Electrochemical treatment of concentrate from reverse osmosis of sanitary landfill leachate

Electrochemical treatment of concentrate from reverse osmosis of sanitary landfill leachate

Journal of Environmental Management 181 (2016) 515e521 Contents lists available at ScienceDirect Journal of Environmental Management journal homepag...

733KB Sizes 13 Downloads 83 Views

Journal of Environmental Management 181 (2016) 515e521

Contents lists available at ScienceDirect

Journal of Environmental Management journal homepage: www.elsevier.com/locate/jenvman

Research article

Electrochemical treatment of concentrate from reverse osmosis of sanitary landfill leachate  Pacheco a, Lazhar Labiadh a, b, Annabel Fernandes a, *, Lurdes Ciríaco a, Maria Jose Abdellatif Gadri b, Salah Ammar b, c, d, Ana Lopes a ~, Portugal FibEnTech/UBI and Department of Chemistry, Universidade da Beira Interior, 6201-001 Covilha D ep. de Chimie, Facult e des Sciences de Gab es, Universit e de Gab es, Cit e Erriadh, 6072 Gab es, Tunisia c  Dep. de Chimie, Facult e des Sciences de Bizerte, Universit e de Carthage, 7021 Jarzouna, Tunisia d Laboratoire Photovoltaïque, Centre de Recherches et des Technologies de l’Energie Technopole Borj Cedria, Bp 95, Hammamm Lif, 2050, Tunisia a

b

a r t i c l e i n f o

a b s t r a c t

Article history: Received 5 February 2016 Received in revised form 29 June 2016 Accepted 30 June 2016 Available online 5 August 2016

Conventional sanitary landfill leachate treatment has recently been complemented and, in some cases, completely replaced by reverse osmosis technology. Despite the good quality of treated water, the efficiency of the process is low and a large volume of reverse osmosis concentrate has to be either discharged or further treated. In this study, the use of anodic oxidation combined with electro-Fenton processes to treat the concentrate obtained in the reverse osmosis of sanitary landfill leachate was evaluated. The anodic oxidation pretreatment was performed in a pilot plant using an electrochemical cell with boron-doped diamond electrodes. In the electro-Fenton experiments, a boron-doped diamond anode and carbon-felt cathode were used, and the influence of the initial pH and iron concentration were studied. For the experimental conditions, the electro-Fenton assays performed at an initial pH of 3 had higher organic load removal levels, whereas the best nitrogen removal was attained when the electrochemical process was performed at the natural pH of 8.8. The increase in the iron concentration had an adverse impact on treatment under natural pH conditions, but it enhanced the nitrogen removal in the electro-Fenton assays performed at an initial pH of 3. The combined anodic oxidation and electro-Fenton process is useful for treating the reverse osmosis concentrate because it is effective at removing the organic load and nitrogen-containing species. Additionally, this process potentiates the increase in the biodegradability index of the treated effluent. © 2016 Elsevier Ltd. All rights reserved.

Keywords: Reverse osmosis concentrate Sanitary landfill leachate Electro-Fenton Anodic oxidation

1. Introduction Sanitary landfilling is the most common method used for municipal solid waste disposal in many countries. This landfilling process has one disadvantage of generating a concentrated effluent, the leachate, which may exhibit acute to chronic toxicity (Lü et al.,

Abbreviations and symbols: AO, anodic oxidation; AOP, advanced oxidation process; BDD, boron-doped diamond; BOD5, biochemical oxygen demand; COD, chemical oxygen demand; DIC, dissolved inorganic carbon; DOC, dissolved organic carbon; EAOP, electrochemical advanced oxidation process; EF, electro-Fenton; Esp, specific energy consumption; HPLC, high performance liquid chromatography; I, applied current intensity; RO, reverse osmosis; ROC, reverse osmosis concentrate; SD, standard deviation; TDC, total dissolved carbon; TN, total nitrogen; U, cell voltage; V, volume of the solution; t, time. * Corresponding author. E-mail address: [email protected] (A. Fernandes). http://dx.doi.org/10.1016/j.jenvman.2016.06.069 0301-4797/© 2016 Elsevier Ltd. All rights reserved.

2008; Renou et al., 2008; Salem et al., 2008). In general, biological and physical-chemical treatments are not good end-of-pipe solutions for treating sanitary landfill leachates; as a result, technologies such as membrane processes have to be utilized to achieve full leachate sanitation. Reverse osmosis (RO) is one approach being implemented as a complement for other treatments or, in some cases, is the only treatment. However, RO generates a large volume of concentrate that has to be discharged or further treated (Renou et al., 2008). Generally, these concentrates are recirculated into sanitary landfills, resulting in the accumulation of pollutants and leading to an increase in the electrical conductivity, chemical oxygen demand (COD) and nitrogen content of the leachates. This process also results in decreased effectiveness of the RO process due to the increased fouling of the RO membranes (Talalaj and Biedka, 2015). Thus, alternative solutions for treating the RO concentrate of the sanitary landfill leachate are required. Electrochemical advanced oxidation processes (EAOPs) have

516

L. Labiadh et al. / Journal of Environmental Management 181 (2016) 515e521

been receiving increasing attention in the approach to treating synthetic and real wastewaters (Anglada et al., 2009; Chen, 2004; s Fernandes et al., 2015; Martínez-Huitle and Brillas, 2009; Sire and Brillas, 2012). The simplest and the most popular EAOP is anodic oxidation (AO), where organics can be oxidized by the following processes: (i) direct electron transfer at the anode surface; (ii) highly reactive hydroxyl radicals ($OH) weakly adsorbed at the anode’s surface that are generated from water oxidation; and (iii) mediators electrogenerated in the bulk solution, such as active chlorine species, ozone, persulfates and hydrogen peroxide (Panizza and Cerisola, 2009a, 2003). According to the literature, AO provides a simple, viable and promising method for remediating sanitary landfill leachates (Fernandes et al., 2015). Under appropriate experimental conditions, AO can remove most of the organic load and almost all ammonia content. Additionally, it significantly reduces the color without resulting in the accumulation of refractory organics, allowing for high treatment efficiency without the disadvantage of sludge production (Fernandes et al., 2015). Removal of 100% of the COD and ammonium nitrogen content has been achieved for the AO in sanitary landfill leachates using borondoped diamond (BDD) anodes (Anglada et al., 2010; Cabeza et al., 2007a, 2007b). Another promising EAOP used in wastewater treatment is the electro-Fenton (EF) process (Atmaca, 2009; Barhoumi et al., 2015; Panizza and Cerisola, 2001). The EF process is a method that promotes the oxidation of organic compounds via an indirect electrochemical oxidation through hydroxyl radicals produced in the Fenton reaction (Eq. (1)).

their greater efficiency for H2O2 generation through Eq. (2) and vez smaller rate of Eq. (3) (El-Ghenymy et al., 2014; Isarain-Cha et al., 2010; Ruiz et al., 2011). In the EF process, the use of an undivided cell with a high O2overpotential anode, such as BDD, accelerates the degradation of organic matter because it can also be oxidized by adsorbed hydroxyl radical (BDD($OH)) and formed as an intermediate of the anodic water discharge (Panizza and Cerisola, 2009a). The BDD anode is better than other typical electrodes, such as Pt and PbO2 (Ciríaco et al., 2009; Flox et al., 2009; Mhemdi et al., 2013), improving the destruction of aromatic compounds and generation of carboxylic acids (Brillas et al., 2009; Mhemdi et al., 2013; Panizza and Cerisola, 2009a). Moreover, the electrogenerated H2O2 is partially oxidized to O2 at the anode, producing the weaker oxidant hydroperoxyl radical (HO2) as an intermediate (Brillas et al., 2009). The aim of this work was to treat a concentrate obtained during RO of a leachate from a municipal sanitary landfill with the combination of anodic oxidation with a BDD anode and EF with a carbon-felt cathode and a BDD anode, as well as to study the influence of the different experimental conditions on the energetic costs of the organic load and ammonium nitrogen removal. The sequence of AO followed by EF was chosen because i) AO is known for its higher ability than EF to treat very complex effluents, even those with high suspended solid contents, such as the raw samples used in this study, and ii) EF is expected to perform organic matter oxidation at lower energetic costs when the sample mainly consists of dissolved matter. The influence of the pH and iron content on the performance of the EF process efficiency was studied.

Fe2þ þ H2O2 / Fe3þ þ HO þ OH

2. Experimental

(1)

In this process, both H2O2 and Fe2þ, can be electrogenerated in situ (Deng and Englehardt, 2006), reducing the cost related to H2O2 consumption, which is the main cost factor for consumables in the setting of traditional AOPs based on Fenton’s reaction (Malato et al., 2009). Hydrogen peroxide can be generated in situ from the twoelectron reduction of O2 on cathodes, such as gas diffusion electrodes, reticulated vitreous carbon or graphite-felt, according to Eq. (2) (Brillas et al., 2009; Oturan et al., 2008; Panizza and Cerisola, 2001): O2 þ 2Hþ þ 2e / H2O2

(2)

An advantage of the EF process over the classical chemical Fenton’s reagent is that Eq. (1) is electrocatalyzed because Fe2þ ions are rapidly regenerated from the reduction of Fe3þ ions at the cathode through Eq. (3) (Brillas et al., 2009; Labiadh et al., 2015). Fe3þ þ e / Fe2þ

(3)

The EF process has been used to treat many organic pollutants, enabling their complete mineralization (Barhoumi et al., 2015; Flox et al., 2009; Mhemdi et al., 2013; Panizza and Cerisola, 2009b), as well as sanitary landfill leachates, and there have been promising results (Atmaca, 2009; Lin and Chang, 2000; Zhang et al., 2006). According to the literature, the optimum pH of the solution for the EF process is 3.0, which is close to the optimum value of 2.8 for the chemical Fenton’s process, and its ability to oxidize the organic load in wastewaters is closely related to the rate of Eq. (1), which depends on the iron concentration and cathode type (Brillas et al., 2009). For the EF process performed with a carbon-felt cathode, optimal Fe2þ or Fe3þ concentrations of 0.1e0.2 mM have been found (Labiadh et al., 2015; Ozcan et al., 2008; Panizza and Cerisola, s et al., 2007), whereas 0.5e1.0 mM Fe2þ and 4.0 mM 2009b; Sire 3þ Fe are the most adequate for O2 or air-diffusion electrodes due to

2.1. Sample characterization The reverse osmosis concentrate (ROC) of the sanitary landfill leachate used in this study was collected in June 2015 from an intermunicipal sanitary landfill facility. This site, which serves a population of over 368,000 inhabitants in 19 municipalities, has an onsite facility capable of treating up to 175 m3 of leachate per day. The treatment applied at this landfill site consists of two reverse osmosis systems followed by a stripping column. With the reverse osmosis process, more than 96 m3 of the leachate concentrate is returned to the sanitary landfill per day. The characteristics of the ROC of the sanitary landfill leachate used in this study are presented in Table 1. 2.2. Electrochemical experiments The AO pretreatment was conducted in a semi-pilot plant operating in batch mode with recirculation at room temperature and natural pH without the addition of background electrolyte. A BDD DiaCell 100 electrochemical cell with an electrode area of 70 cm2 and a DiaCell-PS1500 power supply with automatic polarity reversal were used. During the assays, automatic polarity reversal occurred every minute. Assays were performed for 24 h at an applied current intensity of 4.9 A and a flow rate of 500 L/h using 10 L of ROC of the sanitary landfill leachate. After the AO treatment, the solution was submitted to the electro-Fenton process. The EF experiments were conducted in batch mode with stirring using an open, undivided and cylindrical glass cell and 250 mL of solution over a period of 8 h. A graphite-felt piece (Carbone Loraine) with a thickness of 0.5 cm and an immersed area of 60 cm2 was used as the cathode, and a BDD electrode purchased from Adamant Technologies with an immersed area of 20 cm2 was used as an anode. The anode was centered in the electrochemical cell and surrounded by the cathode, which covered the inner wall of the

L. Labiadh et al. / Journal of Environmental Management 181 (2016) 515e521

cell. H2O2 was produced from reducing the O2 dissolved in the solution (Eq. (2)). Continuous O2 saturation at atmospheric pressure was ensured by bubbling compressed air through a fritted glass diffuser at 1 L/min, starting 10 min before electrolysis, to reach a steady O2 concentration. To optimize the experimental conditions, different iron concentrations between 10 (natural iron concentration) and 27 mg/L and different initial pH conditions, namely, a pH of 3 and 5, were tested for an applied current intensity of 0.3 A. A GW, Lab DC, model GPS-3030D (0e30 V, 0e3 A) was used as the power supply. Electrochemical experiments using the EF experimental set-up were also performed with the AO-pretreated solution at natural pH conditions (8.8). In the assays performed with additional iron, Fe2(SO4)3$5H2O was added to the solution 10 min before starting electrolysis to ensure complete dissolution. When required, pH adjustments were performed by the addition of concentrated H2SO4. All electrochemical assays were performed at least in duplicate. The parameter values for the assays are given as mean values. 2.3. Analytical methods Degradation tests were followed by analysis of the COD, biochemical oxygen demand (BOD5), total dissolved carbon (TDC), dissolved organic carbon (DOC), dissolved inorganic carbon (DIC) and total nitrogen (TN), which were performed according to standard procedures (Eaton et al., 2005). COD determinations were made using the closed reflux titrimetric method. The BOD5 was evaluated by determining the oxygen consumption after 5 days of incubation. The TDC, DOC, DIC and TN were measured using a Shimadzu TOC-VCPH analyzer combined with a TNM-1 unit. Before TDC, DOC, DIC and TN determinations, samples were filtered through 1.2-mm glass microfiber filters. The ammonium, nitrate, nitrite, chloride and chlorate ions concentrations were determined by ion chromatography using a Shimadzu 20A Prominence HPLC system that was equipped with a Shimadzu CDD 10Avp conductivity detector. To determine the NHþ 4, an IC YK-A Shodex (4.6 mm ID  100 mm) column was utilized. The mobile phase consisted of 5.0 mM tartaric acid, 1.0 mM dipicolinic acid and 24 mM boric acid aqueous solution at a flow rate of 1.0 mL/  min. The column temperature was 40  C. To evaluate the NO 3 , NO2 , Cl and ClO levels, an IC I-524A Shodex (4.6 mm ID  100 mm) 3 anion column was employed. The mobile phase consisted of an aqueous solution of 2.5 mM of phthalic acid and 2.3 mM of tris(hydroxymethyl) aminomethane at a flow rate of 1.5 mL/min. The column temperature was 40  C. All solutions for chromatographic analysis were prepared with ultrapure water obtained with Milli-Q® equipment. All eluents were HPLC grade and supplied by Sigma-Aldrich. The iron concentration was determined by flame atomic absorption spectrometry using a Perkin Elmer Apparatus, AAnalyst 800. The sample preparation followed a standard procedure that includes HCleHNO3 acid digestion (Eaton et al., 2005). The pH was measured using a HANNA pH meter (HI 931400). The conductivity was determined using a Mettler Toledo conductivity meter (SevenEasy S30K). 3. Results and discussion The ROC of the sanitary landfill leachate was first submitted to an AO pretreatment, and Table 1 presents the obtained results. A clarified solution was generated, which had approximately 65% of the initial organic load and 83% of the ammonium nitrogen, potentiating the application of a later treatment. With respect to the iron concentration, and because AO pretreatment was performed with automatic polarity reversal, no significant difference

517

Table 1 Physicochemical characteristics of the reverse osmosis concentrate of the sanitary landfill leachate and solution obtained after anodic oxidation pretreatment (AO pretreated). Parameter (g/L)a

Mean value (±SDb) ROC

AO pretreated

COD BOD5 BOD5/COD DOC DIC TN [NHþ 4] [NO 3]  [Cl ] [ClO 3] [Fe] pH Conductivity

9.9 ± 0.4 4.3 ± 0.2 0.43 ± 0.04 3.6 ± 0.3 1.31 ± 0.03 2.4 ± 0.3 3.0 ± 0.2 <0.01 5.3 ± 0.2 <0.01 0.0108 ± 0.0006 8.1 ± 0.1 31 ± 4

6.5 ± 0.4 3.3 ± 0.1 0.51 ± 0.05 2.61 ± 0.05 0.92 ± 0.02 2.16 ± 0.02 2.5 ± 0.1 0.45 ± 0.04 5.1 ± 0.3 0.14 ± 0.01 0.0103 ± 0.0005 8.8 ± 0.1 32 ± 1

a Except for the BOD5/COD and pH values, which are dimensionless, and for conductivity values, which are presented in mS/cm. b SD e Standard deviation.

was found between the iron content in the ROC and pretreated solution. Because ferrous ion concentrations between 6 and 12 mg/L are reported in the literature as optimal for the EF process when a carbon-felt cathode is used (Labiadh et al., 2015; Ozcan et al., 2008; s et al., 2007), the first set of EF Panizza and Cerisola, 2009b; Sire experiments were performed without the addition of external iron. Fig. 1 shows the results obtained for the EF assays performed using the AO pretreated solution with a different initial pH at an applied current intensity of 0.3 A. The following three different initial pH conditions were studied: pH of 3, the optimum pH for the EF process according to the literature (Panizza and Cerisola, 2003); pH of 5, a less costly solution for reagent addition; and natural pH (8.8), the cheapest solution, even though it is not favorable for Fenton’s reaction. As in Fig. 1, acidic pH conditions led to higher COD removals, but the total and ammonium nitrogen removal levels were much higher when the assays were run at a natural initial pH. For instance, approximately 35% of the ammonium was removed after 8 h for natural pH, whereas at pH of 3, only a small change in the ammonium concentration was observed. Additionally, natural pH conditions favor the mineralization of organic compounds. Indeed, the assays performed at different initial pH values yield similar DOC removals, but much lower COD removal was observed with a natural initial pH. This means that oxidation in the assays run at a natural pH leads to higher levels of organic compound combustion instead of converting the compounds into other more oxidized compounds, but that would remain in solution. These results can be explained by considering that at a pH of 8.8, Fenton’s reaction (Eq. (1)) was not favored because, for the experimental conditions of a saturated oxygen concentration, hydrogen peroxide electrochemical production (Eq. (2)) is enhanced at pH 3 (Pimentel et al., 2008), while the regeneration of ferrous ions (Eq. (3)) is inhibited at pH > 4 due to the precipitation of ferric oxyhydroxides (Bigda, 1995; Lin and Lo, 1997; Nesheiwat and Swanson, 2000). Therefore, at a natural pH, the oxidation of organic compounds mainly occurred through the hydroxyl radicals generated from water electrolysis at the BDD anode surface (Eq. (4)). H2O / Hþ þ HO þ e

(4)

In parallel with water electrolysis, and because the chloride content in the pretreated ROC was as high as 5.1 g/L, the direct

518

L. Labiadh et al. / Journal of Environmental Management 181 (2016) 515e521

oxidation of a chloride ion at the anode produced chlorine (Eq. (5)), which can in turn be hydrolyzed, resulting in hypochlorous acid rez et al., 2012). According to the literature, ammonium (Eq. (6)) (Pe degradation mainly occurs as a result of indirect oxidation with rez et al., 2012). Therefore, the produced active chlorine species (Pe hypochlorous acid can react with ammonium (Eq. (7) and (8)), leading to its oxidation and to the formation of gaseous nitrogen rez et al., 2012). forms and nitrate (Pe

was removed for 16 mg/L iron. This decrease in the organic load and nitrogen removals can be explained by the side reactions between the HO and HOCl with the ferrous ion (Eq. (10) and (11)), which are enhanced by the addition of Fe2þ that will consume hydroxyl radicals and hypochlorous acid. As a result, they are not available for the oxidation of organic compounds and ammonium nitrogen. HO þ Fe2þ / Fe3þ þ HO

(10)

2Cl / Cl2 þ 2e

(5)

2Fe2þ þ 3HOCl þ 3H2O / 2Fe(OH)3 þ 3Cl þ 3Hþ

(11)

Cl2 þ H2O / HOCl þ Hþ þ Cl

(6)

 þ 2=3NHþ 4 þ HOCl/1=3N2 þH2 O þ 5=3H þCl

(7)

  þ NHþ 4 þ 4HOCl/NO3 þ H2 O þ 6H þ4Cl

(8)

For the assays performed at an initial pH of 3, the addition of 0.1 mM of Fe2þ did not affect the COD and DOC removal, but it enhanced the ammonium and total nitrogen removal. This result could indicate that Fenton’s reaction was enhanced by the addition of extra iron, resulting in a competition between COD and ammonium removal. Based on this result, the amount of iron was further increased in the solution at an initial pH of 3 to a total iron concentration of 27 mg/L (corresponding to the addition of 0.3 mM of Fe2þ). However, this further increase in the iron concentration resulted in a decrease in the ammonium and total nitrogen removal levels (Fig. 2c and d), which can be explained by the scavenging reaction between the HO and Fe2þ (Eq. (10)) that is further enhanced by the excess of ferrous iron in solution. When the nitrate concentration results are analyzed (inset of Fig. 2d), although the addition of iron did not significantly influence the nitrate formation in the experiments performed at natural pH, the addition of iron decreased the amount of nitrate formed during the assays performed at an initial pH of 3. Conversely, for a pH of 3, an exponential increase in the amount of chlorate formed was observed when the iron concentration was increased to 27 mg/L (Appendix 1b), indicating that chloride is oxidized to chlorate to a greater extent. With respect to the pH evolution in the assays (inset of Fig. 2c), the addition of iron did not have any effect. To analyze the energy consumptions associated with the different experimental conditions, specific energy consumptions, in Wh/g, based on COD or NHþ 4 removal were calculated using Eq. (12), where U is the cell voltage (in V) resulting from the applied current intensity I (in A) and Dt is the duration of the electrolysis (in h). Additionally, V is the volume of the solution (in L), and DX is the removed COD or NHþ 4 (in g/L) during Dt. The results are presented in Table 2.

At pH values of 3 and 5, the electrochemical hydrogen peroxide formation rate (Eq. (2)) was enhanced, favoring Fenton’s reaction (Eq. (1)) as well as the reaction between hydroxyl radicals and hydrogen peroxide (Eq. (9)) (Atmaca, 2009). H2O2 þ HO / HO2 þ H2O

(9)

The increase in the level of oxidizing species present in solution explains the higher COD removal under acidic pH conditions. The low mineralization of the organic compounds under acidic pH conditions can be explained by the fact that HO2 radicals produced as in Eq. (9) present with weaker oxidizing ability compared to the HO radical. Furthermore, the direct oxidation of chloride ions was less favored under acidic pH conditions, and consequently, nitrogen removal decreased. Although no significant differences were found in COD removal for the assays performed at an initial pH of 3 or 5, the total and ammonium nitrogen removal levels increased with increasing initial pH levels. This observation indicates that chlorine generation is depressed with increasing hydrogen peroxide electrochemical production, which is maximal at pH 3 (Pimentel et al., 2008). Conversely, the amount of nitrate formed increased with decreasing initial pH. These results indicate that HO and/or HO2 radicals formed under acidic pH conditions react preferably with the organic nitrogen present in solution rather than with ammonium, oxidizing it to nitrate, or that the formation rate of ammonium from organic nitrogen equals the oxidizing rate of ammonium to nitrate. Regarding pH changes in the assays (inset of Fig. 1b), experiments performed at natural pH did not show significant variations, which can be explained by the buffering effect of these effluents (Chiang et al., 1995). For the assays performed under initial acidic conditions, a decrease in the pH was found, and this decrease has previously been explained by the carboxyl acids formed during the degradation process and oxygen evolution reaction (Zhang et al., 2012). Considering the results obtained in the experiments performed under different initial pH conditions, a second set of assays was performed in which the influence of the iron concentration was studied. Two different initial pH conditions were chosen for this study, natural pH, which results in the best nitrogen removal, and pH 3, which is reportedly optimal for the EF process. As in Fig. 2, when the iron concentration was increased from 10 to 16 mg/L (corresponding to an addition of 0.1 mM of Fe2þ), COD, DOC, ammonium and total nitrogen removal levels decreased for the assays performed at natural pH. Therefore, at the end of the treatments performed at natural pH with a natural iron content (10 mg/L of iron), 24% of the initial COD was destroyed, but only 16%

Esp ¼

U I Dt V DX

(12)

In general, for the experimental conditions studied, the lowest specific energy consumptions regarding COD removal were attained for the assays performed at pH 3, although the same conditions had the highest specific energy consumptions for NHþ 4 removal. By contrast, assays performed at natural pH without added iron presented with slightly higher specific energy consumptions for COD removal, but they had lower specific energy consumptions for NHþ 4 removal. Considering these results and the costs associated with the acidifying process, a natural pH and natural iron concentration were chosen as the most favorable conditions for treating the AO pretreated sample. The biodegradability index, determined as the ratio BOD5/COD, for these more favorable conditions was also assessed, and a value of 0.51 was obtained after the 8 h assay, indicating that the biodegradability index did not change during this treatment. A 24 h assay was performed at a natural iron concentration and natural initial pH at 0.3 A to evaluate the treatment potential. The results presented in Appendix 2 show that after 24 h of reaction, the COD, DOC, NHþ 4 and TN removal rates were approximately 65%,

L. Labiadh et al. / Journal of Environmental Management 181 (2016) 515e521

519

Fig. 1. Influence of the initial pH on the decays of (a) COD, (b) DOC, (inset of b) pH, (c) ammonium concentration and (d) TN and (inset of d) nitrate concentration during the EF process. Sample e ROC pretreated by AO; Iron concentration ¼ 10 mg/L; I ¼ 300 mA; cathode e carbon-felt; and anode e BDD.

Fig. 2. Influence of the initial iron concentration on the decays of (a) COD, (b) DOC, (c) ammonium concentration, (inset of c) pH, and (d) TN and (inset of d) nitrate concentration during the EF process. Sample e ROC pretreated by AO; I ¼ 300 mA; cathode e carbon-felt; and anode e BDD.

520

L. Labiadh et al. / Journal of Environmental Management 181 (2016) 515e521

Table 2 Specific energy consumptions for the assays performed at different initial iron concentrations and different initial pH: Sample e ROC pretreated by AO; I ¼ 300 mA; assay duration e 8 h; cathode e carbon-felt; and anode e BDD. pH

[Fe] (mg/L)

Esp (Wh/gCOD)

Esp (Wh/gNHþ 4)

3

10 16 27 10 10 16

45 42 48 47 53 73

680 205 545 145 108 114

5 Natural (8.8)

55%, 71% and 74%, respectively. Additionally, the nitrite and nitrate concentrations increase until the 20 h assay. Afterwards, the nitrate concentration stabilized and the nitrite concentration decreased. Additionally, the chloride concentration decreased throughout the assay, and chlorate formation increased exponentially. The results of the corresponding 8 h assay are also presented in Appendix 2, and the small differences observed in the results at 8 h assay are mainly from the more exhaustive sample collection during the shorter assay, which reduces the volume of the treated sample. 4. Conclusions Anodic oxidation and electro-Fenton are effective technologies for treating reverse osmosis concentrate from sanitary landfill leachates. In the combined AO and EF treatment approach, utilizing the most favorable EF experimental conditions (initial pH of 3 with 16 mg/L of iron at an applied current intensity of 0.3 A; 8 h), the removal rates of COD, DOC, NHþ 4 and total nitrogen were 60%, 53%, 33% and 22%, respectively. When AO pretreatment was followed by electrochemical assays performed at a natural initial pH of 8.8, a natural iron concentration (10 mg/L), and an applied current intensity of 0.3 A for 8 h, the Fenton reaction was negligible and anodic oxidation dominated. In this case, the COD, DOC, NHþ 4 and total nitrogen removal rates were of 48%, 42%, 50% and 35%, respectively, indicating that the anodic oxidation process is more effective at removing N-containing compounds than the EF process. Therefore, the experimental conditions must be set according to the required objectives. If the priority is organic load removal instead of Ncontaining compound removal or the sample presents with low nitrogen content, combined anodic oxidation followed by an electro-Fenton process should be used, being necessary optimizing the electro-Fenton process, i.e., pH adjustment of the effluent to 3, and, depending on the iron concentration of the sample, correct the iron content. When the primary goal is nitrogen removal, the anodic oxidation process should be utilized because it results in high nitrogen removal without requiring pH adjustment or the addition of iron. An increase in the biodegradability index of the treated effluent from 0.43 to 0.51 was also attained for the different experimental conditions tested. This demonstrated that the combined electrochemical techniques could be used to eliminate the recalcitrant organic load from ROC and increase the biodegradability, enabling its discharge into urban wastewater treatment plants with biological steps. Acknowledgments ~o para a Cie ^ncia e a Tecnologia, FCT, for We thank Fundaça funding the FibEnTech Research Unit and for the grant awarded to A. Fernandes, SFRH/BPD/103615/2014. We thank University of Gabes, in Tunisia, for providing partial financial support.

Appendix A. Supplementary data Supplementary data related to this article can be found at http:// dx.doi.org/10.1016/j.jenvman.2016.06.069. References Anglada, A., Urtiaga, A., Ortiz, I., 2009. Contributions of electrochemical oxidation to waste-water treatment: fundamentals and review of applications. J. Chem. Technol. Biotechnol. 84, 1747e1755. Anglada, A., Urtiaga, A., Ortiz, I., 2010. Laboratory and pilot plant scale study on the electrochemical oxidation of landfill leachate. J. Hazard. Mater 181, 729e735. Atmaca, E., 2009. Treatment of landfill leachate by using electro-Fenton method. J. Hazard. Mater 163, 109e114. Barhoumi, N., Labiadh, L., Oturan, M.A., Oturan, N., Gadri, A., Ammar, S., Brillas, E., 2015. Electrochemical mineralization of the antibiotic levofloxacin by electroFenton-pyrite process. Chemosphere 141, 250e257. Bigda, R.J., 1995. Consider Fenton’s chemistry for wastewater treatment. Chem. Eng. Prog. 91, 62e66. s, I., Oturan, M.A., 2009. Electro-Fenton process and related electroBrillas, E., Sire chemical technologies based on Fenton’s reaction chemistry. Chem. Rev. 109, 6570e6631. Cabeza, A., Urtiaga, A., Ortiz, I., 2007a. Electrochemical treatment of landfill leachates using a boron-doped diamond anode. Ind. Eng. Chem. Res. 46, 1439e1446. Cabeza, A., Urtiaga, A., Rivero, M., Ortiz, I., 2007b. Ammonium removal from landfill leachate by anodic oxidation. J. Hazard. Mater 144, 715e719. Chen, G., 2004. Electrochemical technologies in wastewater treatment. Sep. Purif. Technol. 38, 11e41. Chiang, L., Chang, J., Wen, T., 1995. Indirect oxidation effect in electrochemical oxidation treatment of landfill leachate. Water Res. 29, 671e678. Ciríaco, L., Anjo, C., Correia, J., Pacheco, M.J., Lopes, A., 2009. Electrochemical degradation of Ibuprofen on Ti/Pt/PbO2 and Si/BDD electrodes. Electrochim. Acta 54, 1464e1472. Deng, Y., Englehardt, J.D., 2006. Treatment of landfill leachate by the Fenton process. Water Res. 40, 3683e3694. Eaton, A., Clesceri, L., Rice, E., Greenberg, A., Franson, M.A., 2005. Standard Methods for Examination of Water and Wastewater, twenty-first ed. American Public Health Association, Washington, DC. El-Ghenymy, A., Rodriguez, R.M., Brillas, E., Oturan, N., Oturan, M.A., 2014. ElectroFenton degradation of the antibiotic sulfanilamide with Pt/carbon-felt and BDD/ carbon-felt cells. Kinetics, reaction intermediates, and toxicity assessment. Environ. Sci. Pollut. Res. Int. 21, 8368e8378. Fernandes, A., Pacheco, M.J., Ciríaco, L., Lopes, A., 2015. Review on the electrochemical processes for the treatment of sanitary landfill leachates: present and future. Appl. Catal. B Environ. 176e177, 183e200. Flox, C., Arias, C., Brillas, E., Savall, A., Groenen-Serrano, K., 2009. Electrochemical incineration of cresols: a comparative study between PbO2 and boron-doped diamond anodes. Chemosphere 74, 1340e1347. Isarain-Ch avez, E., Arias, C., Cabot, P.L., Centellas, F., Rodríguez, R.M., Garrido, J.A., Brillas, E., 2010. Mineralization of the drug beta-blocker atenolol by electroFenton and photoelectro-Fenton using an air-diffusion cathode for H2O2 electrogeneration combined with a carbon-felt cathode for Fe2þ regeneration. Appl. Catal. B Environ. 96, 361e369. Labiadh, L., Oturan, M.A., Panizza, M., Hamadi, N.B., Ammar, S., 2015. Complete removal of AHPS synthetic dye from water using new electro-fenton oxidation catalyzed by natural pyrite as heterogeneous catalyst. J. Hazard. Mater 297, 34e41. Lin, S.H., Chang, C.C., 2000. Treatment of landfill leachate by combined electroFenton oxidation and sequencing batch reactor method. Water Res. 34, 4243e4249. Lin, S.H., Lo, C.C., 1997. Fenton process for treatment of desizing wastewater. Water Res. 31, 2050e2056. Lü, F., Zhang, H., Chang, C.-H., Lee, D.-J., He, P.-J., Shao, L.-M., Su, A., 2008. Dissolved organic matter and estrogenic potential of landfill leachate. Chemosphere 72, 1381e1386. ndez-Ib ~ ez, P., Maldonado, M.I., Blanco, J., Gernjak, W., 2009. Malato, S., Ferna an Decontamination and disinfection of water by solar photocatalysis: recent overview and trends. Catal. Today 147, 1e59. Martínez-Huitle, C.A., Brillas, E., 2009. Decontamination of wastewaters containing synthetic organic dyes by electrochemical methods: a general review. Appl. Catal. B Environ. 87, 105e145. di, R., Ammar, S., 2013. ElectroMhemdi, A., Oturan, M.A., Oturan, N., Abdelhe chemical advanced oxidation of 2-chlorobenzoic acid using BDD or Pt anode and carbon felt cathode. J. Electroanal. Chem. 709, 111e117. Nesheiwat, F.K., Swanson, A.G., 2000. Clean contaminated sites using Fenton’s reagent. Chem. Eng. Prog. 96, 61e66. s, I., 2008. Oxidation pathways of malaOturan, M.A., Guivarch, E., Oturan, N., Sire chite green by Fe3þ-catalyzed electro-Fenton process. Appl. Catal. B Environ. 82, 244e254. Ozcan, A., Sahin, Y., Koparal, A.S., Oturan, M.A., 2008. Carbon sponge as a new cathode material for the electro-Fenton process: comparison with carbon felt cathode and application to degradation of synthetic dye basic blue 3 in aqueous medium. J. Electroanal. Chem. 616, 71e78.

L. Labiadh et al. / Journal of Environmental Management 181 (2016) 515e521 Panizza, M., Cerisola, G., 2001. Removal of organic pollutants from industrial wastewater by electrogenerated Fenton’s reagent. Water Res. 35, 3987e3992. Panizza, M., Cerisola, G., 2003. A comparative study on direct and indirect electrochemical oxidation of polyaromatic compounds. Ann. Chim. 93, 977e984. Panizza, M., Cerisola, G., 2009a. Direct and mediated anodic oxidation of organic pollutants. Chem. Rev. 109, 6541e6569. Panizza, M., Cerisola, G., 2009b. Electro-Fenton degradation of synthetic dyes. Water Res. 43, 339e344. rez, G., Saiz, J., Iban ~ ez, R., Urtiaga, A., Ortiz, I., 2012. Assessment of the formation Pe of inorganic oxidation by-products during the electrocatalytic treatment of ammonium from landfill leachates. Water Res. 46, 2579e2590. Pimentel, M., Oturan, N., Dezotti, M., Oturan, M., 2008. Phenol degradation by advanced electrochemical oxidation process electro-Fenton using a carbon felt cathode. Appl. Catal. B Environ. 83, 140e149. Renou, S., Givaudan, J.G., Poulain, S., Dirassouyan, F., Moulin, P., 2008. Landfill leachate treatment: review and opportunity. J. Hazard. Mater 150, 468e493. ndez, J.M., 2011. Ruiz, E.J., Arias, C., Brillas, E., Hern andez-Ramírez, A., Peralta-Herna

521

Mineralization of Acid Yellow 36 azo dye by electro-Fenton and solar photoelectro-Fenton processes with a boron-doped diamond anode. Chemosphere 82, 495e501. Salem, Z., Hamouri, K., Djemaa, R., Allia, K., 2008. Evaluation of landfill leachate pollution and treatment. Desalination 220, 108e114. s, I., Brillas, E., 2012. Remediation of water pollution caused by pharmaceutical Sire residues based on electrochemical separation and degradation technologies: a review. Environ. Int. 40, 212e229. s, I., Garrido, J.A., Rodríguez, R.M., Brillas, E., Oturan, N., Oturan, M.A., 2007. Sire Catalytic behavior of the Fe3þ/Fe2þ system in the electro-Fenton degradation of the antimicrobial chlorophene. Appl. Catal. B Environ. 72, 382e394. Talalaj, I.A., Biedka, P., 2015. Impact of concentrated leachate recirculation on effectiveness of leachate treatment by reverse osmosis. Ecol. Eng. 85, 185e192. Zhang, H., Ran, X., Wu, X., 2012. Electro-Fenton treatment of mature landfill leachate in a continuous flow reactor. J. Hazard. Mater 241e242, 259e266. Zhang, H., Zhang, D., Zhou, J., 2006. Removal of COD from landfill leachate by electro-Fenton method. J. Hazard. Mater 135, 106e111.