Sequential “electrochemical peroxidation – Electro-Fenton” process for anaerobic sludge treatment

Water Research 154 (2019) 277e286

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Sequential “electrochemical peroxidation e Electro-Fenton” process for anaerobic sludge treatment H. Olvera-Vargas, X. Zheng, O. Garcia-Rodriguez, O. Lefebvre* Centre for Water Research, Department of Civil and Environmental Engineering, National University of Singapore, 1 Engineering Drive 2, Singapore, 117576, Singapore

a r t i c l e i n f o

a b s t r a c t

Article history: Received 22 October 2018 Received in revised form 21 January 2019 Accepted 30 January 2019 Available online 15 February 2019

In this study, we present a sequential electrochemical process for integral treatment of anaerobic sludge, combining for the first time electrochemical peroxidation (ECP) and electro-Fenton (EF). In the first step, ECP (consisting of H2O2-assisted electrocoagulation with Fe electrodes) was applied as a conditioning and stabilizing method, whose synergistic electrocoagulation/Fenton oxidation effects considerably reduced the COD, TOC and total suspended solids (TSS) by 89.3%, 75.4% and 85.6%, respectively, under optimized conditions (initial pH of 5, [H2O2]/[Fe2þ] dose ratio of 5, 15.38 mA cm2 and 2 h treatment). Furthermore, total coliforms were completely killed within the first hour of treatment. In the second step, EF was successfully applied to mineralize the remaining organic fraction in the liquid effluent after dewatering, achieving 91.6% and 87.2% of COD and TOC removal, respectively, after 4 h of treatment under optimal conditions (pH 3 and 25 mA cm2), while almost total COD and TOC removal was attained in 8 h. The Fe sludge generated at the end of the ECP treatment was easily dewatered by filtration and 20.9 g of nutrient-rich dry sludge were produced. The overall cost of the ECP-EF treatment was S$ 0.05 L-1 sludge. The combined effects of coagulation and Fenton oxidation during ECP revealed that the treatment efficiency is strongly dependent on the rheological properties of the sludge sample. © 2019 Elsevier Ltd. All rights reserved.

Keywords: Anaerobic sludge Electrochemical peroxidation Electrocoagulation Electro-Fenton Wastewater treatment

1. Introduction The agricultural and agro-industrial sectors produce enormous amounts of wastewater entailing potentially severe pollution issues. Food-processing wastewater treatment is generally conducted by systems composed of conventional physicochemical and biological units (aerobic and anaerobic), where treatment is accompanied by the generation of large amounts of sludge requiring further treatment and disposal (Christensen et al., 2015). The annual production of waste sludge is estimated at about 45 million dry tons per year (Q. Zhang et al., 2017a), out of which the USA, Europe and China contribute together 18.4 million dry tons (Mateo-Sagasta et al., 2015). Because sludge treatment and disposal is difficult and expensive, sludge management represents over 60% of the total cost of conventional wastewater treatment plants (WWTPs) (Zhou et al., 2014). Sludge is generally composed of microbial cells, extracellular polymeric substances (EPS) excreted by bacteria (representing 80% of the mass content), recalcitrant

* Corresponding author. E-mail address: [email protected] (O. Lefebvre). https://doi.org/10.1016/j.watres.2019.01.063 0043-1354/© 2019 Elsevier Ltd. All rights reserved.

pollutants and water (free and bound) (Zhou et al., 2014). Bound water entrapped within the EPS confers high viscosity to the sludge, limiting its dewaterability and settling (Liu and Fang, 2003). The great challenge of sludge treatment is indeed to reduce water content so it can be managed as a solid. Sludge treatment consists of different steps: thickening, stabilization (to decrease sludge mass), conditioning (to increase dewaterability), dewatering (to minimize sludge volume) and final disposal or reuse. Dewatering is a crucial step, impacting directly on transportation and disposal costs (Christensen et al., 2015; Zhou et al., 2014). Incineration, landfilling and land application are the most commonly used sludge disposal options (Q. Zhang et al., 2017a). For land application, wastewater sludge is an excellent source of nutrients for agricultural production. Nevertheless, it is generally loaded with human pathogens that represent an important health threat (Lowman et al., 2013). Consequently, disinfection is compulsory when developing sludge treatment technologies. Conventional physicochemical and biological treatment technologies have shown limited performance due to the complexity of the sludge. Advanced oxidation processes (AOPs) have been used as potent conditioning methods due to their capability to disrupt

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sludge flocs and degrade extracellular polymeric substances, thus enhancing dewaterability (Wang et al., 2017; Zhou et al., 2014). However, some AOPs (e.g., ozonation and ultrasound) are still limited to tertiary treatment applications due to their high energy and chemical demand (Wang et al., 2017). Electrocoagulation (EC) is a powerful electrochemical technology that has found diverse applications for the treatment of highly polluted industrial wastewater (Hakizimana et al., 2017). During EC, the electrochemical production of coagulants from a sacrificial anode promotes contaminant agglomeration and precipitation. Both Al and Fe electrodes have given satisfactory performances; however, Fe has the advantage of being non-toxic and cheaper than Al (Hakizimana et al., 2017). The formation of Fe(OH)3 species by dissolution of an Fe anode follows Eqs. (1) and (2), while H2 evolves at the cathode according to Eq. (3). EC with Fe anodes has been coupled to the chemical Fenton process by external addition of H2O2, which promotes OH-based oxidation of organic pollutants, in addition to the coagulation mechanism. This process is known as electrochemical peroxidation (ECP) (Chiarenzelli et al., 2001). ECP has been applied for the degradation of refractory organics in wastewater via OH, such as industrial wastewater (Kumar et al., 2018), wastewater from a paper recycling plant (Moussavi and Aghanejad, 2014) and pretreated coke wastewater (Ozyonar and Karagozoglu, 2015). Nonetheless, the coagulation effect has generally been neglected and poorly investigated. Fe / Fe2þ þ 2e

(1)

4Fe2þ þ 10H2O þ O2(g) / 4Fe(OH)3(s) þ 8Hþ

(2)

2H2O þ 2e / 2OH þ H2(g)

(3)

Electrochemical advanced oxidation processes (EAOPs) have shown multiple advantages in the treatment of wastewater from many different sources (Garcia-Rodriguez et al., 2018; Martínez
s et al., 2014). These Huitle et al., 2015; Moreira et al., 2017; Sire environmentally friendly methods are characterized by the in-situ generation of OH and the lack of secondary sludge formation. Moreover, they operate under mild conditions and can economi~ izares et al., 2009). The electrocally outcompete other AOPs (Can Fenton process (EF) is one of the most widespread EAOPs and relies on the generation of OH in the bulk solution by means of the Fenton's reaction (Eq. (4)). The cathodic production of the Fenton's reagent (H2O2 and Fe2þ) on carbonaceous electrodes is favoured in acidic solutions containing catalytic amounts of Fe2þ ions, accord
s and Brillas, 2017). In ing to Eqs. (5) and (6) (Brillas et al., 2009; Sire addition, EF efficiency can be enhanced when it is coupled to anodic oxidation (AO) by using a powerful boron-doped diamond (BDD) anode, which produces BDD(OH) at its surface according to Eq. (7) (Sopaj et al., 2016). H2O2 þ Fe2þ / Fe3þ þ OH þ OH

(4)

O2 þ 2Hþ þ 2e / H2O2

(5)

3þ

Fe



þ e / Fe

2þ

BDD þ H2O / BDD(OH) þ Hþ þ e

(6) (7)

Undoubtedly, electrochemical technologies represent promising options for wastewater treatment, especially for decentralised systems (Cid et al., 2018). In fact, electrochemical methods have already been applied to sludge dewatering and stabilization, but results were not conclusive and organic matter removal remained low (Bureau et al., 2012; Li et al., 2016).

In this context, this work presents a sequential treatment strategy based on electrochemical technologies for integral treatment of thick and viscous anaerobic sludge. This kind of sludge generally presents poorer dewaterability than waste activated sludge (Zhang et al., 2015). In the first step, sludge conditioning and stabilization was performed by means of ECP with optimization of the main operating parameters influencing the process (initial pH, [H2O2]/[Fe2þ] ratio and current). The synergistic effects of coagulation and oxidation during ECP of viscous sludge were investigated in detail for the first time. In the second stage, EF was used to achieve complete mineralization of the remaining organic fraction in the wastewater separated from the sludge. An economic assessment is presented in the last section. The ECP-EF process is proposed as an efficient, cost-effective and sustainable solution for integral sludge remediation. To the best of our knowledge, there is no report in the literature of such a comprehensive sludge treatment capable of delivering an effluent with minimal amounts of organic matter, suitable for reuse. 2. Materials and methods 2.1. Sample collection and characteristics The sludge sample originated from a local poultry farm in Singapore, which generates approximately 150 tons per day of anaerobic sludge, requiring proper attention before disposal. Samples were collected and stored in polypropylene containers at 4  C. The chemical characterization of the raw sample is presented in Table 1. It contains very high Chemical Oxygen Demand (COD) and total suspended solid (TSS). 2.2. Experimental setup Prior to electrolysis, K2SO4 (0.1 M) was added to the sample as supporting electrolyte and the pH was adjusted to the desired value using H2SO4. All experiments were conducted in a batch glass electrolytic reactor with 0.5 L capacity, powered by a HAMEG 7042e5 power supply (Germany). For EC and ECP trials, two mild steel plate electrodes (5  6.5  0.5 cm) e referred to as Fe electrodes in the text e were positioned in the middle of the cell facing each other and separated by a distance of 4 cm. The working volume was 400 mL and the experiments were conducted at room temperature with constant stirring at 300 rpm. For ECP experiments, H2O2 (30% w/w) was manually added at a fixed interval of 10 min. H2O2 concentration was calculated following the stoichiometry of the Fenton's reaction and based on the theoretical production of Fe2þ according to the Faraday's law (Eq. (8)): m ¼ ItMw/nF

(8)

Table 1 Chemical composition of the anaerobic sludge. Parameters

Values

pH COD (mg L1) BOD5 (mg L1) TOC (mg L1) TSS (mg L1) TDS (mg L1) PePO4 (mg L1) NeNH4 (mg L1) NeNO3 (mg L1) Total N (mg L1) Total coliforms (CFU mL1) Fe (mg L1)

7.8e8.2 26,250.5 ± 2071.1 3,980.0 ± 492.8 5,235.6 ± 320.4 31,333.3 ± 2494.1 9,400.2 ± 860.1 680.4 ± 31.8 473.6 ± 26.7 80.4 ± 3.4 8,325.5 ± 566.1 2,375.0 ± 75.0 8.1 ± 0.4

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Where m is the mass (g), I is the current (A), t is the time (s), Mw is Fe molecular weight (55.85 g mol1), n is the number of electrons (2) and F is the Faraday constant (96,487 C mol1). Prior to any EC/ECP experiment, the mild steel electrodes were washed in a 1 M HCl solution and sandpaper-polished to remove any oxide layer. The pH of the treated sample was adjusted to 9 at the end of ECP in order to promote the precipitation of Fe(OH)3. The treated solution was vacuum-filtered to separate the solid sludge from the filtrate. The solid sludge was dried in an oven at 55  C overnight for further analysis. 175 mL of the filtrate were recovered and acidified to pH 3 before EF treatment. For EF, the electrochemical cell (250-mL) was equipped with a BDD anode (15  2.5  0.1 cm) centred in the middle surrounded by a carbon brush cathode (polyacrylonitrile (PAN)-based carbon fibres shaped

Total Cost S$ L  1 sludge ¼

In accordance with Singapore Power (SP) 2018 tariffs, energy cost for the combined ECP-EF process was estimated at S$0.16 kWh. For chemicals, industrial grade prices were used: H2SO4 (96% w/w) at S$0.2 kg1, KOH at S$0.2 kg1, K2SO4 at S$0.2 kg1 and H2O2 (30% w/w) at S$0.4 kg1 (www.aikmoh.com.sg). Disposal costs for dehydrated sludge, including transportation and charges for waste disposal, were evaluated at S$0.1 kg1 of wet dehydrated sludge, assuming that these residues are not considered hazardous biosolids (National Environment Agency of Singapore, NEA). The consumption of the sacrificial anode was calculated using Eq. (8), and the cost of the dissolved Fe2þ mass was calculated based on the price of industrial grade mild steel plates (S$0.8 kg1, www. sdsteelpate.cn). The total cost per litre was then calculated according to Eq. (10):

Power cost þ Chemical cost þ Sludge cost þ Anode consumption Volume of sample ðLÞ

in the form of a brush with stainless steel as current collector) around the internal walls of the reactor (Mousset et al., 2017). Compressed air was continuously supplied to the solution through a gas diffuser at a flow rate of 1 L min1. All experiments were conducted in triplicate and the average values and standard deviations are reported in all tables and figures. 2.3. Analytical methods COD, phosphate (P-PO4), nitrite (NeNO2), nitrate (NeNO3) and ammonium (NeNH4) concentrations were determined by colorimetric standard methods using HACH vials tests. Total Organic Carbon (TOC) was measured using a Shimadzu TOC-V CSH analyser. Biochemical oxygen demand at 5 days (BOD5) was determined by respirometric method using an OxiTop® control system (WTW GmbH, Germany). The treatment efficiency was evaluated in terms of TSS, COD and TOC removal yields. For ECP experiments, the samples were analysed after 1-h settling time. The pH was measured with a VWR pHenomenal MU 6100H pH-meter. All samples were filtered with polypropylene 0.45 mm syringe filters prior to analysis. TSS content was measured using standard methods (American Public Health Association, 1999). Total coliforms were determined by filtration following the ISO 9308-1 method. For this purpose, 50 mL of diluted samples were filtered with 0.45 mm membranes and the membranes were spread into plates containing selective eosin methylene blue agar. The plates were incubated at 37  C for 24 h followed by colony counting. Iron determination was carried out by inductively coupled plasma optical emission spectrometry (ICP-OES, PerkinElmer, USA).

279

(10)

3. Results and discussion 3.1. Preliminary tests: ECP mechanisms The primary objective of this section is to investigate the best strategy for sludge conditioning, aiming at disrupting the floc structure of the sludge. For this purpose, three experimental strategies were explored: i) EC with Fe electrodes, ii) Peroxicoagulation (PC) using a graphite cathode and a Fe anode and iii) ECP. The electrolysis tests were performed at 500 mA (corresponding to 15.4 mA cm2) and the TSS, COD and TOC removal efficiencies are presented in Fig. 1. The control experiment with H2O2 but without external current evidence the low oxidative power of H2O2 in the absence of Fe2þ. The results further show limited TSS, COD and TOC removal efficiencies by EC and PC, while the best removal yields were achieved with ECP (85.6, 89.3 and 75.4% of TSS, COD and TOC removal efficiency, respectively). These results can be explained by the

2.4. Cost calculation For feasibility evaluation, an economic analysis was conducted, taking into consideration the costs associated to chemical and energy consumption, sludge disposal and the consumption of the sacrificial anode. The energy consumption was calculated based on Eq. (9) (Brillas et al., 2009): Energy consumption (kWh L1) ¼ (EcellIt/1000Vs)

(9)

Where Ecell is the applied voltage (V), I is the current (A), t is the time (h) and Vs is the sample volume (L).

Fig. 1. Removal efficiency of preliminary experiments. Experimental conditions: V ¼ 400 mL, pH ¼ 5, Na2SO4 ¼ 0.1 M, j ¼ 15.38 mA cm2, total [H2O2] ¼ 0.24 M (for the control and ECP experiments) and 2 h-treatment. The photographs show the sample evolution over time.

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complexity of the sludge sample (high content of organic matter and suspended solids, as well as low filterability). In the case of EC, the coagulation process promoted by the electrochemical release of Fe3þ ions was not efficient enough to disrupt the polymeric materials, break down microbial cells and induce proper coagulation. This behaviour is similar to that reported by (Zhen et al., 2013). For PC, two issues were encountered: first, due to the high viscosity of the sample, limitations in O2 mass transport towards the cathode resulted in low electrochemical production of H2O2 on the surface of graphite (Eq. (5)), which restrained the formation of OH via the Fenton's reaction (Eq. (4)). Second, the PC experiment was conducted at pH 5 because of the nature of the sample, acidification requiring huge amounts of H2SO4 and time, owing to acidhydrolysis reactions (Neyens et al., 2003). Consequently, the rate of the Fenton's reaction was slow at this pH value, due to the limited availability of Fe3þ/Fe2þ ions in solution. Indeed, it is well known that the Fenton's reaction has an optimal operating pH of around 3 according to Fe3þ ions speciation in solution (Pignatello et al., 2006). Yet, one PC experiment was performed at pH 3 (results not shown) and the removal efficiencies were not significantly improved, pointing out that the physicochemical properties of the sample (high viscosity) more than the slow kinetics of the Fenton's reaction at pH 5 were the main factors limiting the efficacy of EF. Thus, it was demonstrated that the electrochemical production of H2O2 on graphite electrodes during PC is not suitable for highly polluted anaerobic sludge with high viscosity. With respect to ECP, the external addition of H2O2 ([H2O2]/ [Fe2þ] ¼ 5) significantly enhanced the removal efficiency, resulting in a dramatic improvement as compared to EC experiments. In fact,  OH produced from Eq. (4) oxidized the organic material e EPS (composed mainly of polysaccharides, proteins and DNA), humiclike substances, lipids and cell membranes (phospholipids and amino acids) e ultimately provoking cell disruption and sludge floc disintegration (Yu et al., 2018). This process released the bound water and exposed the cellular material (EPS) to further OH attacks, leading to mineralization of small molecules, such as amino acids, carbohydrates, proteins, peptides, alcohols, epoxides and carboxylic acids (Erden and Filibeli, 2010; Neyens et al., 2004). Additionally, it has been reported that the cleavage of macromolecules during OH oxidation in the Fenton process diminishes the strength of sludge flocs, resulting in a decrease of viscosity (Pham et al., 2010). In this way, sludge conditioning and stabilization was effectuated by an intricate coagulation-Fenton's oxidation process. Coagulation contributed to 41% and 37% of COD and TOC removal efficiencies, respectively, only slightly lower than Fenton oxidation, which accounted for 48% and 38% of COD and TOC removal, respectively (Fig. 1). Fe-oxyhydoxy precipitates assisted the destabilization of the sludge flocs by compression of the double layer of colloidal particles, charge neutralization and entrapment in the Fe coagulants (Harif et al., 2012). Then, one fraction of the organic material (EPS, cells and by-products from OH cleavage) was removed by adsorption and complexation on Fe-hydroxides species that precipitated following flocculation (Moreno-Casillas et al., 2007). The oxidation and coagulation mechanisms taking place during ECP are schematized in Fig. 2, while the inset panel shows the production of Fe coagulants and OH in the ECP reactor. The inset panel in Fig. 1 depicts the visual appearance of the samples throughout the experimental period. 3.2. ECP optimization Taking into account preliminary experiments, ECP was selected as the most suitable conditioning and stabilizing method. Optimization was then carried out with relation to the main parameters affecting ECP performance, namely: pH, current and [H2O2]/[Fe2þ]

ratio. 3.2.1. Effect of current density Current density is a key operating parameter affecting the performance of electrochemical processes by controlling the rate of electrochemical reactions. In the case of EC and related processes, it is particularly important, as current is directly linked to the release of Fe2þ ions (Eq. (1)), which generally follows the Faraday's law. In general, Fe2þ is quickly oxidized as the pH increases (Eq. (2)) forming Fe(OH)3 as the main coagulant species (Lakshmanan et al., 2009). The amount of Fe(OH)3 species in solution then determines the coagulation efficiency (Garcia-Segura et al., 2017). In this work, no Fe2þ was detected in solution after 2 h of ECP, which can be accounted for by the progressive increase of pH, constant stirring and the presence of H2O2. In addition, current controls the cathodic evolution of H2 (Eq. (3)), which entails an increase of pH during electrolysis. When H2O2 is used (ECP), the amount of dissolved Fe2þ/Feþ3 also determines the production of OH through the Fenton's reaction (Eq. (4)). Thus, the performance of ECP was investigated at different current values (3.08 mA cm2, 15.38 mA cm2 and 30.77 mA cm2) with initial pH 5 and [H2O2]/ [Fe2þ] ¼ 5, following preliminary tests. Results are presented in Fig. 3, showing that ECP performance improved as current increased from 3.08 mA cm2 to 15.38 mA cm2, with the latter value giving the best TSS, COD and TOC removal efficiencies of 85.6%, 89.3% and 75.4%, respectively. This behaviour can be explained by the increase of Fe2þ ions with current density, according to the Faraday's law. Thus, a higher coagulant dose enhanced the coagulation efficiency, while at the same time, more Fe2þ/Fe3þ free ions in solution increased OH production and, consequently, improved the efficiency of Fenton oxidation. However, the application of a higher current value (30.77 mA cm2) did not result in further efficiency improvement, owing to: i) the higher reaction rate of H2O2 decomposition at the anode (Eqs. (11) and (12)), ii) coagulant overdosing that can modify the charge of aggregates, causing their redispersion and decreasing the coagulation/flocculation efficiency (Harif and Adin, 2011), and, most importantly iii) an excess of Fe2þ ions, promoting the competitive waste reactions 13 and 14, consuming OH in detriment of organics oxidation (with Eq. (13) having a higher kinetic constant (k13 ¼ 3.2  108 M1 s1) than Eq. (14) (k ¼ 2.7  103 s1, for the unimolecular decomposition of Fe3þ-hydroxyperoxy complexes formed as intermediates) (Brillas et al., 2009; Pignatello et al., 2006). Moreover, the rise in current density (accompanied by higher potential) entails heat generation and higher energy consumption and thus, the optimal current density of 15.38 mA cm2 was selected for further experiments. H2O2 / HO2 þ Hþ þ e

(11)

HO2 / O2 þ Hþ þ e

(12)

Fe2þ þ OH / Fe3þ þ HO

(13)

Fe3þ þ H2O2 / Fe2þ þ HO2 þ Hþ

(14)

3.2.2. Effect of pH pH plays a fundamental role in ECP as it affects iron speciation in solution, as well as the complexation and redox cycling between Fe ions (Brillas et al., 2009; Lakshmanan et al., 2009). The aqueous chemistry of Fe species determines both coagulation efficiency and the reactivity of the Fenton's reaction. To examine the effect of pH, a series of experiments with pH ranging from 3 to 8 (with 8 being the

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281

Fig. 2. Degradation mechanisms taking place during ECP of anaerobic sludge. Inset panel: production of coagulants and oxidants during ECP using mild steel electrodes.

Fig. 3. Effect of current on TSS, COD and TOC removal efficiencies at pH 5 and [H2O2]/ [Fe2þ] ¼ 5. ECP with Fe electrodes. V ¼ 400 mL, Na2SO4 ¼ 0.1 M, 2 h treatment.

Fig. 4. Effect of pH on TSS, COD and TOC removal efficiencies at 15.38 mA cm2 and [H2O2]/[Fe2þ] ¼ 5. ECP with Fe electrodes. V ¼ 400 mL, Na2SO4 ¼ 0.1 M, 2 h-treatment.

initial pH of the sludge sample) were conducted under current of 15.38 mA cm2 and [H2O2]/[Fe2þ] ¼ 5. The results depicted in Fig. 4 show that ECP displayed the best efficiency under slightly acidic conditions (pH 5) with 85.6%, 89.3% and 75.4% of TSS, COD and TOC

removal efficiencies, respectively. Even though Fenton-based processes generally operate optimally in acidic medium (pH between 2
s and Brillas, 2017), during ECP and 4) (Pignatello et al., 2006; Sire the performance is governed by both Fenton's oxidation and EC. It is known that pH values ranging from 5 to 8 favour the precipitation of Fe(OH)3 species and thus EC efficiency (Lakshmanan et al., 2009). In this way, under acidic conditions (pH 3), coagulation efficiency dropped, which had a dramatic impact on ECP performance. On the other hand, at higher pH (pH 8), the rate constant of the Fenton's reaction was significantly diminished as Fe2þ/Fe3þ species were scarcely available in the solution, hence hindering the production of  OH and decreasing ECP performance. Furthermore, the oxidation potential of OH in alkaline conditions is lower than in acidic solutions (1.90 V vs SHE at pH 7 vs 2.8 V vs SHE at pH 3) (Burbano et al., 2005). Thus, pH 5 was the optimal value that favoured both coagulation and Fenton's oxidation. These findings demonstrate that Fenton's oxidation alone could not deal with the high amount of TSS efficiently and EC alone was hindered by the high organic content of the sludge. Yet, a synergy between both processes made effective treatment possible. In order to understand the efficiency of the Fenton's reaction at pH 5 during ECP, it is important to keep in mind that although the Fenton's reaction performs optimally at acidic values, it can also take place at higher pH values, but with slower kinetics because of the lower reactivity of hydrolysed Fe3þ species. Furthermore, organic substances play a significant role in the complexation of Fe ions both in aqueous phase and on the surface of amorphous aggregates at different pH values. Fe-complexes in solution at nearneutral pH values promote the activation of H2O2 and O2 to form  OH and/or high-valence Fe-oxidizing species (ferrates, Fe (IV)) (Eq. (15)) that contribute to the oxidation of organics (Belanzoni et al., 2009; De Luca et al., 2014). Moreover, Fe species on the surface of coagulant particles can promote heterogeneous Fenton-like reactions in a wide pH range (Wang et al., 2016). Organic ligands also intervene in the reduction of Fe3þ to Fe2þ to maintain the catalytic cycle of the Fenton process (Bolobajev et al., 2015; Fukuchi et al., 2014).

282

Fe2þ-L þ H2O2 / Fe3þ-L þ OH þ OH

H. Olvera-Vargas et al. / Water Research 154 (2019) 277e286

(15)

These results contrast with previous works that reported that 3 was the optimal pH for ECP under the argument that OH oxidation was the prevailing degradation mechanism (Akyol et al., 2013; Guan et al., 2018; Kumar et al., 2018). It should be noted here that the above-mentioned works dealt with non-viscous wastewater with low content of organic matter and suspended solids, where the coagulation effect on organics was not taken into account. These discrepancies show that ECP performance strongly depends on the physicochemical properties of the waste material and the contribution of the dominant degradation mechanisms under optimal operating conditions. 3.2.3. Effect of [H2O2]/[Fe2þ] ratio H2O2 is crucial in Fenton-based processes because OH formation is directly dependent upon Fe2þ and H2O2 concentrations. In general, the efficiency increases as the [H2O2]/[Fe2þ] ratio rises, but the optimal ratio depends on the characteristics of the effluent and the experimental conditions (Akyol et al., 2013; Moussavi and Aghanejad, 2014; Pilli et al., 2015). Thus, the effect of H2O2 concentration ([H2O2]/[Fe2þ] ratio of 1, 5 and 10) was evaluated under current density of 15.38 mA cm2 and pH 5. The results show that the removal efficiency was slightly higher at [H2O2]/[Fe2þ] ¼ 5 (Fig. 5). Moreover, the properties of the Fe sludge obtained at the end of the 2 h treatment were also affected by the amount of H2O2. The dewaterability evaluation of the Fe sludge in terms of sludge filterability showed that Fe sludge obtained with [H2O2]/[Fe2þ] ¼ 1 had lower filterability (38.8 s mL1) than the sludge obtained with [H2O2]/[Fe2þ] ¼ 5 (18.7 s mL1). This phenomenon could be ascribed to the greater level of mineralization of organic matter attained with greater concentrations of OH formed at higher H2O2 to Fe2þ ratios. In this way, the degradation of organic material (EPS and other macromolecules) with [H2O2]/[Fe2þ] ¼ 1 was less efficient and bigger molecules with higher molecular weight were still present in solution (lower mineralization degree), which impacted the rheological properties of the sample (reflected in a higher resistance to filtration). On the contrary, lower-molecular weight compounds were formed with the 5 ratio, resulting in Fe sludge with higher dewaterability. As expected, the difference in organic composition between samples at different [H2O2]/[Fe2þ] ratios was not reflected in the COD and TOC values. The effect of polymeric substances on sludge morphology and compatibility has been reported in past investigations (Erdincler and Vesilind, 2000). Finally, at [H2O2]/[Fe2þ] ¼ 10, higher amounts of H2O2 resulted in decreased efficiency, owing to parasitic reactions that do not contribute to the oxidation of organics, mainly waste consumption

Fig. 5. Effect of [H2O2]/[Fe2þ] ratio on TSS, COD and TOC removal efficiencies at 15.38 mA cm2 and pH 5. ECP with Fe electrodes. V ¼ 400 mL, Na2SO4 ¼ 0.1 M, 2 htreatment.

of OH by excess H2O2, according to Eq. (16). Additionally, at a concentration higher than the optimum value, H2O2 may selfdecompose to H2O and O2, leading to reduced OH production, following Eq. (17). H2O2 is also decomposed via its oxidation at the anode, following Eqs. (11) and (12). H2O2 þ OH / HO2 þ H2O

(16)

2H2O2 / O2 þ H2O

(17)

3.2.4. Fe sludge cake properties After 2 h of ECP treatment under optimal conditions (pH 5, 15.38 mA cm2 of current and [H2O2]/[Fe2þ] ¼ 5), the settled Fe sludge cake was vacuum-filtered (after adjusting the pH to 9), dried at 105  C and further analysed. 21.90 g of dry dewatered sludge were obtained, whose chemical composition is shown in Table 2. It can be seen that the sludge cake contained high amounts or organic carbon, PePO4, NeNH4 and NO 3 , as an indication of good fertilizing properties. These ionic species precipitated with the coagulants. Furthermore, the analysis of heavy metals (Table S1) revealed that all the species that were measured were below the permissible levels for agricultural soil established by the European legislation (Table S1). As expected, Fe was present in the sludge at high concentration of 44.98 mg g1. However, the beneficial use of Fe biosolids in land and agricultural applications has been reported (Brown et al., 2012). Moreover, EC sludge with high contents of Al or Fe has also shown potential as pigments, construction materials and ceramic membranes (Tang and Shih, 2014). Besides the chemical properties of the sludge, it is worthy of note that sludge dewatering was significantly increased by ECP (filterability increased from 2815.63 s mL1 initially to 18.7 s mL1 after 2 h of ECP), resulting in a negligible amount of dry solid that needs to be disposed: only 21.90 g of dry sludge out of 400 mL of sample. 3.4. EF post-treatment ECP alone was able to achieve TSS, COD and TOC removal yields of 85.6%, 89.3% and 75.4%, respectively. However, the filtrate obtained after dewatering still contained a high concentration of COD and TOC (2800.2 and 1287.8 mg L1, respectively). Accordingly, the remaining effluent was further treated by means of EF using a BDD anode and a carbon brush cathode. The aim was to make use of the outstanding EF capacity to mineralize organic pollutants, as stated in previous studies (Lin et al., 2017; Mousset et al., 2017; OlveraVargas et al., 2014). The sample was acidified to pH 3 prior to EF experiments. Interestingly, the Fenton's reaction was catalysed by the remnant amount of Fe2þ/Fe3þ ions from the ECP process (7.35 mg L1), which was close to the optimal amount of Fe2þ catalyst generally reported for EF (between 3.58 and 17.93 mg L1) (Garcia-Rodriguez et al., 2018; Olvera-Vargas et al., 2018). The use of the remaining Fe ions in solution following EC as source of catalyst was reported in earlier works on coupled EC-EF systems (Anfruns-Estrada et al., 2017; Thiam et al., 2014). Fig. 6 depicts the optimization of current density for COD and TOC removal during EF. As expected, the concentration of organic matter decreased with time due to the oxidative attack of OH formed i) homogeneously through the Fenton's reaction (Eq. (4)) and ii) heterogeneously on the BDD surface according to Eq. (7). The performance increased with current density up to 25 mA cm2 (with respect to the anode surface), but further increase to 37.5 mA cm2 did not result in any further enhancement. This behaviour is simply explained by the non-oxidative waste reactions promoted at higher current values: the reduction of H2O2 and the evolution of H2 competing with H2O2 formation at the cathode, and the evolution of O2 at the anode

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283

Table 2 Chemical composition of ECP dry sludge after treatment.

mg g1

Fe

TVSS

NeNH4

NeNO-3

PePO4

TC

44.98 ± 2.50

544.00 ± 0.03

60.00 ± 2.74

8.00 ± 0.34

100.00 ± 5.70

244.67 ± 10.84

Fig. 6. Effect of current on a) COD and b) TOC removal efficiencies during EF treatment of the effluent following ECP. V ¼ 175 mL, pH ¼ 3, BDD anode and carbon brush cathode.

detrimental to BDD(OH) production (Olvera-Vargas et al., 2014; Sopaj et al., 2016). Following 4 h treatment under optimal current density (25 mA cm2), the remaining COD and TOC contents after ECP were reduced by 90% and 63%, respectively, representing an overall removal efficiency of 99.10% and 93.18%, for COD and TOC, respectively, as compared to the initial sample. It is worth mentioning that the solution became clear after only 2 h of EF. Remarkably, after 8 h of EF, the COD and TOC were reduced to 22.6 and 16.3 mg L1, respectively (more than 99% of the initial sample). As a reference, the COD was reduced to below 50 mg L1 after 6 h of EF, allowing discharge in controlled watercourse, as regulated by the NEA of Singapore. The overall evolution of COD and TOC during the sequential ECP-EF process (2 h ECP plus 4 h EF) is displayed in Fig. 7,

Fig. 7. Overall COD and TOC evolution during the sequential ECP-EF process. For ECP: Fe electrodes, V ¼ 400 mL, Na2SO4 ¼ 0.1 M, j ¼ 15.38 mA cm2, [H2O2]/[Fe2þ] ¼ 5, pH ¼ 5 and 2 h-treatment. For EF: V ¼ 175 mL, j ¼ 25 mA cm2, pH ¼ 3 and 4 h-treatment. Inset panel: visual evolution of the sludge sample during the course of treatment.

Table 3 þ 3Evolution of the NO 3 , NH4 eN, PO4 and pH during ECP-EF. Time

NO 3

NHþ 4 eN

PO34 -P

pH

Raw sample ECP 1 h ECP-2 h EF-4 h EF-8 h

80.4 ± 3.4 40.6 ± 2.3 75.4 ± 4.5 165.7 ± 5.8 262.1 ± 10.7

473.6 ± 26.7 2,250.2 ± 153.1 1,530.7 ± 108.9 1,850.5 ± 57.7 1,890.3 ± 81.7

680.4 ± 31.8 164.5 ± 10.7 78.6 ± 3.6 100.1 ± 2.9 128.3 ± 4.2

7.8e8.2 7.7 ± 0.5 8.2 ± 0.6 3.8 ± 0.2 4.0 ± 0.3

alongside photographs of the sample at various treatment stages. Table 3 summarizes the main results obtained with the integrated ECP-EF process. 90% TSS was removed by ECP, while the rest was removed by EF (overall > 90% of TSS removal). Regarding nitrogenous species, it can be seen that the amount of NO 3 decreased during the first hour of ECP, owing to its precipitation with Fe(OH)3 coagulants, followed by a progressive increase, until a final concentration of 262.1 mg L1 was attained after 8 h of EF treatment. In contrast, NHþ 4 eN increased sharply in the first hour of ECP, reaching 2250.2 mg L1, then decreased after 2 h of ECP due to the coagulation/flocculation process (Aoudj et al., 2017). During EF, 1 NHþ after 8 h. 4 eN increased progressively reaching 1890.3 mg L þ The increase of NO and NH eN was due to the continuous 3 4 disintegration of nitrogen-rich compounds such as proteins, nucleic acids and urea promoted by unselective OH attack. Regarding PO34 , a great fraction was precipitated during the ECP process, reaching a low concentration of 78.6 mg L1 after 2 h. PO34 concentration then increased during EF due to the continuous mineralization of the remaining organic substances such as phospholipids present in membrane cells. The release and accumulation of inorganic ions during EF following the mineralization of organics containing heteroatoms such as N, P, S and Cl is well documented (Dirany et al., 2012; Olvera-Vargas et al., 2014). The final effluent can be seen as an important source of P and N. Nutrient recovery could be achieved, for example, by chemical or electrochemical precipitation with magnesium to form struvite fertilizer (Kruk et al., 2014; T. Zhang et al., 2017b). Finally, the analysis of total coliforms revealed that the integrated ECP-EF process was equally efficient for disinfection. Coliforms were totally removed following the first hour of ECP and no more coliforms were detected during the consecutive treatment stages. The analysis of the solid cake obtained after ECP did not show any coliform colonies either. This is a remarkable finding if

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land application is considered for final disposal. These results demonstrate the synergistic effect of EC and Fenton's oxidation to disrupt cells and promote the mineralization of the organic material released and are in agreement with previous works that have stated the disinfection capability of EAOPs (Anfruns-Estrada et al., 2017; Cotillas et al., 2018). 3.5. Cost evaluation and comparison The cost for the overall ECP-EF treatment was estimated according to the method described in section 2.4. The total cost per litre of sludge, including conditioning (pH adjustment and Na2SO4 addition), ECP, sludge dewatering (precipitation of iron excess at pH 9 and vacuum filtration), dry sludge disposal, EF post-treatment (4 h) and final neutralization, was estimated to be S$0.036 L-1 (US$0.027 L-1). As observed in Fig. 8a, the chemical costs represent the largest percentage (64%), while power, sludge management and anode consumption costs account for the remaining fraction (29%, 6% and 2%, respectively). The cost breakdown for each stage depicted in Fig. 8b indicates that the intermediate operations, conditioning and sludge management, corresponded to 19% and 7%, respectively. Of course, it should be reminded here that these calculations are based on a bench-scale system. Larger prototypes may require higher energy costs than chemical costs, owing to higher current inputs needed for larger electrode surfaces in order to maintain the optimal current density (El-Naas et al., 2016). The comparison of different sludge treatment technologies is complicated. Firstly, because treatment depends on the characteristics of the sludge and secondly, because of the properties and fate of the final sludge and/or effluent. Conditioning methods based on AOPs are generally applied as pretreatment techniques aiming at improving the efficiency of anaerobic digestion by reducing TSS and increasing the biodegradability (Wang et al., 2017). Remarkably, the ECP-EF process proposed in this work is a comprehensive treatment strategy yielding a small amount of dry sludge and clean liquid water with minimal content of organic matter. The dry sludge rich in organic and inorganic nutrients and free of hazardous pathogens is suitable for land application, for sustainable nutrient management (Gude and Gnaneswar, 2015). It is already known that EAOPs for wastewater treatment can outcompete other AOPs in ~ izares et al., 2009) and here we show terms of cost efficiency (Can that the same premises are applicable to sludge treatment. For example, the full-scale ozonation pre-treatment of waste activated sludge (WAS) required around V1.30 per ton of TSS and only reduced the production of dry sludge by 10% in a WWTP (Romero et al., 2015). In contrast, we estimate the cost of our ECP-EF system at V2.10 per ton of TSS with the production of dry sludge reduced by 94.5%. Based on the data reported by (Li et al., 2016) on

the treatment of WAS by EC/electrochemically activated persulfate at lab-scale, the cost was around US$4.76 L-1 sludge (vs US$0.026 L1 sludge for ECP-EF in this study), which they claimed to be 42.14% cheaper than conventional Fenton. Even if these approximations should be taken with care, they provide a general idea of the potential and economic advantages of the proposed ECP-EF system for comprehensive treatment of digested sludge. Finally, the ECP-EF process disclosed in this study shows excellent potential as a small-scale decentralised treatment, with which water could be reused for non-potable purposes. On-site water recycling electrochemical systems of this kind have been recently implemented to treat toilet wastewater followed by reuse for flushing (Cid et al., 2018). 4. Conclusions In this study, a sequential electrochemical treatment was applied to treat anaerobic sludge from a poultry farm. The recommended two-step integrated ECP-EF treatment consisted in: i) 2 hECP at pH 5, current density of 15.38 mA cm2 and [H2O2]/ [Fe2þ] ¼ 5, followed by ii) 8 h-EF treatment of the effluent recovered after dewatering (at pH 3 and 25 mA cm2). During the ECP stage, the sludge was efficiently destabilized and conditioned by the synergistic effects of electrocoagulation and Fenton oxidation, which enhanced the dewaterability and reduced the amount of sludge (only 21.9 g of dry Fe sludge rich in C, N and P produced). The residual effluent after dewatering was subjected to EF, during which the excess of organic matter was efficiently mineralized by means of OH. The effluent after 2 h-ECP and 8 h-EF contained only 24.6 and 16.3 mg L1 of COD and TOC, respectively. The estimated total operational treatment cost was S$0.036 L-1 sludge. In conclusion, this promising ECP-EF process is presented as a sustainable alternative for comprehensive sludge remediation, capable of producing a reduced volume of solid sludge suitable for land application, while the recovered water is amenable to discharge or reuse. Acknowledgements The authors would like to acknowledge the Singapore Ministry of Education Academic Research Fund Tier 1 (WBS R302000145112). Appendix A. Supplementary data Supplementary data to this article can be found online at https://doi.org/10.1016/j.watres.2019.01.063. References

Fig. 8. Cost breakdown for the total operational costs per category (a) and total costs per treatment stage of the ECP-EF process (b).

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