H2O2 oxidation processes: Effects of reaction conditions and sludge matrix

Science of the Total Environment 493 (2014) 307–323

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Science of the Total Environment journal homepage: www.elsevier.com/locate/scitotenv

Removal of phenolic endocrine disrupting compounds from waste activated sludge using UV, H2O2, and UV/H2O2 oxidation processes: Effects of reaction conditions and sludge matrix Ai Zhang, Yongmei Li ⁎ State Key Laboratory of Pollution Control and Resources Reuse, College of Environmental Science and Engineering, Tongji University, Shanghai 200092, China

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

• UV/H2O2 is effective in removing EDCs from WAS and improve sludge solubilization. • EDC degradation in sludge fitted well with pseudo-first-order kinetics. • •OH takes the most important role in UV/H2O2 oxidation of EDCs in sludge. • Metal ions in sludge facilitate the removal of EDCs during UV/H2O2 oxidation. • Effect of HA in sludge on EDC removal during UV/H2O2 oxidation is observed.

a r t i c l e

i n f o

Article history: Received 25 February 2014 Received in revised form 25 April 2014 Accepted 31 May 2014 Available online xxxx Editor: Adrian Covaci Keyword: Endocrine disrupting compounds (EDCs) UV/H2O2 oxidation Waste activated sludge (WAS) Metal ions Humic acid (HA) Reaction condition

a b s t r a c t Removal of six phenolic endocrine disrupting compounds (EDCs) (estrone, 17β-estradiol, 17α-ethinylestradiol, estriol, bisphenol A, and 4-nonylphenols) from waste activated sludge (WAS) was investigated using ultraviolet light (UV), hydrogen peroxide (H2O2), and the combined UV/H2O2 processes. Effects of initial EDC concentration, H2O2 dosage, and pH value were investigated. Particularly, the effects of 11 metal ions and humic acid (HA) contained in a sludge matrix on EDC degradation were evaluated. A pseudo-first-order kinetic model was used to describe the EDC degradation during UV, H2O2, and UV/H2O2 treatments of WAS. The results showed that the degradation of the 6 EDCs during all the three oxidation processes fitted well with pseudo-first-order kinetics. Compared with the sole UV irradiation or H2O2 oxidation process, UV/H2O2 treatment was much more effective for both EDC degradation and WAS solubilization. Under their optimal conditions, the EDC degradation rate constants during UV/H2O2 oxidation were 45–197 times greater than those during UV irradiation and 11–53 times greater than those during H2O2 oxidation. High dosage of H2O2 and low pH were favorable for the degradation of EDCs. Under the conditions of pH = 3, UV wavelength = 253.7 nm, UV fluence rate = 0.069 mW cm−2, and H2O2 dosage = 0.5 mol L−1, the removal efficiencies of E1, E2, EE2, E3, BPA, and NP in 2 min were 97%, 92%, 95%, 94%, 89%, and 67%, respectively. The hydroxyl radical (•OH) was proved to take the most important role for the removal of EDCs. Metal ions in sludge could facilitate the removal of EDCs during UV/H2O2 oxidation. Fe, Ag, and Cu ions had more obvious effects compared with other metal ions. The overall role of HA was dependent on the balance between its competition as organics and its catalysis/photosensitization effects. These indicate that the sludge matrix plays an important role in the degradation of EDCs. © 2014 Elsevier B.V. All rights reserved.

⁎ Corresponding author. Tel.: +86 21 65982692; fax: +86 21 65986313. E-mail address: [email protected] (Y. Li).

http://dx.doi.org/10.1016/j.scitotenv.2014.05.149 0048-9697/© 2014 Elsevier B.V. All rights reserved.

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1. Introduction Endocrine disrupting compounds (EDCs) are chemicals that have negative effects on the endocrine systems of humans and wildlife. EDCs include natural estrogens, natural androgens, artificial synthetic estrogens, phytoestrogens, and other industrial compounds (Liu et al., 2009). Among EDCs, the presence of steroid estrogens (such as estrone (E1), 17β-estradiol (E2), estriol (E3), and 17α-ethinylestradiol (EE2)) and phenolic xenoestrogens (such as 4-nonylphenols (4-NP) and bisphenol A (BPA)) deserves particular attention, because the former possesses the highest estrogenicity and the latter has moderate estrogenic potency and widespread applications. Their adverse effects on the reproductive functions of aquatic species and influence on humans have been observed worldwide (Oguz and Kankaya, 2013; Sales et al., 2013). EDCs present in sewage may be transferred to activated sludge in biological treatment systems due to their hydrophobic properties (Li et al., 2011). Ternes et al. (2002) reported that in the activated sludge and digested sludge, concentrations of E1, E2, and EE2 were in the ranges of b2–37 ng g−1 dw, 5–49 ng g−1 dw, and b 2–17 ng g−1 dw, respectively. Nieto et al. (2008) indicated that estrogen concentrations in the sludge exceeded 50–100 ng g−1 dw for E1 and 272–406 ng g−1 dw for E3. Gatidou et al. (2007) reported that BPA level detected in the sludge was 620 ng g−1 dw. Fromme et al. (2002) indicated that BPA concentrations in the sludge were in the range of 4–1363 ng g−1 dw. Reported concentrations of 4-NP in biosolids ranged from a few ug g − 1 dw up to several thousand ug g − 1 dw (La Guardia et al., 2001; Xia et al., 2005). These concentrations are high enough to pose risks to the environment and human health (Muller et al., 2010). The presence of EDCs in the sludge causes further concerns for sludge management (Belgiorno et al., 2007). Lind et al. (2010) indicated that ewes grazing pasture fertilized with sewage sludge exhibited an antiestrogenic effect on their trabecular bone in the form of reduced mineral content and density. Rhind et al. (2011) further reported that liver concentrations of many EDCs in ewes increased significantly and consistently with their increasing duration of exposure to sewage sludge-treated pastures. Therefore, reducing the impact of EDCs in sludge treatment systems is of considerable environmental relevance. Although EDCs have been detected in waste activated sludge (WAS), the occurrence and fate of EDCs during sludge stabilization process are poorly documented (Hamid and Eskicioglu, 2012). This lack of information is largely due to analytical limitations because of the complexity of the WAS matrix. Only a few published works on the mobility and transport of EDCs show that effective removal of EDCs is not achieved during traditional sludge stabilization processes. Andersen et al. (2003) measured E1, E2, and EE2 in specific wastewater treatment plant (WWTP) processes and found a notable increase in the concentrations of E1 and E2 during sludge stabilization. Czajka and Londry (2006) monitored the concentration of EE2 over a 271-day period in the anaerobic digester of a municipal WWTP and suggested that EE2 degradation was minimal even though electron acceptors had been reduced and methanogenesis occurred. Muller et al. (2010) confirmed that full-scale anaerobic digestion did not remove estrogens from the sludge efficiently. They further indicated that final sludge stabilization and dewatering using a thermal-pressurized treatment tended to increase the estrogen content from anaerobically digested sludge to dewatered sludge. Kang et al. (2006) reviewed the biodegradation of BPA by various organisms (bacteria, fungi, planktons, plants and animals) and mentioned that some BPA metabolites could enhance estrogenicity or toxicity. Barnabe et al. (2009) reported that toxic breakdown products resulting from NP biodegradation were of a concern as they exhibited endocrine disrupting properties. All these show that alternative methods are needed to remove the EDCs from sludge prior to their release into receiving media. In recent years, advanced oxidation processes (AOPs) have drawn particular attention for the degradation of emerging micro-pollutants

like EDCs, pharmaceuticals, and personal care products in various aqueous matrices (Klavarioti et al., 2009). Some investigations have been devoted to the ability of AOPs to pre-treat WAS and improve the sludge solubilization (Tokumura et al., 2007, 2009; Erden and Filibeli, 2010; Salihoglu et al., 2012). However, only a few investigations into the ability of AOPs to remove EDCs from sludge have been conducted (Hamid and Eskicioglu, 2012). Bernal-Martinez et al. (2007) used ozone to pre-treat sludge for the removal of polycyclic aromatic hydrocarbon (PAH) and noted an increase in biodegradability of PAH. Carballa et al. (2007) studied the effect of ozone on the removal of E1, E2, and EE2 from sludge and noted high removal efficiencies for natural estrogens. AK et al. (2013) achieved enhancement of both EDC degradation and biogas production during anaerobic sludge digestion by ozone pre-treated feed sludge. Kitis et al. (1999) used Fenton oxidation to remove nonylphenol ethoxylates (NPEs) in sludge and found that the biodegradability of NPEs increased with a higher oxidant dosage. Our previous research revealed that steroid estrogens including E1, E2, E3, and EE2 in WAS could be effectively removed using Fenton oxidation (Li and Zhang, 2014). Salihoglu et al. (2012) used ultraviolet light (UV) to remove PAH from municipal sludge and found that the total concentration of 12 kinds of PAH in the sludge decreased by 2–77% after 24 h of UV application. The combined UV/H2O2 process has been found to be very effective in the degradation of organic micro pollutants in water, because the photolysis of H2O2 generates effective oxidizing species of hydroxyl radicals. The oxidation potential of hydroxyl radical is 2.8 eV, which can completely destroy the pollutants. According to the research of Rosenfeldt and Linden (2004), the removal efficiency of EE2 in water by photolysis was increased from 20% to above 90% by addition of 15 mg L − 1 H2O2 . Zhang et al. (2007) indicated that 98% of E1 and E2 in water disappeared within 1 h under UV irradiation and the degradation rate increased with the addition of H2O2. Chen et al. (2007) also found that UV/H2O2 treatment at H2O2 dose of 10 mg L− 1 and UV fluence of b 1000 mJ cm− 2 was capable of decreasing in vitro estrogenicity of EDC mixture. Although the treatment of EDCs in the aquatic environment using UV/H2O2 has been well-documented, and UV or H2O2 oxidation has also been applied to pre-treat WAS and improve the sludge solubilization (Tokumura et al., 2007, 2009; Erden and Filibeli, 2010; Salihoglu et al., 2012), limited information can be obtained about the effect of UV/H2O2 on the behavior of EDCs in sludge. Unlike relatively simple aqueous matrices, WAS is a rather complex matrix with some species of organic compounds (e.g., humic acid (HA)) and inorganic substances (e.g., metals) in it, which may affect the degradation of micro-pollutants and cause different results from those in water. Niu et al. (2013) indicated that HA can influence the photodegradation of tetracycline. They found that tetracycline photolysis can be enhanced at low HA concentration under solar and xenon lamp irradiation, whereas it can be suppressed at high HA concentration. Sun et al. (1999) reported that transition metals can cause positive effects of accelerating the total degree of H2O2 oxidation reaction as well as the negative effects of increasing the rate of radical coupling reactions during oxidation. Therefore, it is of interest to explore the removal of EDCs during the treatment of WAS using UV/H2O2 and the effect of sludge matrix on the degradation of EDCs. The objective of the present study was to investigate the degradation of some phenolic EDCs (E1, E2, EE2, E3, BPA, and NP) in WAS during UV photolysis, H2O2 oxidation, and UV/H2O2 treatment processes. Degradation kinetic models were developed. The operational parameters influencing the reaction rates were investigated. The role of hydroxyl radical for the degradation of EDCs during UV/H 2 O 2 treatment of sludge was tested. Particularly, the effects of metal ions and HA contained in WAS on the degradation of the selected EDCs were also investigated.

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2. Materials and methods 2.1. Chemicals The six target EDCs were purchased from Sigma-Aldrich (N98% pure, USA). Stock EDC solutions were prepared by dissolving a relevant amount of EDC in methanol and stored at 4 °C. They were spiked into the sludge to produce the desired EDC concentrations. All organic solvents of high-performance liquid chromatography (HPLC) grade, including methanol, dichloromethane (DCM) and acetonitrile, were purchased from Sigma-Aldrich, USA. The water used in the HPLC analysis was purified using a Milli-Q system. Calcium(II) chloride, magnesium(II) sulfate, copper(II) sulfate, zinc(II) chloride, ferric(III) sulfate, ferrous(II) sulfate, silver(I) nitrate, cobalt(II) chloride, barium(II) chloride, manganese(II) sulfate, aluminium(III) chloride, and HA were purchased from Fisher Scientific Co. (N98% pure, UK). HF and HNO3 of super pure grade were purchased from Merck, Germany. Other chemicals including H2O2 (A.R., 30%, w/w) used in the experiments were of analytical grade and all were obtained from Sinopharm Chemical Reagent (Shanghai, China). 2.2. WAS The WAS was collected from the secondary sedimentation tank of a WWTP in Shanghai, China. An activated sludge process including anaerobic–aerobic units for nutrient removal is used in the WWTP. The influent flow rate of the WWTP is 7.5 m3/d, and the service population is 200,000; The sludge retention time (SRT) of the system is 8–10 d. The pH of the collected WAS was in the range of 6.2–6.8. The other main characteristics (average data plus standard deviations of triplicates) were as follows: total solids (TS) 14.5 ± 0.9 g L−1, volatile solids (VS) 11.3 ± 0.2 g L − 1 , total suspended solids (TSS) 13.8 ± 1.5 g L − 1, volatile suspended solids (VSS) 10.8 ± 0.3 g L− 1, soluble chemical oxygen demand (SCOD) 236 ± 86 mg L − 1 , and total chemical oxygen demand (TCOD) 14.8 ± 5.9 g L− 1. 2.3. Experimental set-up If not stated otherwise, the experimental conditions were as follows: the initial EDC concentration in sludge was 0.15 mg g−1 dw, the H2O2 dosage was 0.018 mol L−1, the pH was 7, and the irradiation time was 40 min. Soluble total organic carbon (STOC), TSS, and VSS were analyzed at the beginning and finishing of the treatment processes. All samples were measured in triplicate. The standard deviations of all measurements were less than 10%. A pseudo-first-order kinetic model as shown in Eq. (1) was applied to investigate the degradation kinetics of EDCs during UV irradiation, H2O2 oxidation, and UV/H2O2 oxidation: −kt

C ¼ C0 e

ð1Þ

where C0 and C are the EDC concentrations before treatment (time zero) and during treatment at time t (min), respectively; k is the rate constant (min−1). 2.3.1. UV irradiation The experiments for photo-degradation of EDCs were carried out in a 800 mL UV irradiation reactor equipped with a low pressure Hg lamp (75 W, emission at 253.7 nm, Shanghai ATRA Co., China) (Fig. 1). The lamp was protected by a standard 1 mm quartz sleeve. The UV fluence rate was 0.069 mW cm−2. Homogenization was achieved by continuous recirculation of the sludge using a centrifugal pump. UV dose was calculated by multiplying the UV fluence rate and the irradiation time. 800 mL of sludge spiked with EDC was irradiated under UV light and 20 mL of samples were taken at the preset sampling times with sodium thiosulfate (Na2S2O3) as the terminator. Experiments evaluating the

Fig. 1. Schematic diagram of the reactor used in UV and UV/H2O2 oxidation experiments. 1) power supply, 2) quartz sleeve with UV lamp inside, 3) UV lamp, 4) water outlet, 5) UV sensor, 6) water inlet, 7) peristaltic pump, 8) flexible pipe, 9) shield.

effects of initial EDC concentrations and initial pH values were conducted. Three initial concentrations were tested for each target EDC: 0.86, 0.40, and 0.21 mg g − 1 dw for E1; 0.21, 0.15, and 0.07 mg g − 1 dw for E2; 0.59, 0.31, and 0.17 mg g − 1 dw for EE2; 0.26, 0.15, and 0.08 mg g−1 dw for E3; 0.26, 0.13, and 0.07 mg g−1 dw for BPA; 0.31, 0.14, and 0.08 mg g−1 dw for NP. Five pH values (3, 5, 7, 9, 11) were tested while the initial concentrations of E1, E2, EE2, E3, BPA, and NP were 0.125, 0.035, 0.150, 0.080, 0.070, and 0.080 mg g− 1 dw, respectively. STOC, TSS, and VSS were analyzed at pH 7. 2.3.2. H2O2 oxidation The experiments for EDC degradation using H2O2 oxidation were carried out in 1 L of flasks stirred with magnetic stirrers at room temperature. 800 mL of sludge samples containing target EDCs was transferred to the flasks. The initial pH was first adjusted using NaOH and HNO3 as appropriate. Then H2O2 was immediately added. The reaction was allowed to proceed for a specific length of time. 20 mL of samples was taken at preset sampling times with Na2S2O3 as the terminator. H2O2 solutions were freshly prepared immediately prior to use. Experiments evaluating the effects of initial pH values and H2O2 dosages were conducted. Since H2O2 decomposes under alkaline conditions, 3 pH values (3, 5, 7) were tested with a H2O2 concentration of 0.5 mol L−1; the H2O2 dosage ranged from 0.05 to 2 mol L−1 were carried out at pH values of 3, 5, and 7. STOC, TSS, and VSS were analyzed at experimental conditions of pH 3 and H2O2 dosage of 0.5 mol L−1. 2.3.3. UV/H2O2 oxidation The experiments for EDC degradation using UV/H2O2 oxidation were also carried out in the UV reactor shown in Fig. 1. The initial pH of the sludge was adjusted using NaOH and HNO3 as appropriate.

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H2O2 solutions were pre-added into the reactor. Then the sludge was injected and exposed to UV light as described in experiments using UV irradiation (Section 2.3.1). 20 mL of samples was taken at sampling times with Na2S2O3 as the terminator. The effects of H2O2 dosage and initial pH on the removal of target EDCs were tested. When the initial concentrations of E1, E2, EE2, E3, BPA, and NP were 0.125, 0.035, 0.150, 0.080, 0.070, and 0.080 mg g− 1 dw, respectively, H2O2 dosages of 0.01–0.5 mol L− 1 were investigated at a pH of 3, and three initial pH values (3, 5, and 7) were tested at H2O2 dosage of 0.5 mol L− 1. STOC, TSS, and VSS were analyzed under experimental conditions of pH 3 and H2O2 dosage of 0.5 mol L− 1. In order to verify the role of •OH radicals, 0.5 mol L−1 of t-BuOH was added at H2O2 dosage of 0.02 mol L−1 and a pH of 3. Changes of threedimensional excitation–emission matrix (EEM) of sludge supernatant during UV/H2 O2 oxidation were also observed. The effects of 11 metal ions on the removal of target EDCs during UV/H2O2 oxidation were tested by spiking different concentrations of metal ions (0.05–5 mmol L− 1) to EDC solutions prepared using distilled water at pH 3. Pre-experiments had been conducted to exclude metals that had little effect on degradation of EDCs at a concentration of 0.5 mmol L−1 (data not shown). The effects of HA on the removal of target EDCs during UV/H2O2 oxidation were also tested by spiking different concentrations of HA (1–10,000 mg L−1) to EDC solutions at pH 7. The initial EDC concentration was 5 mg L−1. The reactions lasted for 20 min.

2.4. Analytical methods 2.4.1. EDC analysis The sample preparation procedure of EDCs followed the method described by Zeng et al. (2009). The sludge samples were freeze-dried. The dried pellets were dissolved in 20 mL of methanol and acetate buffer solution (pH 5.0) mixture (90:10, v/v) and ultrasonicated for 30 min and subsequently separated by centrifugation. Repeat this process for three times. The supernatants were mixed and followed by solvent evaporation in a rotatory evaporator. After evaporation, the residue was dissolved in 300 mL of water and was then extracted with SPE according to Zhao et al. (2008). The Supelclean TM LC-18 cartridges (500 mg, 3 mL) were obtained from Supelco (USA). The cartridges were previously conditioned by methanol and deionized water. Water samples passed through the cartridges at a flow rate of 5 mL/min. The cartridges were then washed with 5 mL of a methanol–water (30:70, v/v) solution. After being vacuum-dried, cartridges were eluted with 2 × 5 mL methanol, and the eluates were evaporated to dryness under a gentle stream of nitrogen. Finally, the dried residues were dissolved with 1 mL of methanol. For detection of EDCs, an HPLC equipped with a reversed-phase C-18 column (4.6 mm × 250 mm, 5 μm, Agilent, U.S.) was used. The injection volume was 50 μL and the flow rate was 1 mL/min. The separation was performed under gradient elution conditions using (A) acetonitrile and (B) water. The detection of E1, E2, EE2, and E3 followed the methods described by Li et al. (2013). The solvent program used for E2, EE2, and E3 was as follows: initial conditions 70% B linearly reduced to 30% B over 12 min, then linearly decreased to 0% B over 0.3 min, and then kept isocratic for 10 min. An excitation wavelength of 228 nm and emission wavelength of 316 nm were used for fluorescence detection of E2, EE2, and E3. For E1, a wavelength of 200 nm was used for UV detection at a temperature of 40 °C. The solvent program used for E1 was kept at initial conditions of 50% B for 15 min. For BPA, the mobile phase was 50% A. The excitation and emission wavelengths were 230 and 290 nm, respectively. For 4-NP, the mobile phase was 80% A. The detection wavelengths were 225 nm for the excitation wavelength and 300 nm for the emission wavelength. Under the experimental conditions, the quantification limit of target compounds was 62.5 ng g− 1 dw. The recoveries

were in ranges of 85–92%, 90–101%, 83–94%, 83–93%, 89–99%, and 77–83% for E1, E2, EE2, E3, BPA, and NP, respectively. 2.4.2. EEM analysis EEM was developed for the sludge supernatant before and after UV/H2O2 oxidation using a luminescence spectrometry (FluoroMax-4, HORIBA Jobin Yvon Co., France). The EEM spectra were gathered with the scanning emission (Em) spectra from 200 to 550 nm at 5 nm increments by varying the excitation (Ex) wavelength from 200 to 500 nm at 5 nm increments. Excitation and emission slits were both maintained at 5 nm, and the scanning speed was set at 4800 nm/min for all the measurements. Under the same conditions, the fluorescence spectra for Milli-Q water were subtracted from all the spectra to reduce the background noise. 2.4.3. Metal ion analysis Metal ions were analyzed using an inductive coupled plasma emission spectrometer (ICP, Agilent 720ES) on a sub-set of freezedried WAS. The sludge powders (0.5 g) were totally digested using HF and HNO3 on a hot plate. Dissolved samples were dried out and then re-dissolved in ultrapure water obtained from a Milli-Q purifier system (Millipore Corp., Bedford, MA, USA). 2.4.4. Other analysis TSS and VSS were analyzed before the termination of reactions according to Standard Methods (APHA, 1999). STOC was measured using a TOC-V CPN analyzer (Shimadzu, Japan). TSS 0 , VSS 0 , and STOC0 were referred to as the parameters of WAS before oxidation. The extent of TSS and VSS solubilization was calculated according to Eqs. (2) and (3):

STSS ¼

TSS0 −TSS  100% TSS0

ð2Þ

SVSS ¼

VSS0 −VSS  100%: VSS0

ð3Þ

3. Results and discussion 3.1. Removal of EDCs in WAS using UV irradiation 3.1.1. Effect of initial EDC concentration Concentrations of the target EDCs decreased rapidly over time during UV irradiation (Fig. 2). At the initial concentrations of 0.1–0.2 mg g− 1 dw, about 64% of E1, 67% of E2, 49% of EE2, 54% of E3, 29% of BPA and 13% of NP were degraded after 2 h of irradiation; at the end of 5 h irradiation, about 81% of E1, 95% of E2, 73% of EE2, 72% of E3, 59% of BPA, and 34% of NP were degraded, suggesting that UV irradiation is an effective treatment method. The photo-degradation rates of the six target EDCs in sludge during UV irradiation were well described using pseudo-first-order kinetics (Table 1). The results show clearly that the initial concentration of each target EDC did not affect the degradation rate constants as the k values were almost the same at different initial EDC concentrations (Table 1). It is also noteworthy that the UV photolysis rate constants (k) of EE2 reported by Zhang et al (2010) is 5–10 times higher than the values obtained in this study. Zhang et al. (2007) also investigated the photo-degradation kinetics of E1 and E2 in water and the reported k values of E1 and E2 in their study are 2–7 and 1–4 times of the data in this study. The discrepancy between their results and ours is likely due to the use of aqueous solution as a medium in their researches and WAS as a medium in the present study. The complex matrix of sludge might compete with target compounds, resulting in the decrease of photolysis rate.

A. Zhang, Y. Li / Science of the Total Environment 493 (2014) 307–323

(b) 0.25 C0 = 0.86 mg/g dw

E2 Concentration (mg/g dw)

E1 Concentration (mg/g dw)

(a) 1.0 C0 = 0.40 mg/g dw

0.8

C0 = 0.21 mg/g dw

0.6 0.4 0.2 0.0

C0 = 0.21 mg/g dw C0 = 0.15 mg/g dw

0.20

C0 = 0.07 mg/g dw

0.15 0.10 0.05 0.00

0

50

100

150

200

250

300

0

50

100

Time (min)

E3 Concentration (mg/g dw)

EE2 Concentration (mg/g dw)

C0 = 0.59 mg/g dw C0 = 0.31 mg/g dw

250

300

C0 = 0.17 mg/g dw

0.4 0.3 0.2 0.1

0

50

100

150

200

250

C0 = 0.26 mg/g dw

0.25

C0 = 0.08 mg/g dw

0.20 0.15 0.10 0.05 0.00

300

C0 = 0.15 mg/g dw

0

50

100

Time (min)

150

200

250

300

Time (min)

(e)0.30

(f)

0.25

C0 = 0.13 mg/g dw C0 = 0.07 mg/g dw

0.20 0.15 0.10 0.05

0

50

100

150

200

250

C0 = 0.31 mg/g dw

0.35

C0 = 0.26 mg/g dw

NP Concentration (mg/g dw)

BPA Concentration (mg/g dw)

200

(d) 0.30

0.5

0.00

150

Time (min)

(c) 0.6

0.0

311

300

C0 = 0.14 mg/g dw

0.30

C0 = 0.08 mg/g dw

0.25 0.20 0.15 0.10 0.05 0.00

0

50

Time (min)

100

150

200

250

300

Time (min)

Fig. 2. Removal of the six target EDCs in WAS using UV irradiation with different initial EDC concentrations at pH 7.

3.1.2. Effect of initial pH The removal of EDCs at different pH values are shown in Fig. 3. E2 showed the highest removal efficiency (95%) among the six target EDCs, and there was almost no difference when pH is in the range of 5–11. The removal of other EDCs varied at different pH values. Their degradation rates were further described using pseudo-firstorder kinetic model (R 2 N 0.94), and the results indicate that the rate constants for other five target EDCs were at high values at pH values of 5–7, beyond which the k values started to decrease (Table 1). This suggests an optimum pH range of 5–7 for best performance. The reason that pH could influence EDC degradation rates can be explained as follows:

Firstly, pH affects the formation of •OH. In the presence of dissolved oxygen, water molecules can generate •OH, hydrated electrons (e−aqu), and H• under UV irradiation: −

þ

H2 O þ hv→eaqu þ H  þ  OH þ H :

ð4Þ

Hence in acidic solutions, the formation of •OH will be suppressed as shown in Eq. (4). Moreover, H+ will have a scavenging effect on •OH, causing the decrease in degradation rate. Secondly, pH influences the ionization of reactant compounds and products. The pollutant molecules of target EDCs are neutral in acidic medium and start to become

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Table 1 Pseudo-first-order rate constants for the degradation of six target compounds in sludge during UV irradiation, H2O2 oxidation, and UV/H2O2 oxidation under different experimental conditions. EDCs

H2O2

UV CEDC0 (mg g−1)

k (min−1)

R2

0.21 0.4 0.86

0.006 0.006 0.006

0.94 0.97 0.96

0.07 0.15 0.21

0.01 0.009 0.01

0.93 0.99 0.99

0.17 0.31 0.59

0.004 0.004 0.004

0.98 0.99 0.98

0.08 0.15 0.26

0.005 0.004 0.006

0.91 0.95 0.98

0.07 0.13 0.26

0.003 0.003 0.003

0.92 0.94 0.95

0.08 0.14 0.31

0.001 0.001 0.001

0.94 0.95 0.94

E1

E2

EE2

E3

BPA

NP

pH

k (min−1)

R2

3 5 7 9 11 3 5 7 9 11 3 5 7 9 11 3 5 7 9 11 3 5 7 9 11 3 5 7 9 11

0.0055 0.0053 0.0063 0.0053 0.0049 0.0043 0.01 0.0099 0.0109 0.0104 0.003 0.0037 0.0044 0.0024 0.0022 0.0045 0.0059 0.0057 0.0046 0.0033 0.0028 0.0033 0.0029 0.0022 0.0012 0.0008 0.001 0.0012 0.0006 0.0005

0.98 0.99 0.95 0.99 0.98 0.97 0.99 0.99 0.99 0.98 0.98 0.98 0.97 0.98 0.95 0.99 0.99 0.98 0.99 0.98 0.97 0.98 0.98 0.96 0.95 0.99 0.99 0.97 0.97 0.94

UV/H2O2

pH

k (min−1)

R2

CH2O2 (mol L−1)

k (min−1)

R2

3 5 7

0.016 0.011 0.002

0.93 0.98 0.99

0.01 0.05 0.2

0.025 0.156 0.282

0.99 0.96 0.97

3 5 7

– – –

– – –

0.01 0.05 0.2

0.176 0.391 0.637

0.97 0.99 0.98

3 5 7

0.014 0.008 0.005

0.94 0.96 0.99

0.01 0.05 0.2

0.078 0.211 0.463

0.94 0.9 0.96

3 5 7

0.014 0.008 0.006

0.93 0.98 0.9

0.01 0.05 0.2

0.093 0.389 0.73

0.99 0.99 0.93

3 5 7

0.017 0.009 0.006

0.98 0.95 0.95

0.01 0.05 0.2

0.048 0.166 0.571

0.92 0.96 0.96

3 5 7

0.005 0.004 0.003

0.96 0.96 0.96

0.01 0.05 0.2

0.003 0.018 0.059

0.99 0.94 0.92

negatively charged under alkaline conditions. As a result, alkaline condition is not favorable to the photo-degradation of target EDCs. Similar results were also obtained by Horikoshi et al. (2004) in the photo-degradation of BPA in aqueous solution.

1999; Muruganandham and Swaminathan, 2004). When H2O2 was used in excess, it could act as a scavenger and compete for •OH to inhibit oxidation of the target organic compound, as shown in Eqs. (7) and (8).

3.2. Removal of EDCs in WAS using H2O2 oxidation

H2 O2 þ OH→HO2  þH2 O:

ð7Þ

OH þ HO2  →O2 þ H2 O:

ð8Þ

3.2.1. Effects of H2O2 dosage and initial pH The removal of the target EDCs at different H2O2 dosages and initial pH values are shown in Fig. 4. For E1, EE2, E3, BPA, and NP, the removal efficiencies at different pH values increased with the increase of H2O2 dosage. Low pH value is good for the removal of target EDCs in the tested H2O2 dosage range (0.05–2 mol L−1). At pH 3 and H2O2 dosage of 2 mol L−1, the highest removal efficiencies of E1, EE2, E3, BPA, and NP were 73%, 58%, 58%, 54%, and 46%, respectively, after the reaction for 40 min. The pH value can influence the generation of •OH during H2O2 oxidation. The increase of pH can cause the formation of an inactive hy− droperoxide anion (HO− 2 ) due to H2O2 dissociation. The HO2 can act as an efficient scavenger of •OH radicals as presented in Eq. (5), which leads to the inhibition of degradation (Bouasla et al., 2010). —

—

HO2 þ OH→H2 O þ O2  :

ð5Þ

H2O2 became unstable and decomposed to oxygen when pH was high. It also lost its oxidation ability, as presented in Eq. (6) (Bouasla et al., 2010). 2H2 O2 →O2 þ 2H2 O:

ð6Þ

E2 showed a different removal trend from other EDCs. The removal efficiency decreased as H 2O2 dosage increased from 0.05 to 1 mol L− 1 at pH 3 and from 0.3 to 1 mol L− 1 at pH values of 5 and 7. Excess •OH would also reversely inhibit the oxidation rate (Ince,

In order to explain the different performances of E2 from the other target EDCs in sludge during H2O2 oxidation, the same experiments were conducted in water. As shown in Fig. S1, the removal efficiencies for most target EDCs in water decreased at H2 O2 dosage over 0.05 mol L− 1. The removal trend of E2 was in accordance with those of E1, EE2, E3, and BPA in water, and also in accordance with its removal trend in sludge. This observation indicates that the discrepancy of removal trends between E2 and the other target EDCs in sludge during H2O2 oxidation could be attributed to the influence of the sludge matrix. Different from water matrices, the complex matrix of WAS may compete oxidants with EDCs. Among the six target EDCs, E2 had the highest reaction rate according to the degradation kinetics during UV irradiation and UV/H2O2 oxidation processes (Table 1). The same results was also obtained in the research of Coleman et al. (2004), who indicated that the photocatalysis pseudo-first-order rate constants (min−1) of E2, E1, and EE2 in water were 0.106, 0.086, and 0.086, respectively; moreover, the addition of 5 mg L− 1 H2O2 to the UV system increases the rate of E2 destruction by 130 times and the rate of EE2 destruction by 28 times. Rosenfeldt and Linden (2004) also reported that the rate constants (M−1 s−1) of EDCs with •OH for BPA, EE2, and E2 in water were 1.02 ± 0.06 × 1010, 1.08 ± 0.23 × 1010, and 1.41 ± 0.33 × 1010, respectively. In all these findings, the rate constant of E2 is the highest. Therefore, E2 was less influenced by the sludge

A. Zhang, Y. Li / Science of the Total Environment 493 (2014) 307–323

(b)

80 60 pH = 3 pH = 5 pH = 7 pH = 9 pH = 11

40 20 0

0

50

100

150

200

250

E2 Removal efficiency (%)

E1 Removal efficiency (%)

(a) 100

100 80 60 40

pH = 3 pH = 5 pH = 7 pH = 9 pH = 11

20 0

300

0

50

100

Time (min)

80 60

E3 Removal efficiency (%)

EE2 Removal efficiency (%)

200

250

300

200

250

300

(d) 100 pH = 3 pH = 5 pH = 7 pH = 9 pH = 11

40 20

0

50

100

150

200

250

80 60 40 20 0

300

pH = 3 pH = 5 pH = 7 pH = 9 pH = 11

0

50

100

150

Time (min)

Time (min)

(f) 100

(e) 100 pH = 3 pH = 5 pH = 7 pH = 9 pH = 11

80

60

NP Removal efficiency (%)

BPA Removal efficiency (%)

150

Time (min)

(c) 100

0

313

40

20

pH = 3 pH = 5 pH = 7 pH = 9 pH = 11

80

60

40

20

0

0 0

50

100

150

200

250

300

Time (min)

0

50

100

150

200

250

300

Time (min)

Fig. 3. Removal of the six target EDCs in WAS using UV irradiation at different pH values.

matrix during H2O2 oxidation, resulting in a different removal trend and optimal H2O2 dosage from those of other five target EDCs. 3.2.2. EDC degradation kinetics The removal of the six target EDCs in WAS using H2O2 oxidation at different pH values are shown in Fig. 5. After the reaction for 40 min, the removal efficiencies at pH values of 3, 5, and 7 were 45%, 33%, and 6% for E1, 53%, 70%, and 82% for E2, 41%, 26%, and 17% for EE2, 40%, 26%, and 21% for E3, 48%, 29%, and 21% for BPA, 18%, 14%, and 10% for NP, respectively. The degradation of target EDCs except for E2 in sludge

during H2O2 oxidation was in accordance with pseudo-first-order kinetics, with correlation coefficients greater than 0.9 (Table 1). The degradation rate constants of target EDCs (except for E2) decreased with the increase of pH values from 3 to 7. The rate constants of E1, EE2, E3, BPA, and NP during H2O2 oxidation at pH 3 were calculated as 0.0163, 0.0141, 0.0137, 0.0171, and 0.0054 min−1, respectively, which are much higher than those obtained during UV irradiation. For E2, it was quickly removed to a low concentration in 5 min and then kept stable (Fig. 5). At lower H2O2 dose (0.05 mol L−1 at pH 3; 0.3 mol L−1 at pH values of 5 and 7), E2 removal efficiency even reached

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(b) 100

(a) 80

pH = 3 pH = 5 pH = 7

E2 Removal efficiency (%)

E1 Removal efficiency (%)

100

60 40 20 0 0.0

0.5

1.0

1.5

60 40 20 0 0.0

2.0

pH = 3 pH = 5 pH = 7

80

0.5

pH = 3 pH = 5 pH = 7

60 40 20 0 0.0

0.5

1.0

1.5

80

20

0.5

1.0

1.5

2.0

H2O2 Dosage (mol/L)

(f) 100 pH = 3 pH = 5 pH = 7

NP Removal efficiency (%)

BPA Removal efficiency (%)

pH = 3 pH = 5 pH = 7

40

0 0.0

2.0

(e) 100

60 40 20 0 0.0

2.0

60

H2O2 Dosage (mol/L)

80

1.5

(d) 100 E3 Removal efficiency (%)

EE2 Removal efficiency (%)

(c) 100 80

1.0

H2O2 Dosage (mol/L)

H2O2 Dosage (mol/L)

0.5

1.0

1.5

2.0

H2O2 Dosage (mol/L)

80

pH = 3 pH = 5 pH = 7

60 40 20 0 0.0

0.5

1.0

1.5

2.0

H2O2 Dosage (mol/L)

Fig. 4. Removal of the six target EDCs in WAS using H2O2 oxidation at different H2O2 dosages and pH values.

high values (96–98%) (Fig. 4). These indicate its easy transformation under H2O2 oxidation. Whether E2 was mineralized or just transformed to other compounds needs further investigation. 3.3. Removal of EDCs in WAS using UV/H2O2 oxidation 3.3.1. Effect of H2O2 dosage The removal efficiencies of target EDCs increased with the increase of H2O2 dosage (Fig. 6). The removal efficiencies of EDCs increased dramatically in the first 10 min. Further increasing the reaction time has no significant effect on EDC removal. The removal efficiencies ranged from 21% to 97% for E1, 80% to 95% for E2, 50% to 94% for EE2, 58% to 96% for E3, 35% to 94% for BPA, and 3% to 69% for NP as the

H2O2 dosage ranging from 0.01 to 0.5 mol L− 1. For E2, 0.2 mol L− 1 of H2O2 is enough to get the highest removal efficiency. The highest removal efficiencies of other target EDCs occurred at H2O2 dosage of 0.5 mol L−1. As shown in Table 1, good linear fits (R2 N 0.9) of experimental data indicated pseudo-first-order reaction kinetics of target EDC degradation during the early stage (the first 10 min) of UV/H2O2 oxidation at H2 O 2 dosage of 0.01–0.2 mol L − 1. The rate constants increased with the increase of H 2O 2 dosage. As shown in Fig. S2, UV/H2O2 oxidation was superior to sole UV irradiation or H2O2 oxidation in terms of both EDC removal efficiencies and reaction time. At pH 3 and H2O2 dosage of 0.2 mol L− 1, the rate constants of UV/H2O2 oxidation were 51.2, 64.4, 105.2, 128.1, 196.9, and 44.7 times greater than those of UV irradiation (pH 7) for E1, E2, EE2, E3, BPA, and NP,

A. Zhang, Y. Li / Science of the Total Environment 493 (2014) 307–323

(b) 100 pH = 3 pH = 5 pH = 7

80

E2 Removal efficiency (%)

E1 Removal efficiency (%)

(a) 100

60 40 20 0

0

10

20

30

80 60 40 20 0

40

pH = 3 pH = 5 pH = 7

0

10

pH = 3 pH = 5 pH = 7

80

40

60 40 20

0

10

20

30

60 40 20 0

40

pH = 3 pH = 5 pH = 7

80

0

10

Time (min)

20

30

40

Time (min)

(e) 100

(f) 100 pH = 3 pH = 5 pH = 7

80

NP Removal efficiency (%)

BPA Removal efficiency (%)

30

(d) 100 E3 Removal efficiency (%)

EE2 Removal efficiency (%)

(c) 100

60

40

20

0

20

Time (min)

Time (min)

0

315

pH = 3 pH = 5 pH = 7

80

60

40

20

0 0

10

20

30

40

Time (min)

0

10

20

30

40

Time (min)

Fig. 5. Removal of the six target EDCs in WAS using H2O2 oxidation at different pH values with a H2O2 concentration of 0.5 mol L−1.

respectively, and 17.3, 32.8, 53.3, 33.4, and 10.8 times greater than those of H2O2 oxidation (pH 3 and H2O2 dosage of 0.5 mol L−1) for E1, EE2, E3, BPA, and NP, respectively (Table 1). During UV/H2O2 oxidation, more •OH was generated rapidly than sole UV irradiation or H2O2 oxidation, which could be explained by Eq. (9) (Zhang et al., 2007). hv

H2 O2 → 2  OH:

ð9Þ

Coleman et al. (2004) showed that the addition of 5 mg L−1 of H2O2 to the UV system increased the rate of E2 destruction by 130 times and

the rate of EE2 destruction by 28 times in water. Chen et al. (2007) also reported that UV in combination with H2O2 significantly removed estrogenic activity in vitro and in vivo compared to direct photolysis in water. 3.3.2. Effect of initial pH The removal efficiencies of target EDCs during UV/H2O2 oxidation increased dramatically in the first 2 min at pH 3–7 (Fig. 7). Then the removal efficiencies kept stable or increased little. After the reaction for 2 min, the removal efficiencies at pH values of 3, 5, and 7 were 97%, 76%, and 72% for E1, 92%, 89%, and 89% for E2, 95%, 81%, and 80% for EE2, 94%, 82%, and 74% for E3, 89%, 84%, and 74% for BPA, 67%, 66%, and 62% for NP, respectively. Generally, lower pH is

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(b) 100

80 60 40 20

CH O = 0.01 mol/L 2

0

CH O = 0.05 mol/L

2

2

CH O = 0.2 mol/L 2

0

10

30

2

2

40

50

80 60 40 CH O = 0.01 mol/L

20

2

0

60

0

10

20

2

30

2

40

50

60

(d) 100 E3 Removal efficiency (%)

EE2 Removal efficiency (%)

2

Time (min)

(c) 100 80 60 40 20

CH O = 0.05 mol/L

CH O = 0.01 mol/L 2 2

CH O = 0.2 mol/L 2 2

0

10

20

30

2 2

80 60 40 CH O = 0.01 mol/L

20

2

CH O = 0.5 mol/L

2

CH O = 0.05 mol/L

2

CH O = 0.2 mol/L 2

2

2

CH O = 0.5 mol/L 2

2

2 2

40

50

0

60

0

10

20

Time (min)

30

40

50

60

Time (min)

(f) 100

80

60

40

CH O = 0.01 mol/L

20

2 2

CH O = 0.2 mol/L 2 2

CH O = 0.05 mol/L 2 2

NP Removal efficiency (%)

(e) 100 BPA Removal efficiency (%)

2

CH O = 0.5 mol/L

2

Time (min)

0

CH O = 0.05 mol/L

2

CH O = 0.2 mol/L 2

CH O = 0.5 mol/L

2

20

2

E2 Removal efficiency (%)

E1 Removal efficiency (%)

(a) 100

80

60 CH O = 0.01 mol/L 2 2

CH O = 0.2 mol/L

40

2 2

CH O = 0.05 mol/L 2 2

CH O = 0.5 mol/L 2 2

20

CH O = 0.5 mol/L 2 2

0

0 0

10

20

30

40

50

60

Time (min)

0

10

20

30

40

50

60

Time (min)

Fig. 6. Removal of the six target EDCs in WAS using UV/H2O2 oxidation at different H2O2 dosages (pH = 3).

beneficial to the removal of EDCs. At H2O2 dosage of 0.5 mol L− 1 and pH of 3, the removal efficiencies of target EDCs using UV/H2O2 oxidation for 2 min were much higher than those using sole H2O2 oxidation for 40 min. This indicated that UV/H 2 O2 oxidation was an efficient technology to remove EDCs in WAS. 3.4. Solubilization of WAS during the three oxidation processes Solubilization of WAS during the three oxidation processes was evaluated based on the changes of the sludge characteristics such as STOC, TSS, and VSS. As shown in Fig. 8, the STOC/STOC0, STSS, and SVSS

increased with the increase of reaction time during all the three oxidation processes. The STOC/STOC0, STSS, and SVSS observed during UV/H2O2 oxidation were greater than those during UV irradiation and H2O2 oxidation (Fig. 8). The STSS and SVSS at 60 min during UV/H2O2 oxidation were 28% and 29%, respectively, which were 50 and 17 times higher than those under UV irradiation, respectively. The STOC/STOC0 observed at 60 min was 2.9 during UV/H2O2 oxidation, which was 1.09 times greater than that during UV irradiation. Comparing with the sole UV photolysis, the addition of H2 O 2 greatly increases the production of •OH radicals. The formation of •OH can rupture cell walls and release intracellular material, causing

A. Zhang, Y. Li / Science of the Total Environment 493 (2014) 307–323

(b) 100 E2 Removal efficiency (%)

E1 Removal efficiency (%)

(a) 100 80 60 pH = 3 pH = 5 pH = 7

40 20

80 60 pH = 3 pH = 5 pH = 7

40 20 0

0 0

20

40

60

80

100

0

20

Time (min)

E3 Removal efficiency (%)

EE2 Removal efficiency (%)

60

80

100

(d) 100

80 60 pH = 3 pH = 5 pH = 7

40 20 0 0

20

40

60

80

80 60

20 0

100

pH = 3 pH = 5 pH = 7

40

0

20

Time (min)

40

60

80

100

Time (min)

(f) 100 NP Removal efficiency (%)

(e) 100 BPA Removal efficiency (%)

40

Time (min)

(c) 100

80

60

pH = 3 pH = 5 pH = 7

40

20

0

317

80

60 pH = 3 pH = 5 pH = 7

40

20

0 0

20

40

60

80

100

Time (min)

0

20

40

60

80

100

Time (min)

Fig. 7. Removal of the six target EDCs in WAS using UV/H2O2 oxidation at different pH values (H2O2 dosage = 0.5 mol L−1).

the decrease of TSS and VSS (Mohapatra et al., 2011). In our previous study using Fenton oxidation for sludge treatment (Li and Zhang, 2014), the TSS and VSS solubilization were 13% and 20%, respectively, after 60 min. With the improvement of the sludge solubilization during oxidation pretreatment, organic substances were transferred from the solid phase to the aqueous phase. This transformation resulted in increases of soluble protein and soluble carbohydrate in the aqueous phase (Mohapatra et al., 2010), which would accelerate further biodegradation. Therefore, taking both the EDC removal and sludge solubilization into consideration, the UV/H2O2 oxidation is recommended for sludge pre-treatment in comparison with sole UV irradiation or H2O2 oxidation.

3.5. Verification of the role of •OH radical The removal of target EDCs can be attributed to the absorption of UV radiation and the oxidation of •OH, so the main attributer degrading target EDCs during UV/H2O2 oxidation was tested. Without t-BuOH (a · OH scavenger), the removal efficiencies of E1, E2, EE2, E3, BPA, and NP were 52%, 81%, 61%, 84%, 59%, and 7%, respectively, whereas with 0.5 mol L−1 t-BuOH, they were only 18%, 29%, 21%, 27%, 14%, and 2%, respectively (Fig. 9). The addition of t-BuOH significantly decreased the removal efficiencies of target EDCs. Therefore, the •OH radical takes a dominant role during UV/H2O2 oxidation of EDCs in sludge.

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(a)

8

UV

H2O2

(b)35

UV/H2O2

7

H2O2

UV/H2O2

30

6

25

5

STSS (%)

STOC/STOC0

UV

4 3

20 15 10

2

5

1 0

0 60

100

160

280

400

520

Time (min)

(c) 35

UV

H2O2

60

100

160

280

400

520

Time (min) UV/H2O2

30

SVSS (%)

25 20 15 10 5 0 60

100

160

280

400

520

Time (min) Fig. 8. Effects of UV irradiation, H2O2 oxidation, and UV/H2O2 oxidation on the increase of solubilization expressed as (a) STOC, (b) TSS, and (c) VSS. (Experimental conditions: for UV oxidation, the pH value was 7; for H2O2 and UV/H2O2 oxidation, the pH value was 3 and H2O2 dosage was 0.5 mol L−1.).

3.6. Effects of sludge matrix on degradation of EDCs during UV/H2O2 oxidation 3.6.1. Effects of metal ions WAS is enriched in metal ions. Eleven metals with high concentrations in sludge were detected (Fig. S3). They were Fe, Ca, K, Na, Mg, Al,

100

Removal efficiency (%)

With 0.5 M scavenger Without scavenger

80

60

40

20

0 E1

E2

EE2

E3

BPA

NP

Target compounds Fig. 9. Comparison of the EDC removal efficiencies between without scavenger (t-BuOH) and with 0.5 M scavenger (t-BuOH) in sludge using UV/H2O2 oxidation.

Zn, Ba, Mn, Sr, and Cu. Among these ions, Ca, K, Na, and Mg ions were commonly detected in municipal wastewater; Fe and Al ions exist in WAS because they are usually used as flocculants for the removal of phosphorus. Ba ion is a naturally occurring alkaline earth metal found in surface and ground waters. Sr ion is widely distributed in water of China. Industrial pollution may be the main reason for the high contents of Mn, Cu, and Zn in sludge. As shown in Fig. 10, Ca and Zn ions almost had no effect on the removal of EDCs; Mg and Mn ions could improve the removal of EDCs, but the concentration of these ions had no significant effect on EDC removal efficiencies; Fe and Ba ions improved the removal of EDCs and with the increase of ion concentrations, the removal efficiencies of EDCs increased; for Cu, Al, and Co ions, with the increase of ion concentrations, the removal efficiencies of EDCs increased at low ion concentrations and then decreased at high ion concentrations; the removal efficiencies of EE2 and E3 increased with the addition of Ag+, but the ion concentration had no significant effect on their removal efficiencies; the Ag ion had positive effects on the removals of E1 and E2 at concentrations of 0.05, 1, and 5 mmol L−1, but adverse or little effects at ion concentrations of 0.1 and 0.5 mmol L−1. Compared with other metal ions, Fe, Ag, and Cu ions had more obvious effects. The photochemical behavior of metal ions presents an important contribution to the transformation of organic pollutants in aqueous environment. Zhou et al. (2004) observed that the photo-oxidation efficiencies were dependent on pH values and ferric/oxalate concentration ratios in water. Meng et al. (2006) indicated that oxidation of phenols catalyzed by a copper Schiff base complex in aqueous solution.

A. Zhang, Y. Li / Science of the Total Environment 493 (2014) 307–323

Sun et al. (1999) also reported that transition metal ions such as Mn2+, Cu2+, and Fe3+ may act as electron acceptors catalyzing radical formation. Data in this study is also in agreement with the speculation of Brown and Abbot (1995) that transition metals decompose H2O2 with the subsequent formation of a number of reactive species and these

E3 BPA

E2 E1

E3 BPA

EE2

EDCs Removal efficiencies (%)

EDCs Removal efficiencies (%)

species may in turn react with targets considerably faster than H2O2 itself. The positive impacts of metal ions on the degradation of target EDCs during UV/H2O2 oxidation could be explained by the fact that the presence of metal ions can participate in Fenton-type reactions that produce

(b) 100

(a) 100 80

60

40

20

0

EE2

80

60

40

20

0.05

0.1

0.5

1

5

0

E3 E2 EE2 BPA E1

80

60

40

20

E3 NP

(d)

0.1

0.5

1

5

E2 BPA

EE2 E1

100

EDCs Removal efficiencies (%)

(c) 100

0.05

Zn2+ Concentration (mmol/L)

Ca2+ Concentration (mmol/L)

EDCs Removal efficiencies (%)

E2 E1

0 0

0

80

60

40

20

0 0

0.05

0.1

0.5

1

5

0

Mg2+ Concentration (mmol/L) E3 NP

(e) 100

E2 BPA

EE2 E1

60

40

20

0 0.05

0.1

0.5

1

Fe3+ Concentration (mmol/L)

0.1

E3 NP

(f) 100

80

0

0.05

0.5

1

5

Mn2+ Concentration (mmol/L)

EDCs Removal efficiencies (%)

EDCs Removal efficiencies (%)

319

5

E2 BPA

EE2 E1

80

60

40

20

0 0

0.05

0.1

0.5

1

Fe2+ Concentration (mmol/L)

Fig. 10. Effects of 11 metal concentrations on the removal efficiencies of the six target EDCs during UV/H2O2 oxidation.

5

320

A. Zhang, Y. Li / Science of the Total Environment 493 (2014) 307–323

E3 NP

E2 BPA

EE2 E1

E3 NP

(h) 100 EDCs Removal efficiencies (%)

EDCs Removal efficiencies (%)

(g) 100 80

60

40

20

60

40

20

0 0

0.05

0.1

0.5

1

5

0

Ba2+ Concentration (mmol/L) E3 NP

E2 BPA

0.05

0.1

0.5

1

5

Cu2+ Concentration (mmol/L)

EE2 E1

(j)

100

EDCs Removal efficiencies (%)

(i) 100 EDCs Removal efficiencies (%)

EE2 E1

80

0

80

60

40

20

E3

E2

EE2

BPA

0.05

0.1

0.5

1

80

60

40

20

0

0 0

0.05

0.1

0.5

1

5

0

Al3+ Concentration (mmol/L)

(k) 100 EDCs Removal efficiencies (%)

E2 BPA

E3

E2

EE2

E1

0.05

0.1

0.5

1

5

Co2+ Concentration (mmol/L)

80

60

40

20

0 0

5

Ag+ Concentration (mmol/L) Fig. 10 (continued).

additional •OH radicals as shown in Eqs. (10) and (11), causing the increase in removal efficiencies (Marta, 1999).

ðnþ1Þþ

M

nþ

þ H2 O→M

1 − þ þ O2 þ e þ 2H 2

ð10Þ

nþ

M

ðnþ1Þþ

þ H2 O2 ↔M

−

þ HO  þHO

ð11Þ

where M represents the metals (such as Mg, Mn, Fe, Ba, Cu, Al, Co, and Ag). Generally, reactions in Eqs. (10) and (11) do not occur for all metal ions in the absence of H2O2, but the oxidation rate may be

A. Zhang, Y. Li / Science of the Total Environment 493 (2014) 307–323

influenced if external H2O2 was added (Marta, 1999). As a typical example, impacts of Fe3 + and Fe2 + on the degradation of target EDCs during UV/H2 O2 oxidation were always positive. It was assumed that Fe2 + ions may act as electron donors catalyzing radicals' formation as shown in Eqs. (12) and (13), resulting in the increase of EDC removal efficiencies. 3þ

þ H2 O→Fe

2þ

þ H2 O2 →Fe

Fe

Fe

2þ

1 − þ þ O2 þ e þ 2H 2

3þ

−

þ OH þ OH :

ð12Þ

ð13Þ

Additionally, EDCs may combine with particular metals and form EDC–metal complex (Lazzari et al, 2000). The assumption reaction is presented in Eq. (14): EDCs þ metal↔ EDC–metal complex þ EDCs þ metal:

ð14Þ

Then an attack of molecular oxygen on the EDC–metal complex with the formation of an oxygen complex may occur. An electron is transferred through the metal ion to oxygen and radicals are formed. However, there may be an interaction between metals (such as Cu, Al, and Co) and H2O2 or •OH at high metal ion concentrations. Excess metal ions can also react with H2O2 forming weaker HO2• radicals and trapping •OH radicals, as shown in Eqs. (15) and (16) (Zaki et al., 2013), resulting in the decrease of EDC removal efficiencies. ðnþ1Þþ

þ H2 O2 ↔M

ðnþ1Þþ

þ OH þ OH ↔M

M

M

nþ

−

þ

þ HO2  þH :

nþ

þ H2 O2 :

ð15Þ

ð16Þ

3.6.2. Effects of HA The EEM fluorescence spectroscopies of sludge supernatant samples before and after UV/H2O2 oxidation are shown in Fig. 11. Four EEM peaks were observed for the original sludge supernatant sample. These EEM peaks are associated with humic/fulvic acid-like, or soluble microbial

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byproduct-like organic compounds, according to a location of EEM peaks of many typical chemicals in wastewater or surface water (Chen et al., 2003). The peaks for Ex / Em = (220–250) / (330–380), Ex / Em = (250–280) / (290–380), Ex / Em ≥ 250 / (380–480), and Ex / Em = (220–250) / (380–480) represent aromatic protein, soluble microbial products, HA-like, and fulvic acid, respectively. As shown in Fig. 11, most of the humic/fulvic acid in the sludge supernatant appeared to be decomposed by UV/H2O2 oxidation. Therefore, the organic matters of the humic/fulvic acid fraction may play an important role in the degradation of target EDCs during UV/H2O2 oxidation. The effect of HA (1–10,000 mg L−1) on the removal of target EDCs is shown in Fig. 12. With an increase in HA concentration, the removal efficiencies of E2, EE2, E3, and NP increased gradually. However, for E1 and BPA, the removal efficiencies increased at first and then decreased at HA concentrations beyond 500 and 5000 mg L−1, respectively. Generally, HA might interact with target EDCs through catalytic reactions, sorption, photosensitization, and competition (Tan et al., 2013). In this study, sorption of target EDCs by HA was negligible (b5%) (Fig. S4). The positive influence of HA on the degradation of target EDCs during UV/H2O2 oxidation can be attributed to its catalysis or sensitization effects. Photosensitized reactions involving electronic energy transfer from triplet states of HA to organic molecules as well as photosensitized oxygenations via the singlet oxygen pathway have been widely described (Jiao et al., 2008; López-Peñalver et al., 2012). Niu et al. (2013) also reported that HA can promote tetracycline photolysis by acting as a photosensitizer through the generation of •OH and superoxide radicals. On the other hand, the degradation of target EDCs could decrease due to the competition with HA in WAS for UV energy or reactive oxygen species (López-Peñalver et al., 2012). Thus, the overall role of HA during UV/H2O2 oxidation depends on the balance between the two opposite effects. Li et al. (2013) have reported that the HA accounted for 13.2% of TS in the sludge of a WWTP in Beijing, China. According to their results, the concentration of HA in the sludge was about 500–2000 mg L−1 (TS at 4–16 g L− 1), thus the degradation of most EDCs in the sludge using UV/H2O2 oxidation will be accelerated by HA. 4. Conclusions The degradation of the 6 target EDCs in WAS during UV, H2O2, and UV/H2O2 oxidation processes fitted well with pseudo-first-order kinetics. The UV/H2O2 oxidation was much more effective for EDC degradation and WAS solubilization compared to sole UV irradiation or H2O2 oxidation in terms of EDC removal efficiencies and reaction time. High dosage of H2O2 and low pH were favorable for the degradation of

Fig. 11. The EEM fluorescence spectroscopies of sludge supernatant samples (a) before and (b) after UV/H2O2 oxidation.

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EDC Removal efficiency (%)

100

E3 E2 EE2 NP BPA E1

80

60

40

20

0 1

5

50

100

500 1000 2000 5000 10000

Humic acid concentration (mg L-1) Fig. 12. Effects of humic acid concentration on the removal efficiencies of the six target EDCs during UV/H2O2 oxidation.

EDCs. Under the following reaction conditions for UV/H2O2 treatment of WAS: pH = 3, UV wavelength = 253.7 nm, UV fluence rate = 0.069 mW cm−2, H2O2 dosage = 0.5 mol L−1, the removal efficiencies of E1, E2, EE2, E3, BPA, and NP at the reaction time of 2 min were 97%, 92%, 95%, 94%, 89%, and 67%, respectively; at the reaction time of 60 min, the STOC increased by 3 times, and the extents of TSS and VSS solubilization were close to 28% and 29%, respectively. The •OH oxidation rather than the UV irradiation was the dominant factor resulting in the degradation of EDCs during the UV/H2O2 treatment of WAS. The sludge matrix plays an important role in EDC degradation. Metal ions in the sludge could facilitate the removal of EDCs during UV/H2O2 oxidation. Fe, Ag, and Cu ions had more obvious effects compared with the other metal ions. The overall role of HA to EDC degradation during UV/H2O2 oxidation of WAS depends on the balance between its competition as organics and its catalysis/photosensitization effects. Acknowledgments This work was supported by the National Hi-Tech Research and Development Program of China (863) (2011AA060902). Appendix A. Supplementary data Supplementary data to this article can be found online at http://dx. doi.org/10.1016/j.scitotenv.2014.05.149. References AK MS, Muz M, Komesli OT, GÖkcay CF. Enhancement of bio-gas production and xenobiotics degradation during anaerobic sludge digestion by ozone treated feed sludge. Chem Eng J 2013;230:499–505. American Public Health Association (APHA). Standard methods for the examination of water and wastewater. 20th ed. Washington, DC: American Public Health Association (APHA); 1999. Andersen H, Siegrist H, Halling-Sorensen B, Ternes TA. Fate of estrogens in a municipal sewage treatment plant. Environ Sci Technol 2003;37:4021–6. Barnabe S, Brar SK, Tyagi RD, Beauchesne I, Surampalli RY. Pre-treatment and bioconversion of wastewater sludge to value-added products — fate of endocrine disrupting compounds. Sci Total Environ 2009;407:1471–88. Belgiorno V, Rizzo L, Fatta D, Della Rocca C, Lofrano G, Nikolaou A, et al. Review on endocrine disrupting–emerging compounds in urban wastewater: occurrence and removal by photocatalysis and ultrasonic irradiation for wastewater reuse. Desalination 2007;215:166–76. Bernal-Martinez A, Carrere H, Patureau D, Delgenes JP. Ozone pre-treatment as improver of PAH removal during anaerobic digestion of urban sludge. Chemosphere 2007;68: 1013–9. Bouasla C, Samar ME, Ismail F. Degradation of methyl violet 6B dye by the Fenton process. Desalination 2010;254:35–41.

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