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Dewatering of drinking water treatment sludge using the Fenton-like process induced by electro-osmosis
Accepted Manuscript Dewatering of drinking water treatment sludge using the Fenton-like process induced by electro-osmosis Hang Xu, Kunlun Shen, Tonggang Ding, Jianfeng Cui, Mingmei Ding, Chunhui Lu PII: DOI: Reference:
S1385-8947(16)30107-3 http://dx.doi.org/10.1016/j.cej.2016.02.025 CEJ 14758
To appear in:
Chemical Engineering Journal
Received Date: Revised Date: Accepted Date:
6 November 2015 1 February 2016 8 February 2016
Please cite this article as: H. Xu, K. Shen, T. Ding, J. Cui, M. Ding, C. Lu, Dewatering of drinking water treatment sludge using the Fenton-like process induced by electro-osmosis, Chemical Engineering Journal (2016), doi: http:// dx.doi.org/10.1016/j.cej.2016.02.025
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1
Dewatering of drinking water treatment sludge using the Fenton-like
2
process induced by electro-osmosis
3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 25 26 27 28 29
Hang Xu1,2,*, Kunlun Shen2, Tonggang Ding2, Jianfeng Cui2, Mingmei Ding2, and Chunhui Lu3 1
Key Laboratory of Integrated Regulation and Resource Development on Shallow Lake of Ministry of Education, College of Environment, Hohai University, Nanjing 210098, China 2
Hohai University, College of Environmental Science, Nanjing 210098. China
3
State Key Laboratory of Hydrology-Water Resources and Hydraulic Engineering, Hohai University, Nanjing, China
*Corresponding author: [email protected]
Revised Manuscript Submitted to Chemical Engineering Journal on Feb. 1, 2016
30
Highlights
31 32 33 34 35 36 37 38 39 40
(1) The Fenton-like process in combination with acidification enhances highly the dewaterability of sludge; (2) The Fenton-like process is capable of destroying the extracellular polymeric substances (EPS); (3) The operating parameters to achieve the optimal dewaterability efficiency are obtained.
41
Abstract
42
In the present study, the benefits of a Fenton-like treatment process in enhancing the
43
dewaterability of drinking water treatment sludge have been investigated. It is found
44
that using a vacuum electro-osmosis dewatering reactor often results in a low
45
dewatering efficiency, while it can be improved significantly (by about 63% under the
46
optimal condition) by adding the Fenton-like process in combination with
47
acidification. The results based on the three-dimensional (3D) excitation-emission
48
matrix fluorescence spectroscopy and Fourier-transformed infrared spectroscopy
49
indicated that the sludge treated by the Fenton-like process could make fulvicacid-like
50
substances disappear such that the extracellular polymeric substances (EPS) is broken
51
and the dewaterability efficiency is improved. Furthermore, the scanning electronic
52
microscopy analysis and particle size distribution analysis showed that the Fenton-like
53
process and acidification destroyed the stable sludge flocs such that the water trapped
54
in the flocs could be released easily and the dewaterability efficiency was enhanced.
55
The 3D response surface of the water removal efficiency evidenced that the optimal
56
operating parameters were Fe3+ dosage of 54 mgg-1 sludge, H2O2 dosage of 87 mgg-1
57
sludge and pH of 6.3. The results obtained in this study are expected to provide
58
guidance and support for using the Fenton-like treatment process together with
59
acidification to dewater drinking water treatment sludge.
60 61
Keywords: Drinking water treatment sludge; Dewaterability; Extracellular polymeric
62
substances
63
64 65
1. Introduction In southern China, the increasing production of drinking water treatment sludge
66
has led to a serious problem associated with sludge dewatering. The high organic
67
content of the sludge together with the high summer temperature results in the
68
enhanced activity of sludge microorganisms, causing serious biological pollution [1,2].
69
The extracellular polymeric substances (EPS) produced by the biological activity have
70
been shown capable of reducing the dewaterability of the sludge [3,4], resulting in the
71
drinking water treatment sludge with a high moisture content.
72
For a few decades, efforts have been devoted to improve the drinking water
73
sludge dewaterability [5,6]. The negative charged network EPS occupy a significant
74
fraction of the sludge mass, playing a key role in binding a large amount of bound
75
water. The degradation of EPS and the lysis of the biological cells in sludge
76
significantly enhance the release of bound water from sludge flocs [7,8]. Many
77
conditioning techniques, such as ultrasonic pretreatment, microwave irradiation, acid
78
pretreatment, and thermochemical treatment, have been introduced to degrade EPS
79
and to enhance sludge dewatering [3,4,9].
80
Among various conditioning techniques, the Fenton-like process can degrade
81
EPS, and has been studied and proved effective in enhancing sludge dewatering
82
[10-12]. The main reactions involved in the Fenton-like process are described in the
83
following equations [13,14]:
84
Fe2++H2O2→Fe3++OH-+OH(k=70M-1s-1)
85
Fe3++H2O2→Fe2++H++HO2(k=70M-1s-1) (2)
(1)
86
OH+H2O2→HO2+H2O(k=70M-1s-1)
(3)
87
OH+ Fe2+→Fe3++ OH- (k=70M-1s-1)
(4)
88
Fe3++ HO2→Fe2++O2+H+(k=1.2×106M-1s-1) at pH3
89
Fe2++ HO2+H+→Fe3++ H2O2 (k=1.3×106M-1s-1) at pH3
90
HO2+ HO2→H2O2+O2
(5) (6)
(7)
91
The drinking water treatment sludge has a complicated structure, in which the
92
water exists as free water, pore water, surface adhesion water (held on the surface of
93
solid particles by adsorption and adhesion), and internal combined water (intercellular
94
and chemically bound water) [15-17]. The water in EPS is mainly composed of
95
surface adhesion water and internal combined water, collectively referred to as
96
“bound water”. The Fenton-like process enhances the degradation of EPS and the
97
lysis of biological cells, thereby improving the transformation of internal combined
98
water to free water [18,19]. As shown in equations (1)-(7), the Fenton-like reactions
99
mainly involve the reaction of hydrogen peroxide with ferric ions, and the generated
100
hydroxyl radicals improve dewaterabilty. Acidic conditions favor increasing the
101
content of dissolved ferric irons, and thus the formation of iron hydroxide flocs. In
102
addition, as the main content of the drinking water treatment sludge is Al(OH)3, the
103
acidification of the sludge can expel the water stored within the skeleton of the sludge
104
by dissolving Al(OH)3 (the dominant inorganic species involved in the flocs
105
formation) [20-22].
106 107
As an advanced dewatering technology for the municipal and industrial wastewater treatment sludge, electro-osmosis dewatering has gained increasing
108
attention in recent years and been widely studied. The electro-osmosis dewatering
109
technology is an efficient electro-osmotic dewatering method as it enhances the
110
removal rate of water in the sludge [23-25]. In the process of vacuum electro-osmosis
111
dewatering (VEOD), pore water is discharged from the sludge when the flocs
112
structure of the sludge is broken in the presence of the sludge acidification. Through
113
acidification, pH changes the surface charge of sludge particles and the surface charge
114
changes the Zeta potential, thus enhancing the dewaterability [26,27]. The
115
electro-osmosis dewatering process enhances the removal rate of surface adhesion
116
water, and the Fenton-like process breaks the EPS. In combination, the two
117
mechanisms improve the release of internal combined water. The water in the sludge
118
combines with cations to become cationic hydrate, which tends to adhere to sludge
119
particles. However, under the effect of the electric filed, cationic hydrate migrates to
120
cathode [28,29].
121
In the study described herein, a vacuum electro-osmosis dewatering apparatus
122
has been operated at an existing drinking water treatment plant for more than two
123
years to investigate the optimal condition for vacuum electro-osmosis dewatering.
124
The EPS were examined critically through the use of the Fourier transform infrared
125
(FT-IR) spectroscopy and three-dimensional (3D) excitation-emission matrix (EEM)
126
fluorescence spectroscopy to identify the specific constituents of EPS and their
127
characteristic properties. After the pretreatment of the sludge using the Fenton agent
128
and acidification, the integrated thermal gravity (TG), differential thermal analysis
129
(DTA) and differential scanning calorimetry (DSC) (TG-DTA-DSC) technology were
130
employed to analyze the moisture distribution in the sludge. Furthermore, the sludge
131
particle size distribution was examined to explore the destruction of sludge flocs, and
132
scanning electron microscopy (SEM) was also used to understand the characteristic
133
transformation of sludge particles. The objective of the research was to advance
134
knowledge in the optimization of the Fenton-like process and acidification for sludge
135
dewatering.
136 137
2. Materials and methods
138
2.1. Experimental setup
139
Three identical vacuum electro-osmosis dewatering devices (Fig. 1) were used
140
in this research, located in a drinking water treatment plant in Wuxi, Jiangsu Province,
141
China. Each laboratory-scale apparatus consisted of a cuboid sludge tank (25 cm long,
142
20 cm wide,15 cm high), where a row of three titanium pipe anodes and a row of
143
three titanium pipe cathodes were separated by a distance of 12 cm. The titanium pipe
144
cathodes were incised using a laser cutter to create a kerf for passing water; then the
145
cathodes were covered with a geotechnical liner (5 µm aperture) to prevent the
146
penetration of sludge particles. A directed current (DC) power supply (JX-100A/100V,
147
Huizhou Co., Huizhou, China), and two cylindrical vacuum tanks (150 mm in
148
diameter and 200 mm in height) were also part of the VEOD device. The titanium
149
pipe cathodes were connected to the vacuum tanks for vacuum filtration. A vacuum
150
pressure controller was installed for controlling the operation of the vacuum pump
151
(“On” or “Off”) and maintaining the vacuum pressure of the vacuum tanks at a
152 153
specific value. For each conditioning test, 4,500 mL drinking water treatment sludge (DWTS,
154
described in Tables 1 and 2) was added into the sludge tank, and H2SO4 at the
155
concentration of 1 M was added to a 100 mL portion of sludge to adjust the pH to a
156
defined value. Then, H2O2 was dosed into the sludge tank. After being mechanically
157
stirred for 5 min, an amount of FeCl3 was added to the oxidized sludge, and then
158
stirred for an additional 5 min. When the Fenton-like peroxidation was complete, the
159
sludge in the vacuum electro-osmosis dewatering reactor was dewatered.
160 161
2.2. Analytical methods
162
2.2.1. Response surface methodology for the dewatering effect
163
To determine the optimal condition of the Fenton-like process and pH
164
pretreatment, the central composite design of the response surface methodology
165
(RSM) was used [21]. The independent variables are H2O2 dosage (X1), Fe3+ dosage
166
(X2), and pH value (X3), and the moisture removal efficiency was used as the
167
dependent (response) variable, Y. The dewatering experiment design is presented in
168
Table 2. A second-order polynomial model (equation 8) based on the principle of
169
RSM was used to describe the relationship between the response variable and the
170
independent variables.
171
Y=a0+a1X1+ a2X2+ a3X3+ a12X1X2 + a23X2X3 + a13X1X3 + a11X12+ a22X22+ a33X32 (8)
172
where Y is the predicted response, and a0, a1, a2, a11, a12, a22, a23, and a33 are the model
173
regression coefficients for the independent variables X1, X2, X3 and their products,
174
respectively. Design 7.0 software was used to evaluate the response equation 8 and to
175
conduct the corresponding analysis.
176 177 178
2.2.2. Extraction of EPS from sludge The EPS were extracted from the mixed liquor of the drinking water treatment
179
sludge according to the thermal treatment method described by Xuan et al. [30].
180
Sludge collected from the horizontal sedimentation tank in the treatment plant was
181
centrifuged (3,200 rpm for 30 min) to separate the bound EPS from the sludge pellets.
182
The sludge pellets in the centrifuge tube were then washed twice with saline water
183
(0.9% NaCl solution). The residual granular sludge was ground to a powder having
184
particles smaller than 0.18 mm. The sludge was re-suspended using saline water and
185
placed in a water bath at 100°C for 1h, after which the mixture was centrifuged at
186
3,200 rpm for 30 min. At this step, the organic matter in the supernatant was regarded
187
as the bound EPS. The collected supernatant was filtered through a 0.45 µm cellulose
188
nitrate membrane and then analyzed using the 3D EEM fluorescence spectra.
189 190 191
2.2.3. 3D EEM fluorescence spectroscopy All EEM fluorescence spectra were measured using a luminescence
192
spectrometry instrument (FluoroMax-4, HORIBA JobinYvon Co., Paris, France). The
193
EPS EEM spectra were generated at wavelengths from 200 to 550 nm at 5-nm
194
increments by varying the excitation (Ex) wavelength from 200 to 400 nm at 5-nm
195
increments. Excitation and emission slits on the instrument were both maintained at 5
196
nm, and the scanning speed was set at 4,800 nmmin-1 for all measurements. Under the
197
same conditions, fluorescence spectra for Milli-Q water was generated and subtracted
198
(i.e., separated) from all the EPS spectra to eliminate the effect of Raman scattering
199
due to water and to reduce other background noise. The software Origin 8.5 (Origin
200
8.5, Origin Lab Inc., Hampton, USA) was employed to analyze the EEM data. The
201
EEM spectra were plotted as the elliptical shape of contours.
202 203 204
2.2.4. FT-IR spectroscopy Previous researchers have employed the FT-IR analysis to characterize the
205
major functional groups of organic matter and to predict the major components [31].
206
In the present study, bound EPS extracted from the sludge biomass was analyzed
207
using a FT-IR spectrometer (Nicolet 5700, Thermo Electron Corporation,
208
Massachusetts, USA) to determine the organic substances comprising the EPS.
209 210 211
2.2.5. Determination of water content The total moisture content of DWTS samples was measured gravimetrically
212
(overnight drying at 105°C). Based on DSC and TG/DTA tests [32,33], the
213
distribution of sludge moisture was determined according to the following steps. An
214
approximately 20-mg sludge sample was retrieved using a disposable plastic pipette
215
and put into a special mini-oven. The oven maintained a constant temperature (40°C),
216
constant humidity, and constant gas flow (nitrogen at 400 mlmin-1). Then the sample
217
was placed in the TG-DSC-DTA drying crucible. The moisture content and drying
218
rate of the sludge was measured accurately in real time using the integrated
219
TG-DSC-DTA device (STARe System, METTLER TOLEDO, Zurich, Switzerland).
220
The measurement resolution was 0.001 mg. Data from the drying process were
221
transmitted to a computer, and a sludge drying curve was produced in which the
222
X-axis was the sludge moisture content and the Y-axis was the drying rate.
223
Previous researchers reported that water in DWTS has a specific drying rate
224
[16]. As shown in Fig. 2, between points A and B the drying rate of free water is
225
constant. After all free water has been evaporated, the pore water is removed (points B
226
to C in Fig. 2), and the drying rate decreases because the binding force between pore
227
water and DWTS particles is greater than the binding force between the free water
228
and DWTS particles. The procedure using the integrated TG-DSC-DTA instrument
229
allowed the percentage of free water (points A to B, Fig. 2), pore water (points B to C,
230
Fig. 2), and surface adhesion water (points C to D, Fig. 2) to be determined. To
231
determine the quantity of internal combined water (point D to end, Fig. 2), the drying
232
temperature was increased to 105°C. When the drying rate was reduced to zero after a
233
drying period of 24 h, the percentage of internal combined water could be determined.
234 235
2.2.6. Scanning electron microscopy
236
The microstructure of the DWTS samples was determined using reported
237
procedures [9]. First, a DWTS sample was immobilized onto a Poly-L-Lysine coated
238
glass slide and washed three times with 0.1 M phosphate buffer (pH 7.2). Next, the
239
sample was repeatedly immersed in increasing concentrations of ethanol (20%–96%)
240
and air-dried to achieve complete dehydration. Lastly, the sample was sputter-coated
241
with gold. The surface morphology and structural characteristics were observed via a
242
scanning electron microscope (JSM-6010LA, Japan Electron Optics Laboratory Co.,
243
Tokyo, Japan).
244 245 246
2.2.7. Other analyses The pH of samples was measured using a pH meter (PH-3C, Leici Co.,
247
Shanghai, China). The protein from EPS and filtrate was measured by lowry’s method
248
[12]. The element content of drinking water treatment sludge was determined by
249
ICP-AES (AAnalyst 800, Perkin Elmer Inc., Massachusetts, USA).The measurement
250
of the particle size distribution and Zeta potential distribution was conducted using a
251
Malvern Zeta sizer Nano ZS (ZS 90, Malvern Instruments Ltd. Co., Malvern, UK).
252
The software Origin 8.5 was employed to analyze the particle size distribution data.
253 254
3. Results and discussion
255
3.1. Water removal and moisture distribution
256
As shown in Table 3, the water removal efficiency was used as the response
257
(experimental and predicted, Y1 and Y2, respectively) to the independent variables,
258
and served as an important indicator for the treatment efficiency of the acidification
259
and Fenton-like processes. The following polynomial equation defined the highest
260
water removal efficiency and optimal reaction condition:
261
Y=7.63+12.57X1+6.28X2-3.91X3+4.48X12-5.32X32+7.43X1X2+3.15X1X3-0.75X2X3
262 263
R2=0.956, F=62.44
(9)
Statistical testing of the model was performed with the corresponding analysis
264
of variance (ANOVA). The high coefficient of determination (0.956) indicated that
265
the equation is a reliable model capable of predicting the optimum dewatering
266
conditions for the drinking water treatment sludge [34]. The combination of operating
267
variables to achieve optimal DWTS dewatering was determined to be a Fe3+ dosage of
268
54mgg-1 sludge, an H2O2 dosage of 87 mgg-1 sludge and a pH of 6.3, for which the
269
relevant water removal efficiency was estimated to be 63%.
270
The water removal rates from the sludge were obtained from the surface
271
response plots shown in Fig. 3. These plots describe the Fenton-like process for
272
different operating conditions. The three surface response plots (Figs. 3a, 3b and 3c)
273
indicated that there were clear nonlinear effects on the water removal efficiency,
274
resulting from different combinations of H2O2 dosage and Fe3+ dosage, pH and H2O2
275
dosage, and Fe3+ dosage and pH. As shown in Fig. 3a, 3b and3c, an increasing H2O2
276
dosage improved the water removal efficiency. He et al. [35] observed a similar result
277
when studying the effect of H2O2 dosage on the water removal efficiency of waste
278
water sludge. In the present study, the highest removal efficiency was located on the
279
designated surface boundary in Fig. 3a. The peak efficiencies shown in Figs. 3b and
280
3c were located within the designated boundary of Fe3+ and H2SO4, respectively,
281
indicating that the optimal pH was 6.3 and that Fe3+ had little effect on improving the
282
water removal efficiency without the help of H2O2. However, the sludge at low pH
283
values is not allowed for sludge landfill or resource utilization. As such, further study
284 285
of sludge dewatering with Fenton process at normal pH is required. Fig.4 shows the moisture distribution within the sludge under different
286
operating conditions. Fig.4a presents the TG-DSC-DTA drying plots analyzed to
287
determine the moisture distribution shown in Fig. 4b for different sludge treatment
288
conditions. As shown in Fig. 4b, when compared to that in untreated DWTS, the
289
quantity of internal combined water in the sludge treated by the Fenton process
290
decreased highly, and free water increased highly, consistent with the findings in
291
previous research with other dewatering technologies. Ye et al. [7] reported that
292
ultrasound and the combined reagents Fe2+ and KMnO4 are capable to change the
293
percentage of free water, interstitial water and bound water in the sludge. Feng et al.
294
[25] concluded that the percentage of free water and bound water is always changing
295
during the electro-osmotic dewatering process, indicating that the electric field force
296
has the advantage of removing bound water, while traditional mechanical dewatering
297
techniques only remove free water and pore water [16]. In addition, as shown in Fig.
298
4b, the sludge treated using H2SO4 contained more pore water than did untreated
299
sludge. Because pore water is bound by the Al(OH)3 skeleton that comprises sludge
300
flocs, the H2SO4 has the ability to break the Al(OH)3 skeleton resulting in that pore
301
water is released; similar results have been reported by Mahmoud et al. [21], Guo et al.
302
[22], and Citeau et al. [36].
303 304 305
3.2. 3D EEM fluorescence spectroscopy analysis The EPS are regarded as one of adverse effects on the efficiency of sludge
306
dewatering. To evaluate the contribution of EPS in the sludge to the dewaterability,
307
many studies have been performed [37-39]. In the present study, the EEM
308
fluorescence spectroscopy was used to examine DWTS subjected to various
309
treatments, including no treatment, H2SO4 only, and the Fenton process only, to
310
evaluate possible changes in the EPS composition resulting from treatment. The EEM
311
fluorescence spectroscopy at these conditions was determined to understand whether
312
the Fenton process and acidification have the ability to destroy the EPS of the sludge.
313
Spectral examinations were conducted in triplicate and analyzed quantitatively
314
(Fig. 5). As shown in Fig. 5a, three peaks are clearly identified in the EEM
315
fluorescence spectra for the untreated sludge. The first main peak was identified at the
316
excitation/emission wavelengths (Ex/Em) of 250/330 nm (Peak A), while the second
317
main peak was observed at Ex/Em 285/340 nm (Peak B). The third main peak was
318
observed at Ex/Em 250/375 nm (Peak C). Peak A has been reported as an aromatic
319
protein-like peak [40-42]. Peak B has been described as an SMP-like (soluble
320
microbial product) peak, in which the SMP substance includes tyrosine-, tryptophan-
321
and protein-like components, and Peak C has been described as a fulvic acid-like peak
322
[41,43]. When compared to those in Fig. 5a, the EEM fluorescence spectra shown in
323
Fig. 5b indicate that the sludge treated using only H2SO4 experienced no changes in
324
the presence of Peak A, B, and C. In other words, the acidification of sludge did not
325
destroy EPS. In contrast, the EEM spectra (Fig. 5c) from sludge treated using the
326
Fenton process exhibited no Peak C. Because fulvicacid is always associated with
327
other substances in EPS, the absence of a fulvicacid peak demonstrated that the
328
Fenton process destroyed the EPS structure and increased the dewaterability of
329
sludge.
330
Zhang et al. [44] concluded that peracetic acid pre-oxidation has the ability to
331
destroy the EPS structure and the fulvicacid peak was also absent. Zhen et al. [45]
332
added the Fe(II)-activated persulfate in the activated sludge dewatering pretreatment
333
and got the absence of fulvicacid peak, and in addition, the soluble EPS and bound
334
EPS all have no fulvicacid after the pretreatment. Fluorescence parameters of the
335
spectra, including the excitation/emission wavelengths for the peaks, fluorescence
336
intensity, and the relevant substances emitting the spectral peaks are given in Table 4.
337
The EEM spectra indicate that the main organic substance in EPS was an
338
SMP(soluble microbial products)-like substance, and confirmed that the high activity
339
of microorganisms in the sludge in summer contributed to the unfavorable
340
dewaterability of the sludge.
341 342 343
3.3. FT-IR analysis The FT-IR spectra of the EPS extracted from the sludge and subjected to
344
various treatments were determined in triplicate and are illustrated in Fig. 6. The
345
FT-IR spectra of the drinking water treatment sludge is similar to those of waste
346
activated sludge [46,47]. The weak band at wave number 3,299 cm-1 has been
347
attributed to the O-H stretching vibration in hydroxyl functional groups [48]. The
348
bands at wave numbers 1,645 cm-1, 1,540 cm-1, and 1,457 cm-1 have been shown to
349
correspond to the protein secondary structure, namely C=O (amide I) and N-H (amide
350
II) [48,49]. These three responses indicate that proteins were one of the components
351
of the EPS. In addition, a clear peak at wave number 1,037 cm-1 exhibited the
352
characteristics similar to those of carbohydrates or carbohydrates-like substances [48],
353
suggesting that carbohydrates were present in the EPS of the drinking water treatment
354
sludge.
355
EPS are one of the predominant components in sludge matrixes, mainly
356
consisting of proteins (PN), polysaccharides (PS), lipids, etc. The nature of functional
357
groups present in the flocs may predominantly originate from the EPS compositions.
358
The transmittance of band at 1,457 and 1,037 cm-1 changed in different conditioned
359
EPS spectra, suggesting that the degradation of EPS, which was also proved by the
360
analysis above for EEM.
361 362 363
3.4. SEM analysis To gain insights into the responsible dewatering mechanism in the sludge
364
treated using the Fenton process and H2SO4, SEM was used to examine the
365
micromorphology of the sludge subjected to various treatments (Fig. 7). Fig.7a shows
366
that the untreated sludge matrix was compact and comprised of sludge flocs arranged
367
in sheets or flakes. After treatment using H2SO4, sludge appeared (Fig. 7b) to consist
368
of much more pores and the orientation of the sludge flocs was broken. Because a
369
porous structure in sludge favors the discharge of pore water, Guo et al. [22]
370
concluded that the acidification of sludge broke the flocs of oil treatment sludge,
371
causing pores in the sludge that can improve the dewaterability.
372
The microstructure of sludge treated using the Fenton process (Fig. 7c) showed
373
no obvious differences from that of the untreated sludge. Perhaps some biological
374
detection methods would be helpful in examining the biomass in the sludge and might
375
be able to identify differences. The combination of H2SO4 and the Fenton process as a
376
treatment created more pores in the sludge (Fig. 7d) and the surface of sludge
377
appeared coarser and much less regular than that of untreated sludge. Zhen et al. [50]
378
concluded that the condition of Fe2+ with persulfate oxidation could break sludge
379
flocs to create many pores in the sludge, and break EPS to improve the dewaterability
380
of the waste activated sludge. Notably in the present study, the sludge treatment that
381
combined the Fenton process with H2SO4 pretreatment resulted in the destruction of
382
sludge flocs and the degradation of EPS, so that pore water, surface combined water
383
and internal combined water in the EPS were easily released by the electric field force
384
of the electro-osmosis dewatering device.
385 386 387
3.5. Size distribution The size distribution of sludge particles is a critical factor influencing sludge
388
dewaterability. The relevant determinations of particle sizes were conducted in
389
triplicate, but for clarity, only one of the size distributions for each sludge treatment is
390
shown in Fig. 8. The size distribution in the raw sludge was multimodal, with an
391
average particle size of 3,000 µm. In contrast, the size distribution of treated sludge
392
(regardless of treatment) was unimodal, and the result indicated the sludge was
393
formed under different hydraulic conditions, such as flocculant usage and flocculation
394
time in the flocculation basin. The largest proportion of particles in the sludge treated
395
using H2SO4, using H2SO4 in combination with the Fenton process, and using the
396
Fenton process only was 121, 1,581 and 2,985 nm, respectively. These results
397
revealed that acidified sludge consisted of a high proportion of small particles because
398
the acid broke the sludge flocs and created a small sludge matrix. This was also
399
evidenced in the SEM analysis. The average size of particles in the sludge treated
400
using the Fenton process alone was slightly different from that in the untreated sludge.
401
These results indicated that the Fenton process slightly destroys the drinking water
402
treatment sludge flocs, consistent with observations from the SEM analysis.
403 404 405
4. Conclusions The Fenton-like sludge dewatering process combined with acidification was
406
demonstrated to be efficient in enhancing the dewaterability of drinking water
407
treatment sludge. The optimum operating parameters (Fe3+ dosage of 54 mgg-1 sludge,
408
H2O2 dosage of 87 mgg-1 sludge and pH of 6.3) resulted in the final moisture content
409
of dewatered sludge being decreased to 63%. Both EEM and FT-IR showed that the
410
Fenton-like process led to the destruction of EPS, which in turn improved the sludge
411
dewaterability and increased the proportion of internal combined water that was
412
changed to free water. The SEM and particle size distribution analyses indicated the
413
acidification of sludge had the effect of breaking flocs and creating many pores in the
414
treated sludge, both of which improved dewaterability.
415
416
Acknowledgements
417
This work was supported by the National Natural Science Foundation of
418
China(51308185), Central University business expenses (2013B32314), A Project
419
Funded by the Priority Academic Program Development of Jiangsu Higher Education
420
Institutions and National major water projects (2014ZX07405002). We would like to
421
thank four anonymous reviewers for their constructive comments on the work.
422 423 424 425 426 427 428 429 430 431 432 433 434 435 436 437 438 439 440 441 442 443 444 445 446 447 448 449 450 451 452 453 454 455 456 457 458 459 460 461 462 463 464
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553 554
Table 1
555 556
557 558
Table 1. Characteristics of drinking water treatment sludge Moisture content pH Protein Zeta potential (%) mg/L (mv) 87.47±0.09 7.73±0.15 108.3±27.2 -12.6±1.1
Average particle size(µm) 2860 ± 217
559
Table 2
560 561
Table 2. Element content of the drinking water treatment sludge Element Content
562 563
C (%)
Al (%)
Si (%)
K (%)
Fe (%)
O (%)
45.97±1.12 14.99±1.29 19.62±0.61 1.47±0.88 3.8±0.20 45.97±0.31
564
Table 3
565 566
567 568
Table 3 Experimental design matrix and the observed responses NO. Independent Variables Water removal EXP. efficiency, % 3+ X1 (H2O2 X2 (Fe X3 (pH) Experimental Predicted dosage, dosage, Y1 Y2 mgg-1 mgg-1 sludge) sludge) 1 10 10 1.3 18 13 2 10 30 3.2 35 34 3 10 50 5.3 38 50 4 10 70 7.6 38 37 5 30 10 1.3 41 25 6 30 30 3.2 50 48 7 30 50 5.3 52 50 8 30 70 7.6 48 58 9 50 10 1.3 45 24 10 50 30 3.2 47 46 11 50 50 5.3 49 62 12 50 70 7.6 42 39 13 70 10 1.3 41 28 14 70 30 3.2 44 33 15 70 50 5.3 47 57 16 70 70 7.6 43 40
569
Table 4
570 571
572 573
Table 4 Fluorescence spectral parameters of the extracellular polymeric substances Peaks Ex/Em (nm) Intensity Emitting substances Peak A 225/330 1,204 aromatic protein-like Peak B 285/340 1,673 soluble microbial product-like Peak C 225/400 932 fulvicacid-like
574
Figure Captions
575 576 577 578 579 580 581 582 583 584 585 586 587 588 589 590 591 592 593 594 595 596 597 598 599 600 601 602 603 604 605 606 607 608
Fig. 1. Vacuum electro-osmosis dewatering apparatus: (1) DC power supply; (2) sludge tank; (3) anode; (4) cathode; (5) four-way pipe; (6) vacuum collecting tank; (7) vacuum pump; and (8) vacuum pressure controller. Fig. 2. The universal analytical method used to determine the distribution of moisture in sludge. Fig. 3. Three-dimensional surface plots of water removal efficiency: (a) H2O2 dosage vs. Fe3+; (b) pH vs. H2O2 dosage; and (c) Fe3+ dosage vs. pH. Fig. 4. Moisture distribution in drinking water treatment sludge: (a) drying curves; and (b) moisture distribution in the sludge subjected to various treatments. Fig. 5. Excitation-emission matrix fluorescence spectra of the extracellular polymeric substances in: (a) untreated sludge; (b) sludge treated using H2SO4; and (c) sludge treated using the Fenton process. Fig. 6. Fourier-transform infrared spectra of extracellular polymeric substances from drinking water treatment sludge subjected to different treatments: (a) untreated sludge (black line); (b) sludge treated using H2SO4 (red line); (c) sludge treated using the Fenton process (blue line); and (d) sludge treated using H2SO4 in combination with the Fenton process (magenta line). Fig. 7. Scanning electron microscopy images of drinking water treatment sludge subjected to various treatments: (a) untreated sludge; (b) sludge treated using H2SO4; (c) sludge treated using the Fenton process; and (d) sludge treated using H2SO4 in combination with the Fenton process. Fig. 8. Size distribution of particles in drinking water treatment sludge subjected to different treatments: (a) untreated sludge (red line); (b) sludge treated using H2SO4 (black line); (c) sludge treated using the Fenton process (blue line); and (d) sludge treated using H2SO4 in combination with the Fenton process (magenta line).
609 610 611 612 613
Fig. 1. Vacuum electro-osmosis dewatering apparatus: (1) DC power supply; (2) sludge tank; (3) anode; (4) cathode; (5) four-way pipe; (6) vacuum collecting tank; (7) vacuum pump; and (8) vacuum pressure controller.
614 615 616 617
Fig. 2. The universal analytical method used to determine the distribution of moisture in sludge.
(b)
(a) 70
70 60
(%) water removal
50
40 30 20 10
40 30 20
60
70
107
30
O
3
2
40
5
pH
20
2
20
20
e
Fe 3+ do sag e
g sa do O2 H2
40
60
50
6
40
60
2
0
H
Water removal
50
4
(%)
60
1
0
618
10
(c) 35
25
wat er removal
(%)
30
20 15 10 7
750
6
0
5 0
4
pH
5
60
4
3
30
Fe 20
2
620 621 622
1
10
619
Fig. 3. Three-dimensional surface plots of water removal efficiency: (a) H2O2 dosage vs. Fe3+; (b) pH vs. H2O2 dosage; and (c) Fe3+ dosage vs. pH.
35
untreated H2SO4+Fenton
-0.0001
H2SO4
drying speed (mg/s)
-0.0002
Fenton second speed second speed drop drop
-0.0003
second speed drop
-0.0004 -0.0005
second speed drop
-0.0006
first speed drop
first speed drop
first speed drop
first speed drop
-0.0007
623 624 625 626
0
20
40
60
moisture content (%)
80
100
(b)
mass of water per mass of sludge solid
0.0000
(a)
free water pore water surface adhesion water internal combined water
30
25 20 15 10
5 0 H2SO4+ Fenton
untreated
H2SO4
Fenton
Fig. 4. Moisture distribution in drinking water treatment sludge: (a) drying curves; and (b) moisture distribution in the sludge subjected to various treatments.
450
450 0.000 240.0 480.0 720.0 960.0 1200 1440 1680 1920 2160 2400 2640 2880 3120 3360 3600 3840 4080 4320 4560 4800 5040 5280 5520 5760 6000
400
EX(nm)
350
Peak B
300
250
Peak A
Peak C
200 200
627
250
300
350
400
(b)
350
Peak B
300
250
Peak C
Peak A 200 200
450
EM(nm)
0.000 240.0 480.0 720.0 960.0 1200 1440 1680 1920 2160 2400 2640 2880 3120 3360 3600 3840 4080 4320 4560 4800 5040 5280 5520 5760 6000
400
EX（ nm）
(a)
250
300
350
400
450
EM（ nm）
450 0.000 240.0 480.0 720.0 960.0 1200 1440 1680 1920 2160 2400 2640 2880 3120 3360 3600 3840 4080 4320 4560 4800 5040 5280 5520 5760 6000
(c) 400
EX(nm)
350
300
Peak B
250
Peak A 200 200
628 629 630 631 632
250
300
350
400
450
EM(nm)
Fig. 5. Excitation-emission matrix fluorescence spectra of the extracellular polymeric substances in: (a) untreated sludge; (b) sludge treated using H2SO4; and (c) sludge treated using the Fenton process.
100
0
1037
3299
1654 1540 1457
40
60
treated by Fenton treated by H2SO4+Fenton
20
Transmittance(%)
80
raw sludge treated by H2SO4
4000
633 634 635 636 637 638 639
3500
3000
2500
2000
1500
1000
500
Wavenumbers (cm-1)
Fig. 6. Fourier-transform infrared spectra of extracellular polymeric substances from drinking water treatment sludge subjected to different treatments: (a) untreated sludge (black line); (b) sludge treated using H2SO4 (red line); (c) sludge treated using the Fenton process (blue line); and (d) sludge treated using H2SO4 in combination with the Fenton process (magenta line).
640
641 642 643 644 645 646
Fig. 7. Scanning electron microscopy images of drinking water treatment sludge subjected to various treatments: (a) untreated sludge; (b) sludge treated using H2SO4; (c) sludge treated using the Fenton process; and (d) sludge treated using H2SO4 in combination with the Fenton process.
H2SO4
25
untreated Fenton H2SO4+Fenton
intensity (%)
20
15
10
5
0 0
647 648 649 650 651
1000
2000
3000
4000
5000
6000
7000
size (µm)
Fig. 8. Size distribution of particles in drinking water treatment sludge subjected to different treatments: (a) untreated sludge (red line); (b) sludge treated using H2SO4 (black line); (c) sludge treated using the Fenton process (blue line); and (d) sludge treated using H2SO4 in combination with the Fenton process (magenta line).
S1385-8947(16)30107-3 http://dx.doi.org/10.1016/j.cej.2016.02.025 CEJ 14758
To appear in:
Chemical Engineering Journal
Received Date: Revised Date: Accepted Date:
6 November 2015 1 February 2016 8 February 2016
Please cite this article as: H. Xu, K. Shen, T. Ding, J. Cui, M. Ding, C. Lu, Dewatering of drinking water treatment sludge using the Fenton-like process induced by electro-osmosis, Chemical Engineering Journal (2016), doi: http:// dx.doi.org/10.1016/j.cej.2016.02.025
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1
Dewatering of drinking water treatment sludge using the Fenton-like
2
process induced by electro-osmosis
3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 25 26 27 28 29
Hang Xu1,2,*, Kunlun Shen2, Tonggang Ding2, Jianfeng Cui2, Mingmei Ding2, and Chunhui Lu3 1
Key Laboratory of Integrated Regulation and Resource Development on Shallow Lake of Ministry of Education, College of Environment, Hohai University, Nanjing 210098, China 2
Hohai University, College of Environmental Science, Nanjing 210098. China
3
State Key Laboratory of Hydrology-Water Resources and Hydraulic Engineering, Hohai University, Nanjing, China
*Corresponding author: [email protected]
Revised Manuscript Submitted to Chemical Engineering Journal on Feb. 1, 2016
30
Highlights
31 32 33 34 35 36 37 38 39 40
(1) The Fenton-like process in combination with acidification enhances highly the dewaterability of sludge; (2) The Fenton-like process is capable of destroying the extracellular polymeric substances (EPS); (3) The operating parameters to achieve the optimal dewaterability efficiency are obtained.
41
Abstract
42
In the present study, the benefits of a Fenton-like treatment process in enhancing the
43
dewaterability of drinking water treatment sludge have been investigated. It is found
44
that using a vacuum electro-osmosis dewatering reactor often results in a low
45
dewatering efficiency, while it can be improved significantly (by about 63% under the
46
optimal condition) by adding the Fenton-like process in combination with
47
acidification. The results based on the three-dimensional (3D) excitation-emission
48
matrix fluorescence spectroscopy and Fourier-transformed infrared spectroscopy
49
indicated that the sludge treated by the Fenton-like process could make fulvicacid-like
50
substances disappear such that the extracellular polymeric substances (EPS) is broken
51
and the dewaterability efficiency is improved. Furthermore, the scanning electronic
52
microscopy analysis and particle size distribution analysis showed that the Fenton-like
53
process and acidification destroyed the stable sludge flocs such that the water trapped
54
in the flocs could be released easily and the dewaterability efficiency was enhanced.
55
The 3D response surface of the water removal efficiency evidenced that the optimal
56
operating parameters were Fe3+ dosage of 54 mgg-1 sludge, H2O2 dosage of 87 mgg-1
57
sludge and pH of 6.3. The results obtained in this study are expected to provide
58
guidance and support for using the Fenton-like treatment process together with
59
acidification to dewater drinking water treatment sludge.
60 61
Keywords: Drinking water treatment sludge; Dewaterability; Extracellular polymeric
62
substances
63
64 65
1. Introduction In southern China, the increasing production of drinking water treatment sludge
66
has led to a serious problem associated with sludge dewatering. The high organic
67
content of the sludge together with the high summer temperature results in the
68
enhanced activity of sludge microorganisms, causing serious biological pollution [1,2].
69
The extracellular polymeric substances (EPS) produced by the biological activity have
70
been shown capable of reducing the dewaterability of the sludge [3,4], resulting in the
71
drinking water treatment sludge with a high moisture content.
72
For a few decades, efforts have been devoted to improve the drinking water
73
sludge dewaterability [5,6]. The negative charged network EPS occupy a significant
74
fraction of the sludge mass, playing a key role in binding a large amount of bound
75
water. The degradation of EPS and the lysis of the biological cells in sludge
76
significantly enhance the release of bound water from sludge flocs [7,8]. Many
77
conditioning techniques, such as ultrasonic pretreatment, microwave irradiation, acid
78
pretreatment, and thermochemical treatment, have been introduced to degrade EPS
79
and to enhance sludge dewatering [3,4,9].
80
Among various conditioning techniques, the Fenton-like process can degrade
81
EPS, and has been studied and proved effective in enhancing sludge dewatering
82
[10-12]. The main reactions involved in the Fenton-like process are described in the
83
following equations [13,14]:
84
Fe2++H2O2→Fe3++OH-+OH(k=70M-1s-1)
85
Fe3++H2O2→Fe2++H++HO2(k=70M-1s-1) (2)
(1)
86
OH+H2O2→HO2+H2O(k=70M-1s-1)
(3)
87
OH+ Fe2+→Fe3++ OH- (k=70M-1s-1)
(4)
88
Fe3++ HO2→Fe2++O2+H+(k=1.2×106M-1s-1) at pH3
89
Fe2++ HO2+H+→Fe3++ H2O2 (k=1.3×106M-1s-1) at pH3
90
HO2+ HO2→H2O2+O2
(5) (6)
(7)
91
The drinking water treatment sludge has a complicated structure, in which the
92
water exists as free water, pore water, surface adhesion water (held on the surface of
93
solid particles by adsorption and adhesion), and internal combined water (intercellular
94
and chemically bound water) [15-17]. The water in EPS is mainly composed of
95
surface adhesion water and internal combined water, collectively referred to as
96
“bound water”. The Fenton-like process enhances the degradation of EPS and the
97
lysis of biological cells, thereby improving the transformation of internal combined
98
water to free water [18,19]. As shown in equations (1)-(7), the Fenton-like reactions
99
mainly involve the reaction of hydrogen peroxide with ferric ions, and the generated
100
hydroxyl radicals improve dewaterabilty. Acidic conditions favor increasing the
101
content of dissolved ferric irons, and thus the formation of iron hydroxide flocs. In
102
addition, as the main content of the drinking water treatment sludge is Al(OH)3, the
103
acidification of the sludge can expel the water stored within the skeleton of the sludge
104
by dissolving Al(OH)3 (the dominant inorganic species involved in the flocs
105
formation) [20-22].
106 107
As an advanced dewatering technology for the municipal and industrial wastewater treatment sludge, electro-osmosis dewatering has gained increasing
108
attention in recent years and been widely studied. The electro-osmosis dewatering
109
technology is an efficient electro-osmotic dewatering method as it enhances the
110
removal rate of water in the sludge [23-25]. In the process of vacuum electro-osmosis
111
dewatering (VEOD), pore water is discharged from the sludge when the flocs
112
structure of the sludge is broken in the presence of the sludge acidification. Through
113
acidification, pH changes the surface charge of sludge particles and the surface charge
114
changes the Zeta potential, thus enhancing the dewaterability [26,27]. The
115
electro-osmosis dewatering process enhances the removal rate of surface adhesion
116
water, and the Fenton-like process breaks the EPS. In combination, the two
117
mechanisms improve the release of internal combined water. The water in the sludge
118
combines with cations to become cationic hydrate, which tends to adhere to sludge
119
particles. However, under the effect of the electric filed, cationic hydrate migrates to
120
cathode [28,29].
121
In the study described herein, a vacuum electro-osmosis dewatering apparatus
122
has been operated at an existing drinking water treatment plant for more than two
123
years to investigate the optimal condition for vacuum electro-osmosis dewatering.
124
The EPS were examined critically through the use of the Fourier transform infrared
125
(FT-IR) spectroscopy and three-dimensional (3D) excitation-emission matrix (EEM)
126
fluorescence spectroscopy to identify the specific constituents of EPS and their
127
characteristic properties. After the pretreatment of the sludge using the Fenton agent
128
and acidification, the integrated thermal gravity (TG), differential thermal analysis
129
(DTA) and differential scanning calorimetry (DSC) (TG-DTA-DSC) technology were
130
employed to analyze the moisture distribution in the sludge. Furthermore, the sludge
131
particle size distribution was examined to explore the destruction of sludge flocs, and
132
scanning electron microscopy (SEM) was also used to understand the characteristic
133
transformation of sludge particles. The objective of the research was to advance
134
knowledge in the optimization of the Fenton-like process and acidification for sludge
135
dewatering.
136 137
2. Materials and methods
138
2.1. Experimental setup
139
Three identical vacuum electro-osmosis dewatering devices (Fig. 1) were used
140
in this research, located in a drinking water treatment plant in Wuxi, Jiangsu Province,
141
China. Each laboratory-scale apparatus consisted of a cuboid sludge tank (25 cm long,
142
20 cm wide,15 cm high), where a row of three titanium pipe anodes and a row of
143
three titanium pipe cathodes were separated by a distance of 12 cm. The titanium pipe
144
cathodes were incised using a laser cutter to create a kerf for passing water; then the
145
cathodes were covered with a geotechnical liner (5 µm aperture) to prevent the
146
penetration of sludge particles. A directed current (DC) power supply (JX-100A/100V,
147
Huizhou Co., Huizhou, China), and two cylindrical vacuum tanks (150 mm in
148
diameter and 200 mm in height) were also part of the VEOD device. The titanium
149
pipe cathodes were connected to the vacuum tanks for vacuum filtration. A vacuum
150
pressure controller was installed for controlling the operation of the vacuum pump
151
(“On” or “Off”) and maintaining the vacuum pressure of the vacuum tanks at a
152 153
specific value. For each conditioning test, 4,500 mL drinking water treatment sludge (DWTS,
154
described in Tables 1 and 2) was added into the sludge tank, and H2SO4 at the
155
concentration of 1 M was added to a 100 mL portion of sludge to adjust the pH to a
156
defined value. Then, H2O2 was dosed into the sludge tank. After being mechanically
157
stirred for 5 min, an amount of FeCl3 was added to the oxidized sludge, and then
158
stirred for an additional 5 min. When the Fenton-like peroxidation was complete, the
159
sludge in the vacuum electro-osmosis dewatering reactor was dewatered.
160 161
2.2. Analytical methods
162
2.2.1. Response surface methodology for the dewatering effect
163
To determine the optimal condition of the Fenton-like process and pH
164
pretreatment, the central composite design of the response surface methodology
165
(RSM) was used [21]. The independent variables are H2O2 dosage (X1), Fe3+ dosage
166
(X2), and pH value (X3), and the moisture removal efficiency was used as the
167
dependent (response) variable, Y. The dewatering experiment design is presented in
168
Table 2. A second-order polynomial model (equation 8) based on the principle of
169
RSM was used to describe the relationship between the response variable and the
170
independent variables.
171
Y=a0+a1X1+ a2X2+ a3X3+ a12X1X2 + a23X2X3 + a13X1X3 + a11X12+ a22X22+ a33X32 (8)
172
where Y is the predicted response, and a0, a1, a2, a11, a12, a22, a23, and a33 are the model
173
regression coefficients for the independent variables X1, X2, X3 and their products,
174
respectively. Design 7.0 software was used to evaluate the response equation 8 and to
175
conduct the corresponding analysis.
176 177 178
2.2.2. Extraction of EPS from sludge The EPS were extracted from the mixed liquor of the drinking water treatment
179
sludge according to the thermal treatment method described by Xuan et al. [30].
180
Sludge collected from the horizontal sedimentation tank in the treatment plant was
181
centrifuged (3,200 rpm for 30 min) to separate the bound EPS from the sludge pellets.
182
The sludge pellets in the centrifuge tube were then washed twice with saline water
183
(0.9% NaCl solution). The residual granular sludge was ground to a powder having
184
particles smaller than 0.18 mm. The sludge was re-suspended using saline water and
185
placed in a water bath at 100°C for 1h, after which the mixture was centrifuged at
186
3,200 rpm for 30 min. At this step, the organic matter in the supernatant was regarded
187
as the bound EPS. The collected supernatant was filtered through a 0.45 µm cellulose
188
nitrate membrane and then analyzed using the 3D EEM fluorescence spectra.
189 190 191
2.2.3. 3D EEM fluorescence spectroscopy All EEM fluorescence spectra were measured using a luminescence
192
spectrometry instrument (FluoroMax-4, HORIBA JobinYvon Co., Paris, France). The
193
EPS EEM spectra were generated at wavelengths from 200 to 550 nm at 5-nm
194
increments by varying the excitation (Ex) wavelength from 200 to 400 nm at 5-nm
195
increments. Excitation and emission slits on the instrument were both maintained at 5
196
nm, and the scanning speed was set at 4,800 nmmin-1 for all measurements. Under the
197
same conditions, fluorescence spectra for Milli-Q water was generated and subtracted
198
(i.e., separated) from all the EPS spectra to eliminate the effect of Raman scattering
199
due to water and to reduce other background noise. The software Origin 8.5 (Origin
200
8.5, Origin Lab Inc., Hampton, USA) was employed to analyze the EEM data. The
201
EEM spectra were plotted as the elliptical shape of contours.
202 203 204
2.2.4. FT-IR spectroscopy Previous researchers have employed the FT-IR analysis to characterize the
205
major functional groups of organic matter and to predict the major components [31].
206
In the present study, bound EPS extracted from the sludge biomass was analyzed
207
using a FT-IR spectrometer (Nicolet 5700, Thermo Electron Corporation,
208
Massachusetts, USA) to determine the organic substances comprising the EPS.
209 210 211
2.2.5. Determination of water content The total moisture content of DWTS samples was measured gravimetrically
212
(overnight drying at 105°C). Based on DSC and TG/DTA tests [32,33], the
213
distribution of sludge moisture was determined according to the following steps. An
214
approximately 20-mg sludge sample was retrieved using a disposable plastic pipette
215
and put into a special mini-oven. The oven maintained a constant temperature (40°C),
216
constant humidity, and constant gas flow (nitrogen at 400 mlmin-1). Then the sample
217
was placed in the TG-DSC-DTA drying crucible. The moisture content and drying
218
rate of the sludge was measured accurately in real time using the integrated
219
TG-DSC-DTA device (STARe System, METTLER TOLEDO, Zurich, Switzerland).
220
The measurement resolution was 0.001 mg. Data from the drying process were
221
transmitted to a computer, and a sludge drying curve was produced in which the
222
X-axis was the sludge moisture content and the Y-axis was the drying rate.
223
Previous researchers reported that water in DWTS has a specific drying rate
224
[16]. As shown in Fig. 2, between points A and B the drying rate of free water is
225
constant. After all free water has been evaporated, the pore water is removed (points B
226
to C in Fig. 2), and the drying rate decreases because the binding force between pore
227
water and DWTS particles is greater than the binding force between the free water
228
and DWTS particles. The procedure using the integrated TG-DSC-DTA instrument
229
allowed the percentage of free water (points A to B, Fig. 2), pore water (points B to C,
230
Fig. 2), and surface adhesion water (points C to D, Fig. 2) to be determined. To
231
determine the quantity of internal combined water (point D to end, Fig. 2), the drying
232
temperature was increased to 105°C. When the drying rate was reduced to zero after a
233
drying period of 24 h, the percentage of internal combined water could be determined.
234 235
2.2.6. Scanning electron microscopy
236
The microstructure of the DWTS samples was determined using reported
237
procedures [9]. First, a DWTS sample was immobilized onto a Poly-L-Lysine coated
238
glass slide and washed three times with 0.1 M phosphate buffer (pH 7.2). Next, the
239
sample was repeatedly immersed in increasing concentrations of ethanol (20%–96%)
240
and air-dried to achieve complete dehydration. Lastly, the sample was sputter-coated
241
with gold. The surface morphology and structural characteristics were observed via a
242
scanning electron microscope (JSM-6010LA, Japan Electron Optics Laboratory Co.,
243
Tokyo, Japan).
244 245 246
2.2.7. Other analyses The pH of samples was measured using a pH meter (PH-3C, Leici Co.,
247
Shanghai, China). The protein from EPS and filtrate was measured by lowry’s method
248
[12]. The element content of drinking water treatment sludge was determined by
249
ICP-AES (AAnalyst 800, Perkin Elmer Inc., Massachusetts, USA).The measurement
250
of the particle size distribution and Zeta potential distribution was conducted using a
251
Malvern Zeta sizer Nano ZS (ZS 90, Malvern Instruments Ltd. Co., Malvern, UK).
252
The software Origin 8.5 was employed to analyze the particle size distribution data.
253 254
3. Results and discussion
255
3.1. Water removal and moisture distribution
256
As shown in Table 3, the water removal efficiency was used as the response
257
(experimental and predicted, Y1 and Y2, respectively) to the independent variables,
258
and served as an important indicator for the treatment efficiency of the acidification
259
and Fenton-like processes. The following polynomial equation defined the highest
260
water removal efficiency and optimal reaction condition:
261
Y=7.63+12.57X1+6.28X2-3.91X3+4.48X12-5.32X32+7.43X1X2+3.15X1X3-0.75X2X3
262 263
R2=0.956, F=62.44
(9)
Statistical testing of the model was performed with the corresponding analysis
264
of variance (ANOVA). The high coefficient of determination (0.956) indicated that
265
the equation is a reliable model capable of predicting the optimum dewatering
266
conditions for the drinking water treatment sludge [34]. The combination of operating
267
variables to achieve optimal DWTS dewatering was determined to be a Fe3+ dosage of
268
54mgg-1 sludge, an H2O2 dosage of 87 mgg-1 sludge and a pH of 6.3, for which the
269
relevant water removal efficiency was estimated to be 63%.
270
The water removal rates from the sludge were obtained from the surface
271
response plots shown in Fig. 3. These plots describe the Fenton-like process for
272
different operating conditions. The three surface response plots (Figs. 3a, 3b and 3c)
273
indicated that there were clear nonlinear effects on the water removal efficiency,
274
resulting from different combinations of H2O2 dosage and Fe3+ dosage, pH and H2O2
275
dosage, and Fe3+ dosage and pH. As shown in Fig. 3a, 3b and3c, an increasing H2O2
276
dosage improved the water removal efficiency. He et al. [35] observed a similar result
277
when studying the effect of H2O2 dosage on the water removal efficiency of waste
278
water sludge. In the present study, the highest removal efficiency was located on the
279
designated surface boundary in Fig. 3a. The peak efficiencies shown in Figs. 3b and
280
3c were located within the designated boundary of Fe3+ and H2SO4, respectively,
281
indicating that the optimal pH was 6.3 and that Fe3+ had little effect on improving the
282
water removal efficiency without the help of H2O2. However, the sludge at low pH
283
values is not allowed for sludge landfill or resource utilization. As such, further study
284 285
of sludge dewatering with Fenton process at normal pH is required. Fig.4 shows the moisture distribution within the sludge under different
286
operating conditions. Fig.4a presents the TG-DSC-DTA drying plots analyzed to
287
determine the moisture distribution shown in Fig. 4b for different sludge treatment
288
conditions. As shown in Fig. 4b, when compared to that in untreated DWTS, the
289
quantity of internal combined water in the sludge treated by the Fenton process
290
decreased highly, and free water increased highly, consistent with the findings in
291
previous research with other dewatering technologies. Ye et al. [7] reported that
292
ultrasound and the combined reagents Fe2+ and KMnO4 are capable to change the
293
percentage of free water, interstitial water and bound water in the sludge. Feng et al.
294
[25] concluded that the percentage of free water and bound water is always changing
295
during the electro-osmotic dewatering process, indicating that the electric field force
296
has the advantage of removing bound water, while traditional mechanical dewatering
297
techniques only remove free water and pore water [16]. In addition, as shown in Fig.
298
4b, the sludge treated using H2SO4 contained more pore water than did untreated
299
sludge. Because pore water is bound by the Al(OH)3 skeleton that comprises sludge
300
flocs, the H2SO4 has the ability to break the Al(OH)3 skeleton resulting in that pore
301
water is released; similar results have been reported by Mahmoud et al. [21], Guo et al.
302
[22], and Citeau et al. [36].
303 304 305
3.2. 3D EEM fluorescence spectroscopy analysis The EPS are regarded as one of adverse effects on the efficiency of sludge
306
dewatering. To evaluate the contribution of EPS in the sludge to the dewaterability,
307
many studies have been performed [37-39]. In the present study, the EEM
308
fluorescence spectroscopy was used to examine DWTS subjected to various
309
treatments, including no treatment, H2SO4 only, and the Fenton process only, to
310
evaluate possible changes in the EPS composition resulting from treatment. The EEM
311
fluorescence spectroscopy at these conditions was determined to understand whether
312
the Fenton process and acidification have the ability to destroy the EPS of the sludge.
313
Spectral examinations were conducted in triplicate and analyzed quantitatively
314
(Fig. 5). As shown in Fig. 5a, three peaks are clearly identified in the EEM
315
fluorescence spectra for the untreated sludge. The first main peak was identified at the
316
excitation/emission wavelengths (Ex/Em) of 250/330 nm (Peak A), while the second
317
main peak was observed at Ex/Em 285/340 nm (Peak B). The third main peak was
318
observed at Ex/Em 250/375 nm (Peak C). Peak A has been reported as an aromatic
319
protein-like peak [40-42]. Peak B has been described as an SMP-like (soluble
320
microbial product) peak, in which the SMP substance includes tyrosine-, tryptophan-
321
and protein-like components, and Peak C has been described as a fulvic acid-like peak
322
[41,43]. When compared to those in Fig. 5a, the EEM fluorescence spectra shown in
323
Fig. 5b indicate that the sludge treated using only H2SO4 experienced no changes in
324
the presence of Peak A, B, and C. In other words, the acidification of sludge did not
325
destroy EPS. In contrast, the EEM spectra (Fig. 5c) from sludge treated using the
326
Fenton process exhibited no Peak C. Because fulvicacid is always associated with
327
other substances in EPS, the absence of a fulvicacid peak demonstrated that the
328
Fenton process destroyed the EPS structure and increased the dewaterability of
329
sludge.
330
Zhang et al. [44] concluded that peracetic acid pre-oxidation has the ability to
331
destroy the EPS structure and the fulvicacid peak was also absent. Zhen et al. [45]
332
added the Fe(II)-activated persulfate in the activated sludge dewatering pretreatment
333
and got the absence of fulvicacid peak, and in addition, the soluble EPS and bound
334
EPS all have no fulvicacid after the pretreatment. Fluorescence parameters of the
335
spectra, including the excitation/emission wavelengths for the peaks, fluorescence
336
intensity, and the relevant substances emitting the spectral peaks are given in Table 4.
337
The EEM spectra indicate that the main organic substance in EPS was an
338
SMP(soluble microbial products)-like substance, and confirmed that the high activity
339
of microorganisms in the sludge in summer contributed to the unfavorable
340
dewaterability of the sludge.
341 342 343
3.3. FT-IR analysis The FT-IR spectra of the EPS extracted from the sludge and subjected to
344
various treatments were determined in triplicate and are illustrated in Fig. 6. The
345
FT-IR spectra of the drinking water treatment sludge is similar to those of waste
346
activated sludge [46,47]. The weak band at wave number 3,299 cm-1 has been
347
attributed to the O-H stretching vibration in hydroxyl functional groups [48]. The
348
bands at wave numbers 1,645 cm-1, 1,540 cm-1, and 1,457 cm-1 have been shown to
349
correspond to the protein secondary structure, namely C=O (amide I) and N-H (amide
350
II) [48,49]. These three responses indicate that proteins were one of the components
351
of the EPS. In addition, a clear peak at wave number 1,037 cm-1 exhibited the
352
characteristics similar to those of carbohydrates or carbohydrates-like substances [48],
353
suggesting that carbohydrates were present in the EPS of the drinking water treatment
354
sludge.
355
EPS are one of the predominant components in sludge matrixes, mainly
356
consisting of proteins (PN), polysaccharides (PS), lipids, etc. The nature of functional
357
groups present in the flocs may predominantly originate from the EPS compositions.
358
The transmittance of band at 1,457 and 1,037 cm-1 changed in different conditioned
359
EPS spectra, suggesting that the degradation of EPS, which was also proved by the
360
analysis above for EEM.
361 362 363
3.4. SEM analysis To gain insights into the responsible dewatering mechanism in the sludge
364
treated using the Fenton process and H2SO4, SEM was used to examine the
365
micromorphology of the sludge subjected to various treatments (Fig. 7). Fig.7a shows
366
that the untreated sludge matrix was compact and comprised of sludge flocs arranged
367
in sheets or flakes. After treatment using H2SO4, sludge appeared (Fig. 7b) to consist
368
of much more pores and the orientation of the sludge flocs was broken. Because a
369
porous structure in sludge favors the discharge of pore water, Guo et al. [22]
370
concluded that the acidification of sludge broke the flocs of oil treatment sludge,
371
causing pores in the sludge that can improve the dewaterability.
372
The microstructure of sludge treated using the Fenton process (Fig. 7c) showed
373
no obvious differences from that of the untreated sludge. Perhaps some biological
374
detection methods would be helpful in examining the biomass in the sludge and might
375
be able to identify differences. The combination of H2SO4 and the Fenton process as a
376
treatment created more pores in the sludge (Fig. 7d) and the surface of sludge
377
appeared coarser and much less regular than that of untreated sludge. Zhen et al. [50]
378
concluded that the condition of Fe2+ with persulfate oxidation could break sludge
379
flocs to create many pores in the sludge, and break EPS to improve the dewaterability
380
of the waste activated sludge. Notably in the present study, the sludge treatment that
381
combined the Fenton process with H2SO4 pretreatment resulted in the destruction of
382
sludge flocs and the degradation of EPS, so that pore water, surface combined water
383
and internal combined water in the EPS were easily released by the electric field force
384
of the electro-osmosis dewatering device.
385 386 387
3.5. Size distribution The size distribution of sludge particles is a critical factor influencing sludge
388
dewaterability. The relevant determinations of particle sizes were conducted in
389
triplicate, but for clarity, only one of the size distributions for each sludge treatment is
390
shown in Fig. 8. The size distribution in the raw sludge was multimodal, with an
391
average particle size of 3,000 µm. In contrast, the size distribution of treated sludge
392
(regardless of treatment) was unimodal, and the result indicated the sludge was
393
formed under different hydraulic conditions, such as flocculant usage and flocculation
394
time in the flocculation basin. The largest proportion of particles in the sludge treated
395
using H2SO4, using H2SO4 in combination with the Fenton process, and using the
396
Fenton process only was 121, 1,581 and 2,985 nm, respectively. These results
397
revealed that acidified sludge consisted of a high proportion of small particles because
398
the acid broke the sludge flocs and created a small sludge matrix. This was also
399
evidenced in the SEM analysis. The average size of particles in the sludge treated
400
using the Fenton process alone was slightly different from that in the untreated sludge.
401
These results indicated that the Fenton process slightly destroys the drinking water
402
treatment sludge flocs, consistent with observations from the SEM analysis.
403 404 405
4. Conclusions The Fenton-like sludge dewatering process combined with acidification was
406
demonstrated to be efficient in enhancing the dewaterability of drinking water
407
treatment sludge. The optimum operating parameters (Fe3+ dosage of 54 mgg-1 sludge,
408
H2O2 dosage of 87 mgg-1 sludge and pH of 6.3) resulted in the final moisture content
409
of dewatered sludge being decreased to 63%. Both EEM and FT-IR showed that the
410
Fenton-like process led to the destruction of EPS, which in turn improved the sludge
411
dewaterability and increased the proportion of internal combined water that was
412
changed to free water. The SEM and particle size distribution analyses indicated the
413
acidification of sludge had the effect of breaking flocs and creating many pores in the
414
treated sludge, both of which improved dewaterability.
415
416
Acknowledgements
417
This work was supported by the National Natural Science Foundation of
418
China(51308185), Central University business expenses (2013B32314), A Project
419
Funded by the Priority Academic Program Development of Jiangsu Higher Education
420
Institutions and National major water projects (2014ZX07405002). We would like to
421
thank four anonymous reviewers for their constructive comments on the work.
422 423 424 425 426 427 428 429 430 431 432 433 434 435 436 437 438 439 440 441 442 443 444 445 446 447 448 449 450 451 452 453 454 455 456 457 458 459 460 461 462 463 464
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for pulp mill wastewater treatment using a combination of uniform design and response surface methodology, Water Res. 45. (2011) 5633-5640. [35] D. He, L. Wang, H. Jiang, H. Yu, A Fenton-like process for the enhanced activated sludge dewatering, Chem. Eng. J. 272. (2015) 128-134. [36] M. Citeau, O. Larue, E. Vorobieu, Influence of salt, pH and polyelectrolyte on the pressure electro-dewatering of sewage sludge, Water Res. 45. (2011) 2167-2180. [37] Z. Li, Y. Tian, Y. Ding, H. Wang, L. Chen, Contribution of extracellular polymeric substances (EPS) and their subfractions to the sludge aggregation in membrane bioreactor coupled with worm reactor, Bioresource Technol. 144. (2013) 328-336. [38] F. Orvain, K. Guizien, S. Lefebvre, M. Bréret, C. Dupuy, Relevance of macrozoobenthic grazers to understand the dynamic behaviour of sediment erodibility and microphytobenthos resuspension in sunny summer conditions, J. Sea Res. 92. (2014) 46-55. [39] M. Huo, G. Zheng, L. Zhou, Enhancement of the dewaterability of sludge during bioleaching mainly controlled by microbial quantity change and the decrease of slime extracellular polymeric substances content, Bioresource Technol. 168. (2014) 190-197. [40] Y. Yamashita, E. Tanoue, Chemical characterization of protein-like fluorophores in DOM in relation to aromatic amino acids, Mar. Chem. 82. (2003) 255-271. [41] W. Chen, P. Westerhoff, J.A. Leenheer, K. Booksh, Fluorescence excitation - Emission matrix regional integration to quantify spectra for dissolved organic matter, Environ. Sci. Technol. 37. (2003) 5701-5710. [42] A. Baker, R. Inverarity, Protein-like fluorescence intensity as a possible tool for determining river water quality, Hydrol. Process. 18. (2004) 2927-2945. [43] P.G. Coble, Characterization of marine and terrestrial DOM in seawater using excitation-emission matrix spectroscopy, Mar. Chem. 51. (1996) 325-346. [44] W. Zhang, B. Cao, D. Wang, T. Ma, H. Xia, D. Yu, Influence of wastewater sludge treatment using combined peroxyacetic acid oxidation and inorganic coagulants re-flocculation on characteristics of extracellular polymeric substances (EPS), Water Res. 88. (2016) 728-739. [45] G. Zhen, X. Lu, B. Wang, Y. Zhao, X. Chai, D. Niu, A. Zhao, Y. Li, Y. Song, X. Cao, Synergetic pretreatment of waste activated sludge by Fe(II)–activated persulfate oxidation under mild temperature for enhanced dewaterability, Bioresource Technol. 124. (2012) 29-36. [46] J. Laurent, M. Casellas, H. Carrere, C. Dagot, Effects of thermal hydrolysis on activated sludge solubilization, surface properties and heavy metals biosorption, Chem. Eng. J. 166. (2011) 841-849. [47] O. Gulnaz, A. Kaya, S. Dincer, The reuse of dried activated sludge for adsorption of reactive dye, J. Hazard. Mater. 134. (2006) 190-196. [48] Z. Wang, Z. Wu, S. Tang, Extracellular polymeric substances (EPS) properties and their effects on membrane fouling in a submerged membrane bioreactor, Water Res. 43. (2009) 2504-2512. [49] T. Maruyama, S. Katoh, M. Nakajima, H. Nabetani, T.P. Abbott, A. Shono, K. Satoh, FT-IR analysis of BSA fouled on ultrafiltration and microfiltration membranes, J. Membrane Sci. 192. (2001) 201-207. [50] G. Zhen, X. Lu, B. Wang, Y. Zhao, X. Chai, D. Niu, A. Zhao, Y. Li, Y. Song, X. Cao, Synergetic pretreatment of waste activated sludge by Fe(II)–activated persulfate oxidation under mild temperature for enhanced dewaterability, Bioresource Technol. 124. (2012) 29-36.
553 554
Table 1
555 556
557 558
Table 1. Characteristics of drinking water treatment sludge Moisture content pH Protein Zeta potential (%) mg/L (mv) 87.47±0.09 7.73±0.15 108.3±27.2 -12.6±1.1
Average particle size(µm) 2860 ± 217
559
Table 2
560 561
Table 2. Element content of the drinking water treatment sludge Element Content
562 563
C (%)
Al (%)
Si (%)
K (%)
Fe (%)
O (%)
45.97±1.12 14.99±1.29 19.62±0.61 1.47±0.88 3.8±0.20 45.97±0.31
564
Table 3
565 566
567 568
Table 3 Experimental design matrix and the observed responses NO. Independent Variables Water removal EXP. efficiency, % 3+ X1 (H2O2 X2 (Fe X3 (pH) Experimental Predicted dosage, dosage, Y1 Y2 mgg-1 mgg-1 sludge) sludge) 1 10 10 1.3 18 13 2 10 30 3.2 35 34 3 10 50 5.3 38 50 4 10 70 7.6 38 37 5 30 10 1.3 41 25 6 30 30 3.2 50 48 7 30 50 5.3 52 50 8 30 70 7.6 48 58 9 50 10 1.3 45 24 10 50 30 3.2 47 46 11 50 50 5.3 49 62 12 50 70 7.6 42 39 13 70 10 1.3 41 28 14 70 30 3.2 44 33 15 70 50 5.3 47 57 16 70 70 7.6 43 40
569
Table 4
570 571
572 573
Table 4 Fluorescence spectral parameters of the extracellular polymeric substances Peaks Ex/Em (nm) Intensity Emitting substances Peak A 225/330 1,204 aromatic protein-like Peak B 285/340 1,673 soluble microbial product-like Peak C 225/400 932 fulvicacid-like
574
Figure Captions
575 576 577 578 579 580 581 582 583 584 585 586 587 588 589 590 591 592 593 594 595 596 597 598 599 600 601 602 603 604 605 606 607 608
Fig. 1. Vacuum electro-osmosis dewatering apparatus: (1) DC power supply; (2) sludge tank; (3) anode; (4) cathode; (5) four-way pipe; (6) vacuum collecting tank; (7) vacuum pump; and (8) vacuum pressure controller. Fig. 2. The universal analytical method used to determine the distribution of moisture in sludge. Fig. 3. Three-dimensional surface plots of water removal efficiency: (a) H2O2 dosage vs. Fe3+; (b) pH vs. H2O2 dosage; and (c) Fe3+ dosage vs. pH. Fig. 4. Moisture distribution in drinking water treatment sludge: (a) drying curves; and (b) moisture distribution in the sludge subjected to various treatments. Fig. 5. Excitation-emission matrix fluorescence spectra of the extracellular polymeric substances in: (a) untreated sludge; (b) sludge treated using H2SO4; and (c) sludge treated using the Fenton process. Fig. 6. Fourier-transform infrared spectra of extracellular polymeric substances from drinking water treatment sludge subjected to different treatments: (a) untreated sludge (black line); (b) sludge treated using H2SO4 (red line); (c) sludge treated using the Fenton process (blue line); and (d) sludge treated using H2SO4 in combination with the Fenton process (magenta line). Fig. 7. Scanning electron microscopy images of drinking water treatment sludge subjected to various treatments: (a) untreated sludge; (b) sludge treated using H2SO4; (c) sludge treated using the Fenton process; and (d) sludge treated using H2SO4 in combination with the Fenton process. Fig. 8. Size distribution of particles in drinking water treatment sludge subjected to different treatments: (a) untreated sludge (red line); (b) sludge treated using H2SO4 (black line); (c) sludge treated using the Fenton process (blue line); and (d) sludge treated using H2SO4 in combination with the Fenton process (magenta line).
609 610 611 612 613
Fig. 1. Vacuum electro-osmosis dewatering apparatus: (1) DC power supply; (2) sludge tank; (3) anode; (4) cathode; (5) four-way pipe; (6) vacuum collecting tank; (7) vacuum pump; and (8) vacuum pressure controller.
614 615 616 617
Fig. 2. The universal analytical method used to determine the distribution of moisture in sludge.
(b)
(a) 70
70 60
(%) water removal
50
40 30 20 10
40 30 20
60
70
107
30
O
3
2
40
5
pH
20
2
20
20
e
Fe 3+ do sag e
g sa do O2 H2
40
60
50
6
40
60
2
0
H
Water removal
50
4
(%)
60
1
0
618
10
(c) 35
25
wat er removal
(%)
30
20 15 10 7
750
6
0
5 0
4
pH
5
60
4
3
30
Fe 20
2
620 621 622
1
10
619
Fig. 3. Three-dimensional surface plots of water removal efficiency: (a) H2O2 dosage vs. Fe3+; (b) pH vs. H2O2 dosage; and (c) Fe3+ dosage vs. pH.
35
untreated H2SO4+Fenton
-0.0001
H2SO4
drying speed (mg/s)
-0.0002
Fenton second speed second speed drop drop
-0.0003
second speed drop
-0.0004 -0.0005
second speed drop
-0.0006
first speed drop
first speed drop
first speed drop
first speed drop
-0.0007
623 624 625 626
0
20
40
60
moisture content (%)
80
100
(b)
mass of water per mass of sludge solid
0.0000
(a)
free water pore water surface adhesion water internal combined water
30
25 20 15 10
5 0 H2SO4+ Fenton
untreated
H2SO4
Fenton
Fig. 4. Moisture distribution in drinking water treatment sludge: (a) drying curves; and (b) moisture distribution in the sludge subjected to various treatments.
450
450 0.000 240.0 480.0 720.0 960.0 1200 1440 1680 1920 2160 2400 2640 2880 3120 3360 3600 3840 4080 4320 4560 4800 5040 5280 5520 5760 6000
400
EX(nm)
350
Peak B
300
250
Peak A
Peak C
200 200
627
250
300
350
400
(b)
350
Peak B
300
250
Peak C
Peak A 200 200
450
EM(nm)
0.000 240.0 480.0 720.0 960.0 1200 1440 1680 1920 2160 2400 2640 2880 3120 3360 3600 3840 4080 4320 4560 4800 5040 5280 5520 5760 6000
400
EX（ nm）
(a)
250
300
350
400
450
EM（ nm）
450 0.000 240.0 480.0 720.0 960.0 1200 1440 1680 1920 2160 2400 2640 2880 3120 3360 3600 3840 4080 4320 4560 4800 5040 5280 5520 5760 6000
(c) 400
EX(nm)
350
300
Peak B
250
Peak A 200 200
628 629 630 631 632
250
300
350
400
450
EM(nm)
Fig. 5. Excitation-emission matrix fluorescence spectra of the extracellular polymeric substances in: (a) untreated sludge; (b) sludge treated using H2SO4; and (c) sludge treated using the Fenton process.
100
0
1037
3299
1654 1540 1457
40
60
treated by Fenton treated by H2SO4+Fenton
20
Transmittance(%)
80
raw sludge treated by H2SO4
4000
633 634 635 636 637 638 639
3500
3000
2500
2000
1500
1000
500
Wavenumbers (cm-1)
Fig. 6. Fourier-transform infrared spectra of extracellular polymeric substances from drinking water treatment sludge subjected to different treatments: (a) untreated sludge (black line); (b) sludge treated using H2SO4 (red line); (c) sludge treated using the Fenton process (blue line); and (d) sludge treated using H2SO4 in combination with the Fenton process (magenta line).
640
641 642 643 644 645 646
Fig. 7. Scanning electron microscopy images of drinking water treatment sludge subjected to various treatments: (a) untreated sludge; (b) sludge treated using H2SO4; (c) sludge treated using the Fenton process; and (d) sludge treated using H2SO4 in combination with the Fenton process.
H2SO4
25
untreated Fenton H2SO4+Fenton
intensity (%)
20
15
10
5
0 0
647 648 649 650 651
1000
2000
3000
4000
5000
6000
7000
size (µm)
Fig. 8. Size distribution of particles in drinking water treatment sludge subjected to different treatments: (a) untreated sludge (red line); (b) sludge treated using H2SO4 (black line); (c) sludge treated using the Fenton process (blue line); and (d) sludge treated using H2SO4 in combination with the Fenton process (magenta line).