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Synergetic pretreatment of waste activated sludge by Fe(II)–activated persulfate oxidation under mild temperature for enhanced dewaterability
Bioresource Technology 124 (2012) 29–36
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Synergetic pretreatment of waste activated sludge by Fe(II)–activated persulfate oxidation under mild temperature for enhanced dewaterability Guangyin Zhen a, Xueqin Lu b, Baoying Wang c, Youcai Zhao a,⇑, Xiaoli Chai a, Dongjie Niu a, Aihua Zhao a, Yuyou Li d, Yu Song c, Xianyan Cao c a
The State Key Laboratory of Pollution Control and Resource Reuse, School of Environmental Science and Engineering, Tongji University, Shanghai 200092, PR China School of Environmental and Chemical Engineering, Shanghai University, Shanghai 200444, PR China Shanghai Tongji Construction Co. Ltd., Shanghai 200092, PR China d Graduate School of Environmental Studies, Tohoku University, 6-6-06 Aza-Aoba, Aramaki, Aoba-ku, Sendai, Miyagi 980-8579, Japan b c
h i g h l i g h t s 2 oxidation pretreatment. oxidation in mild temperature. " 94–96% CST reduction efficiency was obtained within the first 5 min. " Tyrosine and tryptophan protein-like matters in EPS mainly affected the dewatering. " The treatment destroyed EPS, releasing EPS-bound water and interstitial water.
" No study carried out on the combined thermal-Fe(II)–S2O8 2
" Sludge dewaterability was enhanced by Fe(II)–S2O8
a r t i c l e
i n f o
Article history: Received 12 June 2012 Received in revised form 6 August 2012 Accepted 10 August 2012 Available online 19 August 2012 Keywords: Waste activated sludge Dewaterability Fe(II)–S2O82 oxidation Extracellular polymeric substances (EPS) Excitation–emission matrix (EEM)
a b s t r a c t The potential benefits of Fe(II)–activated persulfate (S2O82 ) oxidation under mild temperature in enhancing the dewaterability of waste activated sludge were investigated. Capillary suction time (CST) was used to characterize sludge dewatering. Zeta potential, particle size distribution, three-dimensional excitation–emission matrix (EEM) fluorescence spectroscopy, fourier-transformed infrared (FT-IR) spectroscopy and scanning electronic microscopy (SEM) were employed to explore influencing mechanisms. The results indicated that the dewaterability was deteriorated with single thermal treatment, but significantly enhanced in the presence of Fe(II)–S2O82 oxidation and further advanced together with thermal treatment. EEM and FT-IR analysis indicated that combined thermal and Fe(II)–S2O82 oxidation pretreatment led to degrading of tyrosine and tryptophan protein-like substances in extracellular polymeric substances (EPS) and cleavage of linkages in polymeric backbone. SEM images further revealed the rupture of sludge flocs at the colloidal scale, which contributed to the release of EPS-bound water and interstitial water trapped between flocs, and subsequent enhanced dewaterability. Ó 2012 Elsevier Ltd. All rights reserved.
1. Introduction Waste activated sludge is continuously produced from wastewater treatment plants. Dewatering is of major interest in sludge volume reduction, transport and ultimate disposal. Such sludge, however, is often difficult to be dewatered due to the strong hydrophilicity. Hence, in recent years increasing effort has been focused on the study and implementation of alternative dewatering approaches, such as the addition of calcined aluminum salts ⇑ Corresponding author. Address: The State Key Laboratory of Pollution Control and Resource Reuse, Tongji University, 1239 Siping Road, Shanghai 200092, PR China. Tel.: +86 21 6598 0609; fax: +86 21 6598 0041. E-mail address: [email protected] (Y. Zhao). 0960-8524/$ - see front matter Ó 2012 Elsevier Ltd. All rights reserved. http://dx.doi.org/10.1016/j.biortech.2012.08.039
(Zhen et al., 2011), ultrasonication (Feng et al., 2009), freezing and thawing (Diak et al., 2011), electrolysis (Yuan et al., 2011) and acid pretreatment (Devlin et al., 2011). The main goal of these methods was to disintegrate the floc structure of sludge and disrupt the extracellular polymeric substances (EPS), promoting the release of bound water. Nevertheless, none of them are able to efficiently dewater sludge as desired. As a consequence, for enhancing the dewaterability of waste activated sludge, it is highly anticipated to develop a cost-effective technology which can partly or completely substitute the traditional methods. Recent studies have been focused on employing thermal pretreatment for sludge dewatering, with the most common temperatures used ranging from 40 to 200 °C (Neyens et al., 2003a; Bougrier et al., 2006). Bougrier et al. (2008) pointed out a threshold
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temperature of 150 °C for sludge dewatering. The dewaterability could be deteriorated below the temperature but profoundly promoted above it. The study by Liu et al. (2012a) found that thermal treatment at 175 °C for 60 min effectively destroyed cell flocs, making more soluble sugar and protein available for biological degradation, while decreasing viscosity and improving the sludge dewatering performance. Unfortunately, thermal treatment generally consumes higher energy and even has a larger hazardous emission potential as the temperature is higher than 180 °C (Wilson and Novak, 2009), making the process expensive and increasing safety concerns. A couple of studies confirmed that the combined thermo-chemical treatment might act synergistically and give an advantage of lowering the temperature, thus providing an energy-efficient and cost-effective alternative for sludge dewatering. Guan et al. (2012) reported that CaCl2 solution in the low temperature range from 50 to 90 °C could strengthen the bridging between Ca2+ and the flocs and in apparent promote the sludge dewaterability. Abelleira et al. (2012) treated secondary sewage sludge with hydrogen peroxide at mild operating conditions, resulting in a tremendous reduction in the time-to-filter (TTF) and an improvement in the dewaterability. The combination of thermal treatment and H2SO4 was applied to dewater waste sludge (Neyens et al., 2003a), and the dry sludge content of dewatered sludge increased from 22.5% to roughly 70%. The dewaterability of sludge could also be remarkably improved by Ca(OH)2 under 100 °C for 60 min (Neyens et al., 2003b). In our previous study (Zhen et al., 2012a and Zhen et al., 2012b), a novel technique, e.g., Fe(II)–S2O82 oxidation, was firstly set up to condition waste activated sludge, and it was regarded as a simple and efficient method for improving the dewaterability with a higher than 88% capillary suction time (CST) reduction within 1 min. Fe(II)–S2O82 oxidation, through the generation of very reactive species such as sulfate radicals (SO4 ) with oxidation–reduction potentials of 2.6 V, rapidly degraded EPS, disrupted sludge flocs and consequently, strengthened the sludge dewatering characteristics. To our best knowledge, no previous work describing the enhanced dewaterability of waste activated sludge by the mild thermal treatment cooperated with Fe(II)–S2O82 oxidation has been reported today. In this study, the synergetic effectiveness of combined thermal and Fe(II)–S2O82 oxidation pretreatments for waste activated sludge dewatering were systematically elucidated. CST was characterized to evaluate the sludge dewaterability. Zeta potential and particle size distribution were measured to explain the changes observed in sludge conditioning process. Three-dimensional excitation–emission matrix (EEM) fluorescence spectroscopy, fourier-transformed infrared (FT-IR) spectroscopy and scanning electronic microscopy (SEM) were also performed to better understand the possible mechanisms of sludge dewatering. Increasing knowledge on sludge dewaterability will further serve as a theoretical basis for the development of innovative and environmentally friendly technologies addressing the sludge issues in sludge management field.
2. Methods 2.1. Waste activated sludge The waste activated sludge used in this investigation was collected directly from the end of a thickening tank of a local sewage treatment plant in Shanghai, China. The plant treats about 75,000 m3/d of wastewater (93% domestic and 7% industrial sewage) using anaerobic–anoxic–oxic process. The collected samples were transferred to the laboratory within 30 min after sampling and then stored at 4 °C after screened through a 4.0 mm sieve to
remove grit. All experiments should be completed within 48 h. Table 1 tabulates the main characteristics of such sludge. 2.2. Experimental procedure A series of bench-scale experiments were conducted in a thermal reactor equipped with a thermometer and electric stirrer with a sample volume of approximately 300 mL. In each test, the sludge sample was first preheated to an expected temperature (25, 40, 60 and 80 °C) in the reactor. Then the reaction was initiated immediately by adding an appropriate amount of persulfate and Fe(II) (K2S2O8, > 99.5%, FeSO47H2O, > 99.0%, Sinopharm Chemical Reagents Co., China) at a fixed S2O82 /Fe(II) molar ratio of 0.8 according to our previous study (Zhen et al., 2012a). The mixture was continuously homogenized at 200 rpm during the conditioning process to avoid temperature gradients. At predetermined time intervals, a 5.0 mL sample was withdrawn and tested for its dewaterability after cooled to 25 ± 1 °C. Up to four replicates of each experimental condition were performed with the samples analyzed in triplicate to ensure the reproducibility of results, and average values were then reported. 2.3. Analytical methods The CST was determined using a standard CST apparatus (Model 440, Fann, UK) equipped with a 0.535 cm inner diameter funnel and standard Whatman No. 17 chromatography grade paper. Water content, total suspended solids (TSS) and volatile suspended solids (VSS) were estimated following Gravimetric methods. pH and redox potential were tested by a pH meter (PHS-3C, Leici Co., China). Conductivity was measured by a conductivity meter (DDS-307, Leici Co. Ltd., China). The sludge viscosity was determined at the shear rate of 60 s 1 to keep the sludge in suspension for 5 min using a rotating torque cylinder (SNB-1, Nirun Intelligent Technology Co., China).The measurement of zeta-potential was carried out by a Malvern Zetasizer Nano ZS (Malvern Instruments Ltd. Co., UK). Particle size distribution was assessed by an EyeTech Particle Size and Shape Analyzer (Ankersmid Ltd. Co., Netherlands) with a measurement range from 0.1–300.0 lm. The EPS of sludge were extracted according to a revised thermal treatment method by Wang et al. (2010). The sludge concentration was firstly adjusted to about 5.0 g/L after it was taken from the reactor. 50 mL sludge was centrifuged (4000 rpm, 5 min) at 4 °C, in order to separate the soluble EPS in the sludge supernatant. The sludge pellets left in the centrifuge tube were then washed twice with ultrapure water. The granular sludge was grounded to a powder with particles less than 0.18 mm. The sludge was brought up to volume with ultra-pure water and was placed in a water bath at 80 °C for 30 min, then centrifuged at 5000 rpm for 15 min. The organic matter in the supernatant was regarded as the bound EPS. The collected supernatant were filtered through a 0.45 lm cellulose nitrate membrane and then subjected to threedimensional EEM fluorescence spectra analysis. All the EEM fluorescence spectra were measured using a luminescence spectrometry (FluoroMax-4, HORIBA Jobin Yvon Co., France). The EPS EEM spectra were gathered with the scanning emission (Em) spectra from 250 to 550 nm at 5 nm increments by varying the excitation (Ex) wavelength from 250 to 400 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, fluorescence spectra for Milli-Q water was subtracted from all the spectra to eliminate water Raman scattering and to reduce other background noise. The software Origin 8.0 (Origin Lab Inc., USA) was employed to handle the EEM data. EEM spectra are plotted as the elliptical shape of contours.
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G. Zhen et al. / Bioresource Technology 124 (2012) 29–36 Table 1 Fundamental characteristics of waste activated sludge collected from the municipal WWTPa.
a
Water content (%)
pH
CST (s)
Viscosity (mPas)
TSS (g/L)
VSS (g/L)
Conductivity (lS/cm)
Redox potential (mV)
97.2 ± 0.1
5.9 ± 0.1
3006.1 ± 160.0
293.2 ± 19.0
26.4 ± 0.2
20.4 ± 0.2
1575.0 ± 7.1
33.3 ± 2.1
TSS: total suspended solids; VSS: volatile suspended solids.
A sample of 40 mL sludge was filtrated to remove most of water. The filter cake was dried at 105 °C for 24 h, then ground into powder, and dried again under the same conditions. FT-IR spectra were recorded using an ALPHA FT-IR spectrometer (Bruker Optics, Germany) in the range of 4000–400 cm 1. For SEM analysis, sludge sample was fixed with 2.5% (v/v) glutaraldehyde in 0.1 M phosphate buffer (pH 6.8) at 4 °C for 2 h and washed three times with phosphate buffer. After that, the fixed sludge was dehydrated via successive passages through 50, 70, 80 and 90% ethanol followed by freeze drying, then sputter coated with gold for SEM analysis (Hitachi S-520, Japan).
3. Results and discussion 3.1. Sludge dewaterability The sludge dewaterability in terms of CST depicted a definite difference under various pretreatment conditions (Fig. 1a). It is evident that CST varied a little (from 3006.1 ± 160.0 to 3119.2 ± 92.5 s) within 20 min at 25 °C. Whereas as temperature elevated to 60 and 80 °C, CST rapidly reached 4981.5 ± 202.7 and 7074.7 ± 631.9 s, increases of roughly 65.7 and 135.4%, respectively, indicative of a significant deterioration in the dewaterability. This observation correlated well with those of Bougrier et al. (2008), who used thermal treatment of 20–130 °C and found an increase in CST from 1300 to 2030 s, with those of Guan et al. (2012), who reported a 2.0 higher CST after 120 min of holding time at 80 °C, and also with those of Abelleira et al. (2012), who observed that the dewaterability of sludge was deteriorated when subjected to thermal treatments below 125 °C. This might be linked to the progressive cell lysis, releasing intracellular biopolymers (Audrey et al., 2011; Laurent et al., 2011), which decreased dewatering properties. In contrast, when in the presence of Fe(II)–S2O82 oxidation (i.e. S2O82 1.2 mmol/g-VSS and Fe(II) 1.5 mmol/g-VSS), CST apparently decreased from 3006.1 ± 160.0 to 174.2 ± 23.6, 136.9 ± 37.5, 103.3 ± 14.4 and 106.3 ± 12.8 s, decreases of 94.2, 95.4, 96.6 and 96.5% or so at 25, 40, 60 and 80 °C, respectively, within the first 5 min. Afterwards, a stable plateau with less than 40.0 s variation was observed during the following 15 min (Fig. 1a), implying that the dewaterability was improved effectively by Fe(II)–S2O82 oxidation within a relatively short time. The effects of Fe(II)–S2O82 dosages on the dewaterability at various temperature were systematically investigated and main results are illustrated in Fig. 1b. The CST decreased with S2O82 dosage increasing from 0 to 1.2 mmol/g-VSS, then leveled off at higher dosages. Overdosed Fe(II)–S2O82 produced less additional benefits to the dewatering. Therefore, the optimum S2O82 and Fe(II) dosages in the current study were supposed to be 1.2 mmol/g-VSS and 1.5 mmol/g-VSS, respectively. For temperature, as expected, the higher the operational temperature, the better the dewaterability was at a certain Fe(II)–S2O82 dosage. As depicted in the inset of Fig. 1b, when the temperature was raised from 25 to 40, 60 and 80 °C, the corresponding CST jumped from 188.6 ± 9.4 to 157.3 ± 2.1, 131.0 ± 8.6, and 111.5 ± 2.7 s, drops of roughly 16.6, 30.5 and 40.9%, respectively, at S2O82 1.2 mmol/g-VSS and Fe(II) 1.5 mmol/g-VSS. The similar results was also documented by Yang et al. (2010), who confirmed that the degradation rate of azo dye
Fig. 1. CST evolution of sludge as a function of pretreatment time (a) and Fe(II)– S2O82 dosages (b). (During the thermal-Fe(II)–S2O82 oxidation process, the S2O82 /Fe(II) molar ratio was fixed at 0.8. In figures 1–7, only S2O82 dosage was presented).
Acid Orange 7 (AO7) by persulfate increased with the temperature rising. It may be attributable to the fact that the persulfate can be initiated via metal catalyzer as well as heating. The elevated temperatures, in current study, thereby increased the decomposition of persulfate into SO4 radicals and enhanced Fe(II)–S2O82 oxidation process, contributing greatly to the degradation of EPS and cell lysis. As a result of this, the dewaterability of sludge was improved substantially. Obviously, the mild thermal treatment coupled with Fe(II)–S2O82 oxidation provides an alternative technique with a much lower operational temperature to promote the dewaterability, and therefore is more capable for sludge dewatering in the view of efficiency and economy. 3.2. The influence of zeta potential on the dewaterability The zeta potential is a critical factor in influencing sludge dewaterability. Fig. 2a presents the zeta potential evolution of sludge
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slightly and tended to the zero point of charge (0 mV) with further increasing Fe(II)-S2O82 dosages. According to the DLVO theories, the aggregation of sludge flocs is primarily governed by the negative surface charge. Increasing negative surface charge leads to increased electrostatic repulsion and lower interaction energy (Mikkelsen and Keiding, 2002), which in turn is expected to reduce the extent of flocculation and cause poor dewatering properties. Thus, with the decreasing surface charge related to elevating zeta potential, the sludge can aggregate, settle down quickly, and is dewatered more easily. This deduction is well supported by our current findings, in which the best dewaterability occurred at the zeta potential of 0.4 ± 0.1 mV. Liu et al. (2012b) also confirmed that the dewaterability of bioleached sludge was enhanced tremendously when zeta potential increased from 28 mV at the initial stage to close to 0 mV. 3.3. The influence of particle size distribution on the dewaterability
Fig. 2. Zeta potential evolution of sludge as a function of operational temperatures (a) and Fe(II)–S2O82 dosages (b) for 20 min.
flocs under differential operational temperatures. The surfaces of raw sludge were originally negatively charged with the zeta potential of 12.6 ± 1.4 mV at 25 °C. As the temperature increased to 80 °C, the zeta potential was found to sharply decrease to 18.0 ± 1.6 mV. The reduction in the zeta potential possibly originated from the cell disruption under thermal treatment, releasing biopolymers and some inorganic species as already observed by Laurent et al. (2009), Audrey et al. (2011) and Guan et al. (2012). The rise in the biopolymers could cause the increasing negative surface charge. Interestingly, the decline trend with the temperature increasing could be basically controlled by Fe(II)–S2O82 oxidation. The zeta potential showed no obvious shift with increasing temperature in the presence of S2O82 0.8 mmol/g-VSS and Fe(II) 1.0 mmol/g-VSS, while a marked rise was observed beyond S2O82 1.2 mmol/g-VSS and Fe(II) 1.5 mmol/g-VSS. The presence of Fe(II)–S2O82 oxidation altered the zeta potential evolution of the flocs through EPS degradation by SO4 radicals and charge neutralization of ferric and ferrous irons, and higher operational temperature could strengthen the impact. The zeta potential of sludge was significantly affected by Fe(II)– S2O82 dosages, and it shifted towards the less negative charge as Fe(II)–S2O82 dosages rose (Fig. 2b). The zeta potential lifted drastically from 12.6 ± 1.4 to 4.2 ± 0.2 mV at 25 °C, from 12.8 ± 1.4 to 2.8 ± 0.4 mV at 40 °C, from 14.5 ± 2.1 to 2.9 ± 0.4 mV at 60 °C and from 18.0 ± 1.6 to 0.4 ± 0.1 mV at 80 °C as the S2O82 dosage increased from 0 to 1.2 mmol/g-VSS, then elevated
The variations of particle size distribution of sludge flocs are given in Fig. 3 based on cumulative volume distribution. It is noticeable that the particle size distributions of sludge was not effected remarkably by Fe(II)–S2O82 oxidation at 25 °C and 50% of the volume (i.e. dp50) of sludge consisted of particles larger than110.4 ± 5.7 lm (Fig. 3a). For the sludge samples pretreated by single thermal treatment as shown in Fig. 3b, c and d, particle size distribution remained more or less constant whatever the applied operational temperatures. These results revealed that mild thermal treatment did not induce complete microflocs breakage, in accordance with the report by Laurent et al. (2009) who found that thermal treatment caused a progressive cell lysis when increasing temperature from 50–95 °C, whereas, the impact on floc structure was relatively limited. When in the presence of Fe(II)–S2O82 oxidation, however, thermal treatment presented a detectable effect on the granulometric distribution. The floc size distribution curves were shifted to the large size classes as Fe(II)–S2O82 dosages increased (Fig. 3b, c and d), with dp50 of sludge flocs higher than 119.4 ± 5.4 lm at 40 °C, than 124.2 ± 13.0 lm at 60 °C and than 117.6 ± 8.6 lm at 80 °C. These results could be explained by a re-flocculation phenomenon, already hypothesized by Bougrier et al. (2006) and Audrey et al. (2011). The re-flocculation might occur after the combined pretreatment process, mainly due to the release of intracellular and extracellular substances which led to the creation of chemical bonds. In addition, this re-flocculation was also strongly promoted by the coagulation effects of ferric and ferrous ions produced during Fe(II)–S2O82 oxidation process since they, acting as coagulants, could absorb onto the negatively charged sludge flocs and promoted agglomeration of the flocs through charge neutralization, absorption bridging and precipitation catching (Liu et al., 2012b). Decreased CST (inset of Fig. 1b) accompanied by increased dp50 of sludge size (Fig. 3b, c and d) was observed clearly, revealing that floc size is an important controlling factor with respect to sludge dewatering, consistent with those of Higgins and Novak (1997), who similarly observed that the ‘‘supracolloidal’’ flocs in the range of 1–100 lm have the greatest adverse influence on sludge dewaterability, and the dewaterability decreases as the quantities of flocs in this size range increase. The increased floc size could cause an exposure of fewer surfaces, weaken the hydrophilicity of sludge flocs and accordingly contribute to the apparent improvement in the dewaterability. 3.4. EEM fluorescence analysis EPS are regarded as one of the unfavorable elements in sludge dewatering (Feng et al., 2009). To evaluate the contribution of
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Fig. 3. Cumulative particle size distribution (based on volume) of sludge as a function of Fe(II)–S2O82 dosages at the operational temperatures of 25 (a), 40 (b), 60 (c) and 80 °C (d) for 20 min.
EPS to the dewaterability, many investigations have already been performed in literature (Liu et al., 2010; Ye et al., 2012). However, most previous experimental work has not specifically addressed the precise effects of EPS yet and the results were often inconsistent (Yuan et al., 2011; Zhen et al., 2012a; Zhen et al., 2012b). Three-dimensional EEM fluorescence spectroscopy is a potentially useful technique for rapidly determining the fluorescence compounds present in EPS mixtures in sludge flocs. Each EEM gave spectral information regarding the chemical compositions of EPS. Typical EEM fluorescence spectra of EPS and its corresponding fractions extracted from the raw and pretreated sludge are depicted in Fig. 4. Four main peaks could be readily identified from fluorescence spectra of EPS. The first peak (Peak A) was present at the excitation/emission wavelengths (Ex/Em) of 270–280/305– 310 nm in EEM spectra of EPS, which belonged to tyrosine protein-like substances based on the five regions of EEM spectra divided by Chen et al. (2003). The second peak (Peak B) located at the Ex/Em of around 280/350–360 nm was described as tryptophan protein-like substances (Yamashita and Tanoue, 2003; Chen et al., 2003). Compared to the fluorescence peak location of proteins (Ex/Em of 276–281/340–370 nm) described by Baker (2001), the locations of Peak B showed a blue shift along the emission axis. The third peak (Peak C) was found at the Ex/Em of 345/ 435–445 nm, while the fourth peak (Peak D) was observed at the Ex/Em of around 275/435–445 nm. Similar fluorescence signals have also been reported for natural dissolved organic matter (DOM) (Coble, 1996) and DOM from a membrane bioreactor (MBR) (Wang et al., 2009) as well as extracellular substances of waste activated sludge (Zhen et al., 2012b), and were associated with visible humic- and fulvic-like fluorescence (Coble, 1996; Chen
et al., 2003). In comparison with the fluorescence maxima of humic acids and fulvic acids (Ex/Em of 325/452, 320/443 nm) reported by Mobed et al. (1996), the locations of Peak C in current study were red-shifted in terms of excitation wavelengths. According to the research conducted by Wang et al. (2009), two fluorescence peaks, i.e. Peaks A and C, were found in EPS mixtures extracted from a MBR system. Sheng and Yu (2006) observed Peaks B and C together with a new peak at the Ex/Em of 225/340–350 nm related to the aromatic protein in EPS spectra of a conventional activated sludge system. Another research by Liu et al. (2011) indentified six fluorescence peaks including two new peaks at the Ex/Em of 225/300 and 225/340–350 nm, respectively. However, four main fluorescence peaks were observed in the EEM fluorescence spectra of the EPS samples in this study. The differences were presumably due to the fact that the EPS were harvested from the different sludge origins and thus the structure and components in EPS were chemically different. The fluorescence parameters of EEM spectra including peak locations and maximum fluorescence intensity are displayed in Figs. 4 and 5, respectively, which could be used for quantitative analysis. After Fe(II)–S2O82 oxidation at 25 °C for 20 min, the locations of Peaks C and D in soluble and bound EPS were both redshifted (0–5 nm) to longer wavelength than those of raw sludge EPS, while Peak B was blue-shifted to shorter wavelength only in bound EPS by 5 nm compared to that of raw sludge EPS (Fig. 4). The location shift of fluorescence peak provides spectral information on the chemical structure changes of the fluorescence components in EPS samples. A red shift is attributable to the increase of carbonyl, hydroxyl, alkoxyl, amino, and carbonyl groups in fluorophores (Chen et al., 2003; Liu et al., 2011), while a blue shift is as-
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Fig. 4. EEM fluorescence spectra of the soluble and bound EPS fractions from of the raw (a) and pretreated sludge (b) at 1.2 mmol S2O82 /g-VSS, 1.5 mmol Fe(II)/g-VSS, and 25 °C for 20 min.
cribed to an elimination of particular functional groups such as carbonyl, hydroxyl and amine, a reduction in the degree of p-electron systems, a decrease in the number of aromatic rings and conjugated bonds in a chain structure, or a conversion of a linear ring system to a non-linear system (Coble, 1996; Liu et al., 2011; Wang et al., 2009). Intensity reduction of the fluorescence peak before and after pretreatment is an indication for oxidation and removal of fluorescing materials. Fig. 5 exhibits the evolution in the intensities of fluorescence peaks with Fe(II)–S2O82 dosages under different temperatures. It is clear that as the temperature increased, more bound EPS, particularly tyrosine and tryptophan protein-like substances indicated respectively by Peaks A and B, were solubilized and converted into soluble EPS. For example, when pretreated by thermal treatment alone, the intensities of Peaks A and B in soluble EPS respectively increased about 215.0 and 896.8% as the temperature increased from 25 to 80 °C, while the corresponding intensities in bound EPS reduced 80.7 and 88.9%, respectively. In contrast, the intensities of Peaks C and D related to humic- and fulvic-like substances remained nearly constant, regardless of applied Fe(II)–S2O82 dosages and temperatures. These observations suggested that thermal treatment caused flocs disruption and cell lysis, releasing the extracellular and intracellular protein-like biopolymers, consistent with Audrey et al. (2011) and Laurent et al. (2009). The rise in soluble EPS could often lead to the increasing surface charge and poorer dewaterability (Zhen et al., 2012a). However in the presence of Fe(II)–S2O82 oxidation, to the opposite, better dewaterability was observed as already described in Fig. 1, mainly attributed to the degradation of fluorescence compo-
nents in EPS by active SO4 . As depicted in Fig. 5, the intensities of Peaks A and B in soluble EPS declined approximately 71.9 and 76.2% at 25 °C, 78.8 and 96.0% at 40 °C, 85.3 and 95.0% at 60 °C, 64.3 and 88.9% at 80 °C, respectively as the S2O82 dosage increased from 0 to 1.2 mmol/g-VSS. This finding implied that the sludge dewaterability was mainly governed by tyrosine and tryptophan protein-like fluorophores in EPS. Zhen et al. (2012b) found that sludge dewaterability was influenced together by aromatic protein-, tryptophan protein-, humic- and fulvic-like substances in EPS. The results observed in this study, however, further corroborated that in EPS, tyrosine and tryptophan protein-like substances rather than humic- and fulvic-like substances contributed more to the dewaterability. The results of our study coincided well with those of Zhen et al. (2012a), who showed that sludge dewaterability was highly related to soluble EPS, and a large amount of soluble EPS would cause poor performance in sludge dewatering, with those of Liu et al. (2011), who similarly observed a close relationship between protein-like substances of EPS and filtration resistance of membrane in MBRs, and also with those of Wang et al. (2009), who found that protein-like substances rose, the specific cake resistance increased, and this consequently resulted in the rise of trans-membrane pressure (TMP). Based on the aforesaid results, it is thus reasonable to address that the enhanced dewaterability by the combined thermal and Fe(II)–S2O82 oxidation pretreatment lies in the cleavage of linkages in the polymeric backbone (e.g. bound EPS) and microbial cells rupture, which induced the rise of tyrosine and tryptophan protein-like substances in soluble EPS, most of which could be degraded by SO4 . As a result of that, a large quantity of EPS-bound
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Fig. 5. The fluorescence intensities of peaks of the soluble (S) and bound (B) EPS fractions under the various S2O82 –Fe(II) dosages and operational temperatures for 20 min.
water and interstitial water was released into the free water, and eventually sludge dewaterability was promoted greatly.
supported by the analysis as shown for EEM spectra (Figs. 4 and 5), i.e., the intensities of protein-like biopolymers decreased with the increase of Fe(II)–S2O82 dosages.
3.5. FT-IR analysis Fig. S1 in Supplementary material illustrates the FT-IR spectra of sludge flocs previously subjected to various pretreatment at 25 °C, similar to those of Gulnaz et al. (2006), Laurent et al. (2011) and Pei et al. (2010). The strong band at 3299 cm 1 was attributed to the O–H stretching vibration. The bands at 2923 and 2852 cm 1 corresponded to the asymmetric and symmetric stretching vibrations of CH2 of aliphatic structures and lipids, respectively. The band at 1716 cm 1 was linked to the C@O vibration of carboxylic acids. The typical bands located at 1654 and 1540 cm 1 were assigned to the stretching and deformation vibrations of C@O, C–N (amide I) and N–H (amide II) peptidic bond of proteins. The band at 1457 cm 1 was the CH2 deformation vibration. The band present at 1418 cm 1 was related to the C@O stretching vibrations of carboxylates and OH deformation vibration of alcohols and phenols. The bands at 1037–1088 cm 1 was due to the C–O–C and C–O vibration of polysaccharides. While the bands at <1000 cm 1 were the ‘‘fingerprint’’ zone of phosphate or sulfur functional groups. EPS are one of the predominant components in sludge matrixes, mainly consisting of proteins (PN), polysaccharides (PS), smaller amounts of DNA, lipids, etc. Therefore, the nature of functional groups present in the flocs may predominantly originate from the EPS compositions (Laurent et al., 2011). The relative intensities of bands at 1654 and 1540 cm 1 in the six spectra gradually decreased following Fe(II)–S2O82 oxidation, indicating the degradation and removal of protein-like substances in EPS as proved by the decreased negative surface charge of flocs (Fig. 2). It was also
3.6. SEM analysis To get insight into the responsible dewatering mechanisms, SEM analysis of sludge with different pretreatment was performed. The differences in sludge appearance were profoundly significant (Fig. S2 in Supplementary material). The surface of un-pretreated sludge flocs was relatively round and smooth (Fig. S2a), while pretreated by Fe(II)–S2O82 oxidation at 25 °C, the flocs of waste activated sludge appeared to be much less regular, forming a structure with many pores (Fig. S2b). When the temperature increased to 80 °C, however, there were nearly no round sludge flocs to be detected (Fig. S2d), indicating the total collapse of the flocs. Notably, the combined thermal and Fe(II)–S2O82 oxidation pretreatment resulted in the disruption of sludge flocs by degrading EPS, also EPS bound-water and interstitial water trapped between the flocs was released into the free water, which enhanced the dewaterability of waste activated sludge. 4. Conclusions Fe(II)–S2O82 oxidation under mild temperature (25–80 °C) was demonstrated to be efficient in enhancing the dewaterability of sludge. Increasing particle size, and decreasing negative zeta potential and EPS content were mainly responsible for the observed changes in the dewaterability. EEM and FT-IR analysis showed that combined thermal and Fe(II)–S2O82 oxidation pretreatment led to degrading of tyrosine and tryptophan protein-like substances of
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EPS and cleavage of linkages in polymeric backbone. SEM images revealed the rupture of sludge flocs at the colloidal scale, inducing the release of EPS-bound water, also, interstitial water trapped between the flocs were released, which enhanced the dewaterability. Acknowledgements The authors wish to thank the Science and Technology Commission of Shanghai Municipality (No. 08 DZ 1202802, 09 DZ 1204105 and 09 DZ 2251700) and the Fundamental Research Funds for the Central Universities (No. 0400219152) for the partial support of this study. 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.2012.08. 039. References Abelleira, J., Pérez-Elvira, S.I., Sánchez-Oneto, J., Portela, J.R., Nebot, E., 2012. Advanced thermal hydrolysis of secondary sewage sludge: a novel process combining thermal hydrolysis and hydrogen peroxide addition. Resour. Conserv. Recycl. 59, 52–57. Audrey, P., Julien, L., Christophe, D., Patrick, L., 2011. Sludge disintegration during heat treatment at low temperature: a better understanding of involved mechanisms with a multiparametric approach. Biochem. Eng. J. 54 (3), 178– 184. Bougrier, C., Albasi, C., Delgenès, J.P., Carrère, H., 2006. Effect of ultrasonic, thermal and ozone pre-treatments on waste activated sludge solubilisation and anaerobic biodegradability. Chem. Eng. Process. 45, 711–718. Bougrier, C., Delgenes, J.P., Carrere, H., 2008. Effects of thermal treatments on five different waste activated sludge samples solubilisation, physical properties and anaerobic digestion. Chem. Eng. J. 139 (2), 236–244. Baker, A., 2001. Fluorescence excitation–emission matrix characterization of some sewage-impacted rivers. Environ. Sci. Technol. 35 (5), 948–953. Chen, W., Westerhoff, P., Leenheer, J.A., Booksh, K., 2003. Fluorescence excitation– emission matrix regional integration to quantify spectra for dissolved organic matter. Environ. Sci. Technol. 37, 5701–5710. Coble, P.G., 1996. Characterization of marine and terrestrial DOM in seawater using excitation–emission matrix spectroscopy. Mar. Chem. 51 (4), 325–346. Diak, J., Ormeci, B., Proux, C., 2011. Freeze-thaw treatment of RBC sludge from a remote mining exploration facility in subarctic Canada. Water Sci. Technol. 63 (6), 1309–1313. Devlin, D.C., Esteves, S.R.R., Dinsdale, R.M., Guwy, A.J., 2011. The effect of acid pretreatment on the anaerobic digestion and dewatering of waste activated sludge. Bioresour. Technol. 102, 4076–4082. Feng, X., Deng, J.C., Lei, H.Y., Bai, T., Fan, Q.J., Li, Z.X., 2009. Dewaterability of waste activated sludge with ultrasound conditioning. Bioresour. Technol. 100 (3), 1074–1081. Guan, B.H., Yu, J., Fu, H.L., Guo, M.H., Xu, X.H., 2012. Improvement of activated sludge dewaterability by mild thermal treatment in CaCl2 solution. Water Res. 46, 425–432. Gulnaz, O., Kaya, A., Dincer, S., 2006. The reuse of dried activated sludge for adsorption of reactive dye. J. Hazard. Mater. B134, 190–196. Higgins, M.J., Novak, J.T., 1997. The effect of cations on the settling and dewatering of activated sludges. Water Environ. Res. 6, 225–232.
Liu, X., Wang, W., Gao, X.B., Zhou, Y.J., Shen, R.J., 2012a. Effect of thermal pretreatment on the physical and chemical properties of municipal biomass waste. Waste Manage. (Oxford) 32, 249–255. Laurent, J., Casellas, M., Carrère, H., Dagot, C., 2011. Effects of thermal hydrolysis on activated sludge solubilization, surface properties and heavy metals biosorption. Chem. Eng. J. 166, 841–849. Laurent, J., Pierra, M., Casellas, M., Dagot, C., 2009. Fate of cadmium in activated sludge after changing its physico-chemical properties by thermal treatment. Chemosphere 77 (6), 771–777. Liu, F.W., Zhou, L.X., Zhou, J., Song, X.W., Wang, D.Z., 2012b. Improvement of sludge dewaterability and removal of sludge-borne metals by bioleaching at optimum pH. J. Hazard. Mater. 221–222 (30), 170–177. Liu, X.M., Sheng, G.P., Luo, H.W., Zhang, F., Yuan, S.J., Xu, J., Zeng, R.J., Wu, J.G., Yu, H.Q., 2010. Contribution of extracellular polymeric substances (EPS) to the sludge aggregation. Environ. Sci. Technol. 44, 4355–4360. Liu, T., Chen, Z.L., Yu, W.Z., You, S.J., 2011. Characterization of organic membrane foulants in a submerged membrane bioreactor with pre-ozonation using threedimensional excitation–emission matrix fluorescence spectroscopy. Water Res. 45, 2111–2121. Mikkelsen, L.H., Keiding, K., 2002. Physico-chemical characteristics of full scale sewage sludges with implications to dewatering. Water Res. 36, 2451–2462. Mobed, J.J., Hemmingsen, S.L., Autry, J.L., McGown, L.B., 1996. Fluorescence characterization of IHSS humic substances: total luminescence spectra with absorbance correction. Environ. Sci. Technol. 30 (10), 3061–3065. Neyens, E., Baeyens, J., Weemaes, M., De heyder, B., 2003a. Hot acid hydrolysis as a potential treatment of thickened sewage sludge. J. Hazard. Mater. 98 (1–3), 275–293. Neyens, E., Baeyens, J., Creemers, C., 2003b. Alkaline thermal sludge hydrolysis. J. Hazard. Mater. 97 (1–3), 295–314. Pei, H.Y., Hu, W.R., Liu, Q.H., 2010. Effect of protease and cellulose on the characteristic of activated sludge. J. Hazard. Mater. 178, 397–403. Sheng, G.P., Yu, H.Q., 2006. Characterization of extracellular polymeric substances of aerobic and anaerobic sludge using three-dimensional excitation and emission matrix fluorescence spectroscopy. Water Res. 40, 1233–1239. Wilson, C.A., Novak, J.T., 2009. Hydrolysis of macromolecular components of primary and secondary wastewater sludge by thermal hydrolytic pretreatment. Water Res. 43 (18), 4489–4498. Wang, X., Zhang, B., Shen, Z.Q., Qiu, Z.G., Chen, Z.L., Jin, M., Li, J.W., Wang, J.F., 2010. The EPS characteristics of sludge in an aerobic granule membrane bioreactor. Bioresour. Technol. 101 (21), 8046–8050. Wang, Z.W., Wu, Z.C., Tang, S.J., 2009. Characterization of dissolved organic matter in a submerged membrane bioreactor by using three-dimensional excitation and emission matrix fluorescence spectroscopy. Water Res. 43, 1533–1540. Yuan, H.P., Yan, X.F., Yang, C.F., Zhu, N.W., 2011. Enhancement of waste activated sludge dewaterability by electro-chemical pretreatment. J. Hazard. Mater. 187, 82–88. Yang, S.Y., Wang, P., Yang, X., Shan, L., Zhan, W.Y., Shao, X.T., Niu, R., 2010. Degradation efficiencies of azo dye Acid Orange 7 by the interaction of heat, UV and anions with common oxidants: persulfate, peroxymonosulfate and hydrogen peroxide. J. Hazard. Mater. 179, 552–558. Ye, F.X., Liu, X.W., Li, Y., 2012. Effects of potassium ferrate on extracellular polymeric substances (EPS) and physicochemical properties of excess activated sludge. J. Hazard. Mater. 199–200, 158–163. Yamashita, Y., Tanoue, E., 2003. Chemical characterization of protein-like fluorophores in DOM in relation to aromatic amino acids. Mar. Chem. 82, 255–271. Zhen, G.Y., Yan, X.F., Zhou, H.Y., Chen, H., Zhao, T.T., Zhao, Y.C., 2011. Effects of calcined aluminum salts on the advanced dewatering and solidification/ stabilization of sewage sludge. J. Environ. Sci. 23 (7), 1225–1232. 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.
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Bioresource Technology journal homepage: www.elsevier.com/locate/biortech
Synergetic pretreatment of waste activated sludge by Fe(II)–activated persulfate oxidation under mild temperature for enhanced dewaterability Guangyin Zhen a, Xueqin Lu b, Baoying Wang c, Youcai Zhao a,⇑, Xiaoli Chai a, Dongjie Niu a, Aihua Zhao a, Yuyou Li d, Yu Song c, Xianyan Cao c a
The State Key Laboratory of Pollution Control and Resource Reuse, School of Environmental Science and Engineering, Tongji University, Shanghai 200092, PR China School of Environmental and Chemical Engineering, Shanghai University, Shanghai 200444, PR China Shanghai Tongji Construction Co. Ltd., Shanghai 200092, PR China d Graduate School of Environmental Studies, Tohoku University, 6-6-06 Aza-Aoba, Aramaki, Aoba-ku, Sendai, Miyagi 980-8579, Japan b c
h i g h l i g h t s 2 oxidation pretreatment. oxidation in mild temperature. " 94–96% CST reduction efficiency was obtained within the first 5 min. " Tyrosine and tryptophan protein-like matters in EPS mainly affected the dewatering. " The treatment destroyed EPS, releasing EPS-bound water and interstitial water.
" No study carried out on the combined thermal-Fe(II)–S2O8 2
" Sludge dewaterability was enhanced by Fe(II)–S2O8
a r t i c l e
i n f o
Article history: Received 12 June 2012 Received in revised form 6 August 2012 Accepted 10 August 2012 Available online 19 August 2012 Keywords: Waste activated sludge Dewaterability Fe(II)–S2O82 oxidation Extracellular polymeric substances (EPS) Excitation–emission matrix (EEM)
a b s t r a c t The potential benefits of Fe(II)–activated persulfate (S2O82 ) oxidation under mild temperature in enhancing the dewaterability of waste activated sludge were investigated. Capillary suction time (CST) was used to characterize sludge dewatering. Zeta potential, particle size distribution, three-dimensional excitation–emission matrix (EEM) fluorescence spectroscopy, fourier-transformed infrared (FT-IR) spectroscopy and scanning electronic microscopy (SEM) were employed to explore influencing mechanisms. The results indicated that the dewaterability was deteriorated with single thermal treatment, but significantly enhanced in the presence of Fe(II)–S2O82 oxidation and further advanced together with thermal treatment. EEM and FT-IR analysis indicated that combined thermal and Fe(II)–S2O82 oxidation pretreatment led to degrading of tyrosine and tryptophan protein-like substances in extracellular polymeric substances (EPS) and cleavage of linkages in polymeric backbone. SEM images further revealed the rupture of sludge flocs at the colloidal scale, which contributed to the release of EPS-bound water and interstitial water trapped between flocs, and subsequent enhanced dewaterability. Ó 2012 Elsevier Ltd. All rights reserved.
1. Introduction Waste activated sludge is continuously produced from wastewater treatment plants. Dewatering is of major interest in sludge volume reduction, transport and ultimate disposal. Such sludge, however, is often difficult to be dewatered due to the strong hydrophilicity. Hence, in recent years increasing effort has been focused on the study and implementation of alternative dewatering approaches, such as the addition of calcined aluminum salts ⇑ Corresponding author. Address: The State Key Laboratory of Pollution Control and Resource Reuse, Tongji University, 1239 Siping Road, Shanghai 200092, PR China. Tel.: +86 21 6598 0609; fax: +86 21 6598 0041. E-mail address: [email protected] (Y. Zhao). 0960-8524/$ - see front matter Ó 2012 Elsevier Ltd. All rights reserved. http://dx.doi.org/10.1016/j.biortech.2012.08.039
(Zhen et al., 2011), ultrasonication (Feng et al., 2009), freezing and thawing (Diak et al., 2011), electrolysis (Yuan et al., 2011) and acid pretreatment (Devlin et al., 2011). The main goal of these methods was to disintegrate the floc structure of sludge and disrupt the extracellular polymeric substances (EPS), promoting the release of bound water. Nevertheless, none of them are able to efficiently dewater sludge as desired. As a consequence, for enhancing the dewaterability of waste activated sludge, it is highly anticipated to develop a cost-effective technology which can partly or completely substitute the traditional methods. Recent studies have been focused on employing thermal pretreatment for sludge dewatering, with the most common temperatures used ranging from 40 to 200 °C (Neyens et al., 2003a; Bougrier et al., 2006). Bougrier et al. (2008) pointed out a threshold
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temperature of 150 °C for sludge dewatering. The dewaterability could be deteriorated below the temperature but profoundly promoted above it. The study by Liu et al. (2012a) found that thermal treatment at 175 °C for 60 min effectively destroyed cell flocs, making more soluble sugar and protein available for biological degradation, while decreasing viscosity and improving the sludge dewatering performance. Unfortunately, thermal treatment generally consumes higher energy and even has a larger hazardous emission potential as the temperature is higher than 180 °C (Wilson and Novak, 2009), making the process expensive and increasing safety concerns. A couple of studies confirmed that the combined thermo-chemical treatment might act synergistically and give an advantage of lowering the temperature, thus providing an energy-efficient and cost-effective alternative for sludge dewatering. Guan et al. (2012) reported that CaCl2 solution in the low temperature range from 50 to 90 °C could strengthen the bridging between Ca2+ and the flocs and in apparent promote the sludge dewaterability. Abelleira et al. (2012) treated secondary sewage sludge with hydrogen peroxide at mild operating conditions, resulting in a tremendous reduction in the time-to-filter (TTF) and an improvement in the dewaterability. The combination of thermal treatment and H2SO4 was applied to dewater waste sludge (Neyens et al., 2003a), and the dry sludge content of dewatered sludge increased from 22.5% to roughly 70%. The dewaterability of sludge could also be remarkably improved by Ca(OH)2 under 100 °C for 60 min (Neyens et al., 2003b). In our previous study (Zhen et al., 2012a and Zhen et al., 2012b), a novel technique, e.g., Fe(II)–S2O82 oxidation, was firstly set up to condition waste activated sludge, and it was regarded as a simple and efficient method for improving the dewaterability with a higher than 88% capillary suction time (CST) reduction within 1 min. Fe(II)–S2O82 oxidation, through the generation of very reactive species such as sulfate radicals (SO4 ) with oxidation–reduction potentials of 2.6 V, rapidly degraded EPS, disrupted sludge flocs and consequently, strengthened the sludge dewatering characteristics. To our best knowledge, no previous work describing the enhanced dewaterability of waste activated sludge by the mild thermal treatment cooperated with Fe(II)–S2O82 oxidation has been reported today. In this study, the synergetic effectiveness of combined thermal and Fe(II)–S2O82 oxidation pretreatments for waste activated sludge dewatering were systematically elucidated. CST was characterized to evaluate the sludge dewaterability. Zeta potential and particle size distribution were measured to explain the changes observed in sludge conditioning process. Three-dimensional excitation–emission matrix (EEM) fluorescence spectroscopy, fourier-transformed infrared (FT-IR) spectroscopy and scanning electronic microscopy (SEM) were also performed to better understand the possible mechanisms of sludge dewatering. Increasing knowledge on sludge dewaterability will further serve as a theoretical basis for the development of innovative and environmentally friendly technologies addressing the sludge issues in sludge management field.
2. Methods 2.1. Waste activated sludge The waste activated sludge used in this investigation was collected directly from the end of a thickening tank of a local sewage treatment plant in Shanghai, China. The plant treats about 75,000 m3/d of wastewater (93% domestic and 7% industrial sewage) using anaerobic–anoxic–oxic process. The collected samples were transferred to the laboratory within 30 min after sampling and then stored at 4 °C after screened through a 4.0 mm sieve to
remove grit. All experiments should be completed within 48 h. Table 1 tabulates the main characteristics of such sludge. 2.2. Experimental procedure A series of bench-scale experiments were conducted in a thermal reactor equipped with a thermometer and electric stirrer with a sample volume of approximately 300 mL. In each test, the sludge sample was first preheated to an expected temperature (25, 40, 60 and 80 °C) in the reactor. Then the reaction was initiated immediately by adding an appropriate amount of persulfate and Fe(II) (K2S2O8, > 99.5%, FeSO47H2O, > 99.0%, Sinopharm Chemical Reagents Co., China) at a fixed S2O82 /Fe(II) molar ratio of 0.8 according to our previous study (Zhen et al., 2012a). The mixture was continuously homogenized at 200 rpm during the conditioning process to avoid temperature gradients. At predetermined time intervals, a 5.0 mL sample was withdrawn and tested for its dewaterability after cooled to 25 ± 1 °C. Up to four replicates of each experimental condition were performed with the samples analyzed in triplicate to ensure the reproducibility of results, and average values were then reported. 2.3. Analytical methods The CST was determined using a standard CST apparatus (Model 440, Fann, UK) equipped with a 0.535 cm inner diameter funnel and standard Whatman No. 17 chromatography grade paper. Water content, total suspended solids (TSS) and volatile suspended solids (VSS) were estimated following Gravimetric methods. pH and redox potential were tested by a pH meter (PHS-3C, Leici Co., China). Conductivity was measured by a conductivity meter (DDS-307, Leici Co. Ltd., China). The sludge viscosity was determined at the shear rate of 60 s 1 to keep the sludge in suspension for 5 min using a rotating torque cylinder (SNB-1, Nirun Intelligent Technology Co., China).The measurement of zeta-potential was carried out by a Malvern Zetasizer Nano ZS (Malvern Instruments Ltd. Co., UK). Particle size distribution was assessed by an EyeTech Particle Size and Shape Analyzer (Ankersmid Ltd. Co., Netherlands) with a measurement range from 0.1–300.0 lm. The EPS of sludge were extracted according to a revised thermal treatment method by Wang et al. (2010). The sludge concentration was firstly adjusted to about 5.0 g/L after it was taken from the reactor. 50 mL sludge was centrifuged (4000 rpm, 5 min) at 4 °C, in order to separate the soluble EPS in the sludge supernatant. The sludge pellets left in the centrifuge tube were then washed twice with ultrapure water. The granular sludge was grounded to a powder with particles less than 0.18 mm. The sludge was brought up to volume with ultra-pure water and was placed in a water bath at 80 °C for 30 min, then centrifuged at 5000 rpm for 15 min. The organic matter in the supernatant was regarded as the bound EPS. The collected supernatant were filtered through a 0.45 lm cellulose nitrate membrane and then subjected to threedimensional EEM fluorescence spectra analysis. All the EEM fluorescence spectra were measured using a luminescence spectrometry (FluoroMax-4, HORIBA Jobin Yvon Co., France). The EPS EEM spectra were gathered with the scanning emission (Em) spectra from 250 to 550 nm at 5 nm increments by varying the excitation (Ex) wavelength from 250 to 400 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, fluorescence spectra for Milli-Q water was subtracted from all the spectra to eliminate water Raman scattering and to reduce other background noise. The software Origin 8.0 (Origin Lab Inc., USA) was employed to handle the EEM data. EEM spectra are plotted as the elliptical shape of contours.
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G. Zhen et al. / Bioresource Technology 124 (2012) 29–36 Table 1 Fundamental characteristics of waste activated sludge collected from the municipal WWTPa.
a
Water content (%)
pH
CST (s)
Viscosity (mPas)
TSS (g/L)
VSS (g/L)
Conductivity (lS/cm)
Redox potential (mV)
97.2 ± 0.1
5.9 ± 0.1
3006.1 ± 160.0
293.2 ± 19.0
26.4 ± 0.2
20.4 ± 0.2
1575.0 ± 7.1
33.3 ± 2.1
TSS: total suspended solids; VSS: volatile suspended solids.
A sample of 40 mL sludge was filtrated to remove most of water. The filter cake was dried at 105 °C for 24 h, then ground into powder, and dried again under the same conditions. FT-IR spectra were recorded using an ALPHA FT-IR spectrometer (Bruker Optics, Germany) in the range of 4000–400 cm 1. For SEM analysis, sludge sample was fixed with 2.5% (v/v) glutaraldehyde in 0.1 M phosphate buffer (pH 6.8) at 4 °C for 2 h and washed three times with phosphate buffer. After that, the fixed sludge was dehydrated via successive passages through 50, 70, 80 and 90% ethanol followed by freeze drying, then sputter coated with gold for SEM analysis (Hitachi S-520, Japan).
3. Results and discussion 3.1. Sludge dewaterability The sludge dewaterability in terms of CST depicted a definite difference under various pretreatment conditions (Fig. 1a). It is evident that CST varied a little (from 3006.1 ± 160.0 to 3119.2 ± 92.5 s) within 20 min at 25 °C. Whereas as temperature elevated to 60 and 80 °C, CST rapidly reached 4981.5 ± 202.7 and 7074.7 ± 631.9 s, increases of roughly 65.7 and 135.4%, respectively, indicative of a significant deterioration in the dewaterability. This observation correlated well with those of Bougrier et al. (2008), who used thermal treatment of 20–130 °C and found an increase in CST from 1300 to 2030 s, with those of Guan et al. (2012), who reported a 2.0 higher CST after 120 min of holding time at 80 °C, and also with those of Abelleira et al. (2012), who observed that the dewaterability of sludge was deteriorated when subjected to thermal treatments below 125 °C. This might be linked to the progressive cell lysis, releasing intracellular biopolymers (Audrey et al., 2011; Laurent et al., 2011), which decreased dewatering properties. In contrast, when in the presence of Fe(II)–S2O82 oxidation (i.e. S2O82 1.2 mmol/g-VSS and Fe(II) 1.5 mmol/g-VSS), CST apparently decreased from 3006.1 ± 160.0 to 174.2 ± 23.6, 136.9 ± 37.5, 103.3 ± 14.4 and 106.3 ± 12.8 s, decreases of 94.2, 95.4, 96.6 and 96.5% or so at 25, 40, 60 and 80 °C, respectively, within the first 5 min. Afterwards, a stable plateau with less than 40.0 s variation was observed during the following 15 min (Fig. 1a), implying that the dewaterability was improved effectively by Fe(II)–S2O82 oxidation within a relatively short time. The effects of Fe(II)–S2O82 dosages on the dewaterability at various temperature were systematically investigated and main results are illustrated in Fig. 1b. The CST decreased with S2O82 dosage increasing from 0 to 1.2 mmol/g-VSS, then leveled off at higher dosages. Overdosed Fe(II)–S2O82 produced less additional benefits to the dewatering. Therefore, the optimum S2O82 and Fe(II) dosages in the current study were supposed to be 1.2 mmol/g-VSS and 1.5 mmol/g-VSS, respectively. For temperature, as expected, the higher the operational temperature, the better the dewaterability was at a certain Fe(II)–S2O82 dosage. As depicted in the inset of Fig. 1b, when the temperature was raised from 25 to 40, 60 and 80 °C, the corresponding CST jumped from 188.6 ± 9.4 to 157.3 ± 2.1, 131.0 ± 8.6, and 111.5 ± 2.7 s, drops of roughly 16.6, 30.5 and 40.9%, respectively, at S2O82 1.2 mmol/g-VSS and Fe(II) 1.5 mmol/g-VSS. The similar results was also documented by Yang et al. (2010), who confirmed that the degradation rate of azo dye
Fig. 1. CST evolution of sludge as a function of pretreatment time (a) and Fe(II)– S2O82 dosages (b). (During the thermal-Fe(II)–S2O82 oxidation process, the S2O82 /Fe(II) molar ratio was fixed at 0.8. In figures 1–7, only S2O82 dosage was presented).
Acid Orange 7 (AO7) by persulfate increased with the temperature rising. It may be attributable to the fact that the persulfate can be initiated via metal catalyzer as well as heating. The elevated temperatures, in current study, thereby increased the decomposition of persulfate into SO4 radicals and enhanced Fe(II)–S2O82 oxidation process, contributing greatly to the degradation of EPS and cell lysis. As a result of this, the dewaterability of sludge was improved substantially. Obviously, the mild thermal treatment coupled with Fe(II)–S2O82 oxidation provides an alternative technique with a much lower operational temperature to promote the dewaterability, and therefore is more capable for sludge dewatering in the view of efficiency and economy. 3.2. The influence of zeta potential on the dewaterability The zeta potential is a critical factor in influencing sludge dewaterability. Fig. 2a presents the zeta potential evolution of sludge
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slightly and tended to the zero point of charge (0 mV) with further increasing Fe(II)-S2O82 dosages. According to the DLVO theories, the aggregation of sludge flocs is primarily governed by the negative surface charge. Increasing negative surface charge leads to increased electrostatic repulsion and lower interaction energy (Mikkelsen and Keiding, 2002), which in turn is expected to reduce the extent of flocculation and cause poor dewatering properties. Thus, with the decreasing surface charge related to elevating zeta potential, the sludge can aggregate, settle down quickly, and is dewatered more easily. This deduction is well supported by our current findings, in which the best dewaterability occurred at the zeta potential of 0.4 ± 0.1 mV. Liu et al. (2012b) also confirmed that the dewaterability of bioleached sludge was enhanced tremendously when zeta potential increased from 28 mV at the initial stage to close to 0 mV. 3.3. The influence of particle size distribution on the dewaterability
Fig. 2. Zeta potential evolution of sludge as a function of operational temperatures (a) and Fe(II)–S2O82 dosages (b) for 20 min.
flocs under differential operational temperatures. The surfaces of raw sludge were originally negatively charged with the zeta potential of 12.6 ± 1.4 mV at 25 °C. As the temperature increased to 80 °C, the zeta potential was found to sharply decrease to 18.0 ± 1.6 mV. The reduction in the zeta potential possibly originated from the cell disruption under thermal treatment, releasing biopolymers and some inorganic species as already observed by Laurent et al. (2009), Audrey et al. (2011) and Guan et al. (2012). The rise in the biopolymers could cause the increasing negative surface charge. Interestingly, the decline trend with the temperature increasing could be basically controlled by Fe(II)–S2O82 oxidation. The zeta potential showed no obvious shift with increasing temperature in the presence of S2O82 0.8 mmol/g-VSS and Fe(II) 1.0 mmol/g-VSS, while a marked rise was observed beyond S2O82 1.2 mmol/g-VSS and Fe(II) 1.5 mmol/g-VSS. The presence of Fe(II)–S2O82 oxidation altered the zeta potential evolution of the flocs through EPS degradation by SO4 radicals and charge neutralization of ferric and ferrous irons, and higher operational temperature could strengthen the impact. The zeta potential of sludge was significantly affected by Fe(II)– S2O82 dosages, and it shifted towards the less negative charge as Fe(II)–S2O82 dosages rose (Fig. 2b). The zeta potential lifted drastically from 12.6 ± 1.4 to 4.2 ± 0.2 mV at 25 °C, from 12.8 ± 1.4 to 2.8 ± 0.4 mV at 40 °C, from 14.5 ± 2.1 to 2.9 ± 0.4 mV at 60 °C and from 18.0 ± 1.6 to 0.4 ± 0.1 mV at 80 °C as the S2O82 dosage increased from 0 to 1.2 mmol/g-VSS, then elevated
The variations of particle size distribution of sludge flocs are given in Fig. 3 based on cumulative volume distribution. It is noticeable that the particle size distributions of sludge was not effected remarkably by Fe(II)–S2O82 oxidation at 25 °C and 50% of the volume (i.e. dp50) of sludge consisted of particles larger than110.4 ± 5.7 lm (Fig. 3a). For the sludge samples pretreated by single thermal treatment as shown in Fig. 3b, c and d, particle size distribution remained more or less constant whatever the applied operational temperatures. These results revealed that mild thermal treatment did not induce complete microflocs breakage, in accordance with the report by Laurent et al. (2009) who found that thermal treatment caused a progressive cell lysis when increasing temperature from 50–95 °C, whereas, the impact on floc structure was relatively limited. When in the presence of Fe(II)–S2O82 oxidation, however, thermal treatment presented a detectable effect on the granulometric distribution. The floc size distribution curves were shifted to the large size classes as Fe(II)–S2O82 dosages increased (Fig. 3b, c and d), with dp50 of sludge flocs higher than 119.4 ± 5.4 lm at 40 °C, than 124.2 ± 13.0 lm at 60 °C and than 117.6 ± 8.6 lm at 80 °C. These results could be explained by a re-flocculation phenomenon, already hypothesized by Bougrier et al. (2006) and Audrey et al. (2011). The re-flocculation might occur after the combined pretreatment process, mainly due to the release of intracellular and extracellular substances which led to the creation of chemical bonds. In addition, this re-flocculation was also strongly promoted by the coagulation effects of ferric and ferrous ions produced during Fe(II)–S2O82 oxidation process since they, acting as coagulants, could absorb onto the negatively charged sludge flocs and promoted agglomeration of the flocs through charge neutralization, absorption bridging and precipitation catching (Liu et al., 2012b). Decreased CST (inset of Fig. 1b) accompanied by increased dp50 of sludge size (Fig. 3b, c and d) was observed clearly, revealing that floc size is an important controlling factor with respect to sludge dewatering, consistent with those of Higgins and Novak (1997), who similarly observed that the ‘‘supracolloidal’’ flocs in the range of 1–100 lm have the greatest adverse influence on sludge dewaterability, and the dewaterability decreases as the quantities of flocs in this size range increase. The increased floc size could cause an exposure of fewer surfaces, weaken the hydrophilicity of sludge flocs and accordingly contribute to the apparent improvement in the dewaterability. 3.4. EEM fluorescence analysis EPS are regarded as one of the unfavorable elements in sludge dewatering (Feng et al., 2009). To evaluate the contribution of
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Fig. 3. Cumulative particle size distribution (based on volume) of sludge as a function of Fe(II)–S2O82 dosages at the operational temperatures of 25 (a), 40 (b), 60 (c) and 80 °C (d) for 20 min.
EPS to the dewaterability, many investigations have already been performed in literature (Liu et al., 2010; Ye et al., 2012). However, most previous experimental work has not specifically addressed the precise effects of EPS yet and the results were often inconsistent (Yuan et al., 2011; Zhen et al., 2012a; Zhen et al., 2012b). Three-dimensional EEM fluorescence spectroscopy is a potentially useful technique for rapidly determining the fluorescence compounds present in EPS mixtures in sludge flocs. Each EEM gave spectral information regarding the chemical compositions of EPS. Typical EEM fluorescence spectra of EPS and its corresponding fractions extracted from the raw and pretreated sludge are depicted in Fig. 4. Four main peaks could be readily identified from fluorescence spectra of EPS. The first peak (Peak A) was present at the excitation/emission wavelengths (Ex/Em) of 270–280/305– 310 nm in EEM spectra of EPS, which belonged to tyrosine protein-like substances based on the five regions of EEM spectra divided by Chen et al. (2003). The second peak (Peak B) located at the Ex/Em of around 280/350–360 nm was described as tryptophan protein-like substances (Yamashita and Tanoue, 2003; Chen et al., 2003). Compared to the fluorescence peak location of proteins (Ex/Em of 276–281/340–370 nm) described by Baker (2001), the locations of Peak B showed a blue shift along the emission axis. The third peak (Peak C) was found at the Ex/Em of 345/ 435–445 nm, while the fourth peak (Peak D) was observed at the Ex/Em of around 275/435–445 nm. Similar fluorescence signals have also been reported for natural dissolved organic matter (DOM) (Coble, 1996) and DOM from a membrane bioreactor (MBR) (Wang et al., 2009) as well as extracellular substances of waste activated sludge (Zhen et al., 2012b), and were associated with visible humic- and fulvic-like fluorescence (Coble, 1996; Chen
et al., 2003). In comparison with the fluorescence maxima of humic acids and fulvic acids (Ex/Em of 325/452, 320/443 nm) reported by Mobed et al. (1996), the locations of Peak C in current study were red-shifted in terms of excitation wavelengths. According to the research conducted by Wang et al. (2009), two fluorescence peaks, i.e. Peaks A and C, were found in EPS mixtures extracted from a MBR system. Sheng and Yu (2006) observed Peaks B and C together with a new peak at the Ex/Em of 225/340–350 nm related to the aromatic protein in EPS spectra of a conventional activated sludge system. Another research by Liu et al. (2011) indentified six fluorescence peaks including two new peaks at the Ex/Em of 225/300 and 225/340–350 nm, respectively. However, four main fluorescence peaks were observed in the EEM fluorescence spectra of the EPS samples in this study. The differences were presumably due to the fact that the EPS were harvested from the different sludge origins and thus the structure and components in EPS were chemically different. The fluorescence parameters of EEM spectra including peak locations and maximum fluorescence intensity are displayed in Figs. 4 and 5, respectively, which could be used for quantitative analysis. After Fe(II)–S2O82 oxidation at 25 °C for 20 min, the locations of Peaks C and D in soluble and bound EPS were both redshifted (0–5 nm) to longer wavelength than those of raw sludge EPS, while Peak B was blue-shifted to shorter wavelength only in bound EPS by 5 nm compared to that of raw sludge EPS (Fig. 4). The location shift of fluorescence peak provides spectral information on the chemical structure changes of the fluorescence components in EPS samples. A red shift is attributable to the increase of carbonyl, hydroxyl, alkoxyl, amino, and carbonyl groups in fluorophores (Chen et al., 2003; Liu et al., 2011), while a blue shift is as-
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Fig. 4. EEM fluorescence spectra of the soluble and bound EPS fractions from of the raw (a) and pretreated sludge (b) at 1.2 mmol S2O82 /g-VSS, 1.5 mmol Fe(II)/g-VSS, and 25 °C for 20 min.
cribed to an elimination of particular functional groups such as carbonyl, hydroxyl and amine, a reduction in the degree of p-electron systems, a decrease in the number of aromatic rings and conjugated bonds in a chain structure, or a conversion of a linear ring system to a non-linear system (Coble, 1996; Liu et al., 2011; Wang et al., 2009). Intensity reduction of the fluorescence peak before and after pretreatment is an indication for oxidation and removal of fluorescing materials. Fig. 5 exhibits the evolution in the intensities of fluorescence peaks with Fe(II)–S2O82 dosages under different temperatures. It is clear that as the temperature increased, more bound EPS, particularly tyrosine and tryptophan protein-like substances indicated respectively by Peaks A and B, were solubilized and converted into soluble EPS. For example, when pretreated by thermal treatment alone, the intensities of Peaks A and B in soluble EPS respectively increased about 215.0 and 896.8% as the temperature increased from 25 to 80 °C, while the corresponding intensities in bound EPS reduced 80.7 and 88.9%, respectively. In contrast, the intensities of Peaks C and D related to humic- and fulvic-like substances remained nearly constant, regardless of applied Fe(II)–S2O82 dosages and temperatures. These observations suggested that thermal treatment caused flocs disruption and cell lysis, releasing the extracellular and intracellular protein-like biopolymers, consistent with Audrey et al. (2011) and Laurent et al. (2009). The rise in soluble EPS could often lead to the increasing surface charge and poorer dewaterability (Zhen et al., 2012a). However in the presence of Fe(II)–S2O82 oxidation, to the opposite, better dewaterability was observed as already described in Fig. 1, mainly attributed to the degradation of fluorescence compo-
nents in EPS by active SO4 . As depicted in Fig. 5, the intensities of Peaks A and B in soluble EPS declined approximately 71.9 and 76.2% at 25 °C, 78.8 and 96.0% at 40 °C, 85.3 and 95.0% at 60 °C, 64.3 and 88.9% at 80 °C, respectively as the S2O82 dosage increased from 0 to 1.2 mmol/g-VSS. This finding implied that the sludge dewaterability was mainly governed by tyrosine and tryptophan protein-like fluorophores in EPS. Zhen et al. (2012b) found that sludge dewaterability was influenced together by aromatic protein-, tryptophan protein-, humic- and fulvic-like substances in EPS. The results observed in this study, however, further corroborated that in EPS, tyrosine and tryptophan protein-like substances rather than humic- and fulvic-like substances contributed more to the dewaterability. The results of our study coincided well with those of Zhen et al. (2012a), who showed that sludge dewaterability was highly related to soluble EPS, and a large amount of soluble EPS would cause poor performance in sludge dewatering, with those of Liu et al. (2011), who similarly observed a close relationship between protein-like substances of EPS and filtration resistance of membrane in MBRs, and also with those of Wang et al. (2009), who found that protein-like substances rose, the specific cake resistance increased, and this consequently resulted in the rise of trans-membrane pressure (TMP). Based on the aforesaid results, it is thus reasonable to address that the enhanced dewaterability by the combined thermal and Fe(II)–S2O82 oxidation pretreatment lies in the cleavage of linkages in the polymeric backbone (e.g. bound EPS) and microbial cells rupture, which induced the rise of tyrosine and tryptophan protein-like substances in soluble EPS, most of which could be degraded by SO4 . As a result of that, a large quantity of EPS-bound
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Fig. 5. The fluorescence intensities of peaks of the soluble (S) and bound (B) EPS fractions under the various S2O82 –Fe(II) dosages and operational temperatures for 20 min.
water and interstitial water was released into the free water, and eventually sludge dewaterability was promoted greatly.
supported by the analysis as shown for EEM spectra (Figs. 4 and 5), i.e., the intensities of protein-like biopolymers decreased with the increase of Fe(II)–S2O82 dosages.
3.5. FT-IR analysis Fig. S1 in Supplementary material illustrates the FT-IR spectra of sludge flocs previously subjected to various pretreatment at 25 °C, similar to those of Gulnaz et al. (2006), Laurent et al. (2011) and Pei et al. (2010). The strong band at 3299 cm 1 was attributed to the O–H stretching vibration. The bands at 2923 and 2852 cm 1 corresponded to the asymmetric and symmetric stretching vibrations of CH2 of aliphatic structures and lipids, respectively. The band at 1716 cm 1 was linked to the C@O vibration of carboxylic acids. The typical bands located at 1654 and 1540 cm 1 were assigned to the stretching and deformation vibrations of C@O, C–N (amide I) and N–H (amide II) peptidic bond of proteins. The band at 1457 cm 1 was the CH2 deformation vibration. The band present at 1418 cm 1 was related to the C@O stretching vibrations of carboxylates and OH deformation vibration of alcohols and phenols. The bands at 1037–1088 cm 1 was due to the C–O–C and C–O vibration of polysaccharides. While the bands at <1000 cm 1 were the ‘‘fingerprint’’ zone of phosphate or sulfur functional groups. EPS are one of the predominant components in sludge matrixes, mainly consisting of proteins (PN), polysaccharides (PS), smaller amounts of DNA, lipids, etc. Therefore, the nature of functional groups present in the flocs may predominantly originate from the EPS compositions (Laurent et al., 2011). The relative intensities of bands at 1654 and 1540 cm 1 in the six spectra gradually decreased following Fe(II)–S2O82 oxidation, indicating the degradation and removal of protein-like substances in EPS as proved by the decreased negative surface charge of flocs (Fig. 2). It was also
3.6. SEM analysis To get insight into the responsible dewatering mechanisms, SEM analysis of sludge with different pretreatment was performed. The differences in sludge appearance were profoundly significant (Fig. S2 in Supplementary material). The surface of un-pretreated sludge flocs was relatively round and smooth (Fig. S2a), while pretreated by Fe(II)–S2O82 oxidation at 25 °C, the flocs of waste activated sludge appeared to be much less regular, forming a structure with many pores (Fig. S2b). When the temperature increased to 80 °C, however, there were nearly no round sludge flocs to be detected (Fig. S2d), indicating the total collapse of the flocs. Notably, the combined thermal and Fe(II)–S2O82 oxidation pretreatment resulted in the disruption of sludge flocs by degrading EPS, also EPS bound-water and interstitial water trapped between the flocs was released into the free water, which enhanced the dewaterability of waste activated sludge. 4. Conclusions Fe(II)–S2O82 oxidation under mild temperature (25–80 °C) was demonstrated to be efficient in enhancing the dewaterability of sludge. Increasing particle size, and decreasing negative zeta potential and EPS content were mainly responsible for the observed changes in the dewaterability. EEM and FT-IR analysis showed that combined thermal and Fe(II)–S2O82 oxidation pretreatment led to degrading of tyrosine and tryptophan protein-like substances of
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EPS and cleavage of linkages in polymeric backbone. SEM images revealed the rupture of sludge flocs at the colloidal scale, inducing the release of EPS-bound water, also, interstitial water trapped between the flocs were released, which enhanced the dewaterability. Acknowledgements The authors wish to thank the Science and Technology Commission of Shanghai Municipality (No. 08 DZ 1202802, 09 DZ 1204105 and 09 DZ 2251700) and the Fundamental Research Funds for the Central Universities (No. 0400219152) for the partial support of this study. 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.2012.08. 039. References Abelleira, J., Pérez-Elvira, S.I., Sánchez-Oneto, J., Portela, J.R., Nebot, E., 2012. Advanced thermal hydrolysis of secondary sewage sludge: a novel process combining thermal hydrolysis and hydrogen peroxide addition. Resour. Conserv. Recycl. 59, 52–57. Audrey, P., Julien, L., Christophe, D., Patrick, L., 2011. Sludge disintegration during heat treatment at low temperature: a better understanding of involved mechanisms with a multiparametric approach. Biochem. Eng. J. 54 (3), 178– 184. Bougrier, C., Albasi, C., Delgenès, J.P., Carrère, H., 2006. Effect of ultrasonic, thermal and ozone pre-treatments on waste activated sludge solubilisation and anaerobic biodegradability. Chem. Eng. Process. 45, 711–718. Bougrier, C., Delgenes, J.P., Carrere, H., 2008. Effects of thermal treatments on five different waste activated sludge samples solubilisation, physical properties and anaerobic digestion. Chem. Eng. J. 139 (2), 236–244. Baker, A., 2001. Fluorescence excitation–emission matrix characterization of some sewage-impacted rivers. Environ. Sci. Technol. 35 (5), 948–953. Chen, W., Westerhoff, P., Leenheer, J.A., Booksh, K., 2003. Fluorescence excitation– emission matrix regional integration to quantify spectra for dissolved organic matter. Environ. Sci. Technol. 37, 5701–5710. Coble, P.G., 1996. Characterization of marine and terrestrial DOM in seawater using excitation–emission matrix spectroscopy. Mar. Chem. 51 (4), 325–346. Diak, J., Ormeci, B., Proux, C., 2011. Freeze-thaw treatment of RBC sludge from a remote mining exploration facility in subarctic Canada. Water Sci. Technol. 63 (6), 1309–1313. Devlin, D.C., Esteves, S.R.R., Dinsdale, R.M., Guwy, A.J., 2011. The effect of acid pretreatment on the anaerobic digestion and dewatering of waste activated sludge. Bioresour. Technol. 102, 4076–4082. Feng, X., Deng, J.C., Lei, H.Y., Bai, T., Fan, Q.J., Li, Z.X., 2009. Dewaterability of waste activated sludge with ultrasound conditioning. Bioresour. Technol. 100 (3), 1074–1081. Guan, B.H., Yu, J., Fu, H.L., Guo, M.H., Xu, X.H., 2012. Improvement of activated sludge dewaterability by mild thermal treatment in CaCl2 solution. Water Res. 46, 425–432. Gulnaz, O., Kaya, A., Dincer, S., 2006. The reuse of dried activated sludge for adsorption of reactive dye. J. Hazard. Mater. B134, 190–196. Higgins, M.J., Novak, J.T., 1997. The effect of cations on the settling and dewatering of activated sludges. Water Environ. Res. 6, 225–232.
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