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Page 1 ofEnvironmental 36 Science & Technology

ACS Paragon Plus Environment

Environmental Science & Technology

Page 2 of 36

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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31

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


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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80

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


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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


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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

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Sludge dewaterability was assessed by capillary suction time (CST) value, 8 ACS Paragon Plus Environment


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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


plasma‒atomic

emission

Page 13 of 36


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


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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325

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



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


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


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


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


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


594

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Yuan, H. P.; Cheng, X. B.; Chen, S. P.; Zhu, N. W.; Zhou, Z. Y., New sludge pretreatment method

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Liao, B. Q.; Allen, D. G.; Droppo, I. G.; Leppard, G. G.; Liss, S. N., Surface properties of sludge

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Neyens, E.; Baeyens, J.; Weemaes, M.; De heyder, B., Hot acid hydrolysis as a potential treatment

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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.



Page 32 of 36

-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


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)


cake moisture (%)

Page 33 of 36

-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


S-EPS 400

LB-EPS

RAW

Page 34 of 36

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


Page 35 of 36


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


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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