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Page 1 ofEnvironmental 36 Science & Technology
ACS Paragon Plus Environment
Environmental Science & Technology
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1
Understanding mechanisms of synergy between
2
acidification
3
activated sludge dewatering: From bench to pilot‒
4
scale investigation
and
ultrasound
treatments
for
5 6
Mei‒Qiang Cai†, Jian‒Qiang Hu†, George Wells‡, Youngwoo Seo⊥, Richard Spinney§,
7
Shih-Hsin Ho#, Dionysios D. Dionysiou║, Jie Su†, Ruiyang Xiao∇,*, and Zongsu
8
Wei†,○,*
9 10
†
11
Hangzhou, 310018, China
12
‡
13
Evanston, Illinois, 60208, U.S.A.
14
⊥
15
Ohio, 43606, U.S.A.
16
§
17
Ohio, 43210, U.S.A.
18
#
19
Technology, Harbin, 150090, China
20
║
21
Cincinnati, Ohio, 45221, U.S.A.
22
∇Institute
23
Central South University, Changsha, 410083, China
24
○
25
and Environmental Engineering, University of California, Los Angeles, California,
26
90095, U.S.A.
School of Environmental Science and Engineering, Zhejiang Gongshang University, Department of Civil and Environmental Engineering, Northwestern University, Department of Civil and Environmental Engineering, University of Toledo, Toledo,
Department of Chemistry and Biochemistry, The Ohio State University, Columbus, State Key Laboratory of Urban Water Resource and Environment, Harbin Institute of Environmental Engineering and Science Program, University of Cincinnati, of Environmental Engineering, School of Metallurgy and Environment,
Laboratory for the Chemistry of Construction Materials (LC2), Department of Civil
27 28 29 30
* To whom correspondence should be addressed. R.X. Phone: +86‒731‒88830511; fax: +86‒731‒88710171; email address:
[email protected]; Z.W. Phone: +1‒213‒ 705‒8331; fax: +1‒310‒206‒2222; email address:
[email protected]. 1 ACS Paragon Plus Environment
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ABSTRACT
32
Enhancing activated sludge dewaterability is of scientific and engineering
33
importance in the face of accelerated urbanization and stringent environmental
34
regulations. In this study, we investigated the integration of acidification and
35
ultrasound (A/US) treatments for improving sludge dewaterability at both bench‒ and
36
pilot‒scales. Our results showed that the A/US system exhibited significantly
37
improved sludge dewatering performance, characterized by capillary suction time,
38
cake moisture, and water/solid content of sludge cake. Synergistic dewatering
39
mechanisms were elucidated with a suite of macro and spectroscopic evidence.
40
Characterization of treated sludge revealed that US‒induced thermal, mechanical
41
shearing force, and radical oxidation effects disrupted floc cells and accelerated the
42
decomposition of extracellular polymeric substances (EPS), releasing bound water
43
into the bulk phase. In addition to enhancing hydrolysis of EPS, the acidic pH
44
environment caused the protonation of functional groups on EPS, facilitating the
45
reflocculation of US decomposed sludge for improved filterability. Our bench‒scale
46
and pilot‒scale investigations provide a mechanistic basis for better understanding of
47
the A/US process, and enable development of a viable and economical dewatering
48
technology.
49 50 51 52 53
Keywords: activated sludge; dewaterability; ultrasound; acidification; extracellular
54
polymeric substances
2 ACS Paragon Plus Environment
Environmental Science & Technology
55
INTRODUCTION
56
Addressing the environmental concern of excess sludge is a key challenge in
57
municipal wastewater treatment plants (MWTP) for populated urban areas.1 Sludge is
58
a complex colloidal system in which highly dispersed and fine particles around 1 µm
59
form stable suspension.2 Water in sludge can be classified into free water, interstitial
60
water, vicinal water, and hydration water.3, 4 The bound water typically refers to the
61
sum of interstitial, vicinal, and hydration waters that are arduous to remove, because
62
they are tightly bound to extracellular polymeric substances (EPS) and other sludge
63
components stick by adhesive forces and/or chemical bonds.5
64
During sewage sludge treatment, the dewatering process is of practical
65
importance in reducing water content and sludge volume for the purpose of
66
transportation and safe disposal.6 Usually, thickened sludge is conditioned through
67
physical disruption and chemical addition (e.g., acid, ferric chloride, and lime),
68
followed by mechanical dewatering techniques (e.g., press filters and centrifuges).
69
The cost of such dewatering processes represents a significant amount of the total
70
operational cost in typical MWTPs. However, water content in filtered sludge cake
71
still ranges from 75 to 85% (w/w), failing to meet the increasingly stringent criteria of
72
subsequent disposal.7,
73
challenge. Thus, other processes have been implemented to improve the performance
74
of sludge dewatering. For example, Liu et al. reported that microwave irradiation
75
improved dewaterability, which is defined as “the ability of a sludge to be
76
concentrated into a more manageable and less voluminous form”,9 by decomposing
77
both EPS and microbial cells in the sludge.10 Although the dewaterability has been
78
suggested to be associated with the changes in soluble EPS (S‒EPS), loosely bound
79
EPS (LB‒EPS), and tightly bound EPS (TB‒EPS),3 there is no consistent correlation
8
Further reducing bound water from EPS still remains a
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between EPS fractions and sludge
81
technologies.11
dewaterability for different treatment
82
Ultrasound (US), an emerging technology used in the sludge disintegration
83
process for enhanced digestion, is recognized as an efficient approach for improving
84
dewaterability.12, 13 US with a series of compression and rarefaction cycles generates
85
cavitation bubbles in water. The cavitation bubbles implode and yield localized
86
temperatures (as high as 5000 K), mechanical shear forces (i.e., shock‒waves and
87
micro‒jets), and free radicals (e.g. •H, •OH and HO2• ).14 The mechanical shearing
88
effect causes destruction of microbial cells and release of LB‒EPS and TB‒EPS.15
89
Thermolysis and radical oxidation of the released organics free EPS‒bound water into
90
bulk phase, thereby enhancing sludge dewaterability.12, 16-18 Nevertheless, a high US
91
energy input was observed to deteriorate sludge dewaterability, since intensive US
92
waves disrupt large flocs into smaller sizes with high surface area, resulting in an
93
increase of bound water through re‒adsorption. Previous studies also suggested that a
94
thin layer of released intracellular substances formed on the new particle surface acts
95
as a barrier against water release, consequently reducing the dewaterability.12, 19, 20
96
However, these dewatering mechanisms were proposed based on speculation and need
97
to be experimentally confirmed. Indeed, many fundamental questions regarding the
98
US dewatering mechanisms still remain unanswered: How does radical oxidation
99
quantitatively contribute to the overall sludge dewaterability? Does US‒induced
100
thermolysis play a significant role in the dewatering process? How can the mechanical
101
shearing force effect be controlled to prevent the deterioration of sludge
102
dewaterability? Such mechanistic and fundamental questions have important
103
implications for application of US dewatering process in engineering practice.
104
Pretreatments of sludges under acidic conditions can be beneficial to further 4 ACS Paragon Plus Environment
Environmental Science & Technology
105
improve US dewatering performance.11, 19, 21, 22 At low pH, protonation of functional
106
groups on EPS reduces the electrostatic repulsion, resulting in enhanced sludge
107
aggregation and subsequent improvement of dewaterability.3, 23 Acidic environments
108
also favor hydrolysis of EPS.10, 19 For example, Liu et al. reported that the bound
109
water content was reduced 50% with a decrease of capillary suction time (CST) from
110
37.7 to 9.2 s in a microwave‒acid treatment.10 In addition, by characterizing organic
111
components in EPS matrix, Xiao et al. successfully demonstrated that US followed by
112
acidification is more effective in enhancing sludge dewaterability than acidification
113
alone.11 Based on previous studies, the combination of US and acidification
114
potentially integrates complementary roles for these two discrete technologies. But
115
US prior to acidification treatment may overlook the fact that the application of US in
116
advance can cause deterioration of sludge dewaterability,12 and the subsequent
117
acidification step cannot yield optimal performance (see Text S1 and Figure S1 in the
118
supporting information, SI). Therefore, further efforts are required to optimize the
119
synergism between US and acidification treatments, such as the sequence of applying
120
acidification and improvement of US devices.
121
Unfortunately, there are limited engineered investigations for sludge dewatering
122
processes due to the lack of industrial investment into current technologies. Full‒scale
123
treatment necessitates fulfilling the following requirements: 1) operations are simple
124
and effective; 2) dewatering units are compatible with filtration devices; and 3)
125
dewatering processes must feature low cost, operational stability, and minimal
126
chemical use. Therefore, in this study, we investigated the integration of acidification
127
and US (A/US) treatments for enhancing sludge dewaterability at both bench‒scale
128
and pilot‒scale. The objectives of the present study were to (1) investigate the effect
129
of A/US treatments on sludge dewatering performance; (2) scale up the benchtop 5 ACS Paragon Plus Environment
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experiments to pilot‒scale with high‒power US units for testing engineering
131
feasibility; and (3) provide mechanistic insights into the synergism between US and
132
acidification processes. Our bench‒scale and pilot‒scale investigations will provide a
133
mechanistic basis for better understanding the A/US process and enable development
134
of a viable and economical dewatering technology.
135
136
MATERIALS AND METHOD
137
Sewage sludge and reagents
138
The waste activated sludge was collected from in a MWTP in Hangzhou, China.
139
The plant treats approximately 600,000 m3 municipal wastewater daily by an
140
anaerobic/anoxic/oxic process. Sludge samples were stored at 4 °C before use. All
141
samples were kept for a maximum of two days. The characteristics of raw sewage
142
sludge are summarized in Table S1 of the SI.
143
Information for all the chemicals used in this study such as purity and
144
manufacturer is tabulated in Table S2. H2SO4 and NaOH were used to adjust pH of
145
sludge samples.
146 147
Sludge conditioning and dewatering
148
Bench‒scale experiments
149
The bench‒scale tests for sludge conditioning and dewatering were conducted in
150
batch mode using a 500 mL cylindrical flask (12 cm in depth, 9 cm in I.D., and 763
151
mL in volume). Samples of activated sludge (300 mL) were first added into the flasks. 6 ACS Paragon Plus Environment
Environmental Science & Technology
152
The pH of the sludge samples was then adjusted to the designated values (2 ~ 7) by
153
adding 5M H2SO4 or NaOH slowly while being continuously stirred. Sludge pH was
154
measured by submerging a PHS‒3C pH meter (Yoke, China) into the sludge slurry.
155
After the pH of sludge stabilized (~ 30 s), US irradiation was applied using a high
156
amplitude power US probe on the basis of the barbell horn ultrasound (BHU)
157
technology. This unique design has been shown to be reliable, and can be directly
158
scaled up to pilot‒scale sonochemical processes without reducing the US amplitude.24
159
The US probe has a maximum power of 3 kW and a maximum cross‒sectional output
160
diameter about 40 mm at a frequency of 20 kHz (TJS‒300, Hangzhou Success
161
Ultrasonic Equipment, China). Detailed design of the US horn is illustrated in Figure
162
1a. Bench‒scale experiments were carried out at different US power densities (0.0 to
163
10.0 W mL‒1) and durations (0 to 2 min). The selection of sonication duration assured
164
low energy consumption because treatment tests were short. The titanium probe was
165
immersed at 1 cm below solution surface. A HAAKE A80 cooling bath and a
166
thermocouple connected to a digital thermometer (DS18B20, Dallas Crop, U.S.) were
167
used to maintain the temperature of sonicated sludge samples in the flask at 25 ± 2 °C.
168
A set of control experiments was conducted to quantitatively evaluate the contribution
169
of thermal, radical oxidation and shear force effects to sludge dewatering. Specifically,
170
the change of cake moisture, UV254 and CST were examined under four different
171
scenarios, i.e., US with temperature control (scenario A), US without temperature
172
control (scenario B), US with t‒butanol25, 26 as radical (i.e., •OH) scavenger (scenario
173
C), and acidification at elevated temperature (scenario D). To check the potential
174
influence of t‒butanol on cavitation,27-29 we conducted control experiments with and
175
without the scavenger reagent. The result show thats addition of 50 mM t‒butanol
176
only caused ~3% variation for cake moisture and ~5% for CST value (Figure S2). 7 ACS Paragon Plus Environment
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Therefore, the concentration of 50 mM was used in our experiments to minimize the
178
potential interference exerted by t-butanol. All experiments were conducted in
179
duplicate or triplicate.
180 181
Pilot‒scale treatment
182
An online pilot‒scale experiment at the MWTP was used to test and verify the
183
findings from bench‒scale experiments. This upscaling effort enables us to evaluate
184
the viability of the A/US process and estimate the economic benefits. As illustrated in
185
Figure 1b, the pilot sludge conditioning and dewatering system was operated in a
186
flow‒through mode. The pilot system consists of a 3000 L conditioning tank, a BHU
187
system, and a spring filter press (Zhejiang OuKeMei Filtration Equipment, China).
188
The activated sludge was first treated with H2SO4 to achieve pH 3 or 5 with the
189
resident time of 4.5 h. Next, the conditioned sludge was pumped to the BHU system
190
by an air pump with a flow rate of 25 L min‒1, and then into a neutralization tank
191
where sludge pH was adjusted to 7 for corrosion protection. Finally, the conditioned
192
sludge was pumped into a pressure filtration. The dewatering of conditioned sludge in
193
the pressure filtration system consisted of two phases: a 70 min feeding phase with a
194
pressure of 1.5 MPa and a 30 min pressing phase with a pressure of 1.6 MPa. A
195
polypropylene filament filter cloth in dimension of 1640 × 1640 mm2 was used in the
196
pressure filtration system.
197
Analytical methods
198
Sludge Dewaterability Tests
199
Sludge dewaterability was assessed by capillary suction time (CST) value, 8 ACS Paragon Plus Environment
Environmental Science & Technology
200
moisture content, bound water content, and total solids content.11, 30, 31 Specifically, 5
201
mL of conditioned sludge suspension was sampled from the cylindrical reactor and
202
measured in a CST instrument (Type 304M, Triton Electronics Ltd, UK) with an 18
203
mm inner radius and a Whatman No. 17 chromatography grade filtration paper.
204
Likewise, sludge dewatering was performed by the vacuum filtration, during which
205
100 mL conditioned sludge was poured into a 9 cm standard Buchner funnel fitted
206
with 1.2 µm pre-wetted Whatman filter paper under a constant vacuum pressure of
207
34.5 kPa until all filtrate was removed. The dewatered sludge “cake” on filter paper
208
was dried at 105 °C for 24 h and weighed to determine water and solid contents of
209
sludge samples according to Standard Method 2540‒B of “Total Solids Dried at 103‒
210
105 °C”.30 Bound water content measurement was determined using a differential
211
scanning calorimetry (DSC) analyzer equipped with a liquid nitrogen cooling system
212
(Diamond, PerkinElmer). An approximate of 8 mg sludge sample was retrieved and
213
placed into the crucible of a special mini-oven. The temperature was first cooled to –
214
20 °C, at which free water in the sample was frozen, and then was increased to 10 °C
215
at a rate of 2 °C min‒1. Details for the DSC analytical procedure can be found in
216
Katsiris and Kouzeli-Katsiri (1987)32 and Zhang et al. (2014).33 The amount of free
217
water (Wf) was calculated by the heat absorption determined by integrating the
218
endothermic curve area. Then, the bound water content was calculated as the
219
difference between the known total water of the sludge sample and the amount of free
220
water:
221
Wb =Wt –∆H/∆H0
(1)
222
where Wb and Wt are the bound water content and total water content of the sludge
223
samples, respectively; ∆H and ∆H0 is the amount of energy absorbed by sludge and
224
the standard melting heat of ice (345.4 J g‒1), respectively. 9 ACS Paragon Plus Environment
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EPS extraction and analysis
226
A heat extraction method was modified to extract different EPS fractions from
227
waste activated sludge.34, 35 Sludge suspensions were first centrifuged at 3000×g and
228
4 °C for 15 min to collect the supernatant fluids containing S‒EPS. Then, solids at the
229
bottom of the centrifuge tubes were re‒suspended to the original volume by adding
230
0.05% (w/w) NaCl solution and mixed using a vortex mixer (VA08G1‒24, Mo Bio
231
Labs, U.S.) for 1 min at 50 °C. The suspension was then immediately centrifuged at
232
4000×g and 4 °C for 10 min to remove any suspended biomass left over in the
233
supernatant. The supernatant fluids containing LB‒EPS were collected. The solids
234
were again re‒suspended to the initial volume with 0.05% (w/w) NaCl and held at
235
60 °C for 15 min to extract TB‒EPS. The supernatant containing TB‒EPS were
236
obtained by centrifuging the suspensions at 4000×g and 4 °C for 15 min. All
237
supernatant fluids were filtered through a 0.45 µm cellulose acetate membrane filter
238
(Whatman) and stored at 4 °C before analysis.
239
Three‒dimensional excitation emission matrix (3D‒EEM) spectra of extracted
240
EPS were measured on a F‒4600 fluorescence spectrophotometer (Hitachi, Japan)
241
with an excitation range from 200 to 400 nm and an emission range from 280 to 500
242
nm. The spectra were recorded at a scanning speed of 12,000 nm min‒1 using
243
excitation and emission slit widths of 10 nm. Each scan had 37 emission and 27
244
excitation wavelengths. The protein content in extracted EPS was determined with the
245
Lowry–Folin method.36 It should be noted that the presence of humic like substances
246
from the sludge samples could possibly yield error in the quantification level of
247
proteins estimated by the Lowry‒Folin method.37 Polysaccharide concentrations were
248
measured with the anthrone method using glucose as standard.38 Supernatant pH was
249
adjusted to 7 prior to analysis to exclude pH interference. 10 ACS Paragon Plus Environment
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Sludge flocs characterization
251
The surface morphology of different sludge samples was characterized by
252
scanning electron microscope (SEM; S‒4800, Hitachi, Japan). Optical microscopic
253
pictures were also obtained using an optical microscope (Leica DM4 B) equipped
254
with a digital camera (Moticam Pro. 205A). ZetaSizer Nano ZS (Malvern, UK) was
255
used for measuring Zeta potential of sludge flocs, while a MasterSizer 2000 equipped
256
with a Hydro2000MU auto sampler (Malvern, UK) was applied for measuring the
257
sludge floc size. The dissolved organic carbon (DOC) in supernatant was analyzed
258
using a total organic carbon (TOC) analyzer (TOC‒VCPN 5000A, Shimadzu, Japan).
259
UV254, which is a measure of aromatics, was determined by a UV/Vis
260
spectrophotometer (UV2600, Purkinje, China). Both DOC and UV254 were measured
261
after filtration through 0.45 µm acetate fiber membranes. The biochemical oxygen
262
demand (BOD) and soluble chemical oxygen demand (COD) in the supernatant of
263
sludge samples were measured with a BOD meter (2173B, Hach, USA) and a
264
spectrophotometer (DR/2000, Hach, USA), respectively. To measure heavy metal
265
concentrations, sludge was initially filtered and the dried solid cake was digested in a
266
Teflon vessel using a mixture of HNO3, HCl, and HF for 4 h and diluted at least 10
267
times
268
spectroscopy (ICP-AES, AAnalyst 800, Perkin Elmer Inc., USA) was used to measure
269
the typical inorganic elements (i.e., Cu, Zn, Cd, Pb, Mn, Cr, Ni, Al, Ca, and Mg) in
270
the raw sludge and treated sludge samples according to APHA method 3030I.
271
Duplicate samples were prepared and treated for the heavy metal analysis. We
272
concluded that the levels of organic matter and heavy metals in the recycled filtrate do
273
not exhibit a detrimental effect on the operation of the wastewater treatment plant (see
274
Text S2 and Table S3 in the SI).
with
deionized
water.
Inductively coupled
11 ACS Paragon Plus Environment
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emission
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275 276
RESULTS AND DISCUSSION
277
Characterization of dewaterability under different treatments
278
Batch experiments were performed to compare the dewatering performance by
279
US, acidification, and A/US. As illustrated in Figure 2a, the cake moisture was
280
reduced from 77.6±3.2% to 73.3±2.9%, 75.8±1.5% and 68.6±3.2% after US,
281
acidification and A/US treatments, respectively. The 9.0% moisture reduction by
282
A/US was higher than US (4.3%) or acidification treatment (1.8%) alone. The DSC
283
results show that the original sludge contained approximately 2.85± 0.12 g g‒1 DS
284
bound water content, which was similar to the value of 2.32 g g‒1 DS reported by
285
Qian et al. (2016)39 but lower than 6.3 g g‒1 DS by Lee and Lee (1995)40 and 11.52 ±
286
3.32 g‒1 DS by Feng et al. (2014)41. This discrepancy may possibly arise from the
287
sludge source. As shown in Figure 3, bound water content significantly decreased
288
from 2.85 ± 0.12 to 2.59 ± 0.13, 1.68 ± 0.10 and 0.82 ± 0.11 g g‒1 DS after US,
289
acidification, and A/US treatments, respectively. This result indicated the released
290
bound water can transform to free water and thus accelerate the dewatering process.
291
Figure 2a also demonstrates the sludge filterability reflected by CST change
292
under different treatments. The CST values by US treatment increased from 36.0 to
293
78.9 s, indicating a deteriorated filterability when sludge was treated with US alone.
294
This result is consistent with the previous findings.16, 42, 43 Chu et al. attributed the
295
deteriorated dewaterability by 20 min sonication to the breakdown of sludge flocs into
296
fine flocculi.16 In contrast, the CST value after acidification was observed to decrease
297
to 14.3 s, which is also consistent with the report by Zhang et al.18 With the
298
acidification treatment, the protonation of negatively charged functional groups is
299
thought to reduce intermolecular binding capacity and electrostatic interactions, 12 ACS Paragon Plus Environment
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300
resulting in destabilization and flocculation of sludge particles and ultimately a
301
decrease in CST value.3, 17, 23 By comparison, decreased CST and moisture content
302
indicated that the A/US treatment is capable of improving the sludge dewaterability
303
while still maintaining filterability. The A/US treatment integrated complementary
304
roles for these two discrete processes.
305
Figure 2a also shows reduced weight of sludge cake and total solids content after
306
US, acidification and A/US treatments. Figure 2b shows the effect of different
307
treatments on the release of organics, quantified by UV254 (for aromatic compounds)
308
and DOC of the supernatant in treated sludges. Student’s t‒test result suggests that
309
there is no significant difference in UV254 values of raw and acidification treated
310
sludges, as well as the US and acidification treated sludges. By comparison, DOC
311
values are significantly different by various treatments. Overall, DOC increased after
312
US, acidification, and A/US treatments. The increase of DOC is consistent with the
313
reduced weight of sludge cake and total solids content in Figure 2a.
314
Although acidification favored the reduction of sludge mass, the DOC and UV254
315
remained relatively stable, indicating that protonation-induced flocculation plays a
316
more profound role in sludge dewatering than acid‒assisted hydrolysis of organics.
317
US treatment also decreased sludge mass, but also led to a significant increase in
318
DOC and UV254 and an undesirable increase in CST. As expected, combined A/US
319
treatments provided the optimal sludge dewater performance compared to either
320
acidification or US alone, due the fact that this treatment integrates the advantages of
321
these two discrete processes. In particular, the decomposition of organic components
322
at acidic pH (evidenced by elevated UV254 and DOC in supernatant) was significantly
323
improved by US (Figure 2b), resulting in greater release of bound water (i.e., lower
324
cake moisture in Fig. 2a) relative to US or acidification alone. Further, the 13 ACS Paragon Plus Environment
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acidification promoted the flocculation of US destructed flocs and thus increased
326
sludge dewaterability, as evidenced by the low CST value and water content (Figure
327
2a).
328
Dewatering performance of A/US treatments
329
Effect of pH
330
To optimize A/US treatments, the effect of pH on sludge dewaterability was
331
investigated. As showed in Figure 4a, the cake moisture was reduced from 73.3±0.8%
332
to 68.5±0.8%, the total solid content from 0.70±0.2 g to 0.55±0.1 g, and the CST
333
value from 80.5±1.4 s to 15.5±2.2 s with a pH drop from 7 to 2, respectively. Figure
334
4b shows the zeta potential of sludge particles as function of pH under A/US
335
treatment. Zeta potential significantly increased from ‒29.1 to ‒6.3 mV as pH
336
decreased from 7 to 2 indicating a decrease in net surface charge on flocs. Sludge
337
particles are negatively charged due to the presence of carboxylate and phosphate
338
groups.3,
339
electrostatically repulsive, preventing destabilization and flocculation of sludge
340
particles.3, 23 But at lower pH, protonation of functional groups of EPS reduces the
341
surface charge as well as denatures the protein’s tertiary structure. This in turn reduces
342
solubility and increases protein aggregation, leading to flocculation of the colloidal
343
sludge, as evidenced by the floc size increase from d50% = 29.6±1.3 µm to 68.6±1.4
344
µm as pH decreased from 7 to 2 (Figure 4b). This observation supports
345
the idea that increased size of sludge flocs is important to improve the filterability of
346
activated sludge and consequently sludge dewaterability. Zhang et al. also observed a
347
maximal floc size around pH 3.18 In Figure 4b, the gradual increase of floc size from
7, 23
At higher pH, these negatively charged sludge particles are
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348
pH 3 to 2 was probably due to the more complete hydrolysis of the protein facilitated
349
by strong acid.18
350
To further characterize shifts in EPS fractions after A/US treatment, 3D‒EEM
351
fluorescence spectroscopy was employed. The effects of pH on 3D‒EEM spectra of
352
different EPS fractions in sludge flocs are presented in Figure 5. Five fluorescent
353
peaks were detected in the filtrate of sludge samples. Peaks A and B in the range of
354
excitation wavelength (Ex) < 250 nm and emission wavelength (Em) < 380 nm were
355
associated with proteins containing aromatic amino acids tryptophan and tyrosine
356
(Regions I and II), respectively.44-47 A decrease in their intensity can be attributed to
357
the increasing hydrolysis of the proteins at lower pH. Peak C in the range of Ex < 250
358
nm and Em > 380 nm (Region III) has been shown to be related to fulvic acid‒like
359
substances. Peak D in the range of 250 nm < Ex < 280 and Em > 430 nm (Region IV)
360
corresponded to soluble microbial products (SMP)‒like substances, and Peak E in the
361
range of 250 nm < Ex < 300 nm and Em < 380 nm (Region V) was characterized as
362
humic substances.44-47 Figure 5 clearly identified these five peaks in the 3D‒EEM
363
fluorescence spectra for the raw sludge, indicating the presence of the organic
364
fractions listed above in EPS. The fluorescence peaks located in different fluorescent
365
regions decreased significantly when pH was reduced to 3, and such pH influence of
366
on fractions of different EPS was summarized in Table S4. It was observed that the
367
cumulative fluorescent intensities for aromatic amino acids, SMP‒like substances,
368
humic‒ and fulvic‒like substances of S‒EPS and TB‒EPS decreased with a decrease
369
of pH from 7 to 3, whereas the reduction of LB‒EPS was not obvious until pH 3.
370
These results indicate that A/US treatment effectively decomposes the S‒EPS, LB‒
371
EPS and even TB‒EPS in the inner layer of the sludge flocs. Decomposition of EPS
372
in turn is likely linked to release of bound water, resulting in a decrease in sludge 15 ACS Paragon Plus Environment
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water content and improvement in sludge dewaterability, as demonstrated in Figure 2a
374
and 3. It should be noted that some destruction of S‒EPS and LB‒EPS by US was
375
evident at pH = 7 (i.e., without acidification). Further, acidification alone appeared not
376
to influence UV254 values in aqueous phase (Figure 2b). There are two possible
377
reasons to account for our observations. First, to be consistent with sonication tests,
378
only a 2 min treatment time was selected for the acidification process. Such a short
379
time probably did not allow sufficient release of organic matter from sludge flocs to
380
bulk solution. Second, under acidic conditions, release and hydrolysis of organic
381
matter occurred simultaneously. The comprehensive effects of release and degradation
382
of organic matter resulted in the constant DOC and UV254. Similarly, the LB‒EPS
383
release and TB‒EPS degradation by acidification were also observed simultaneously
384
by He et al. (2017).48 Wang et al. (2017)49 also reported unchanged polysaccharides
385
(PS) and humic-like substances (HS) in the total EPS with decreasing pH. Therefore,
386
the diminishment of EPS spectra from pH 7 to 3 was in general due to the presence of
387
US, indicating that US catalyzes the hydrolysis of EPS under acidic conditions.
388
Effect of US power density
389
To better constrain the operational cost of the A/US process, we investigated the
390
US power density (PD) applied to the waste activated sludge.13 The effect of PD on
391
cake moisture, water content, total solids (TS) content, and CST values was examined
392
to better characterize dewaterability, as shown in Figure 6a. The cake moisture was
393
reduced from 75.8±1.5% to 67.6±1.5% as PD increased to 2.5 W mL‒1, but then
394
increased with further increasing PD to 10 W mL‒1, consistent with sludge
395
dewaterability results in Feng et al.50 In Figure 6a, the CST value increased slightly as
396
PD increased 0 to 2.5 W mL‒1, and then sharply increased to 80 s with further PD 16 ACS Paragon Plus Environment
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397
increase. Figure 6b shows that the mean floc diameter of sludge fell from 110±3.2 to
398
65.8±4.1 µm as PD increased from 0 to 2.5 W mL‒1, and then decreased gradually
399
with further PD increase. The re‒adsorption of bound water on the newly formed
400
particle surface seemed not be favored, otherwise sharp changes in CST values or
401
cake moisture corresponding to floc size reduction, especially at PD < 2.5 W mL-1,
402
should be observed in Figure 6a. By comparison, the zeta potential remained
403
unchanged with increasing PD, suggesting that US power input had no impact on the
404
surface charge of sludge flocs.16
405
The 3D‒EEM spectra of EPS in the supernatant for treated sludge at different PD
406
values are illustrated in Figure S3. The fluorescence intensities of peaks located in
407
five fluorescent regions for different EPS fractions in Figure S3 are presented in Table
408
S5. For the S‒EPS, the fluorescence intensities of these peaks in 3D‒EEM spectra
409
decreased with a PD increase up to 2.5 W mL‒1, and then gradually increased as PD
410
continued to increase. EPS degradation by A/US treatments resulted in the diminished
411
3D‒EEM peaks at PD < 2.5 W mL‒1. Increasing 3D-EEM peaks at PD > 2.5 W mL-1
412
is likely due to release of more protein/amino acids, SMP, humic‒ and fulvic‒like
413
substances into supernatant. Similar trends were observed for LB‒EPS and TB‒EPS
414
fractions. The decomposition of organic fractions in the supernatant was also
415
observed after sonication at frequencies of 40, 68, and 160 kHz by Zhou et al.51
416
Interestingly, the fluorescence intensities of TB‒EPS at the highest PD (10 W ml-1)
417
were higher than raw sludge, suggesting an intensive destruction of microbial cells by
418
the A/US treatments. Overall, the varying EPS fractions with PD aligned well with the
419
trends of cake moisture and water content in sludge cake in Figure 6a. Such
420
corroboration indicates that the decomposition of EPS likely caused the release of
421
bound water to the bulk phase as free water. However, the increased EPS fractions at 17 ACS Paragon Plus Environment
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422
higher PD likely resulted in decreased sludge dewaterability,52, 53 thereby increasing
423
the CST values in Figure 6a. The fluorescence characteristics at different PD values
424
suggest a simultaneous release, dissolution and degradation of organics during the
425
A/US treatments, confirming the importance of EPS, especially protein, humic‒ and
426
fulvic‒like substances in the sludge dewatering process.
427
Dewatering Mechanisms
428
Role of acidification
429
EPS readily undergoes hydrolysis in acidic condition.10, 18 Thus, acids added to
430
the cell suspension can react with both EPS and cell wall components, leading to the
431
release of bound water and subsequently enhanced dewaterability.54 The results we
432
present for the role of acidification on sludge dewaterability can be related to
433
morphology and structure of flocs that consist of microorganisms and EPS.1 Figure 7
434
shows SEM micrographs of sludge before and after treatment. The raw sludge existed
435
in the form of filamentous structure with organic fibers and fine particles filling the
436
voids (Figure 7a). After acidification, the floc structure exhibited a relatively
437
smoother surface (Figure 7b). Optical microscopic pictures magnified by 100 times
438
showed that the floc structure was broken–down and appeared as a loose structure
439
after acidification conditioning (Figure S4). Therefore, it seems that the smooth
440
surface resulted from the release of hydrolyzed EPS into bulk solution during the
441
acidification treatment.55 At 20 °C there is insufficient time and energy available for
442
acid catalyzed hydrolysis to significantly degrade the protein structure.10, 18 Thus, the
443
floc structure of sludge filaments was not significantly disrupted, as shown in Figure
444
7b.
445
Figure S5 shows the concentrations of proteins (PN) and polysaccharides (PS) as 18 ACS Paragon Plus Environment
Environmental Science & Technology
446
the major EPS components in the supernatants of raw and treated sludges. After
447
acidification treatment, the S‒EPS, LB‒EPS and TB‒EPS fractions changed from
448
26.5, 53.5, 58.6 mg L‒1 to 36.0, 36.5 and 26.6 mg L‒1, respectively. It should be noted
449
that acidification facilitated the release of PS in S‒EPS but reduced PN concentration
450
for all EPS fractions. PS are generally neutral and composed of long chains of
451
monosaccharides such as D‒glucose or L‒fructose that may be hydrolyzed under acid
452
conditions.56 In contrast to PS, PN (e.g., tyrosine and tryptophan secreted from the
453
disrupted flocs) declined upon acid treatment for all EPS fractions, which is consistent
454
with the 3D‒EEM profiles (Regions I and II). It appears that acidification treatment
455
could not only destroy EPS structure and free bound water, but also cause protonation
456
of anionic functional groups of EPS, resulting in the destabilization of sludge
457
particles.57 The lower particle surface charge also leads to a decrease in solubility and
458
hence an increase in aggregation of the proteins. The increased particle size due to
459
flocculation at acidic pH was favorable to improve the sludge dewaterability,
460
evidenced by the reduced CST value and moisture content at pH 2 and 3 (Figure 4a).
461
Role of thermal, radical oxidation and shear force effects
462
The release of EPS by US appears to be the primary reason for the increase in
463
sludge dewaterability after A/US treatment, which was reflected by a suite of
464
evidence including 3D‒EEM spectra, DOC and UV254 measurements. In Figure 7c,
465
the sludge microstructure conditioned by US had smaller filaments with less filled
466
particles and its flocs surface was plate‒like with irregular pores. As shown in Figure
467
S5, the S‒EPS, LB‒EPS and TB‒EPS fractions in the filtered sludge samples were all
468
remarkably increased after US treatment from 25.5, 53.5 and 58.6 mg L‒1 to 173.1,
469
134.4 and 79.4 mg L‒1, respectively. This result further confirmed that US treatment
470
can break down the flocs directly in sludge, resulting in the release of PN and PS into 19 ACS Paragon Plus Environment
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471
the bulk phase (Figure S5). Yang et al. also reported that tryptophan PN were the main
472
solubilization products after US treatment.58 The DOC and UV254 values in Figure 2b
473
also support that US irradiation was destructive to EPS (e.g., aromatic protein) and
474
cells in the floc.59
475
US cavitation features unique physical and chemical effects.60 First, shear forces
476
in the form of shock‒waves and micro‒jets generated by ultrasonic cavitation cause
477
the breakage of sludge flocs, while the radical oxidation effect also disrupts the
478
relatively rigid cell membranes that may further reduce the floc size.61-63 Generally,
479
the shear forces predominate under lower frequency, while the radical oxidation is the
480
main mechanism under higher frequency.64 In our system, 20 kHz ultrasonic devices
481
were used. Thus, we assume the role of radical oxidation in disruption of sludge flocs
482
was minor. In this case, we consider that the particle size reduction observed was
483
mainly due to shock‒waves, micro‒jets, or acoustic mixing from elongation of
484
ultrasonic waves. We believe that the high speed microjets caused erosion/pitting of
485
solid surfaces and particle fracture. For our 20 kHz ultrasonic system, microjets
486
tended to occur on a solid surface with size greater than 150 µm bubble size,65, 66
487
which is applicable to the aggregated sludge flocs. The shock wave has been shown to
488
cause inter-particle collisions that is responsible for particle size reduction.67 It should
489
be noted that the shear forces induced movement of sludge flocs in the solution,
490
which in turn results in an attenuation of the ultrasonic energy and a reduction of the
491
cavitation performance. In the past, hydrophone system has been used to measure the
492
acoustic pressure and shear forces for describing the attenuation of ultrasonic wave
493
propagation.68, 69
494
Figure S6 demonstrates the thermal and radical oxidation effects of US treatment
495
on sludge dewaterability. As compared to US under temperature control (scenario A), 20 ACS Paragon Plus Environment
Environmental Science & Technology
496
the temperature rise induced by US without temperature control (scenario B) was
497
about 9 °C, resulting in the enhanced UV254 value of 2.08 cm‒1 but reduced CST value
498
of 15.7 s (Figure S6). Other studies found that temperature did not have significant
499
impact on the sludge disintegration.70 The difference may be due to the initial acidic
500
condition (pH = 3) used in this study, which promoted the hydrolysis of EPS (i.e.,
501
increased UV254) at higher temperature. As shown in Figure S6, the addition of 50
502
mM t‒butanol (scenario C) decreased the dewatering performance, with CST value
503
and cake moisture increasing from 17.8 s to 22.5 s and 68.5% to 71.4%, respectively.
504
The decreased UV254 values from 1.91 to 1.27 cm‒1 by t‒butanol implied that radical
505
species such as •OH play an important role in decomposing EPS. Similar to the A/US
506
treatments without temperature control (scenario B), it was also found that the thermal
507
effect is beneficial to improve sludge dewaterability in acidification treatment alone
508
(scenario D): the CST value decreased from 15.7 to 12.8 s, and UV254 from 2.08 cm‒1
509
to 1.27 cm‒1. However, the temperature rise seems to increase the cake moisture for
510
both scenarios B and D. The increased temperature likely aids hydrolysis of EPS to
511
smaller and more soluble proteins which are still able to aggregate and thus include
512
more water into the cake. From these results, it can be inferred that US‒induced
513
radical oxidation and thermal effects played significant roles in the sludge
514
dewaterability enhancement in response to A/US treatment.
515
Conceptual model of dewatering mechanisms in A/US treatment
516
A/US treatments of sludge are expected to integrate benefits of the two
517
independent processes. As illustrated in Figure 7d, more crannies and porous structure
518
were observed after the A/US treatments in the floc microstructure, possessing a good
519
water permeability and dewaterabilty.33, 71 In Figure S5, the S‒EPS, LB‒EPS and TB‒
520
EPS fractions were 40.5, 40.2 and 29.6 mg L‒1 after the A/US treatments, respectively, 21 ACS Paragon Plus Environment
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521
slightly higher than those of the acid treated sludge. From these results, a clear
522
dewatering mechanism can be elucidated in Table of Content. First, the addition of
523
acid initiated the decomposition of EPS through hydrolysis reactions. Then, the flocs
524
were broken down by US‒induced mechanical shearing forces (i.e., shock‒waves
525
and/or micro‒jets), releasing more EPS into bulk phase. Meanwhile, the disintegration
526
of sludge by US is beneficial to improve the acidification treatment through enhanced
527
mass transfer and associated increased kinetics of hydrolysis reactions. Therefore, the
528
hydrolysis and US disruption, together with oxidation by cavitation generated •OH,
529
produced the intensive decomposition of EPS resulting in release of the bound water.
530
Further, the acidic pH increased the surface charge of the reduced sized sludge
531
particles, thereby facilitating the reflocculation of the colloidal sludge system for
532
improved filterability. These two major processes of A/US are complementary with
533
each other in the sludge dewaterability enhancement.
534
Pilot‒scale Investigation and Economic Analysis
535
In order to evaluate the feasibility and operational cost for industrial treatment, a
536
pilot‒scale A/US process was constructed for the practice of sludge dewatering. As
537
shown in Figure 1b, after acidification pretreatment in a conditioning tank, a plug
538
flow pipe reactor equipped with three series of barbell horn US probes was adapted
539
for continuous sludge treatment. This system can be easily further scaled up by adding
540
more US probes. After US conditioning, the sludge mixture was neutralized with
541
NaOH. Finally, the sludge was dewatered in a pilot‒scale press filtration system
542
whose pressure was adjusted by changing the hydraulic pressure as compared to the
543
vacuum filtration system in bench‒scale tests. The results of pilot‒scale investigation
544
are tabulated in Table S6. The residence time of sludge is about 2 s under PD of 100 22 ACS Paragon Plus Environment
Environmental Science & Technology
545
W mL‒1 in the sonication process. US energy per unit volume of sludge was also
546
calculated. With an energy input of 32.4 kJ L‒1, the water contents of the sludge cakes
547
conditioned were 58.9% and 63.1% at pH 3 and 5, respectively. These results were
548
lower than the results from bench‒scale experiments. The difference may due to the
549
application of a press filtration system (rather than vacuum filtration) in the pilot‒
550
scale operation. The dry solids content was reduced from 35 to 16.8 g L‒1, which is
551
consistent with the results from the bench‒scale experiments.
552
To estimate the economics of the A/US process, the cost for sludge conditioning
553
was calculated. Variable costs including chemical and electrical consumption were
554
determined for the A/US process, as summarized in Table S7. Chemical costs
555
involved acidification necessitating H2SO4 and NaOH for pH adjustments. Electricity
556
was assumed to cost 0.12 USD per kWh, and is needed for the pumps, US system, and
557
filtration system. The final cost of the A/US process is estimated to be 29.3 USD per
558
ton of dry solids (DS), which is 45.1% lower than the 65 USD per ton of DS by the
559
traditional polyacrylamide technology.33 The calculation of the cost estimation was
560
detailed in Table S7. An economic assessment of a pilot‒scale Fenton’s reagent
561
treatment of sludge indicates a cost of 165 USD per ton of DS, considering the fixed
562
and variable costs as well as the energy savings for incinerating the sludge after
563
dewatering. Further, the water content in the pilot treatment is less than 60%, the DS
564
per liter reduced by about 50%, and the volume reduced by about 75%, thus saving
565
considerable cost on sludge transport and disposal expenses. However, corrosion‒
566
resistant construction materials for the system are suggested, since the acidification
567
treatment requires acid/base for pH adjustments. Based on the Chinese Discharge
568
Standard of Pollutants for Municipal Wastewater Treatment Plant (GB 18918‒2008),
569
sludge moisture content less than 60% can be directly disposed. Thus, the sludge 23 ACS Paragon Plus Environment
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570
moisture content after A/US pretreatment in our study meets the standard for direct
571
disposal. However, it should be noted that it is possible to further reduce the sludge
572
moisture content by subsequent drying. Although the water content can be reduced to
573
5 ‒ 10% after drying, there are other concerns and costs involved, such as specific
574
drying equipment and processes (i.e., one‒step or two‒step drying) and heat sources
575
(e.g., coal, natural gas, or steam). Thus, the drying process increases the operational
576
cost for wastewater treatment plants. Further, it is reported that the energy
577
consumption peaked for the sludge with water contents ranging from 35 ‒ 65%, while
578
the energy consumption is much lower when sludge water content is less than 35% or
579
greater than 65%.72 This is because at the range of 35 ‒ 65% water content, the sludge
580
nature is in the viscoplastic phase, which is similar to that of glue.72 In this phase,
581
sludge is very difficult to be dried.
582
Both the pilot‒scale and batch‒scale investigations indicated that A/US
583
treatments can effectively enhance the dewaterability of waste activated sludge and
584
reduce dry solid weight. The cost estimation showed that the A/US conditioning
585
process is more economical than other typical dewatering techniques in the
586
engineering practice. In future work, it would be of practical importance to further
587
improve sludge dewaterability by A/US treatments at higher pH for the purpose of
588
minimizing chemical use.
589
590
ACKNOWLEDGEMENTS
591
This work was supported by the National Natural Science Foundation of China
592
(No. 2167070159 and No. 21507167) and Zhejiang Provincial Natural Science
593
Foundation of China (LY16B050001). 24 ACS Paragon Plus Environment
Environmental Science & Technology
594
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affect extracellular polymeric substances (EPSs) and improve waste activated sludge dewatering. Process Biochem. 2015, 50, 438–446. 22. Apul, O. G.; Dogan, I.; Sanin, F. D., Can capillary suction time be an indicator for sludge disintegration? J. Residuals Sci. Tech. 2009, 6 (3), 99-104. 23. Wong, J. W. C.; Zhou, J.; Kurade, M. B.; Murugesan, K., Influence of ferrous ions on extracellular polymeric substances content and sludge dewaterability during bioleaching. Bioresour. Technol. 2015, 179, 78–83. 24. Bystryak, S.; Santockyte, R.; Peshkovsky, A. S., Cell disruption of S. cerevisiae by scalable high-intensity ultrasound. Biochem. Eng. J. 2015, 99, 99-106. 25. Rae, J.; Ashokkumar, M.; Eulaerts, O.; von Sonntag, C.; Reisse, J.; Grieser, F., Estimation of ultrasound induced cavitation bubble temperatures in aqueous solutions. Ultrason. Sonochem. 2005, 12 (5), 325-329. 26. Tauber, A.; Mark, G.; Schuchmann, H. P.; von Sonntag, C., Sonolysis of tert-butyl alcohol in aqueous solution. J. Chem. Soc., Perkin Trans. 2 1999, 6, 1129-1135. 27. Senanayake, P. C.; Gee, N.; Freeman, G. R., Viscosity and density of isomeric butanollwater mixtures as functions of composition and temperature. Can. J. Chem. 1987, 65 (10), 2441-2446. 28. Cheong, W. J.; Carr, P. W., The surface tension of mixtures of methanol, acetonitrile, tetrahydrofuran, isopropanol, tertiary butanol and dimethyl-sulfoxide with water at 25°C. J. Liq. Chromatogr. 1987, 10 (4), 561-581. 29. Torres-Palma, R. A.; Gibson, J.; Droppo, I. G.; Seto, P.; Farnood, R., Surfactant-assisted sono-breakage of wastewater particles for improved UV disinfection. Water Air Soil Pollut. 2017, 228 (106), 1-10. 30. APHA, Standard Methods for the Examination of Water and Wastewater, 20th ed. American Public Health Association, American Water Works Association and Water Environmental Federation: Washington D.C., 1998. 31. Kim, M. S.; Lee, K. M.; Kim, H. E.; Lee, H. J.; Lee, C. S.; Lee, C. H., Disintegration of waste-activated sludge by thermally-activated persulfates for enhanced dewaterability. Environ. Sci. Technol. 2016, 50 (13), 7106-7115. 32. Katsiris, N.; Kouzelikatsiri, A., Bound water content of biological sludges in relation to filtration and dewatering. Water Res. 1987, 21 (11), 1319-1327. 33. Zhang, H.; Yang, J. K.; Yu, W. B.; Luo, S.; Peng, L.; Shen, X. X.; Shi, Y. F.; Zhang, S. N.; Song, J.; Ye, N.; Li, Y.; Yang, C. Z.; Liang, S., Mechanism of red mud combined with Fenton's reagent in sewage sludge conditioning. Water Res. 2014, 59, 239-247. 34. Morgan, J. W.; Forster, C. F.; Evison, L., A comparative study of the nature of biopolymers extracted from anaerobic and activated sludges. Water Res. 1990, 24 (6), 743-750. 35. Li, X. Y.; Yang, S. F., Influence of loosely bound extracellular polymeric substances (EPS) on the flocculation, sedimentation and dewaterability of activated sludge. Water Res. 2007, 41 (5), 1022-1030. 36. Lowry, O. H.; Rosebrough, N. J.; Farn, A. L.; Randall, R. J., Protein measurement with the Folin phenol reagent. J. Biol. Chem. 1951, 193 (1), 265–275. 37. Redmile-Gordon, M. A.; Armenise, E.; White, R. P.; Hirsch, P. R.; Goulding, K. W. T., A comparison of two colorimetric assays, based upon Lowry and Bradford techniques, to estimate total protein in soil extracts. Soil Biol. Biochem. 2013, 67, 166-173. 38. Herbert, D.; Philipps, P. J.; Strange, R. E., Carbohydrate analysis. Methods Enzymol. B 1971, 5, 265–277.
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39. Qian, X.; Wang, Y. L.; Zheng, H. L., Migration and distribution of water and organic matter for activated sludge during coupling magnetic conditioning horizontal electro-dewatering (CM-HED). Water Res. 2016, 88, 93-103. 40. Lee, D. J.; Lee, S. F., Measurement of bound water content in sludge - the use of differential scanning calorimetry (DSC). J. Chem. Technol. Biot. 1995, 62 (4), 359-365. 41. Feng, J.; Wang, Y. L.; Ji, X. Y., Dynamic changes in the characteristics and components of activated sludge and filtrate during the pressurized electro-osmotic dewatering process. Sep. Purif. Technol. 2014, 134, 1-11. 42. Gonze, E.; Pillot, S.; Valette, E.; Gonthier, Y.; Bernis, A., Ultrasonic treatment of an aerobic activated sludge in a batch reactor. Chem. Eng. Process 2003, 42 (12), 965-975. 43. Wolski, P.; Zawieja, I., Effect of ultrasound field on dewatering of sewage sludge. Arch. Environ. Prot. 2012, 38 (2), 25-31. 44. Baker, A., Fluorescence excitation-emission matrix characterization of some sewage-impacted rivers. Environ. Sci. Technol. 2001, 35 (5), 948-953. 45. Chen, W.; Westerhoff, P.; Leenheer, J. A.; Booksh, K., Fluorescence excitation - Emission matrix regional integration to quantify spectra for dissolved organic matter. Environ. Sci. Technol. 2003, 37 (24), 5701-5710. 46. Guo, L.; Lu, M. M.; Li, Q. Q.; Zhang, J. W.; Zong, Y.; She, Z. L., Three-dimensional fluorescence excitation-emission matrix (EEM) spectroscopy with regional integration analysis for assessing waste sludge hydrolysis treated with multi-enzyme and thermophilic bacteria. Bioresour. Technol. 2014, 171, 22-28. 47. Sheng, G. P.; Yu, H. Q., Characterization of extracellular polymeric substances of aerobic and anaerobic sludge using three-dimensional excitation and emission matrix fluorescence spectroscopy. Water Res. 2006, 40 (6), 1233-1239. 48. He, D. Q.; Zhang, Y. J.; He, C. S.; Yu, H. Q., Changing profiles of bound water content and distribution in the activated sludge treatment by NaCl addition and pH modification. Chemosphere 2017, 186, 702-708. 49. Wang, H. F.; Ma, Y. J.; Wang, H. J.; Hu, H.; Yang, H. Y.; Zeng, R. J., Applying rheological analysis to better understand the mechanism of acid conditioning on activated sludge dewatering. Water Res. 2017, 122, 398-406. 50. Feng, X.; Lei, H. Y.; Deng, J. C.; Yu, Q.; Li, H. L., Physical and chemical characteristics of waste activated sludge treated ultrasonically. Chem. Eng. Process 2009, 48 (1), 187-194. 51. Zhou, Z. W.; Yang, Y. L.; Li, X., Effects of ultrasound pretreatment on the characteristic evolutions of drinking water treatment sludge and its impact on coagulation property of sludge recycling process. Ultrason. Sonochem. 2015, 27, 62-71. 52. Kanmani, P.; Kumar, R. S.; Yuvaraj, N.; Paari, K. A.; Pattukumar, V.; Arul, V., Production and purification of a novel exopolysaccharide from lactic acid bacterium Streptococcus phocae PI80 and its functional characteristics activity in vitro. Bioresour. Technol. 2011, 102 (7), 4827-4833. 53. Jia, F. X.; Yang, Q.; Liu, X. H.; Li, X. Y.; Li, B. K.; Zhang, L.; Peng, Y. Z., Stratification of extracellular polymeric substances (EPS) for aggregated anammox microorganisms. Environ. Sci. Technol. 2017, 51 (6), 3260-3268. 54. Erdincler, A.; Vesilind, P. A., Effect of sludge cell disruption on compactibility of biological sludges. Water Sci. Technol. 2000, 42 (9), 119-126. 55. Guo, S. H.; Li, G.; Qu, J. H.; Liu, X. L., Improvement of acidification on dewaterability of oily
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770
771 772
Figure 1: (a) Schematic diagram of barbell horn ultrasound (BHU) system used at the
773
bench scale, (1) US generator, (2) pH meter, (3) thermometer, (4) ultrasonic probe, (5)
774
cooling water, (6) receiving ring, (7) reactor. (b) Pilot‒scale setup of sludge
775
conditioning and dewatering system with the A/US treatments.
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24
water in sludge cake total solid content cake moisture CST
(a)
80 60 40
CST (s)
cake moisture (%)
6 4 2 0
DOC (mg L-1)
90 85 80 75 70 65
weight (g)
777
20
raw sludge US acidification A/US
0
UV254
(b)
DOC
20
UV254 (cm-1)
16 2.0 1.5 1.0 0.5 0.0
raw sludge US
acidification A/US
778 779
Figure 2: Comparison of sludge treatments after US, acidification, and A/US. (a)
780
cake moisture, capillary suction time (CST), weight of sludge cake and total solids
781
content with 300 mL sludge; (b) UV254 and DOC. pH = 6.7 for raw sludge and US,
782
pH = 3 for acidification and A/US under conditions of power density (PD) = 2.5 W
783
mL‒1 and 2 min sonication.
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-1
Bound water (g g DS)
3
2
1
0
raw sludge
US
acidification
A/US
784 785
Figure 3: Bound water contents after different treatments based on the DSC result.
786 787
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water in sludge cake total solid content cake moisture CST
76 72 68 64 60 4 3 2 1 0
2
3
4
0 Zeta potential (mV)
(a)
pH
5
6
7 (b)
-5
-10
90 80 70 60 50 40 30 20 10 0
70 60
-15 -20
50
d50% (µm)
weight (g)
80
CST (s)
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cake moisture (%)
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-25 40
-30 Zeta potential d50%
-35 -40 788
2
3
4
30
pH
5
6
7
789
Figure 4: The effect of pH on sludge dewaterability. (a) cake moisture, CST, water
790
content in sludge cake, and total solids content with 300 mL sludge; (b) Zeta potential
791
and mass median diameter of sludge flocs (d50%) under conditions of power density
792
(PD) = 2.5 W mL‒1 and 2 min sonication. 32 ACS Paragon Plus Environment
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S-EPS 400
LB-EPS
RAW
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TB-EPS
RAW
0.000
RAW
200.0
350
400.0
E
D
600.0
300
250
200 400
800.0 Region V
A
B
Region I
Region II
300
350
1000
Region IV
C
1200 1400 1500
Region III
400
450
500
300
350
400
450
500
pH=7
pH=7
300
350
400
450
500 0.000
pH=7
200.0
Excitation wavelength (nm)
350
400.0
300 600.0
250
200 400
800.0 1000
300
350
400
450
500
pH=5
300
350
400
450
500
pH=5
300
350
400
450
500 0.000
pH=5
200.0
350
400.0
300 600.0
250
800.0 1000
200 400
300
350
400
450
500
pH=3
300
350
400
450
500
300
350
400
450
500 0.000
pH=3
pH=3
200.0
350
400.0
300 600.0
250
800.0 1000
200
300
350
400
450
500
300
350
400
450
500
300
350
400
450
500
793
Emission wavelength (nm)
794
Figure 5: Influence of pH on 3D‒EEM profile of EPS fractions (S‒EPS, LB‒EPS,
795
and TB‒EPS) under conditions of power density (PD) = 2.5 W mL‒1 and 2 min
796
sonication. Raw = untreated waste activated sludge. Note, the unit of EEM output is
797
arbitrary.
798
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cake moisture CST water in sludge cake total solid content
90 80
40 20
0 0.4 1.7 2.5 3.3 6.7 10 power density (W mL-1)
Zeta potential (mV)
d50% Zeta potential
-2
0
(b) 110 100
-3
90
-4 -5
80
-6
70
-7
60
-8
CST (s)
60 4 3 2 1 0
80 60
70
-1
799
(a)
0
2 4 6 8 10 power density (W mL-1)
d50% (µm)
weight (g) cake moisture (%)
100
50
800
Figure 6: Effect of US power density (PD, W mL‒1) on sludge dewaterability. (a)
801
cake moisture, CST, water content in sludge cake, and total solids content with 300
802
mL sludge; (b) Zeta potential and mass median diameter of sludge flocs (d50%) under
803
conditions of pH = 3 and 2 min sonication. 34 ACS Paragon Plus Environment
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804
(b)
(a)
(d)
(c)
805
Figure 7: SEM of raw sludge (a), acidified sludge (b), US treated sludge (c), and
806
A/US treated sludge (d) under conditions of PD = 2.5 W mL‒1 and 2 min sonication.
807
Note, pH = 6.7 for (a) and (c), pH = 3 for (b) and (d).
808
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