and release the H+, resulting in the rapid decrease in the solution pH (Eq. (4)). This process can positively accelerate

M AN U

the struvite dissolution (Eq. (5)). Gunay et al. (2008) reported minimum struvite solubility at pH 8.8–9.4, whereas at pH <6.5, 98% of the struvite could be dissolved. In this study, the solution pH reached 6.2 at the electrolysis time of 50 min. Therefore, the struvite was basically completely dissolved in the first stage. According to the literature (Musvoto et al., 2000; Mijangos et al., 2004), newberyite begins to form at pH <7 and significantly crystallizes at an approximate pH of 6,

TE D

with high magnesium and phosphate concentrations. In fact, newberyite may be formed during the first reaction stage and in a small quantity because the solution pH did not reach the optimal value. Nevertheless, when the reaction was at the second stage, the solution pH was maintained at

EP

approximately 6, which could significantly facilitate newberyite formation. The XRD patterns of the solids collected in the second stage confirmed that the main component was newberyite (Fig.6),

AC C

which concurred with the analysis of this study. As the electrolysis proceeded, the solution pH further decreased and the reaction moved into the third stage. At this stage, the solution pH was approximately 5.4. Under this condition, H2PO4– was the dominant phosphate species in the solution phase and the newberyite might become transformed to other phosphate-based compounds. The insoluble precipitates collected in the third stage were characterized based on the XRD and SEM images. The SEM micrograph indicated that the precipitates were amorphous compounds with a 3–5 µm diameter, and the XRD patterns revealed that the main component of the insoluble compounds was cattiite [Mg3(PO4)2·22H2O]. Cattiite is a type of magnesium phosphate that is precipitated by the combination of PO43– and Mg2+ in the solution system (Hirooka and Ichihashi, 2013). The results of this study were supported by Mijangos et al., (2004), 11

ACCEPTED MANUSCRIPT who reported that large magnesium phosphate quantities could be formed at a solution pH of 4–6 when the total magnesium concentration was >0.1 molar per liter. Fig.5. here Fig.6. here

RI PT

3.3. Removal of TAN by recycling struvite decomposition product 3.3.1. Optimal recycling conditions

The results of recycling the struvite decomposition product at different pHs and the reaction times are shown in Fig. 7. Fig. 7a demonstrated that, at a given reaction time, the removal

SC

efficiency of the TAN increased with increasing pH from 8 to 9, hitting a peak at pH 9, followed by the gradual decline in the pH range of 9–9.5. Moreover, at an identical pH, the TAN-removal

M AN U

efficiency increased rapidly during the first reaction time of 60 min, followed by a comparatively slow increase until a plateau was reached. As noted in Fig. 7a, at the solution pH of 9 and reaction time of 110 min, the TAN-removal efficiency was > 90%, which was comparable to those by using fresh pure magnesium and phosphate salts (Li and Zhao, 2003; Yetilmezsoy and Sapci-Zengin, 2009). Fig. 7b shows the changes in the remaining PT concentrations with reaction

TE D

time at different pHs. The remaining PT concentration was seen to reduce with an increase in the pH. At the reaction time-range tested, the remaining PT concentration under different pH conditions remained stable. When the solution pH was 9, the remaining PT concentration was

EP

approximately 7 mg/L. Solution pH is an important parameter which influences recycling of the struvite decomposition product. He et al. (2007) reported that the optimal pH for TAN removal

AC C

from landfill leachate by recycling the NaOH pyrogenation product of the struvite was 9. Zhang et al. (2009) found that the optimum pH for TAN removal from coking wastewater by reusing struvite pyrolysate was 9.5. In the present study, the optimum pH for recycling the electrochemical decomposition product of struvite was considered to be 9, which is consistent with the findings of struvite precipitation by using pure chemicals in other studies (Di Iaconi et al., 2010; Nelson et al., 2003; Li et al., 1999). Struvite formation was extremely rapid and could be completed within approximately 30 s (Türker and Çelen, 2007). Therefore, it was commonly accepted that the reaction time was not a vital controlling factor for struvite crystallization (Stratful et al., 2001). However, in this study, the struvite crystallization was found to be significantly affected by the reaction time, which is 12

ACCEPTED MANUSCRIPT inconsistent with the published findings. To clarify this issue, the mechanism for recycling the electrochemical decomposition product of struvite for TAN removal was elucidated as follows. The results of the electrochemical decomposition mechanism revealed that the dissolved phosphate accounted for approximately 25% of the total phosphate in the decomposition product,

RI PT

suggesting that approximately 75% phosphate was present in the form of Mg3(PO4)2·22H2O in the decomposition product. Hence, recycling of the electrochemical decomposition product may involve the following reaction equations:

Mg 2+ + NH 4 + + HPO 4 2 − + 6H 2 O → MgNH 4 PO 4 ⋅ 6H 2 O + H +

SC

Mg 3 (PO 4 ) 2 ⋅ 22H 2 O + 2H + → 3Mg 2+ + 2HPO 4 2− + 22H 2 O

(8)

(9)

M AN U

At the initial reaction stage, the dissolved phosphate rapidly reacted with the NH4+ and Mg2+ to form struvite (Eq.(8)), resulting in the rapid increase in TAN removal. However, after the dissolved

phosphate

was

exhausted,

the

struvite

was

formed

according

to

the

dissolution-precipitation mechanism of Mg3(PO4)2·22H2O (Eq.(9)), which was significantly influenced by the reaction time (Sugiyama et al., 2005).

TE D

Fig. 7. here

3.3.2. Multiple recycling of decomposition product Two recycling modes of the struvite electrochemical decomposition product were performed to remove the TAN from the supernatant in the R1 reactor. In the first mode, the

EP

electrochemical decomposition product was directly recycled for 110 min at pH 9, without supplementation using additional phosphate sources. In the second mode, approximately 80 mg of

AC C

the struvite precipitates collected from the R1 reactor was supplemented to the R2 reactor per recycle time, and the decomposition product was then recycled according to the first mode. The experimental results are shown in Fig. 8. The TAN-removal efficiency in the first mode was obviously decreased with the increase in the recycle time, while, in the second mode, it remained stable (at about 91%). The decrease in the TAN-removal efficiency of the first mode may have been mainly caused by the loss of phosphate per recycle time (Fig. 8). Certainly, the accumulations of K+ and Ca2+ in the supernatant may be another reason for this decrease (He et al., 2007; Huang et al., 2011b). Nevertheless, when the additional phosphate source was replenished in the recycle process, these negative influences could be effectively counteracted. In similar 13

ACCEPTED MANUSCRIPT investigations by Türker and Çelen (2007), the TAN-removal efficiency by recycling the struvite pyrolysate without supplementation with the phosphate and magnesium sources was initially 92%, followed by a rapid drop to 77% in the fifth recycle cycle. Zhang et al. (2009) supplemented the MgCl2·6H2O and Na2HPO4·12H2O at Mg2+:NH4+:PO43– molar ratio of 0.05:1:0.05 in the recycling

RI PT

process of the struvite pyrolysate to achieve stable TAN removal. Compared to the reports in the literature, the second recycle mode proposed in this study did not utilize any external phosphate source, which did not show any further increase in the TAN-treatment cost as well. Fig.8. here

SC

3.4. Continuous operation of the proposed process

To evaluate the overall performance of the proposed struvite recovery and recycling

M AN U

processes, a pilot-scale test was performed for 30 operation cycles under the optimal conditions obtained from the investigations mentioned above. The results shown in Fig. 9 demonstrate that the remaining concentrations of PT and TAN in the effluent were controlled at approximately 7 mg/L and 30 mg/L, respectively. The recovery/removal efficiencies of PT and TAN were calculated to be approximately 93%. In the continuous treatment process, all the struvite

TE D

precipitates recovered from the R1 reactor were almost reused for the TAN removal, suggesting that the finally recovered struvite was obtained from the R3 reactor. The struvite recovered from the R3 reactor after 10 operation cycles was characterized by the XRD and SEM, and its chemical

EP

components were analyzed by dissolving it in 0.5% nitric acid solution. The characterization results showed that the main component of the precipitates was struvite, with a crystal diameter of

AC C

30–50 µm. The chemical analysis results showed that the purity of the struvite precipitates was only 87.3% (± 0.8), and the contents of Mg, N, P, K, and Ca in it were 102.3 ± 1.5, 49.9 ± 0.4, 126.8 ± 0.8, 3.1 ± 0.4, and 5.5 ± 0.9 mg/g, respectively. Based on these results, it can be confirmed that the PT and TAN can be simultaneously removed and recovered from swine wastewater by the proposed recovery and recycling process of struvite. Fig.9. here 4. Conclusions The present study investigated the feasibility of the simultaneous recovery of PT and removal of TAN from swine wastewater by a coupled electrochemical process. The experimental results showed that using magnesium alloy as the electrochemical magnesium source of struvite 14

ACCEPTED MANUSCRIPT crystallization is feasible to recover phosphate from swine wastewater, and the current density and the reaction time are the key factors influencing the recovery of phosphate. The electrochemical decomposition of struvite is significantly influenced by the initial pH of solution, the current density, and the electrolysis time. Struvite could be completely decomposed as dissolved

RI PT

phosphate and magnesium ions and insoluble Mg3(PO4)2·22H2O. When the electrochemical decomposition product of struvite was recycled for the removal of TAN, a TAN-removal efficiency of >90% could be achieved. A pilot-scale test confirmed that a stable recovery and removal of phosphate and TAN from swine wastewater can be achieved by the proposed coupled

SC

electrochemical process. Based on the conclusions mentioned above, it can be confirmed that the proposed electrochemical process is suitable for the treatment of swine wastewater.

M AN U

Acknowledgments

This work was financially supported by the National Natural Science Foundation of China (Grant No. 51408529), the Natural Science Foundation of Hebei Province (Grant Nos. E2014203080), the Outstanding Young Scholars Project of Colleges and Universities of Hebei province (Grant No. BJ2014059) and China Postdoctoral Science Foundation Funded Project

TE D

(Grant No. 2015M580215).

References

An, L., Gu, G., 1993. The treatment of saline wastewater using a two-stage contact oxidation

EP

method. Water Sci. Technol. 28 (7), 31–37.

APHA, Standard Methods for the Examination of Water and Wastewater, Nineteenth ed. Public

Health

Association/American

Water

Works

Association/Water

AC C

American

Environment Federation, Washington, DC, USA, 1998.

Ben Moussa, S., Maurin, G., Gabrielli, C., Ben Amor, M., 2006. Electrochemical precipitation of struvite. Electrochem. Solid St. 9, 97–101.

Cardoso, A. M., Horn, M. B., Ferret, L. S., Azevedo, C.M.N., Pires, M., 2015. Integrated synthesis of zeolites 4A and Na-P1 using coal fly ash for application in the formulation of detergents and swine wastewater treatment. J. Hazard. Mater. 287, 69–77. Cao, F., Shi, Z., Hofstetter, J., Uggowitzer, P. J., Song, G., Liu, M., Atrens, A., 2013. Corrosion of ultra-high-purity Mg in 3.5% NaCl solution saturated with Mg(OH)2, Corros. Sci. 75, 78–99. Çelen, I., Buchanan, J.R., Burns, R.T., Robinson, R.B., Raman, D.R., 2007. Using a chemical 15

ACCEPTED MANUSCRIPT equilibrium model to predict amendments required to precipitate phosphorus as struvite in liquid swine manure. Water Res. 41, 1689–1696. Chen, G., 2004. Electrochemical technologies in wastewater treatment. Sep. Purif. Technol. 38, 11–41.

calcium phosphate. Water Sci. Technol. 64 (12), 2460–2467.

RI PT

Crutchik, D., Garrido, J.M., 2011. Struvite crystallization versus amorphous magnesium and

Demirer, S.U., Demirer, G.N., Frear, C., Chen, S., 2008. Anaerobic digestion of dairy manure with enhanced ammonia removal. J. Environ. Manage. 86, 193–200.

SC

Di Iaconi, C., Pagano, M., Ramadori, R., Lopez, A., 2010. Nitrogen recovery from a stabilized municipal landfill leachate. Bioresour. Technol. 101, 1732–1736.

M AN U

Gilbert, N., 2009. The disappearing nutrient. Nature 461, 716–718.

Gunay, A., Karadag, D., Tosun, I., Ozturk, M., 2008. Use of magnesit as a magnesium source for ammonium removal from leachate. J. Hazard. Mater. 156, 619–623. Hao, X., Wang, C., Mark C., van Loosdrecht, M., Hu, Y., 2013. Looking Beyond Struvite for P-Recovery. Environ. Sci. Technol. 47, 4965−4966.

TE D

He, S., Zhang, Y., Yang, M., Du, W., Harada, H., 2007. Repeated use of MAP decomposition residues for the removal of high ammonium concentration from landfill leachate. Chemosphere 66, 2233–2238.

EP

Hirooka, K., Ichihashi, O., 2013. Phosphorus recovery from artificial wastewater by microbial fuel cell and its effect on power generation. Bioresour. Technol. 137, 368–375.

AC C

Hug, A., Udert, K.M., 2013. Struvite precipitation from urine with electrochemical magnesium dosage. Water Res. 47, 289–299.

Huang, H., Song, Q., Xu, C., 2011a. Removal of ammonium from aqueous solutions using the residue obtained from struvite pyrogenation. Water Sci. Technol. 2011, 64(12), 2508–2514.

Huang H., Xu, C., Zhang, W., 2011b. Removal of nutrients from piggery wastewater using struvite precipitation and pyrogenation technology. Bioresour. Technol. 102, 2523–2528. Huang, H., Jiang, Y., Ding, L., 2014. Recovery and removal of ammonia–nitrogen and phosphate from swine wastewater by internal recycling of struvite chlorination product. Bioresour. 16

ACCEPTED MANUSCRIPT Technol. 172, 253–259. Ichihashi, O., Hirooka, K., 2012. Removal and recovery of phosphorus as struvite from swine wastewater using microbial fuel cell. Bioresour. Technol. 114, 303–307. Kemacheevakul, P., Polprasert, C., Shimizu, Y., 2011. Phosphorus recovery from human urine

RI PT

and anaerobically treated wastewater through pH adjustment and chemical precipitation. Environ. Technol. 32, 693–698.

Kruk, D. J., Elektorowicz, M., Oleszkiewicz, J. A., 2014. Struvite precipitation and phosphorus removal using magnesium sacrificial anode. Chemosphere 101, 28−33.

SC

Lahav, O., Telzhensky, M., Zewuhn, A., Gendel, Y., Gerth, J., Calmano,W., Birnhack, L., 2013. Struvite recovery from municipal-wastewater sludge centrifuge supernatant using seawater

M AN U

NF concentrate as a cheap Mg(II) source. Sep. Purif. Technol. 108, 103–110. Lee, S.I., Weon, S.Y., Lee, C.W., Koopman, B., 2003. Removal of nitrogen and phosphate from wastewater by addition of bittern. Chemosphere 51, 265–271.

Li, X.Z., Zhao, Q.L., Hao, X.D., 1999. Ammonium removal from landfilll leachate by chemical precipitation. Waste Manage. 19, 409–415.

TE D

Li, X. Z., Zhao, Q.L., 2003. Recovery of ammonium-nitrogen from landfill leachate as a multi-nutrient fertilizer. Ecol. Eng. 20, 171–181. Lim, S. J., Kim, T. H., 2015. Combined treatment of swine wastewater by electron beam

EP

irradiation and ion-exchange biological reactor system. Sep. Purif. Technol. 146, 42–49. Liu, Y.H., Kumar, S., Kwang, J.H., Kim, J.H., Kim, J.D., Ra, C.S., 2011. Recycle of

AC C

electrolytically dissolved struvite as an alternative to enhance phosphate and nitrogen recovery from swine wastewater. J. Hazard. Mater. 195, 175–181.

Long, T., Anas, F., Al-Harbawi, H., Zhai, Q., 2007. Comparison between biological treatment and chemical precipitation for nitrogen removal from old landfill leachate. Am. J. Environ. Sci. 3 (4), 183–187.

Mijangos, F., Kamel, M., Lesmes, G., Muraviev, D.N., 2004. Synthesis of struvite by ion exchange isothermal supersaturation technique. React. Funct. Polym. 60, 151–161. 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. 17

ACCEPTED MANUSCRIPT Desalination 285, 1–13. Musvoto, E.V., Wentzel, M.C., Ekama, G.A., 2000. Integrated chemical-physical processes modelling-II. Simulating aeration treatment of anaerobic digester supernatants. Water Res. 34, 1868–1880.

RI PT

Nelson, N.O., Mikkelsen, R.L., Hesterberg, D.L., 2003. Struvite precipitation in anaerobic swine lagoon liquid: effect of pH and Mg:P ratio and determination of rate constant. Bioresour. Technol. 89, 229–236.

Okano, K., Uemoto, M., Kagami, J., Miura, K., Aketo, T., Toda, M., Honda, K., Ohtake, H., 2013.

silicate hydrates (A-CSHs). Water Res. 47, 2251–2259.

SC

Novel technique for phosphorus recovery from aqueous solutions using amorphous calcium

M AN U

Pérez, G., Saiz, J., Ibañez, R., Urtiaga, A.M., Ortiz, I., 2012. Assessment of the formation of inorganic oxidation by-products during the electrocatalytic treatment of ammonium from landfill leachates. Water Res. 46, 2579–2590.

Qiu, G., Ting, Y., 2014. Direct phosphorus recovery from municipal wastewater via osmotic membrane bioreactor (OMBR) for wastewater treatment. Bioresour. Technol. 170, 221–229.

TE D

Rathod, M., Mody, K., Basha, S., 2014. Efficient removal of phosphate from aqueous solutions by red seaweed, Kappaphycus alverezii. J. Clean. Prod. 84, 484–493. Romero-Güiza, M.S., Astals, S., Mata-Alvarez, J., Chimenos, J.M., 2015. Feasibility of coupling

EP

anaerobic digestion and struvite precipitation in the same reactor: Evaluation of different magnesium sources. Chem. Eng. J. 270, 542–548.

AC C

Saidou, H., Korchef, A., Ben Moussa, S., Ben Amor, M., 2009. Struvite precipitation by the dissolved CO2 degasification technique: Impact of the airflow rate and pH. Chemosphere 74, 338–343.

Song, G., Atrens, A., 2003. Understanding magnesium corrosion–a framework for improved alloy performance. Adv. Eng. Mater. 5, 837–858.

Stratful, I., Scrimshaw, M.D., Lester, J.N., 2001. Conditions influencing the precipitation of magnesium ammonium phosphate. Water Res. 35, 4191–4199. Sugiyama, S., Yokoyama, M., Ishizuka, H., Sotowa, K., Tomida, T., Shigemoto, N., 2005. Removal of aqueous ammonium with magnesium phosphates obtained from the ammonium-elimination of magnesium ammonium phosphate. J. Colloid Inter. Sci. 292, 18

ACCEPTED MANUSCRIPT 133–138. Thomas, S., Medhekar, N.V., Frankel, G.S., Birbilis, N., 2015. Corrosion mechanism and hydrogen evolution on Mg. Curr. Opin. Solid St. Mater. Sci. 19, 85–94. Tran, A.T.K., Zhang, Y., De Corte, D., Hannes, J., Ye, W., Mondal, P., Jullok, N., Meesschaert,

RI PT

B., Pinoy, L., Bruggen, B.V., 2014. P-recovery as calcium phosphate fromwastewater using an integrated selectrodialysis/crystallization process. J. Clean. Prod. 77, 140–151.

Türker, M., Çelen, I., 2007. Removal of ammonia as struvite from anaerobic digester effluents and recycling of magnesium and phosphate. Bioresour. Technol. 98, 1529–1534.

SC

Vanlangendonck, Y., Cornisier, D., Lierde, A.V., 2005. Influence of operating conditions on the ammonia electro-oxidation rate in wastewaters from power plants. Water Res. 39,

M AN U

3028–3034.

Wang, J., Song, Y., Yuan, P., Peng, J., Fan, M., 2006. Modeling the crystallization of magnesium ammonium phosphate for phosphorus recovery. Chemosphere 65, 1182–1187. Wang, C. C., Hao, X. D., Guo, G. S., Van Loosdrecht, M.C.M., 2010. Formation of pure struvite at neutral pH by electrochemical deposition. Chem. Eng. J. 159, 280–283.

TE D

Wilsenach, J.A., Schuurbiers, C. A. H., van Loosdrecht, M. C. M., 2007. Phosphate and potassium recovery from source separated urine through struvite precipitation. Water Res. 41, 458–466. Wu, Y., Zhou, S., 2012. Improving the prediction of ammonium nitrogen removal through struvite

EP

precipitation. Environ. Sci. Pollu. Res. 19, 347–360. Xie, Z.M., Li, X.Y., Chan, K.Y., 2006. Nitrogen removal from the saline sludge liquor by

AC C

electrochemical denitrification. Water Sci. Technol. 54, 171–179. Ye, Z., Chen, S., Wang, S., Lin, L., Yan, Y., Zhang, Z., Chen, J., 2010. Phosphorus recovery from synthetic swine wastewater by chemical precipitation using response surface methodology. J. Hazard. Mater. 176, 1083–1088.

Yetilmezsoy, K., Sapci-Zengin, Z., 2009. Recovery of ammonium nitrogen from the effluent of UASB treating poultry manure wastewater by MAP precipitation as a slow release fertilizer. J. Hazard. Mater. 166, 260–269. Zhang, Y., Desmidt, E., Van Looveren, A., Pinoy, L., Meesschaert, B., Van der Bruggen, B., 2013. Phosphate separation and recovery from wastewater by novel electrodialysis. Environ. Sci. Technol. 47, 5888−5895 19

ACCEPTED MANUSCRIPT Zhang, T., Ding, L., Ren, H., Xiong, X., 2009. Ammonium nitrogen removal from coking

AC C

EP

TE D

M AN U

SC

RI PT

wastewater by chemical precipitation recycle technology. Water Res. 43, 5209–5215.

20

ACCEPTED MANUSCRIPT Table captions Table 1. The characteristics of pretreated swine wastewater used in the experiments. Table 2. The main characteristics of the filtered seawater used in the experiments. Figure captions Fig.1. Schematic illustration of the combined electrochemical process for the simultaneous

RI PT

removal of PT and TAN from swine wastewater (R1: PT recovery reactor; R2: struvite decomposition reactor; R3: TAN removal reactor; A: steel mesh; B: Mg alloy plate; C: graphite plate).

Fig. 2. Variations in solution pH (a), PT-recovery efficiency (b), and struvite purity (c) as a function of current density versus electrolysis time.

SC

Fig. 3. XRD diffractogram (a) and SEM micrograph (b) of the precipitates obtained in the swine wastewater treatment at a current density of 2 mA/cm2 for 45 min.

M AN U

Fig. 4. The effects of the initial solution pH and current density on struvite decomposition, (a) solution pH during decomposition, and (b) ammonium decomposition rate at different initial solution pHs and decomposition reactions (current density, 84 mA/cm2), (c) solution pH during decomposition, and (d) ammonium decomposition rate at different current densities and decomposition reactions (initial solution pH of 8).

Fig. 5. (a) Changes in the TAN concentration and pH of the electrolysis solution and

TE D

decomposition rate of ammonium in struvite with reaction time, and (b) comparison of the actual concentration of PT in the electrolysis solution and calculated PT concentration at different reaction times.

Fig. 6. Diagrammatic representation of the electrochemical decomposition stages of struvite: 1)

EP

first stage, the rapid dissolution of struvite during the initial reaction time of 50 min, 2) second stage, the rapid formation of MgHPO4·3H2O within 50–110 min, and 3) third stage, the formation of Mg3(PO4)2·22H2O between 110–180 min.

AC C

Fig. 7. Changes in the TAN-removal efficiency (a) and remaining PT concentration (b) by recycling the electrochemical decomposition product of struvite with reaction times at different pHs.

Fig. 8. TAN-removal efficiency and remaining PT concentration by recycling the electrochemical decomposition product of struvite at different time periods. Fig. 9. The overall performance for the PT recovery and TAN removal by the struvite recycling process proposed.

21

ACCEPTED MANUSCRIPT Table 1

Table 1. The characteristics of pretreated swine wastewater used in the experiments. Average values plus standard deviation

pH COD (mg/L) Alkalinity (as Na2CO3) (mg/L) PT (total orthophosphate, mg/L)

7.8 ± 0.06 4105 ± 327 3135 ± 293 103 ± 9.8

TAN (total ammonia nitrogen, mg/L) K (mg/L) Ca (mg/L) Mg (mg/L ) Fe (mg/L) Cu (mg/L) Zn (mg/L)

426 ± 21 293 ± 18 64 ± 7.2 13 ± 2.5 1.6 ± 0.5 0.8 ± 0.1 0.5 ± 0.1

M AN U

SC

RI PT

Parameter

Table 2

Parameter

TE D

Table 2. The main characteristics of the filtered seawater used in the experiments. Average values plus standard deviation 18355 ± 1508 9970 ± 701 369 ± 52

Mg2+ (mg/L) K+ (mg/L)

1521 ± 118 493 ± 39

AC C

EP

Cl– (mg/L) Na+ (mg/L) Ca2+ (mg/L)

22

ACCEPTED MANUSCRIPT

TE D

M AN U

SC

RI PT

Figure 1

EP

Fig.1. Schematic illustration of the combined electrochemical process for the simultaneous removal of PT and TAN from swine wastewater (R1: PT recovery reactor; R2: struvite

AC C

decomposition reactor; R3: TAN removal reactor; A: steel mesh; B: Mg alloy plate; C: graphite plate).

23

ACCEPTED MANUSCRIPT

AC C

EP

TE D

M AN U

SC

RI PT

Figure 2

Fig. 2. Variations in solution pH (a), PT-recovery efficiency (b), and struvite purity (c) as a function of current density versus electrolysis time.

24

ACCEPTED MANUSCRIPT

M AN U

SC

RI PT

Figure 3

Fig. 3. XRD diffractogram (a) and SEM micrograph (b) of the precipitates obtained in the swine

AC C

EP

TE D

wastewater treatment at a current density of 2 mA/cm2 for 45 min.

25

ACCEPTED MANUSCRIPT

AC C

EP

TE D

M AN U

SC

RI PT

Figure 4

Fig. 4. The effects of the initial solution pH and current density on struvite decomposition, (a) solution pH during decomposition, and (b) ammonium decomposition rate at different initial solution pHs and decomposition reactions (current density, 84 mA/cm2), (c) solution pH during decomposition, and (d) ammonium decomposition rate at different current densities and decomposition reactions (initial solution pH of 8).

26

ACCEPTED MANUSCRIPT

AC C

EP

TE D

M AN U

SC

RI PT

Figure 5

Fig. 5. (a) Changes in the TAN concentration and pH of the electrolysis solution and decomposition rate of ammonium in struvite with reaction time, and (b) comparison of the actual concentration of PT in the electrolysis solution and calculated PT concentration at different reaction times.

27

ACCEPTED MANUSCRIPT

AC C

EP

TE D

M AN U

SC

RI PT

Figure 6

Fig. 6. Diagrammatic representation of the electrochemical decomposition stages of struvite: 1) first stage, the rapid dissolution of struvite during the initial reaction time of 50 min, 2) second stage, the rapid formation of MgHPO4·3H2O within 50–110 min, and 3) third stage, the formation of Mg3(PO4)2·22H2O between 110–180 min.

28

ACCEPTED MANUSCRIPT

AC C

EP

TE D

M AN U

SC

RI PT

Figure 7

Fig. 7. Changes in the TAN-removal efficiency (a) and remaining PT concentration (b) by recycling the electrochemical decomposition product of struvite with reaction times at different pHs.

29

ACCEPTED MANUSCRIPT

TE D

M AN U

SC

RI PT

Figure 8

EP

Fig. 8. TAN-removal efficiency and remaining PT concentration by recycling the electrochemical

AC C

decomposition product of struvite at different time periods.

30

ACCEPTED MANUSCRIPT

M AN U

SC

RI PT

Figure 9

AC C

EP

TE D

Fig. 9. The overall performance for the PT recovery and TAN removal by the struvite recycling process proposed.

31

ACCEPTED MANUSCRIPT Highlights

A coupled electrochemical process was developed to treat swine wastewater.




Struvite of high purity can be obtained by electrochemical precipitation.




About 93% nutrients were recovered from swine wastewater by the proposed process.

AC C

EP

TE D

M AN U

SC

RI PT