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Page 1 ofEnvironmental 36 Science & Technology

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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

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Enhancing activated sludge dewaterability is of scientific and engineering

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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

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pilot‒scales. Our results showed that the A/US system exhibited significantly

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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

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INTRODUCTION

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Addressing the environmental concern of excess sludge is a key challenge in

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municipal wastewater treatment plants (MWTP) for populated urban areas.1 Sludge is

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a complex colloidal system in which highly dispersed and fine particles around 1 µm

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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

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sum of interstitial, vicinal, and hydration waters that are arduous to remove, because

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they are tightly bound to extracellular polymeric substances (EPS) and other sludge

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components stick by adhesive forces and/or chemical bonds.5

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During sewage sludge treatment, the dewatering process is of practical

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importance in reducing water content and sludge volume for the purpose of

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transportation and safe disposal.6 Usually, thickened sludge is conditioned through

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physical disruption and chemical addition (e.g., acid, ferric chloride, and lime),

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followed by mechanical dewatering techniques (e.g., press filters and centrifuges).

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The cost of such dewatering processes represents a significant amount of the total

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operational cost in typical MWTPs. However, water content in filtered sludge cake

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still ranges from 75 to 85% (w/w), failing to meet the increasingly stringent criteria of

72

subsequent disposal.7,

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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

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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

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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

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technologies.11

dewaterability for different treatment

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Ultrasound (US), an emerging technology used in the sludge disintegration

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process for enhanced digestion, is recognized as an efficient approach for improving

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dewaterability.12, 13 US with a series of compression and rarefaction cycles generates

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cavitation bubbles in water. The cavitation bubbles implode and yield localized

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temperatures (as high as 5000 K), mechanical shear forces (i.e., shock‒waves and

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micro‒jets), and free radicals (e.g. •H, •OH and HO2• ).14 The mechanical shearing

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effect causes destruction of microbial cells and release of LB‒EPS and TB‒EPS.15

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Thermolysis and radical oxidation of the released organics free EPS‒bound water into

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bulk phase, thereby enhancing sludge dewaterability.12, 16-18 Nevertheless, a high US

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energy input was observed to deteriorate sludge dewaterability, since intensive US

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waves disrupt large flocs into smaller sizes with high surface area, resulting in an

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increase of bound water through re‒adsorption. Previous studies also suggested that a

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thin layer of released intracellular substances formed on the new particle surface acts

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as a barrier against water release, consequently reducing the dewaterability.12, 19, 20

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However, these dewatering mechanisms were proposed based on speculation and need

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to be experimentally confirmed. Indeed, many fundamental questions regarding the

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US dewatering mechanisms still remain unanswered: How does radical oxidation

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quantitatively contribute to the overall sludge dewaterability? Does US‒induced

100

thermolysis play a significant role in the dewatering process? How can the mechanical

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shearing force effect be controlled to prevent the deterioration of sludge

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dewaterability? Such mechanistic and fundamental questions have important

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implications for application of US dewatering process in engineering practice.

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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

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groups on EPS reduces the electrostatic repulsion, resulting in enhanced sludge

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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

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water content was reduced 50% with a decrease of capillary suction time (CST) from

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37.7 to 9.2 s in a microwave‒acid treatment.10 In addition, by characterizing organic

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components in EPS matrix, Xiao et al. successfully demonstrated that US followed by

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acidification is more effective in enhancing sludge dewaterability than acidification

113

alone.11 Based on previous studies, the combination of US and acidification

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potentially integrates complementary roles for these two discrete technologies. But

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US prior to acidification treatment may overlook the fact that the application of US in

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advance can cause deterioration of sludge dewaterability,12 and the subsequent

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acidification step cannot yield optimal performance (see Text S1 and Figure S1 in the

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supporting information, SI). Therefore, further efforts are required to optimize the

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synergism between US and acidification treatments, such as the sequence of applying

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acidification and improvement of US devices.

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Unfortunately, there are limited engineered investigations for sludge dewatering

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processes due to the lack of industrial investment into current technologies. Full‒scale

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treatment necessitates fulfilling the following requirements: 1) operations are simple

124

and effective; 2) dewatering units are compatible with filtration devices; and 3)

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dewatering processes must feature low cost, operational stability, and minimal

126

chemical use. Therefore, in this study, we investigated the integration of acidification

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and US (A/US) treatments for enhancing sludge dewaterability at both bench‒scale

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and pilot‒scale. The objectives of the present study were to (1) investigate the effect

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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

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feasibility; and (3) provide mechanistic insights into the synergism between US and

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acidification processes. Our bench‒scale and pilot‒scale investigations will provide a

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mechanistic basis for better understanding the A/US process and enable development

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of a viable and economical dewatering technology.

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136

MATERIALS AND METHOD

137

Sewage sludge and reagents

138

The waste activated sludge was collected from in a MWTP in Hangzhou, China.

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The plant treats approximately 600,000 m3 municipal wastewater daily by an

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anaerobic/anoxic/oxic process. Sludge samples were stored at 4 °C before use. All

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samples were kept for a maximum of two days. The characteristics of raw sewage

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sludge are summarized in Table S1 of the SI.

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Information for all the chemicals used in this study such as purity and

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manufacturer is tabulated in Table S2. H2SO4 and NaOH were used to adjust pH of

145

sludge samples.

146 147

Sludge conditioning and dewatering

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Bench‒scale experiments

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The bench‒scale tests for sludge conditioning and dewatering were conducted in

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batch mode using a 500 mL cylindrical flask (12 cm in depth, 9 cm in I.D., and 763

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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

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adding 5M H2SO4 or NaOH slowly while being continuously stirred. Sludge pH was

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measured by submerging a PHS‒3C pH meter (Yoke, China) into the sludge slurry.

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After the pH of sludge stabilized (~ 30 s), US irradiation was applied using a high

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amplitude power US probe on the basis of the barbell horn ultrasound (BHU)

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technology. This unique design has been shown to be reliable, and can be directly

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scaled up to pilot‒scale sonochemical processes without reducing the US amplitude.24

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The US probe has a maximum power of 3 kW and a maximum cross‒sectional output

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diameter about 40 mm at a frequency of 20 kHz (TJS‒300, Hangzhou Success

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Ultrasonic Equipment, China). Detailed design of the US horn is illustrated in Figure

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1a. Bench‒scale experiments were carried out at different US power densities (0.0 to

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10.0 W mL‒1) and durations (0 to 2 min). The selection of sonication duration assured

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low energy consumption because treatment tests were short. The titanium probe was

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immersed at 1 cm below solution surface. A HAAKE A80 cooling bath and a

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thermocouple connected to a digital thermometer (DS18B20, Dallas Crop, U.S.) were

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used to maintain the temperature of sonicated sludge samples in the flask at 25 ± 2 °C.

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A set of control experiments was conducted to quantitatively evaluate the contribution

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of thermal, radical oxidation and shear force effects to sludge dewatering. Specifically,

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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

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control (scenario B), US with t‒butanol25, 26 as radical (i.e., •OH) scavenger (scenario

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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

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without the scavenger reagent. The result show thats addition of 50 mM t‒butanol

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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

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potential interference exerted by t-butanol. All experiments were conducted in

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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

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flow‒through mode. The pilot system consists of a 3000 L conditioning tank, a BHU

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system, and a spring filter press (Zhejiang OuKeMei Filtration Equipment, China).

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The activated sludge was first treated with H2SO4 to achieve pH 3 or 5 with the

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resident time of 4.5 h. Next, the conditioned sludge was pumped to the BHU system

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by an air pump with a flow rate of 25 L min‒1, and then into a neutralization tank

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where sludge pH was adjusted to 7 for corrosion protection. Finally, the conditioned

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sludge was pumped into a pressure filtration. The dewatering of conditioned sludge in

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the pressure filtration system consisted of two phases: a 70 min feeding phase with a

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pressure of 1.5 MPa and a 30 min pressing phase with a pressure of 1.6 MPa. A

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polypropylene filament filter cloth in dimension of 1640 × 1640 mm2 was used in the

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pressure filtration system.

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Analytical methods

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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

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mL of conditioned sludge suspension was sampled from the cylindrical reactor and

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measured in a CST instrument (Type 304M, Triton Electronics Ltd, UK) with an 18

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mm inner radius and a Whatman No. 17 chromatography grade filtration paper.

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Likewise, sludge dewatering was performed by the vacuum filtration, during which

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100 mL conditioned sludge was poured into a 9 cm standard Buchner funnel fitted

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with 1.2 µm pre-wetted Whatman filter paper under a constant vacuum pressure of

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34.5 kPa until all filtrate was removed. The dewatered sludge “cake” on filter paper

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was dried at 105 °C for 24 h and weighed to determine water and solid contents of

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sludge samples according to Standard Method 2540‒B of “Total Solids Dried at 103‒

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105 °C”.30 Bound water content measurement was determined using a differential

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scanning calorimetry (DSC) analyzer equipped with a liquid nitrogen cooling system

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(Diamond, PerkinElmer). An approximate of 8 mg sludge sample was retrieved and

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placed into the crucible of a special mini-oven. The temperature was first cooled to –

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20 °C, at which free water in the sample was frozen, and then was increased to 10 °C

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at a rate of 2 °C min‒1. Details for the DSC analytical procedure can be found in

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Katsiris and Kouzeli-Katsiri (1987)32 and Zhang et al. (2014).33 The amount of free

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water (Wf) was calculated by the heat absorption determined by integrating the

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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

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water:

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Wb =Wt –∆H/∆H0

(1)

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where Wb and Wt are the bound water content and total water content of the sludge

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samples, respectively; ∆H and ∆H0 is the amount of energy absorbed by sludge and

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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

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A heat extraction method was modified to extract different EPS fractions from

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waste activated sludge.34, 35 Sludge suspensions were first centrifuged at 3000×g and

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4 °C for 15 min to collect the supernatant fluids containing S‒EPS. Then, solids at the

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bottom of the centrifuge tubes were re‒suspended to the original volume by adding

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0.05% (w/w) NaCl solution and mixed using a vortex mixer (VA08G1‒24, Mo Bio

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Labs, U.S.) for 1 min at 50 °C. The suspension was then immediately centrifuged at

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4000×g and 4 °C for 10 min to remove any suspended biomass left over in the

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supernatant. The supernatant fluids containing LB‒EPS were collected. The solids

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were again re‒suspended to the initial volume with 0.05% (w/w) NaCl and held at

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60 °C for 15 min to extract TB‒EPS. The supernatant containing TB‒EPS were

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obtained by centrifuging the suspensions at 4000×g and 4 °C for 15 min. All

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supernatant fluids were filtered through a 0.45 µm cellulose acetate membrane filter

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(Whatman) and stored at 4 °C before analysis.

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Three‒dimensional excitation emission matrix (3D‒EEM) spectra of extracted

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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

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excitation and emission slit widths of 10 nm. Each scan had 37 emission and 27

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excitation wavelengths. The protein content in extracted EPS was determined with the

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Lowry–Folin method.36 It should be noted that the presence of humic like substances

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from the sludge samples could possibly yield error in the quantification level of

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proteins estimated by the Lowry‒Folin method.37 Polysaccharide concentrations were

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measured with the anthrone method using glucose as standard.38 Supernatant pH was

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adjusted to 7 prior to analysis to exclude pH interference. 10 ACS Paragon Plus Environment

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Sludge flocs characterization

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The surface morphology of different sludge samples was characterized by

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scanning electron microscope (SEM; S‒4800, Hitachi, Japan). Optical microscopic

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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

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used for measuring Zeta potential of sludge flocs, while a MasterSizer 2000 equipped

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with a Hydro2000MU auto sampler (Malvern, UK) was applied for measuring the

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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).

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UV254, which is a measure of aromatics, was determined by a UV/Vis

260

spectrophotometer (UV2600, Purkinje, China). Both DOC and UV254 were measured

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after filtration through 0.45 µm acetate fiber membranes. The biochemical oxygen

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demand (BOD) and soluble chemical oxygen demand (COD) in the supernatant of

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sludge samples were measured with a BOD meter (2173B, Hach, USA) and a

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spectrophotometer (DR/2000, Hach, USA), respectively. To measure heavy metal

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concentrations, sludge was initially filtered and the dried solid cake was digested in a

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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

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the typical inorganic elements (i.e., Cu, Zn, Cd, Pb, Mn, Cr, Ni, Al, Ca, and Mg) in

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the raw sludge and treated sludge samples according to APHA method 3030I.

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Duplicate samples were prepared and treated for the heavy metal analysis. We

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concluded that the levels of organic matter and heavy metals in the recycled filtrate do

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not exhibit a detrimental effect on the operation of the wastewater treatment plant (see

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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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RESULTS AND DISCUSSION

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Characterization of dewaterability under different treatments

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Batch experiments were performed to compare the dewatering performance by

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US, acidification, and A/US. As illustrated in Figure 2a, the cake moisture was

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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

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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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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

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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

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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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sludge from flotation. Chem. Eng. J. 2011, 168 (2), 746-751. 56. Wingender, J.; Jäger, K. E.; Flemming, H. C., Interaction between extracellular polysaccharides and enzymes. In Microbial Extracellular Polymeric Substances; Wingender, J.; Neu, T. R., Eds.; Springer: Heidelberg, 1999; pp 1–19. 57. Zhang, W. J.; Cao, B. D.; Wang, D. S.; Ma, T.; Yu, D. H., Variations in distribution and composition of extracellular polymeric substances (EPS) of biological sludge under potassium ferrate conditioning: Effects of pH and ferrate dosage. Biochem. Eng. J. 2016, 106, 37-47. 58. Yang, S. S.; Guo, W. Q.; Meng, Z. H.; Zhou, X. J.; Feng, X. C.; Zheng, H. S.; Liu, B.; Ren, N. Q.; Cui, Y. S., Characterizing the fluorescent products of waste activated sludge in dissolved organic matter following ultrasound assisted ozone pretreatments. Bioresour. Technol. 2013, 131, 560-563. 59. Jin, Y. Y.; Li, H.; Mahar, R. B.; Wang, Z. Y.; Nie, Y. F., Combined alkaline and ultrasonic pretreatment of sludge before aerobic digestion. J. Environ. Sci. China 2009, 21 (3), 279-284. 60. Wei, Z. Characterizing Ultrasonic Systems for Improved Remediation of Contaminated Sediments. PhD Dissertation, The Ohio State University, Columbus, OH, 2015. 61. de La Rochebrochard, S.; Naffrechoux, E.; Drogui, P.; Mercier, G.; Blais, J. F., Low frequency ultrasound-assisted leaching of sewage sludge for toxic metal removal, dewatering and fertilizing properties preservation. Ultrason. Sonochem. 2013, 20, (1), 109-117. 62. Pham, T. T. H.; Tyagi, R. D.; Brar, S. K.; Surampalli, R. Y., Effect of ultrasonication and Fenton oxidation on biodegradation of bis(2-ethylhexyl) phthalate (DEHP) in wastewater sludge. Chemosphere 2011, 82 (6), 923-928. 63. Tyagi, V. K.; Lo, S. L.; Appels, L.; Dewil, R., Ultrasonic treatment of waste sludge: A review on mechanisms and applications. Crit. Rev. Env. Sci. Tec. 2014, 44 (11), 1220-1288. 64. Wood, R. J.; Lee, J.; Bussemaker, M. J., A parametric review of sonochemistry: Control and augmentation of sonochemical activity in aqueous solutions. Ultrason. Sonochem. 2017, 38, 351-370. 65. Doktycz, S. J.; Suslick, K. S., Interparticle collisions driven by ultrasound. Science 1990, 247 (4946), 1067-1069. 66. Suslick, K. S., Ultrasound: Its Chemical, Physical, and Biological Effects. VCH Publishers: New York, 1988. 67. Suslick, K. S.; Casadonte, D. J.; Doktycz, S. J., The effects of ultrasound on nickel and copper powders. Solid State Ionics 1989, 32-3, 444-452. 68. Bandelin, J.; Lippert, T.; Drewes, J. E.; Koch, K., Cavitation field analysis for an increased efficiency of ultrasonic sludge pretreatment using a novel hydrophone system. Ultrason. Sonochem. 2018, 42, 672–678. 69. Wei, Z. S.; Kosterman, J. A.; Xiao, R. Y.; Pee, G. Y.; Cai, M. Q.; Weavers, L. K., Designing and characterizing a multi-stepped ultrasonic horn for enhanced sonochemical performance. Ultrason. Sonochem. 2015, 27, 325-333. 70. Cheung, H. M.; Kurup, S., Sonochemical destruction of CFC-11 and CFC-113 in dilute aqueous solution. Environ. Sci. Technol. 1994, 28 (9), 1619-1622. 71. Yu, W. B.; Yang, J. K.; Shi, Y. F.; Song, J.; Shi, Y.; Xiao, J.; Li, C.; Xu, X. Y.; He, S.; Liang, S.; Wu, X.; Hu, J. P., Roles of iron species and pH optimization on sewage sludge conditioning with Fenton's reagent and lime. Water Res. 2016, 95, 124-133. 72. Wang, P.; Ke, Z.; Tian, X., A review on sewage sludge drying technologies. Development & Innovation of Machinery & Electrical Products 2003, 2, 9-11. (In Chinese)

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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)

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cooling water, (6) receiving ring, (7) reactor. (b) Pilot‒scale setup of sludge

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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,

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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

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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

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Region IV

C

1200 1400 1500

Region III

400

450

500

300

350

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pH=7

pH=7

300

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pH=7

200.0

Excitation wavelength (nm)

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400.0

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pH=5

300

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pH=5

300

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pH=5

200.0

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300 600.0

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pH=3

300

350

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300

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500 0.000

pH=3

pH=3

200.0

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300

350

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450

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300

350

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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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