Microwave-assisted chemical oxidation of biological waste sludge: Simultaneous micropollutant degradation and sludge solubilization

Bioresource Technology 146 (2013) 126–134

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Microwave-assisted chemical oxidation of biological waste sludge: Simultaneous micropollutant degradation and sludge solubilization Nalan Bilgin Oncu, Isil Akmehmet Balcioglu ⇑ Bogazici University, Institute of Environmental Sciences, Bebek, 34342 Istanbul, Turkey

h i g h l i g h t s

g r a p h i c a l a b s t r a c t

 Comparison of MW/H2O2 and

a r t i c l e

i n f o

Article history: Received 28 May 2013 Received in revised form 5 July 2013 Accepted 11 July 2013 Available online 19 July 2013 Keywords: Biological waste sludge Antibacterial micropollutants Microwave Advanced oxidation processes Sludge treatment

-MW

Waste sewage sludge

Non-regulated micropollutants in sewage sludge: Antibiotics O

O

F N

O

OH

O

H2N

N

OH OH

OH OH

HO N

HN

Ciprofloxacin CIP

O

OH

Oxytetracycline OTC

Antibiotic degradation

Sludge treatment

Sludge solubilization

MW=S2 O2 8 for sludge treatment.  Oxidant dosing is crucial for recalcitrant antibiotic degradation. 2  MW=S2 O8 offers effective treatment at shorter periods and lower temperatures.  40–75% metal solubilization and enhanced filterability of sludge with MW=S2 O2 8 . 2  MW=S2 O8 oxidizes solubilized ammonia to nitrate.

-MW/H2O2 -MW/S2O82-

OTC degradation

100%

CIP degradation

80% 60%

SCOD/TCOD

40%

NH4+/TKN

20%

PO43-/TP

0%

MeS/MeT

2MW h2o2 s2o8 2 MW/H 2 MW/H 8 82MW/H22OO22 MW/S MW/S2O MW 2O

OH•

SO4-•

a b s t r a c t Microwave-assisted hydrogen peroxide (MW/H2O2) treatment and microwave-assisted persulfate ðMW=S2 O2 8 Þ treatment of biological waste sludge were compared in terms of simultaneous antibiotic degradation and sludge solubilization. A 23 full factorial design was utilized to evaluate the influences of temperature, oxidant dose, and holding time on the efficiency of these processes. Although both yielded P97% antibiotic degradation with 1.2 g H2O2 and 0.87 g S2 O2 per MW/H2O2 and MW=S2 O2 8 8 gram total solids, respectively, at 160 °C in 15 min, MW=S2 O2 8 was found to be more promising for efficient sludge treatment at a lower temperature and a lower oxidant dosage, as it allows more effective 2 gives 48% more activation of persulfate to produce the SO 4 radical. Relative to MW/H2O2, MW=S2 O8 overall metal solubilization, twofold higher improvement in dewaterability, and the oxidation of solubilized ammonia to nitrate in a shorter treatment period. Ó 2013 Elsevier Ltd. All rights reserved.

1. Introduction The presence of a wide variety of emerging micropollutants in sewage sludge (Oncu Bilgin and Balcioglu Akmehmet, 2013a; Stasinakis, 2012; Clarke and Smith, 2011) has increased concerns regarding sludge disposal on land. Challenges associated with the ever-growing volumes of sludge produced from biological wastewater treatment processes in combination with the motivation to recycle the valuable constituents of sludge have brought land application as a means of reusing biosolids in a beneficial way to ⇑ Corresponding author. Tel.: +90 212 359 70 36; fax: +90 212 257 50 33. E-mail addresses: [email protected] (N. Bilgin Oncu), balciogl@boun. edu.tr (I. Akmehmet Balcioglu). 0960-8524/$ - see front matter Ó 2013 Elsevier Ltd. All rights reserved. http://dx.doi.org/10.1016/j.biortech.2013.07.043

the forefront. To minimize the potential risk of pathogens and heavy metals in the environment, waste sludge needs to be stabilized prior to land application. However, conventional stabilization methods such as anaerobic or aerobic digestion and composting cannot destroy many nonregulated organic micropollutants as a result of strong sorption on particulate matter (Stasinakis, 2012). Currently, studies dealing with waste sludge management are aimed at improving the efficiency of the biological stabilization process (Carrere et al., 2010) and the physical properties of the waste sludge (Li et al., 2008) and at recovering valuable nutrients in the sludge (Tyagi and Lo, 2013). For these purposes, various mechanical, thermal, and chemical processes have been used. However, removal of micropollutants from waste sludge has only been investigated in a limited number of studies (Oncu Bilgin and

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Balcioglu Akmehmet, 2013b; McNamara et al., 2012; Carrere et al., 2006), and it requires significant care. During sludge treatment, the degradation efficiency of organic micropollutants is known to be rather low and dependent on their thermal stability (McNamara et al., 2012) and their strong tendency to sorb on sludge (Oncu Bilgin and Balcioglu Akmehmet, 2013b; Carrere et al., 2006). Among various organic micropollutants, antibacterial substances deserve specific attention because they spread into the environment in an uncontrolled way, and this is thought to increase the rate of the development of antibacterial resistance in microorganisms (Oncu Bilgin and Balcioglu Akmehmet, 2013a). Therefore, secondary pollution is created. Moreover, antibiotics, and in particular those from the tetracycline and fluoroquinolone groups, have been detected in sludge at concentrations of milligram per kilogram (Oncu Bilgin and Balcioglu Akmehmet, 2013a). As for ciprofloxacin, an antibiotic from the fluoroquinolone group, a sludge concentration of 97.5 mg/kg was detected (SFT, 2007), and long-term persistence in soil was reported (Golet et al., 2003). A recent study demonstrated that environmental concentration of ciprofloxacin and tetracycline could exert selective pressure and increase the prevalence of resistant bacteria in soil (Tello et al., 2012). Of the sludge treatment processes, microwave (MW) technology has attracted much interest lately, as it can rapidly and homogeneously heat a sample, which allows effective sludge disintegration, conditioning, and pathogen destruction (Wu, 2008). This technology can be utilized as a stand-alone pretreatment process for sludge solubilization or in combination with chemical oxidation to provide further improvement in sludge disintegration and nutrient recovery (Tyagi and Lo, 2013). In a number of studies, it was demonstrated that hydrogen peroxide, the most commonly used reagent for oxidation, synergistically enhances the solubilization of sludge in MW/H2O2 treatment (Tyagi and Lo, 2013; Eskicioglu et al., 2008; Wong et al., 2007). Without MW assistance, it is possible to use hydrogen peroxide and persulfate to enhance sludge dewaterability (Zhen et al., 2012a–c); however, as far as it is known, neither the influence of MW treatment on the fate of micro-organic pollutants in waste sludge nor the combined application of persulfate with MW for sludge treatment have been investigated. Given that the effect of MW irradiation on the desorption of organic pollutants (Wu et al., 2008) is known and that the strong oxidative power of hydrogen peroxide and persulfate on the degradation of organics in solid matrices is also known (Uslu Otker and Balcioglu Akmehmet, 2009), the benefits of a combined process are clear. In this study, the degradation of sorbed antibacterial microorganic pollutants in sewage sludge was investigated during MW/H2O2 and MW=S2 O2 treatments, which were performed in 8 laboratory-scale experiments under different conditions. For this purpose, two antibacterial micropollutants that are commonly detected in sludge, that is, the antibiotics oxytetracycline (OTC) and ciprofloxacin (CIP), were added to the sludge to study the effects of selected operational parameters and their interactions on the efficiency of the applied processes. The efficiency of MW/H2O2 and MW=S2 O2 on the solubilization of metals as a regulated 8 micropollutant group and on the solubilization of organic matter and nutrients was also studied. Furthermore, efforts were made to recognize the contribution of a radical mechanism during the combined MW treatment and chemical oxidation.

2. Methods 2.1. Preparation of antibiotic-contaminated sewage sludge The secondary sewage sludge used in this study was obtained from the recirculation line of a municipal wastewater treatment plant (500,000 population equivalent) located in Istanbul, Turkey.

The plant employs an anaerobic–anoxic–oxic biological process and operates with a sludge age of about 20 days. The physicochemical parameters of the raw sludge samples and their mean values over the course of the experiments are listed in Table 1. After concentrating the sludge by centrifugation, the total solid concentration was adjusted to 10.0 ± 0.1 g/L by following a previously described procedure (Oncu Bilgin and Balcioglu Akmehmet, 2013b). In the treatment experiments, synthetically contaminated secondary sewage sludge was used, and the sludge was spiked with both antibiotics, OTC (C22H24N2O9HCl, >95%) and CIP (C17H18FN3O3HCl, 99%), by taking into account the concentrations of these antibiotics already present in the sludge. In the majority of the experiments, the concentration of the antibiotic was 2 mg/g total solids (TS). However, the performance of the treatment processes was also tested with an environmentally relevant antibiotic concentration of 0.08 mg/g TS. Prior to the sludge treatment experiments, the antibiotic-spiked sludge was equilibrated to provide >95% CIP and OTC sorption, which was verified by dissolved and total antibiotic analyses. 2.2. Treatment of sewage sludge with MW/H2O2 and MW=S2 O2 8 The treatment of sewage sludge was performed with a benchscale microwave irradiation system (Berghof, Speedwave MWS-3, 2.54 GHz), which had the capacity to accommodate up to 12 TFM vessels (each with a volume of 60 mL) and could be operated at a maximum temperature, power, and pressure of 230 °C, 1450 W, and 4000 kPa, respectively. After the addition of either hydrogen peroxide, H2O2 (30% w/w), or sodium persulfate, Na2S2O8 (>98%), to 25 mL of the sludge sample, the samples were treated in closed vessels under predetermined experimental conditions. The temperature of the sludge samples placed in the MW system was increased at a rate of 10 °C/min. Separate experiments in which no oxidant was added were also performed under the same MW conditions. Of the treatments applied to the sludge, only the MW/H2O2 treatment consisted of a preheating stage. Before dosing hydrogen peroxide, all of the samples were heated at 120 °C for 15 min to destruct the biological enzymes in the sludge, whereby undesirable consumption of hydrogen peroxide would be prevented (Wang et al., 2009). In all of the experiments, no effort was made to adjust the pH of the sludge. At the end of the treatment period and before carrying out subsequent analyses, the vessels were inserted in an ice bath without opening the caps and cooled to room temperature to avoid evaporation. 2.3. Experimental design To explore the effects of the selected variables on the efficiency of the MW/H2O2 and MW=S2 O2 treatments, a 23 full factorial 8 Table 1 Characteristics of raw secondary sewage sludge. Sludge properties

Metals

Parameter

Value

Metals

Sludge concentrations (mg/kg DS)

TS (g/L) VS/TS (%) TCOD (g/L) SCOD (mg/L) TKN (mg/L) NHþ 4 NO 3 TP (mg/L)

12.2 ± 0.3 57 ± 5 10.4 ± 0.1 105 ± 7 430 ± 7 1.0 ± 0.4 4.2 ± 0.3 847 ± 32 61.0 ± 9.5

Ni Cr Cu Zn Cd Pb Fe Mn

321 ± 8 605 ± 12 544 ± 85 909 ± 98 BLD BLD 13,877 ± 97 563 ± 67

PO3 4 (mg/L) pH Alkalinity (g/L)

6.5–7.0 1.5

BLD, below limit of detection.

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design with three independent variables (A: temperature, B: oxidant dose, C: MW holding time) was performed. Notably, preliminary experiments were accomplished to determine the extreme values (i.e., corner points) of the variables. Each MW/H2O2 and MW=S2 O2 8 process consisted of 19 runs carried out in random order, and these runs included duplicates at the corner points and triplicates at the center point. The design matrix was established, and the testing was performed with a Minitab 15 (Minitab Inc.). Statistical analysis of the data would allow the main and significant effects of the variables and their interactions to be explained by conducting a minimal number of experiments with real sludge samples. The experimental range and the levels of the independent variables are given in Table 2. As can be seen from Table 2, each one of the three variables received two values, as indicated by the plus and minus signs, whereas the center value is indicated by 0. The minus sign for the oxidant dose variable is indicative of its absence. 2.4. Analytical methods 2.4.1. Extraction and analysis of the antibiotics The quantitative analysis of OTC and CIP in raw and treated sludge samples was based on published methods with slight modifications (Blackwell et al., 2004; Turiel et al., 2006). For this purpose, the sludge samples (10–30 mL) were mixed with an aqueous solution of Mg(NO3)2 (1 M, pH 8.1) for CIP and with an extraction buffer consisting of methanol/0.1 M EDTA/McIlvaine buffer (60 mL of 0.2 M citric acid + 40 mL of 0.4 M Na2HPO4) (50:25:25) for OTC in a vortex for 30 s. Subsequent extraction of the samples by ultrasonication (Sonorex super RK 510–640 W, Morfelden-Walldorf) for 30 min was followed by centrifugation at 5000g for 10 min. Thereafter, the supernatant was withdrawn, and the residue was re-extracted two times as described above. The combined supernatant was filtered through a 0.45 lm membrane syringe filter (regenerated cellulose, Sartorius). For OTC analysis, the filtered samples were purified and concentrated by solidphase extraction, for which SAX (6 mL/500 mg, Phenomenex) and HLB cartridges (6 mL/200 mg, Waters, Milford) were utilized in tandem, and the elution of OTC was accomplished with methanol (4 mL). Consequent analysis of the samples was carried out by high-pressure liquid chromatography, HPLC system (Agilent Technol. 1100 series), and the analytes were separated on a C18 column (YMC-Pack ODS-AQ; 3 lm, 50  4.0 mm). Gradient elution was performed by using acetonitrile (C2H3N, HPLC grade) and water (Milli-Q), both containing 0.1% formic acid (CH2O2, HPLC grade). OTC was detected at 360 nm with a diode array detector, whereas CIP was detected with a fluorescence detector at excitation and emission wavelengths of 280 and 450 nm, respectively. The recovery rates of CIP and OTC from the sludge were 86 ± 2% and 81 ± 3%, respectively. Limits of detection for CIP and OTC with HPLC (S/N = 3) were 0.004 and 0.002 mg/g of dry weight, respectively. Antibiotic degradation was calculated as the percent change in antibiotic concentration after sludge treatment. For each experiment before treatment of the sludge, the initial antibiotic concentration was determined to eliminate the probable influence of changes in sludge characteristics on the antibiotic recoveries.

The observed degradation rates for the antibiotics spiked into the sludge at low concentrations were confirmed by qualitative LC–MS/MS analysis, which was performed with a triple quadrupole mass spectrometer (Agilent Technol. 6460 series). The mass spectrometer was operated in the positive electrospray ionization (ESI) mode. Gradient elution at a flow rate of 0.5 mL/min was performed with oxalic acid solution (0.002 M) and acetonitrile containing 0.2% and 0.1% formic acid, respectively. Chromatographic separation was performed with a Zorbax SB-C18 column (3.5 lm, 3  75 mm) operated at a temperature of 40 °C. The parent mass and product mass range values were 461 and 426–444 for OTC and 332 and 288–314 for CIP, respectively. 2.4.2. Other analyses The raw and treated sludge samples were characterized by the parameters determined by the standard methods (Table 1) (APHA, 2005). The soluble components of the sludge were measured after filtering the centrifuged samples (5000g for 10 min; Eppendorf Centrifuge 5804) through a 0.45 lm filter (Sartorius), and these data were used to calculate the extent of solubilization of the sludge components. Total chemical oxygen demand (TCOD) analysis for the raw and treated sludge samples was carried out after alkaline hydrolysis, and the solubilization of organic carbon was expressed as the percentage of the ratio of soluble chemical oxygen demand (SCOD) to TCOD of the treated sludge. Solubilization of phosphorus, nitrogen, and metals was defined as the percentage þ ratio of released PO3 4 , NH4 , and dissolved metal concentration (Mes) to TP (total phosphorus), TKN (total Kjeldahl nitrogen), and the total metal concentration (Met), respectively. Assessment of the sludge nutrient components was performed spectrophotometrically with analytical test kits (Hach Company) on the basis of the procedures of the manufacturer by utilizing the ascorbic acid method for TP and PO3 analysis, the Nessler method for TKN 4  and NHþ 4 analysis, and the cadmium reduction method for NO3 analysis (APHA, 2005). Total metal concentrations were analyzed after acid digestion of the raw sludge in the MW system according to the EPA Method 3052 (USEPA, 1995), and the metals released in the supernatant of the treated samples were detected by ICP-OES (Perkin-Elmer Optima 2100 DV). Sludge dewaterability was evaluated by the capillary suction time (CST) (APHA, 2005). Analysis of the radical probe anisole (CH3OC6H5, 99%) was carried out in the supernatants of the centrifuged and filtered sludge samples by using an HPLC instrument fitted with a YMC-Pack ODSAQ column (3 lm, 50  4.0 mm) and a diode array detector. The mobile phase was a mixture of acetonitrile and water (60:40 v/v) delivered at a flow rate of 0.7 mL/min. Anisole was detected at 254 nm. The recovery rate of anisole in the supernatant was 67 ± 0%. The concentrations of hydrogen peroxide and persulfate in the samples were determined by the iodometric method (Kolthoff and Carr, 1953). Given that residual hydrogen peroxide exerts considerable COD interference, a photometric method based on the formation of a copper(II)-2,9-dimethyl-1,10-phenanthroline (DMP) complex (Kosaka et al.,1998) was used to confirm the low concentrations of hydrogen peroxide. Then, the COD equivalent of the residual hydrogen peroxide was subtracted from the

Table 2 Independent variables and their levels used in the 23 full factorial design for sewage sludge treatment with MW/H2O2 and MW=S2 O2 8 . Experimental factors

Symbol

Temperature (°C) Oxidant dose (g/g TS) MW holding time (min)

A B C

MW=S2 O2 8

MW/H2O2 Level ()

Level (0)

Level (+)

Level ()

Level (0)

Level (+)

120 0.0 5

140 0.6 10

160 1.2 15

120 0.00 5

140 0.44 10

160 0.87 15

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N. Bilgin Oncu, I. Akmehmet Balcioglu / Bioresource Technology 146 (2013) 126–134 Table 3 23 full factorial design table for the coded factors and the responses obtained for sewage sludge treatment with MW/H2O2. Experiment number

1 13 5 7 2 12 3 4 10 11 19 16 17 14 18 8 9 6 15

Process parameter level

Responses Antibiotic degradation (%)

Solubilization (%)

A

B

C

OTC

CIP

Organic carbon (SCOD/TCOD  100)

Nitrogen ðNHþ 4 =TKNÞ

Phosphorus ðPO3 4 =TPÞ

Overall metal (Mes/Met)

        0 0 0 + + + + + + + +

    + + + + 0 0 0     + + + +

  + +   + + 0 0 0   + +   + +

40.3 41.6 43.2 44.2 83.6 84.8 86.0 87.3 91.4 91.6 91.5 81.6 82.3 83.8 84.6 98.0 98.2 98.8 98.9

7.6 7.2 8.7 9.1 31.6 32.9 35.2 36.5 60.9 60.3 60.5 9.8 9.9 12.4 11.9 88.6 89.2 98.1 98.3

28.1 27.8 31.4 32.1 24.2 23.3 30.8 31.2 36.8 37.6 38.2 34.5 34.0 40.2 39.8 40.4 38.6 26.5 24.7

2.6 2.9 4.1 4.4 17.7 18.1 23.9 24.9 11.0 12.3 11.9 7.4 7.7 12.6 13.2 40.7 42.4 46.7 44.5

8.4 8.7 9.1 9.4 16.5 17.2 20.3 19.9 29.2 28.9 29.3 11.5 11.8 13.5 14.5 49.6 49.0 51.3 53.3

3.0 2.9 3.0 2.9 11.6 11.9 13.0 12.6 19.5 18.9 19.3 3.6 3.6 3.9 3.8 22.3 23.3 25.3 25.9

A, temperature; B, hydrogen peroxide dose; C, MW holding time.

measured COD value to eliminate its positive interference (Kang et al., 1999).

3.1. Comparison of the MW/H2O2 and MW=S2 O2 8 treatments for antibiotic degradation in sewage sludge

3. Results and discussion

The preheating stage required for the MW/H2O2 treatment resulted in a prolonged MW exposure time. Thus, different OTC and CIP degradation efficiencies were obtained when MW treatment was performed in the absence of hydrogen peroxide (exp. 1, 13 in Table 3) and persulfate (exp. 6, 14 in Table 4). An additional experiment to evaluate a single-stage MW/H2O2 treatment (1.2 g H2O2/g TS, 160 °C, 15 min) performance yielded degradation rates of only 81% for OTC and 71% for CIP, whereas with a preheating stage, a degradation rate greater than 98% was obtained for both

The results of antibiotic degradation and sludge solubilization for the 19 experiments performed with the MW/H2O2 and MW=S2 O2 8 treatments are shown in Tables 3 and 4, respectively. The statistical significance of the effects of the selected independent variables on the efficiency of these treatments was evaluated, and the significant factors (p < 0.05) listed in decreasing order of effects are displayed in Fig. 1a–f as Pareto plots.

Table 4 23 full factorial design table for the coded factors and the responses obtained for sewage sludge treatment with MW=S2 O2 8 . Experiment number

6 14 9 17 3 19 2 12 5 13 7 10 11 15 16 4 18 1 8

Process parameter level

Responses Antibiotic degradation (%)

Solubilization (%)

A

B

C

OTC

CIP

Organic carbon (SCOD/TCOD  100)

Nitrogen ðNHþ 4 =TKNÞ

Phosphorus ðPO3 4 =TPÞ

Overall metal (Mes/Met)

        0 0 0 + + + + + + + +

    + + + + 0 0 0     + + + +

  + +   + + 0 0 0   + +   + +

31.2 33.5 36.7 38.9 82.3 83.3 85.9 88.0 97.4 97.1 96.4 77.4 77.6 78.3 79.1 96.2 97.0 98.5 99.6

0.3 0.2 0.7 0.8 58.3 59.8 60.6 62.1 88.2 87.0 87.3 3.3 4.8 7.0 6.9 94.2 93.9 96.6 97.3

25.0 24.4 25.8 26.4 23.8 23.8 23.9 24.1 26.3 25.7 26.1 30.0 29.6 32.8 33.2 20.4 19.9 22.1 21.5

0.5 0.4 1.9 1.5 8.1 9.7 12.1 13.2 13.5 12.9 12.1 2.2 2.5 3.1 2.9 13.8 14.1 14.9 15.5

3.1 3.0 7.2 6.9 25.6 26.2 27.5 28.2 29.3 29.3 29.4 8.1 8.3 10.6 11.0 41.2 41.9 43.9 44.2

2.5 2.4 2.8 2.6 39.1 38.7 52.0 52.2 57.0 56.5 56.1 3.0 3.0 2.5 3.0 65.7 65.4 73.6 73.4

A, temperature; B, persulfate dose; C, MW holding time.

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N. Bilgin Oncu, I. Akmehmet Balcioglu / Bioresource Technology 146 (2013) 126–134

OTC degradation (%)

(a)

CIP degradation (%)

(b) B

B

A A

AB

AB

C BC

MW/H2O2

C

AC B

MW/H2O2

ABC

A

B A

AB

AB C

C

2-

-50

-30

AC

MW/S2O8

-10 10 30 50 Standardized effects

70

0

90

50

100

150

2-

200

250

Standardized effects

SCOD/TCOD×100 (%)

(c)

MW/S2O8

AC

+

NH4 /TKN (%)

(d) B

A ABC

A

AC

AB

BC

C

B

ABC

AB

MW/H2O2

BC

MW/H2O2

C

B

B

A

AB C

C

AB

A

AC

AC

-40

-30

-20

2-

2-

MW/S2O8

BC

-10 0 10 Standardized effects

20

30

MW/S2O8

BC

-20

40

0

3-

PO4 /TP (%)

(e)

80

B

A

A

AB

AB C

C

BC

MW/H2O2

BC

MW/H2O2

AC

B

B

A

A

AB

AB BC

C

C

BC

AC

2MW/S2O8

ABC

-30

60

Me S /Me T (%)

(f)

B

20 40 Standardized effects

0

30

60

90

120

150

180

210

Standardized effects

2-

MW/S2O8

ABC

-100

0

100 200 300 Standardized effects

400

Fig. 1. Pareto charts for (a) OTC and (b) CIP degradation, solubilization of (c) organics, (d) phosphorus, (e) nitrogen, and (f) metals during sewage sludge treatment with MW/ H2O2 and MW=S2 O2 8 (A: Temperature; B: Oxidant dose; C: MW holding time).

antibiotics. The relatively poor antibiotic degradation efficiency obtained in the absence of a preheating stage could be the result of the undesired consumption of hydrogen peroxide by the nontarget components of the sludge, as explained previously. In accordance with this hypothesis, <1% of the initially added hydrogen peroxide remained in the sludge at the end of the single-stage MW/H2O2 treatment period, whereas this amount increased to 17% with the integration of the preheating stage (Supplementary

data, exp. 6, 15 in Table S1). The higher availability of hydrogen peroxide in a two-stage MW/H2O2 treatment can be beneficial for contaminant desorption, which is a characteristic of catalyzed hydrogen peroxide treatment systems (Watts et al., 1999). Statistical analysis of the results presented in Tables 3 and 4 clearly indicates that substantial antibiotic degradation depends on the dosing of the oxidants, which was the most important parameter in the MW treatment of the sludge (Fig. 1a and b). At

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160 °C with a MW holding time of 15 min, the addition of 1.2 g H2O2 and 0.87 g S2 O2 8 per gram of TS improved the average degradation rate of CIP from 12% to 98% and from 7% to 97%, respectively. At a high temperature, although the amount of antibiotic removed was almost the same regardless of the type of oxidant at the center point of the experimental design, the effect on the degradation of the antibiotics was greater for S2 O2 than for 8 H2O2. In contrast, the control experiments performed with these oxidants in the absence of MW irradiation resulted in negligible antibiotic degradation (<10%) within a 30 min treatment period (Fig. 2). Therefore, it can be deduced that the addition of the oxidant and the MW irradiation synergistically promote the degradation of the antibiotics. However, the degradation of OTC was affected by oxidant dosing to a lesser extent than the degradation of CIP. Even in the absence of the oxidant, 79% degradation of OTC was obtained by MW treatment at 160 °C for 15 min (exp. 15, 16 in Table 4). As deduced from Fig. 1, the MW temperature was the second most important factor in the degradation of the antibiotic during the MW/H2O2 and MW=S2 O2 8 treatments, yet the interaction of the oxidant dose and the temperature did not induce a positive influence on the degradation of OTC, which is in contrast to CIP, owing to the instability of OTC at elevated temperatures. The thermal stabilities of the antibiotics were confirmed by performing experiments in deionized water under experimental conditions that were similar to those applied to the sludge (Fig. 2). Even if CIP is relatively stable at high temperature (Fig. 2) and even if almost complete persulfate consumption took place at 120 °C with a MW=S2 O2 8 treatment holding time of 5 min (Supplementary data, Table S2), the increase in the MW temperature from 120 to 160 °C enhanced the degradation of CIP by 36% (exp. 2, 12– 1, 8 in Table 4). This result can be attributed to the dramatic decrease in the pH of the sludge (Supplementary data, Table S2), which could promote the desorption of the antibiotics from the sludge as a result of increased electrostatic repulsions between the sludge particles and the zwitterionic antibiotics (Golet et al., 2002). Hence, dissolved components of the sludge could be more vulnerable to oxidation than those sorbed on the particulate matter (Oncu Bilgin and Balcioglu Akmehmet, 2013b). The influence of sludge pH on antibiotic degradation during MW treatment is supported by additional experiments performed at an initial pH of 2 (adjusted with 1 M H2SO4) and at either 120 or 160 °C for 15 min. The results of these experiments (Supplementary data, Table S3) demonstrated that although acidic pH values caused considerable desorption of both antibiotics, particularly for CIP (OTC = 12%, CIP = 58% at 120 °C), and that an increase in the temperature exerted a slight contribution to their desorption, MW

OTC (2 mg/g TS)

treatment without the addition of any oxidant did not provide efficient degradation of the antibiotics. As a result, the overall degradation rates of OTC and CIP at pH 2 at 160 °C were only 81% and 12%, respectively. Despite being a significant factor (p < 0.05) in both MW/H2O2 and MW=S2 O2 8 , the effect of holding time was not very pronounced in the degradation of the antibiotics (Fig. 1a and b). For MW=S2 O2 8 , this result can be attributed to the consumption of the oxidant (Supplementary data, Table S2). In contrast, a higher concentration of residual hydrogen peroxide was observed by increasing the holding time in the MW/H2O2 treatment (Supplementary data, Table S1), which suggests the production of hydrogen peroxide as a result of radical reactions. Nevertheless, neither the remaining hydrogen peroxide in the sludge nor a prolonged treatment time improved the degradation rate of the antibiotics at 120 and 140 °C, and the performance of the MW/H2O2 treatment was lower than that of the MW=S2 O2 8 treatment. Therefore, the small positive effect of the holding time on the degradation of the antibiotic may suggest increased competitive influence exerted by the solubilized components of the sludge with a prolonged treatment period. The higher performance of the MW=S2 O2 8 treatment at lower temperature is indicative of more effective activation of persulfate to generate the sulfate radical, which exhibits higher selectivity towards certain (e.g., carboxylic, anilinic, and phenolic) functional groups of organics (Neta et al., 1977). Regardless of the spiking concentrations of the antibiotics in the sludge, the degradation rates were almost the same for both the MW/H2O2 and MW=S2 O2 treatments at 120 and 160 °C (Fig. 2). 8 It was previously shown that the degradation of sorbed antibiotics, particularly at low concentrations, was influenced to a greater extent by competition with solubilized organic constituents during the ozonation of waste sludge (Oncu Bilgin and Balcioglu Akmehmet, 2013b). However, as shown in this study the desorbing ability of MW irradiation promoted the concurrent extraction and degradation of the antibiotics in the presence of the oxidant. However, it was not possible to analyze the antibiotics in the aqueous phase because of their fast degradation, which took place simultaneously with the desorption of the antibiotics. 3.2. Comparison of MW/H2O2 and MW=S2 O2 8 for sewage sludge solubilization As clearly seen from the results in Tables 3 and 4, although the use of MW irradiation alone did not provide substantial antibiotic degradation, effective COD solubilization was achieved under mild

CIP (2 mg/g TS)

OTC (0.08 mg/g TS)

CIP (0.08 mg/g TS)

150 Secondary sludge

Antibiotic degradation (%)

Water 125 120°C

160°C

120°C

160°C

100

Ambient temperature

120°C

160°C

120°C

160°C

120°C

160°C

120°C

160°C

120°C

160°C

120°C

160°C 2-

75 50 25 0 MW

H2O2 H2 O2

S2O8 2-

S 2 O8

MW

MW/H2 O2

MW/S 2 O8

Fig. 2. Comparison of MW, MW/H2O2, and MW=S2 O2 8 treatment efficiencies for antibiotic degradation at two different spiking concentrations in water and sewage sludge. Holding time: 15 min, oxidant dose: 1.2 g H2O2/g TS and 0.87 g S2 O2 8 =g TS, treatment time of control experiments carried out at room temperature: 30 min.

N. Bilgin Oncu, I. Akmehmet Balcioglu / Bioresource Technology 146 (2013) 126–134

TCOD

SCOD

TCOD decrease 60

10000 MW

TCOD and SCOD (mg/L)

Control 8000

MW/S2O82-

MW/H2O2

50 2H2O2 S2O8

160°C

120°C

160°C

120°C

120°C

160°C

40 6000 30 4000 20 2000

TCOD decrease (%)

132

10

0

0 5

15

5

15

5

15

5

15

5

15

5

15

MW holding time (min) 2 Fig. 3. Comparison of the influences of MW, MW/H2O2, and MW=S2 O2 8 treatments on TCOD, SCOD, and decrease in TCOD. Oxidant dose: 1.2 g H2O2/g TS and 0.87 g S2 O8 =g TS.

conditions, that is, low temperature and short reaction time. At 120 °C with a MW treatment holding time of 15 min, the solubilization percent of organic carbon was 26% (exp. 9, 17 in Table 4), which increased to 32% in the two-step MW treatment with an initial preheating stage (exp. 5, 7 in Table 3). A further increase in this ratio to 40% was achieved by increasing the temperature to 160 °C (exp. 14, 18 in Table 3). Although COD solubilization greatly depends on various factors including sludge source, TS concentration of the sludge, and MW treatment conditions, which vary considerably in different studies, the SCOD/TCOD ratios obtained in the present study were in the range of those reported in other studies (Tyagi and Lo, 2013). Similar to previous results (e.g., Eskicioglu et al., 2008), the addition of the oxidant to the sludge in the MW treatment resulted in simultaneous solubilization of the organics and the oxidation of the solubilized organic components of the sludge, as realized by a decrease in both the TCOD and SCOD values (Fig. 3). Consequently, the oxidant dose and its interaction with other factors exhibit a negative influence (Fig. 1c). A higher MW temperature and a prolonged treatment period enhanced the oxidation of the solubilized components. Hence, a negative three-factor interaction (temperature, oxidant dose, exposure time) was revealed for the SCOD/ TCOD ratio (Fig. 1c). However, as opposed to MW=S2 O2 8 , MW/ H2O2 treatment resulted in a greater decrease in the value of TCOD, regardless of the temperature and holding time (Fig. 3). It is evident that the MW treatment conditions that favor the degradation of the antibiotics caused a deterioration in the solubilization of organic carbon, particularly for the MW/H2O2 treatment, but the solubilization of the nutrients from the sludge was positively influenced by the addition of the oxidant (Fig. 1d and e). In accordance with the results of a previous study (Wong et al., 2007), the effect of the addition of the oxidant on the solubilization of the nutrients was greater than that of the temperature. The higher release of ammonia with the addition of the oxidant can be attributed to the higher oxidation rate of the proteins released from the bacterial cells during MW irradiation, as it is known that MW irradiation alone is not able to degrade proteins (Toreci et al., 2010). Relative to the addition of persulfate to the sludge, the addition of hydrogen peroxide exerted a higher influence on the release of ammonia at both the low and high temperatures used in the MW treatment. Still, persulfate induced the oxidation of ammonia to nitrate, and a nitrate concentration that was almost threefold higher was detected in the sludge samples that were treated under the extreme conditions of the experimental design. Although the solubilization rates of phosphorus in both MW/H2O2 and

MW=S2 O2 8 were comparable, the latter treatment resulted in the oxidation of ammonia to nitrate, which could decrease the potential loss of nitrogen from the sludge through ammonia volatilization. One of the dramatic differences between MW/H2O2 and MW=S2 O2 is their influence on the overall solubilization of the 8 metal (e.g., exp. 6, 15 in Table 3; exp. 1, 8 in Table 4). Whereas the MW treatment yielded only about 3% solubilization of the metal at 120 °C with a holding time of 15 min, the addition of hydrogen peroxide and persulfate enhanced the overall solubilization of the metal to 13% and 52%, respectively, and this effect became more remarkable at higher temperatures. The high efficiency of metal release during the MW=S2 O2 treatment can be attributed 8 to the enhanced solubilization of the sludge, as well as to the influence exerted on metal speciation by the severe decrease in the pH of the sludge. This suggestion can be supported by 32% overall solubilization of the metal obtained by MW treatment at pH 2 without the addition of any oxidant. Solubilization rates of individual metals in MW/H2O2 and MW=S2 O2 8 as well as in MW treatment at pH 2 are given in Tables S1, S2, and S3, respectively (Supplementary data). In the treatment of the sludge, metals that are associated with organics are generally not easily solubilized, and Cu is a typical example (Beauchesne et al., 2007). However, MW=S2 O2 treatment was effective for the solubilization of this 8 metal as well as for overall metal solubilization. Consequently, remarkable solubilization was attained under the acidic conditions that developed during the MW=S2 O2 8 treatment. 3.3. Comparison of MW/H2O2 and MW=S2 O2 8 for sewage sludge dewaterability Dewatering properties are important for sludge management and dewaterability can influence the ease of separation of solubilized metals. In this study, the influence of MW/H2O2 and MW=S2 O2 8 on sludge dewaterability was assessed by the CST. In Fig. 4, the results obtained only at the extreme conditions of the processes (160 °C, 15 min, and oxidant doses of 1.2 g H2O2/g TS and 0.87 g S2 O2 8 =g TS) are given as CST per gram of suspended solids to better compare the two processes. Generally, both treatments improved sludge dewaterability, but the enhancement was twofold higher for the MW=S2 O2 treat8 ment. The difference between the performances of the treatments can be attributed to the considerably higher amount of solubilized iron (Supplementary data, Table S2), which can act as a coagulant. Detailed investigations to further elucidate the influence of the

133

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1CST/SS Series2

Series3 SS

10

H2 O2 þ MW irradiation ! 2OH

ð1Þ

 S2 O2 8 þ MW irradiation ! 2SO4

ð2Þ

8

6

6

4

4

2

2

0

0

Raw initial secondary raw sludge sludge

2-

MW/H2O2 MW/H 2 O2

In addition to MW irradiation, the metal content of the sludge (Table 1), especially the metals released during sludge treatment (Tables 3 and 4), may contribute to the activation of the oxidants to produce reactive radicals in both the MW/H2O2 and MW=S2 O2 treatments according to Eqs. (2) and (3) (Siegrist 8 et al., 2011):

SS (sg/L)

8 CST/SS (sec/g/L)

and the oxidation of the organic matter, as well as the release of the nutrients and metals:

Series4

10

MW/S2O82MW/S 2 O8

H2 O2 þ Menþ ! Meðnþ1Þþ þ OH þ OH

ð3Þ

 nþ S2 O2 ! Meðnþ1Þþ þ SO2 8 þ Me 4 þ SO4

ð4Þ

Metal activation could be especially important in MW=S2 O2 8 treatment. A severe decrease in the pH took place, in particular at 160 °C (Supplementary data, Table S2). The acidic conditions could have further increased the activation of persulfate to produce the sulfate radicals, as shown in Eq. (5) (Peyton, 1993), which created an even higher concentration of the sulfate radical, and in turn, higher amounts of the radical probe or higher antibiotic degradation rates could have been achieved:

Fig. 4. Influences of MW/H2O2 and MW=S2 O2 8 treatments on sludge dewaterability. Oxidant dose: 1.2 g/g TS and 0.87 g/g TS, temperature: 160 °C, MW holding time: 15 min.

processes on sludge dewaterability and to evaluate the biodegradability of sludge after treatment with MW/H2O2 and MW=S2 O2 8 are currently underway.

  2 þ þ S2 O2 8 þ H ! HS2 O8 ! SO4 þ SO4 þ H

3.4. Investigation of the contribution of a radical mechanism in MW=S2 O2 8 and MW/H2O2

ð5Þ

Considering the presence of hydrogen peroxide residues in the treated sludge (Supplementary data, Table S1) that could act as an indicator for the formation of hydroxyl radicals in the MW/ H2O2 treatment, additional MW irradiation experiments were performed to detect the formation of hydrogen peroxide in water dosed with 356 mM H2O2 (corresponding to the 1.2 g/g TS dosage in sludge) at different temperatures and holding times (Supplementary data, Fig. S1). In these experiments, the formation of hydrogen peroxide was confirmed, and the elevated MW temperature and the prolonged holding time resulted in a higher concentration of hydrogen peroxide in the water. Certainly, the concentration of hydrogen peroxide generated under the MW conditions would be lower in the sludge as a result of the high oxidant demand of the complex matrix and the competitive influence exerted by the sludge constituents that were solubilized during the preliminary heating stage (Table 3). Hence, the degradation rate of anisole in the MW/H2O2 treatment at 120 °C was not higher than

To explain the probable contribution of free radicals on the efficiencies of the MW/H2O2 and MW=S2 O2 8 treatments, anisole was 9  9 used as a radical probe ½kðSO 4 Þ ¼ 4:9  10 , kðOH Þ ¼ 7:8  10 ; Liang and Su, 2009], and the degradation rate of anisole at an initial concentration of 0.06 mM was investigated at 120 and 160 °C without applying a holding time because of the rapid degradation of the radical probe (Fig. 5). Relative to the results obtained from the control experiments performed in the absence of either MW irradiation or the oxidant, the degradation rate of the radical probe increased in both the MW/H2O2 and MW=S2 O2 treatments. Therefore, it can be sug8 gested that, in the current study, during MW irradiation both hydrogen peroxide and persulfate can generate reactive radicals (Eqs. (1) and (2)), which improves the degradation of the antibiotics

160

H2O2

Anisole degradation (%)

MW

Control

140 120

120°C 160°C

S2O82-

MW/S2O82-

MW/H2O2 120°C

160°C

120°C 100

160°C 100

99

100

84

82 74

80

57

60 41

47

40 20

20 1

3

1

3

0.6

1.2

0.44

0.87

0 0

0 0.6 1.2 0.6 Oxidant dose (g/g TS)

1.2

0.44

0.87

0.44

0.87

Fig. 5. Radical probe compound degradation during sewage sludge treatment with MW, MW/H2O2, and MW=S2 O2 8 . Initial anisole concentration: 0.06 mM, treatment time of control experiment carried out at room temperature: 20 min.

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that in the MW=S2 O2 8 treatment. Therefore, it can be inferred that, at this temperature, the activation of hydrogen peroxide to produce hydroxyl radicals in the sludge is not efficient, probably because of the lack of sufficiently acidic conditions (Supplementary data, Table S1). Notably, the differences in the degradation rates of anisole in the MW/H2O2 and MW=S2 O2 treatments at 120 °C 8 were not as high as the differences in the degradation rates of CIP, as anisole was mainly present in the dissolved phase and was more available for oxidation. In contrast, at 160 °C, the MW/ H2O2 and MW=S2 O2 8 treatments provided degradation rates of anisole and the antibiotics that were almost the same, although the oxidation mechanisms of the sulfate and hydroxyl radicals and their rates of reaction with the inorganic anions produced in the treated sludge are considerably different (Neta et al., 1977). 4. Conclusion This study offers an approach to combine MW irradiation with chemical oxidation for the treatment of waste sludge to degrade sorbed antibiotics that can be problematic for the safe reuse or disposal of biosolids. Relative to the conditions developed during MW/H2O2, the treatment conditions developed during MW=S2 O2 8 were more effective for the degradation of the thermally more stable and strongly sorbed antibiotic, CIP, in a shorter treatment period, which also resulted in lower TCOD degradation. Oxidant dosing was necessary for antibiotic degradation and for efficient nutrient and metal solubilization. Further optimization and evaluation of MW=S2 O2 8 for sludge treatment is promising. Acknowledgements The authors would like to thank the Scientific Research Council of Bogazici University for funding this investigation (Project No.: 11Y00P7) and the Pendik Veterinary Control and Research Institute for allocating their facilities to conduct the LC–MS/MS analyses. Appendix A. Supplementary data Supplementary data associated with this article can be found, in the online version, at http://dx.doi.org/10.1016/j.biortech.2013. 07.043. References APHA, 2005. Standard Methods for the Examination of Water and Wastewater, 20th ed. American Public Health Association, Washington, DC, USA. Beauchesne, I., Ben, C.R., Mercier, G., Blais, J.F., Ouarda, T., 2007. Chemical treatment of sludge: in-depth study on toxic metal removal efficiency, dewatering ability and fertilizing property preservation. Water Res. 41, 2028–2038. Blackwell, P.A., Lutzhoft, H.C.H., Ma, H.P., Halling-Sorensen, B., Boxall, A.B.A., Kay, P., 2004. Ultrasonic extraction of veterinary antibiotics from soils and pig slurry with SPE clean-up and LC–UV and fluorescence detection. Talanta 64, 1058–1064. Carrere, H., Bernal-Martinez, A., Patureau, D., Delgenes, J.P., 2006. Parameters explaining removal of PAHs from sewage sludge by ozonation. AlChE 52, 3612– 3620. Carrere, H., Dumad, C., Battimelli, A., Batstone, D.J., Delgenes, J.P., Steyer, J.P., Ferrer, I., 2010. Pretreatment methods to improve sludge anaerobic degradability: a review. J. Hazard. Mater. 183, 1–15. Clarke, B.O., Smith, S.R., 2011. Review of ‘emerging’ organic contaminants in biosolids and assessment of international research priorities for the agricultural use of biosolids. Environ. Int. 37, 226–247. Eskicioglu, C., Prorot, A., Marin, J., Droste, R.L., Kennedy, K.J., 2008. Synergetic pretreatment of sewage sludge by microwave irradiation in presence of H2O2 for enhanced anaerobic digestion. Water Res. 42, 4674–4682. Golet, E.M., Strehler, A., Alder, A.C., Giger, W., 2002. Determination of fluoroquinolone antibacterial agents in sewage sludge and sludge-treated soil using accelerated solvent extraction followed by solid-phase extraction. Anal. Chem. 74, 5455–5462.

Golet, E.M., Xifra, I., Siegrist, H., Alder, A.C., Giger, W., 2003. Environmental exposure assessment of fluoroquinolone antibacterial agents from sewage to soil. Environ. Sci. Technol. 37, 3243–3249. Kang, Y.W., Cho, M.J., Hwang, K.Y., 1999. Correction of hydrogen peroxide interference on standard chemical oxygen demand test. Water Res. 33, 1247– 1251. Kolthoff, I.M., Carr, E.M., 1953. Volumetric determination of persulfate in the presence of organic substances. Anal. Chem. 25, 298–301. Kosaka, K., Yamada, H., Matsui, S., Echigo, S., Shishida, K., 1998. Comparison among the methods for hydrogen peroxide measurements to evaluate advanced oxidation processes: application of a spectrophotometric method using copper(II) ion and 2,9-dimethyl-1,10-phenanthroline. Environ. Sci. Technol. 32, 3821–3824. Li, H., Jin, Y.Y., Mahar, R., Wang, Z.Y., Nie, Y.F., 2008. Effects and model of alkaline waste activated sludge treatment. Bioresour. Technol. 99, 5140–5144. Liang, C., Su, H.W., 2009. Identification of sulfate and hydroxyl radicals in thermally activated persulfate. Ind. Eng. Chem. Res. 48, 5558–5562. McNamara, P.J., Wilson, C.A., Wogen, M.T., Murthy, S.N., Novak, J.T., Novak, P.J., 2012. The effect of thermal hydrolysis pretreatment on the anaerobic degradation of nonylphenol and short-chain nonylphenol ethoxylates in digested biosolids. Water Res. 46, 2937–2946. Neta, P., Madhavan, V., Zemel, H., Fessenden, R.W., 1977. Rate constants and mechanisms of reaction of SO4 with aromatic compounds. J. Am. Chem. Soc. 99, 163–164. Oncu Bilgin, N., Balcioglu Akmehmet, I., 2013a. Antimicrobial contamination removal from environmentally relevant matrices: a literature review and a comparison of three processes for drinking water treatment. Ozone Sci. Eng. 35, 73–85. Oncu Bilgin, N., Balcioglu Akmehmet, I., 2013b. Degradation of ciprofloxacin and oxytetracycline antibiotics in waste sewage sludge by ozonation. J. Adv. Oxid. Technol. 16, 107–116. Peyton, G.R., 1993. The free-radical chemistry of persulfate-based total organic carbon analyzers. Marine Chem. 41, 91–103. SFT, 2007. Occurrence of selected pharmaceuticals in wastewater effluents from hospitals (Ulleval and Rikshospitalet) and VEAS wastewater treatment works. Norwegian Pollutant Control Authority SFT Report, TA-2246/2007. Siegrist, R.L., Crimi, M., Simpkin, T.J. (Eds.), 2011. In Situ Chemical Oxidation for Groundwater Remediation. Springer Science + Business Media, New York. Stasinakis, A.S., 2012. Review on the fate of emerging contaminants during sludge anaerobic digestion. Bioresour. Technol. 121, 432–440. Tello, A., Austin, B., Telfer, T.C., 2012. Selective pressure of antibiotic pollution on bacteria of importance to public health. Environ. Health Perspect. 120, 1100– 1106. Toreci, I., Kennedy, K.J., Droste, R.L., 2010. Effect of high-temperature microwave irradiation on municipal thickened waste activated sludge solubilization. Heat Transfer Eng. 31, 766–773. Turiel, E., Martin-Esteban, A., Tadeo, J.L., 2006. Multiresidue analysis of quinolones and fluoroquinolones in soil by ultrasonic-assisted extraction in small columns and HPLC–UV. Anal. Chim. Acta 562, 30–35. Tyagi, V.K., Lo, S.L., 2013. Microwave irradiation: a sustainable way for sludge treatment and resource recovery. Renew. Sust. Energy Rev. 18, 288–305. USEPA, 1995. Method 3052: microwave assisted acid digestion of siliceous and organically based matrices. Available from: . Uslu Otker, M., Balcioglu Akmehmet, I., 2009. Simultaneous removal of oxytetracycline and sulfamethazine antibacterials from animal waste by chemical oxidation processes. J. Agric. Food Chem. 57, 11284–11291. Wang, Y., Wei, Y., Liu, J., 2009. Effect of H2O2 dosing strategy on sludge pretreatment by microwave-H2O2 advanced oxidation process. J. Hazard. Mater. 169, 680– 684. Watts, R.J., Bottenberg, B.C., Hess, T.F., Jensen, M.D., Teel, A.L., 1999. Role of reductants in the enhanced desorption and transformation of chloroaliphatic compounds by modified Fenton’s reactions. Environ. Sci. Technol. 33, 3432– 3437. Wong, W.T., Lo, K.V., Liao, P.H., 2007. Factors affecting nutrient solubilization from sewage sludge using microwave-enhanced advanced oxidation process. J. Environ. Sci. Health A 42, 825–829. Wu, T.N., 2008. Environmental perspectives of microwave applications as remedial alternatives: review. Pract. Period. Hazard. Toxic Radioact. Waste Manage. 12, 102–115. Zhen, G.Y., Lu, X.Q., Zhao, Y.C., Chai, X.L., Niu, D.J., 2012a. Enhanced dewaterability of sewage sludge in the presence of Fe(II)-activated persulfate oxidation. Bioresour. Technol. 116, 259–265. Zhen, G.Y., Lu, X.Q., Li, Y.Y., Zhao, Y.C., Wang, B.Y., Song, Y., Chai, X.L., Niu, D.J., Cao, X.Y., 2012b. Novel insights into enhanced dewaterability of waste activated sludge by Fe(II)-activated persulfate oxidation. Bioresour. Technol. 119, 7–14. Zhen, G.Y., Lu, X.Q., Wang, B.Y., Zhao, Y.C., Chai, X.L., Niu, D.J., Zhao, A.H., Li, Y.Y., Song, Y., Cao, X.Y., 2012c. Synergetic pretreatment of waste activated sludge by Fe(II)-activated persulfate oxidation under mild temperature for enhanced dewaterability. Bioresour. Technol. 124, 29–36.