Heterogeneous Fenton oxidation of trichloroethylene catalyzed by sewage sludge biochar: Experimental study and life cycle assessment

Journal Pre-proof Heterogeneous Fenton oxidation of trichloroethylene catalyzed by sewage sludge biochar: Experimental study and life cycle assessment Yu-Fong Huang, Yu-Yang Huang, Pei-Te Chiueh, Shang-Lien Lo PII:

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DOI:

https://doi.org/10.1016/j.chemosphere.2020.126139

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CHEM 126139

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Received Date: 16 September 2019 Revised Date:

5 February 2020

Accepted Date: 5 February 2020

Please cite this article as: Huang, Y.-F., Huang, Y.-Y., Chiueh, P.-T., Lo, S.-L., Heterogeneous Fenton oxidation of trichloroethylene catalyzed by sewage sludge biochar: Experimental study and life cycle assessment, Chemosphere (2020), doi: https://doi.org/10.1016/j.chemosphere.2020.126139. This is a PDF file of an article that has undergone enhancements after acceptance, such as the addition of a cover page and metadata, and formatting for readability, but it is not yet the definitive version of record. This version will undergo additional copyediting, typesetting and review before it is published in its final form, but we are providing this version to give early visibility of the article. Please note that, during the production process, errors may be discovered which could affect the content, and all legal disclaimers that apply to the journal pertain. © 2020 Published by Elsevier Ltd.

Credit Author Statement:

Yu-Fong Huang: Conceptualization, Methodology, Investigation, Data Curation, Writing - Original Draft, Writing - Review & Editing, Visualization

Yu-Yang Huang: Conceptualization, Methodology, Software, Formal analysis, Investigation, Visualization,

Pei-Te Chiueh: Validation, Resources, Writing - Original Draft, Writing - Review & Editing, Supervision, Project administration, Funding acquisition

Shang-Lien Lo: Validation, Resources, Writing - Original Draft, Writing - Review & Editing

Heterogeneous Fenton oxidation of trichloroethylene catalyzed by sewage sludge biochar: experimental study and life cycle assessment

Yu-Fong Huang, Yu-Yang Huang, Pei-Te Chiueh*, Shang-Lien Lo

Graduate Institute of Environmental Engineering, National Taiwan University, No. 1, Sec. 4, Roosevelt Rd., Taipei 106, Taiwan, ROC

*

Corresponding author. Tel.: +886 2 3366 2798; fax: +886 2 2392 8830 Email address: [email protected] (P.-T. Chiueh).

Graphical abstract:

1

Abstract

2 3

Heterogeneous Fenton oxidation of trichloroethylene (TCE) catalyzed by sewage

4

sludge biochar was studied. The highest TCE removal efficiency was 83% at pH 3.1,

5

catalyzed by 300 W biochar. The biochars produced at higher microwave power levels

6

provided better catalytic effect, due to higher iron contents and specific surface areas.

7

Reactivity of sewage sludge biochar maintained after several uses, which provides an

8

advantage for using as a permeable reactive barrier to remediate groundwater pollution.

9

Chromium, copper, nickel, lead, and zinc were found in the leachate generated from

10

sewage sludge biochar, and most of the concentrations were lower than the standards for

11

non-drinking water use. Besides, copper, zinc, and iron were found in the reaction

12

solutions of Fenton oxidation. Because of the highest dosage required for Fenton

13

oxidation, the environmental impact caused by 200 W biochar is highest. The

14

environmental impact caused by 300 W biochar is lowest. Among the four endpoint

15

impact categories in the life cycle assessment (LCA), human health is the highest

16

concern, whereas ecosystem quality is the least. According to experimental and LCA

17

results, the optimum microwave power level would be 300 W. The primary impact

18

source is microwave pyrolysis because of high energy usage.

19 20 21

Keywords: Heterogeneous Fenton oxidation; Trichloroethylene; Sewage sludge biochar; Catalyst; Life cycle assessment

22 23

1. Introduction

24

1

25

Municipal wastewater treatment is essential for environmental protection and

26

public health, especially in highly populated urban areas. However, it is inevitable that

27

wastewater treatment plants generate large amounts of sewage sludge which is the

28

largest in volume among the byproducts of wastewater treatments and may contain

29

pathogenic and toxic substances to a great degree (Werthera and Ogada, 1999; Fytili and

30

Zabaniotou, 2008; Samaras et al., 2008). The processing and disposal of sewage sludge

31

is considered as one of the most complicated environmental problems in this sector

32

(Werthera and Ogada, 1999). Although sewage sludge is generally processed by

33

thickening, stabilization, conditioning and dewatering on site to reduce water and

34

organic contents and to eliminate pathogens, its final disposal is still a matter of great

35

concern. Sewage sludge is traditionally disposed by land filling and ocean dumping.

36

Nowadays, the reuse and recycling of sewage sludge has attracted increasing attention

37

(Bridle and Pritchard, 2004; Stasta et al., 2006; Khwairakpam and Bhargava, 2009;

38

Tyagi and Lo, 2013; Smol et al., 2015). In the sewage sludge management, circular

39

economy could be achievable through resource recycling and energy recovery using

40

appropriate technologies (Smol et al., 2015). There are several recently developed AOP

41

systems, including catalytic, cavitation-based and peroxone-based processes, which can

42

be alternatives to the Fenton process (Boczkaj et al., 2017; Boczkaj et al., 2018; Gągol

43

et al., 2018a; Fernandes et al., 2019a; Gągol et al., 2019). Besides, the hydroxyl and

44

sulfate radical related AOPs as well as the processes based on cavitation can be useful

45

(Boczkaj and Fernandes, 2017; Gągol et al., 2018b). There are also promising types of

46

AOPs based on persulfates (Shah et al., 2018; Fernandes et al., 2018; Fernandes et al.,

47

2019b; Yuan et al., 2020).

48

The biochar produced by pyrolysis or carbonization of sewage sludge can be

2

49

utilized as a source of catalyst for the heterogeneous Fenton and Fenton-like oxidation

50

of organic contaminants in aqueous solution (Tu et al., 2012; Gu et al., 2013; Yuan and

51

Dai, 2014; Nidheesh, 2015; Munoz et al., 2015; Yuan and Dai, 2017). The Fenton

52

reaction is an advanced oxidation process (AOP) that uses hydrogen peroxide (H2O2)

53

and ferrous ions (Fe2+) for the production of a powerful oxidant, the hydroxyl radicals

54

(HO•) (Nidheesh, 2015). However, there are several disadvantages to limit the practical

55

applications of the classical homogeneous Fenton process, including the low and narrow

56

pH range where applicable (commonly 3–5), generating iron sludge, and the difficulty

57

in catalyst recovery. The heterogeneous Fenton process has therefore attracted much

58

interest, since it degrades organic contaminants in a wider pH range and possesses less

59

catalyst loss than the homogeneous Fenton process (Tu et al., 2012; Gu et al., 2013;

60

Nidheesh, 2015; Munoz et al., 2015). In addition to silicon and calcium, sewage sludge

61

is also rich in iron species (Gu et al., 2013). The pyrolysis of sewage sludge can

62

concentrate heavy metals in the solid residue and convert organic matter into renewable

63

biofuels and green chemicals (Yuan and Dai, 2017). The material derived from sewage

64

sludge can be utilized as a highly active and stable heterogeneous catalyst for Fenton

65

oxidation (Nidheesh, 2015; Yuan and Dai, 2014).

66 67

Trichloroethylene (TCE) is a chlorinated solvent widely used for metal degreasing,

68

dry cleaning, and chemical extraction. Contamination of soil and groundwater by TCE

69

has become a serious environmental problem and health risk (Teel et al., 2001; Xiu et al.,

70

2010; Che et al., 2011; Choi and Lee, 2012; Chiu et al., 2013; Yuan et al., 2014). TCE is

71

carcinogenic to humans by all exposure routes and also potentially hazardous to human

72

health for non-cancer toxicity (Chiu et al., 2013). The Fenton process is effective for the

3

73

degradation of TCE (Che et al., 2011; Choi and Lee, 2012; Yuan et al., 2014). Compared

74

with 91% TCE degradation accomplished by the classical homogeneous Fenton

75

oxidation at pH 3, the heterogeneous Fenton oxidation using goethite as catalyst can

76

provide an almost complete mineralization of TCE (>99%) (Teel et al., 2001). This

77

study aimed at the heterogeneous Fenton oxidation of TCE over the biochar produced

78

by microwave pyrolysis of sewage sludge. Microwave pyrolysis is faster and more

79

efficient than conventional pyrolysis (Yin et al., 2012). The effect of microwave power

80

level on TCE degradation was investigated. In addition to the importance of technical

81

feasibility, it is also necessary to consider the potential impacts to the environment when

82

using sewage sludge biochar for heterogeneous Fenton oxidation of TCE. Therefore, life

83

cycle assessment (LCA) was carried out to evaluate the environmental impacts of this

84

technique.

85 86

2. Material and methods

87 88

2.1. Materials

89 90

Sewage sludge used in this study was provided by the Dihua Sewage Treatment

91

Plant, Taipei, Taiwan. The as-received sewage sludge was air dried for several months

92

and then oven dried at 105 °C for 1 hr prior to microwave pyrolysis experiments and

93

characterization tests. The reagent grade TCE and hydrogen peroxide were obtained

94

from Sigma-Aldrich.

95 96

2.2. Experimental details

4

97 98

2.2.1. Microwave pyrolysis

99 100

The microwave pyrolysis of sewage sludge was carried out by using a

101

laboratory-scale microwave oven which provides single-mode microwave irradiation at

102

2.45 GHz. A schematic diagram of the microwave heating system can be found

103

elsewhere (Huang et al., 2015). Sample crucible and reaction tube were both made of

104

quartz. A three-stub tuner was placed in the middle of microwave propagation pathway

105

to regulate MW incident angle and to let the peak of microwave be located at the center

106

of sample crucible. There was a short-circuit plunger placed at the end of microwave

107

propagation pathway to modify the wavelength phase of microwave. After starting

108

microwave irradiation, the three-stub tuner and short-circuit plunger were both carefully

109

adjusted to let the reflected microwave power level be as low as possible. The heat

110

generated by the reflected microwave was absorbed by a water load device, whose

111

temperature was controlled by using a thermostat. Reflected microwave power level

112

was continuously monitored by using a power meter.

113 114

In each experiment, about 10–15 g sewage sludge sample was filled in the quartz

115

crucible, and then it was placed at the center of reaction cavity. The inert atmosphere of

116

reaction cavity was maintained by purging the pure nitrogen gas (99.99%) at 100

117

mL/min flow rate. When the nitrogen purging is enough to keep the reaction atmosphere

118

inert, the microwave heating system was switched on to the designated microwave

119

power level for 30 min process time. In order to prevent the hazard from non-ionizing

120

radiation, both three-stub tuner and short-circuit plunger were adjusted to minimize the

5

121

reflected microwave power level. In this study, sewage sludge sample was heated at

122

microwave power levels of 200, 300, and 400 W. These microwave power levels were

123

chosen because lower or higher than the microwave power level range would result in

124

incomplete pyrolysis or overheating of sewage sludge, respectively. The temperatures of

125

sewage sludge during the microwave heating experiments were measured by using a

126

K-type thermocouple sensor placed at the bottom of quartz crucible. In this study, each

127

experiment was repeated in triplicate to obtain mean and standard deviation values for

128

the experimental result.

129 130

2.2.2. Fenton oxidation

131 132

The Fenton oxidation of 10 mg/L TCE in 100 mL aqueous solution was carried out

133

by using 20 mM hydrogen peroxide and 0.05 g sewage sludge biochar. The temperature

134

of the reaction system was controlled at 25 °C. The initial solution pH was adjusted to

135

approximately 3.1, 4.8, and 6.8. The TCE solution was poured into a 125-mL

136

Erlenmeyer flask which was sealed with a rubber cap to minimize vaporization loss. A

137

glass tube was inserted into the rubber cap for syringe sampling. During the experiment,

138

the TCE solution was sampled for 1 mL at 0, 2, 5, 10, 20, 30, 50, 80, and 120 min. the

139

solution was filtered by using a 0.45-µm PTFE membrane filter to separate sewage

140

sludge biochar from the solution and to prevent further oxidation. All samples were

141

tightly sealed before characterization. The reactor and sampling vials were coated by

142

aluminum foil to prevent the photolysis of TCE. The Fenton oxidation experiment was

143

repeated in triplicate to obtain mean and standard deviation values for the experimental

144

result. In addition, to observe the change in catalytic activity of sewage sludge biochar,

6

145

The Fenton oxidation of TCE was repeated for five times.

146 147

2.3. Analytical methods

148 149

The proximate analyses of raw and pyrolyzed sewage sludge were carried out

150

based on the standard test methods D7582 and D3172 published by the American

151

Society for Testing and Materials (ASTM). The proximate analyses were performed by

152

using a thermogravimetric analyzer (TA Instruments SDT Q600). The specific surface

153

area was determined by using Micromeritics ASAP 2020M. The point of zero charge

154

(PZC) was determined by using a zeta potentiometer (Malvern Instrument Zetasizer

155

2000). The metal compositions of sewage sludge and biochar were determined by using

156

an Agilent inductively coupled plasma–optical emission spectrometer (ICP–OES 700

157

series). Toxicity characteristic leaching procedure (TCLP), which is a test to measure

158

the leachability of toxic elements (e.g., heavy metals) from hazardous waste, was

159

carried out based on a standard test method (NIEA R201.14C). The mercury released

160

during the microwave pyrolysis of sewage sludge was calculated by the difference

161

between the mercury contents of raw and pyrolyzed sewage sludge. The mercury

162

content was determined by using the ICP–OES as aforementioned. The quantitative

163

analysis of TCE was performed by using a gas chromatography analyzer (Agilent GC

164

7890) equipped with electron capture detector (ECD). Static headspace method was

165

used to skip the extraction process of TCE. The TCE contained solution with a volume

166

of 5 mL was filled into 20-mL headspace vials equipped with PTFE septa and screw

167

caps. The distribution of TCE in aqueous and gas phases at 25 °C was balanced for

168

more than 30 min. The volume of gas sample injected to the GC system was 1 mL for

7

169

each test. The temperatures of inlet and detector were set at 180 and 300 °C,

170

respectively. The temperature of oven was controlled at 80 °C in 6.5 min. The carrier

171

gas was ultra-pure nitrogen at a flow rate of 5 mL/min. The total hydrocarbons (THCs)

172

content of gaseous product produced by microwave pyrolysis of sewage sludge (i.e., the

173

non-condensable fraction of organic vapor released by biomass pyrolysis) was

174

determined by using the GC analyzer equipped with flame ionization detector (FID) and

175

Supelco Equity–5 capillary column. The standard used for THCs analysis was methane

176

(concentration balanced by pure nitrogen gas). The temperatures of inlet and detector

177

were set at 200 and 250 °C, respectively. The temperature of oven was controlled at 40

178

°C in 5 min. The flow rates of air and pure hydrogen were 200 and 100 mL/min. The

179

limit of detection (LOD), limit of quantification (LOQ) and linearity (R2) were 0.02

180

mg/L, 0.07 mg/L and 0.9997 for TCE analysis, and 0.07%, 0.23% and 0.9993 for THCs

181

analysis, respectively. The scanning electron microscope (SEM) images of sewage

182

sludge biochar were obtained by using a field emission microscope (Hitachi S-4800)

183

operating at 15.0 kV accelerating voltage. Static headspace analysis was used to skip

184

extraction process. The operational parameters for separating TCE and dechlorinated

185

intermediate were based on the literature (Lien and Zhang, 2001).

186 187

2.4. Life cycle assessment

188 189

In this study, LCA was performed on Simapro 8.0 platform with Ecoinvent 3.0

190

database. Process inventory was based on literature reviewed, experiment data, and

191

database. IMPACT 2002+ was used as an impact assessment model to transform

192

materials, energy inputs, and emissions into environmental impact points. Time scale of

8

193

global warming was set to 100 years. The Dihua Sewage Treatment Plant (Taipei,

194

Taiwan) was chosen to be case study site. The scenario of sewage sludge biochar used

195

as a catalyst for TCE degradation was built based on experiment results. Functional unit

196

is the degradation of 0.1 mg TCE within 2 hours. The System boundary of sewage

197

sludge recycling for TCE degradation is shown in Fig. 1. Sewage sludge was considered

198

as a burden-free material because the environmental impact caused by sewage sludge

199

production was included in wastewater treatment. The life time of sewage sludge

200

biochar starts from thermal drying process in Dihua Sewage Treatment Plant. To

201

simplify the analysis, the chemical input for pH adjustment was not considered, and the

202

primary product of the Fenton oxidation of TCE was assumed to be non-toxic chloride.

203 204

Figure 1

205 206

The thermal drying process, which decreased water content of sewage sludge to

207

less than 10%, was powered by digestion gas utilization and electricity. The dried

208

sewage sludge was then sent to microwave pyrolysis to produce biochar. Electricity was

209

consumed by direct circuit (DC) power supply, circulated water cooler, and microwave

210

generator. The emissions of total hydrocarbons (THCs) and mercury to air during the

211

microwave pyrolysis of sewage sludge were also taken into account. The input of

212

hydrogen peroxide and the emission of heavy metals were considered in the TCE

213

degradation stage. Heavy metal emission was based on the TCLP result. The items and

214

sources of the inventory data are described in Table 1, and the full inventory data are

215

listed in the Appendix (Table A.1). The damage assessment model used in this study

216

was IMPACT 2002+. The original version, IMPACT 2002, was developed by Swiss

9

217

Federal Institute of Technology for human toxicity and aquatic and terrestrial

218

ecotoxicity assessment. IMPACT 2002+ was modified by adapting characterization

219

factors from existing methods, such as Eco-indicator 99, CML 2001, IPCC 2001 and

220

2007, and Cumulative energy demand (Hischier et al., 2010).

221 222

Table 1

223 224

3. Results and discussion

225 226

3.1. Microwave pyrolysis

227 228

The maximum temperatures and solid yields of microwave pyrolysis of sewage

229

sludge at different microwave power levels are listed in Table 2. Each microwave power

230

level was repeated in triplicate, and the standard deviation was less than 3% of the

231

microwave power level designated. The maximum temperature of microwave pyrolysis

232

of sewage sludge at 200 W was only approximately 193 °C, much lower than the

233

maximum temperatures of approximately 412 and 430 °C at 300 and 400 W,

234

respectively. On the contrary, the solid yield at 200 W was approximately 63%, much

235

higher than the solid yields of approximately 38% and 35% at 300 and 400 W,

236

respectively. The experimental results show that the solid yield decreased with

237

increasing microwave power level, which may be attributable to the more intense

238

thermal decomposition occurred under the stronger microwave irradiation. However, the

239

result of microwave pyrolysis of sewage sludge at 200 W was substantially different

240

from those operated at 300 and 400 W. The mercury and THCs emissions from

10

241

microwave pyrolysis of sewage sludge at different microwave power levels are

242

illustrated in Fig. 2. In general, both mercury and THCs emissions increased with

243

increasing microwave power level. The microwave pyrolysis of 1 g raw sewage sludge

244

released approximately 0.18 mg mercury and 0.03 mg THCs at 200 W, and

245

approximately 0.3 mg mercury and 0.17 mg THCs at 300 W. The mercury released at

246

400 W was comparable to that at 300 W. However, the THCs emission at 400 W was

247

approximately 0.9 mg per g raw sewage sludge, much higher than those processed at

248

200 and 300 W.

249 250

Table 2

251 252

Figure 2

253 254

3.2. Characteristics of sewage sludge biochar

255 256

The proximate compositions, specific surface areas, and PZCs of raw sewage

257

sludge and microwave pyrolysis biochars are listed in Table 3. According to the volatile

258

matter contents of biochars, the microwave power level of 200 W may not be efficient

259

for the production of stable sewage sludge biochar, since the volatile matter content of

260

the biochar was approximately 39.1 wt%, much higher than the volatile matter contents

261

of approximately 11.5 and 11.8 wt% of the biochars produced at 300 and 400 W,

262

respectively. The lower volatile matter contents could be attributable to the more intense

263

thermal decomposition occurred at higher microwave power levels as aforementioned.

264

The specific surface areas of biochars produced at 300 and 400 W were approximately

11

265

37.4 and 33.1 m2/g, whereas the specific surface area of biochar produced at 200 W was

266

only 0.9 m2/g. Therefore, 200 W seems to be not sufficient to increase the porosity of

267

sewage sludge biochar, which may be attributable to the relatively low devolatilization

268

extent under the low microwave power level. The PZCs of 200, 300, and 400 W sewage

269

sludge biochar were approximately 3.9, 5.5, and 5.3, respectively. At low pH (below the

270

PZC), the biochar surface is positively charged to attract anions and thus to affect the

271

Fenton oxidation.

272 273

Table 3

274 275

The SEM images of biochars produced at different microwave power levels are

276

shown in Fig. 3, and the energy dispersive x-ray (EDX) analysis results are listed in the

277

Appendix (Table A.2). The SEM images illustrate that the particle and pore sizes of 200

278

W biochar were much bigger than those of 300 and 400 W biochars. The plain surface

279

of 400 W biochar indicates that there could be occurrence of melting or sintering

280

phenomenon during the microwave pyrolysis of sewage sludge at the relatively high

281

microwave power level. In general, due to the more intense pyrolysis accomplished by

282

the higher microwave power level, the carbon content decreased but the contents of

283

oxygen and metal elements increased with increasing power level. The surface iron

284

content of 400 W biochar can be as high as approximately 20 wt%. The specific surface

285

areas and the SEM images indicate that there was little carbonization occurred during

286

the microwave pyrolysis of sewage sludge at 200 W, which can be confirmed by the

287

relatively high volatile matter content of the biochar. Because of the hotspot formation

288

by microwave heating, the sintering phenomenon may occur during the microwave

12

289

pyrolysis of sewage sludge at 300 and 400 W. Besides, there would be more sintering

290

effect at 400 W due to the stronger microwave irradiation, resulting in the lower specific

291

surface area than that of 300 W biochar. The metal compositions of raw and pyrolyzed

292

sewage sludge are listed in Table 4. Arsenic was not found in all tests. The highest metal

293

content of both raw and pyrolyzed sewage sludge were iron. The iron content was

294

directly proportional to microwave power level. The mercury content substantially

295

decreased after microwave pyrolysis, which shows that mercury would be released

296

during the process. On the contrary, the other metals remained in the sewage sludge

297

biochars. This may imply that the volatilization of these metals is limited during the

298

microwave pyrolysis of sewage sludge. In addition to the metals listed in Table 4, the

299

sewage sludge could also contain silicon, calcium, and aluminum (Domínguez et al.,

300

2003), which would not cause catalytic effect on the Fenton oxidation.

301 302

Figure 3

303 304

Table 4

305 306

3.3. Fenton oxidation

307 308

The Fenton oxidation experiments of TCE using different sewage sludge biochars

309

as catalysts were attempted at pH 3.1, 4.8, and 6.8, from acidic to neutral conditions.

310

The removal efficiencies of TCE after the processing time of 120 min are illustrated in

311

Fig. 4, and the changes of TCE concentrations by time are shown in the Appendix (Fig.

312

A.1–3). The experimental results of non-biochar blank indicate that there was still some

13

313

TCE degraded by H2O2 oxidant without the presence of biochar, and the highest TCE

314

removal efficiency (approximately 43%) occurred at pH 4.8. This may imply that there

315

is more oxidizing power of H2O2 alone for the degradation of TCE at the pH value,

316

without the formation of free radicals. When the 200 W biochar was used at all pH

317

values, the TCE removal efficiencies were substantially lower than those using 300 and

318

400 W biochar. Therefore, the biochars produced at higher microwave power levels can

319

provide better catalytic effect and thus better TCE removal efficiency, which may be

320

attributable to their higher iron contents and specific surface areas as aforementioned.

321

The iron contents of 300 and 400 W biochars were approximately 36 and 42 mg/g,

322

whereas the specific surface areas were approximately 37 and 33 m2/g, respectively.

323

Since the TCE removal efficiency of 300 W biochar was higher than that of 400 W

324

biochar, specific surface area may play a more important role than iron content.

325

However, this phenomenon needs to be further justified. Besides, the TCE removal

326

efficiency was in reverse proportion to pH value, which matches the requirement of

327

acidic condition for Fenton oxidation. The highest TCE removal efficiency

328

(approximately 83%) was obtained by using 300 W biochar at pH 3.1. The optimal pH

329

for the TCE removal in the H2O2/biochar system was similar to the conventional Fenton

330

oxidation process, so the chemical cost for pH adjustment would not be saved when

331

applying the sewage sludge biochar as a heterogeneous catalyst. However, to prevent

332

the loss of catalyst, heterogeneous Fenton oxidation would still be a better choice. There

333

were only approximately 37% and 24% TCE removal efficiencies when using the 200

334

W biochar at pH 4.8 and 6.8, respectively, lower than those of non-biochar blank. This

335

may be attributable to the chlorinated organic compounds produced by microwave

336

pyrolysis of sewage sludge (Fonts et al., 2009), which were released as or converted

14

337

into TCE, resulting in the lower removal efficiency. It has been reported that the TCE

338

degradation can be almost completed by Fenton reaction in pyrite suspension (Che et

339

al., 2011). However, the pyrite concentration was as high as 0.21–12.82 g/L. Although

340

the TCE removal efficiency in this study was not such high, it would be expectable that

341

the efficiency would be higher with the addition of more sewage sludge biochar.

342 343

Figure 4

344 345

The experimental results of non-oxidant blank with the presence 400 W biochar

346

indicate that the removal of TCE may come from the adsorption onto sewage sludge

347

biochar. Comparing the removal efficiencies of non-oxidant blank with those of Fenton

348

oxidation using 400 W biochar, the adsorption phenomenon would be of great

349

importance to remove TCE from aqueous solution. Therefore, when using the same

350

sewage sludge biochar, the relatively high TCE removal efficiencies under the neutral

351

condition could be due to its adsorption onto the biochar. It is difficult to determine and

352

compare the relative contributions of adsorption and oxidation by H2O2 and free

353

radicals. However, since the oxidizing power of free radicals is much stronger than that

354

of H2O2, the TCE removal could be primarily come from the contribution of free radical

355

oxidation. Besides, in the heterogeneous Fenton oxidation system, it may be possible

356

that TCE is quickly oxidized by free radicals before being adsorbed onto sewage sludge

357

biochar. It has been reported that TCE may be converted into lower molecular weight

358

acids, CO2, and chloride by oxidative degradation (Pham et al., 2009). Che et al. (2011)

359

pointed out that, because of the high mass balance of chlorine, the dechlorination of

360

TCE would be much faster than the formation of acids and non-chlorinated products.

15

361 362

To test the reproducibility of TCE degradation, the H2O2 oxidation using 400 W

363

sewage sludge biochar at pH 7.5 was repeated for five times. The experimental results

364

are shown in Fig. A.4. It can be seen that the TCE removal efficiencies fluctuated from a

365

low of 18% (third test) to a high of 36% (second test), and the removal efficiency of

366

final test was 32%. Therefore, it may be concluded that the reactivity of sewage sludge

367

biochar can be maintained after several uses. Besides, the reproducibility of TCE

368

removal was not as satisfactory as it should be. This could be attributable to the

369

complicated composition and unstable property of sewage sludge biochar. The

370

durability of sewage sludge biochar provides an advantage for using as a permeable

371

reactive barrier to remediate groundwater pollution, in comparison with the

372

performance decrease with time of zero-valent iron (ZVI), the most common material

373

used up to date (Obiri-Nyarko et al., 2014). It has been reported that H2O2 can be

374

activated by biochar, because the persistent free radicals (PFRs) contained in biochar

375

could provide a single-electron transfer from PFRs to H2O2 to produce hydroxyl

376

radicals (Fang et al., 2014). The PFRs could be generated by the electron transfer from

377

phenolic compounds to transition metals loaded on biochar (Fang et al., 2015).

378

Therefore, in addition to acting as a catalyst, sewage sludge biochar may activate H2O2

379

by another mechanism.

380 381

3.4. Toxicity characteristic leaching procedure

382 383 384

There were some hazardous heavy metals found in the raw sewage sludge and its derived biochar as aforementioned (Table 4), so it would be necessary to test the

16

385

leachability of the metals. The TCLP results of sewage sludge biochar produced at

386

different microwave power levels are compared with the groundwater pollution control

387

standards of Taiwan, as listed in Table 5. It can be seen that zinc possessed the highest

388

leachability which was directly proportional to the microwave power level. The leachate

389

from the 400 W biochar had the highest zinc concentration (approximately 32 mg/L),

390

much higher than the standard for drinking water use but still lower than that for

391

non-drinking water use. The zinc concentration in the leachate of 200 W biochar was

392

only approximately 3.2 mg/L, much lower than those of 300 and 400 W biochar. The

393

relationship between the zinc concentration in leachate and the microwave power level

394

was not found for other heavy metals. Like zinc, the leachate concentrations of

395

chromium, copper, and nickel were higher than the standards for drinking water use but

396

lower than that for non-drinking water use. However, the lead concentrations leached

397

out from the three biochars were all slightly higher than the standard for non-drinking

398

water use. Arsenic, cadmium, and mercury were not found in the leachate. Therefore, to

399

apply the sewage sludge biochar for in-situ remediation of TCE-contaminated

400

groundwater, the potential release of zinc, chromium, copper, nickel and lead needs to

401

be prevented or mitigated.

402 403

Table 5

404 405

After Fenton oxidation experiments, the metals in the reaction solutions were

406

determined to observe the metal leachability during the process. The metal

407

concentrations in the reaction solutions of heterogeneous Fenton oxidation under

408

different conditions are illustrated in Fig. 5. It can be seen that only copper, iron, and

17

409

zinc were found in the reaction solutions. The highest copper, iron, and zinc

410

concentrations were 0.10, 0.14, and 0.35 mg/L, respectively. These were all in the

411

reaction solutions processed at pH 3. Copper was not found at pH 5 and pH 7, which

412

indicates that it would only be leached out from sewage sludge biochar at acidic

413

condition. In most cases, the concentrations of iron and zinc increased with decreasing

414

pH value. Therefore, the effect of pH value on the metal leachability is significant. On

415

the other hand, it is difficult to find out a substantial relationship between metal

416

leachability and sewage sludge biochar. Furthermore, the lower metal concentrations at

417

higher pH values may be owing to the formation of hydroxides or the adsorption

418

phenomenon of metal ions onto the biochar.

419 420

Figure 5

421 422

3.5. Life cycle assessment

423 424

The midpoint and endpoint environmental impacts caused in the life cycles of

425

different sewage sludge biochar were assessed. The midpoint is considered to be a link

426

in the cause-and-effect chain of impact pathway prior to the endpoint (Bare et al., 2000).

427

The midpoint impact is shown as equivalent amount of reference species, and it is

428

transformed into endpoint impact through normalization and weighting processes. The

429

endpoint environmental impact is classified into four categories: resources, climate

430

change, ecosystem quality, and human health. Environmental impact in the midpoint

431

categories of carcinogens, non-carcinogen, ionizing radiation, ozone layer depletion,

432

and photochemical oxidation contribute to human health impact in endpoint. Aquatic

18

433

ecotoxicity, terrestrial ecotoxicity, aquatic acidification, aquatic eutrophication,

434

terrestrial acidification/nutrification, and land occupation are classified as ecosystem

435

quality. Global warming is connected to climate change. Non-renewable energy and

436

mineral extraction are categories linked to resources in endpoint. The midpoint impacts

437

of sewage sludge biochar are listed in Table A.3. The impacts were transformed into

438

impact scores as illustrated in Fig. A.5. In the category of human health, the item of

439

respiratory inorganics is the main impact contributor and carcinogens impact is of

440

second importance. The impacts of ecosystem quality and resources categories

441

primarily come from terrestrial ecotoxicity and non-renewable energy, respectively.

442 443

The endpoint environmental impact of sewage sludge biochar is shown in Fig. 6. It

444

is clear that 200 W biochar causes the highest environmental impact. This is because of

445

the highest 200 W biochar dosage required for the heterogeneous Fenton oxidation of

446

TCE, which is owing to its lower reactivity as aforementioned. The 300 W biochar

447

causes the lowest environmental impact, so it could be most environmentally-friendly.

448

The share percentages of environmental impact to the resources, climate change,

449

ecosystem quality, and human health categories are approximately 27.7%, 27.3%, 3.4%,

450

and 41.7%, respectively. Among the four endpoint impact categories, human health is

451

the highest concern when utilizing sewage sludge biochar as a catalyst for

452

heterogeneous Fenton oxidation of TCE. This is because of the relatively high

453

environmental impacts in the midpoint categories of respiratory inorganics and

454

carcinogens (Fig. A.5). Compared with other impact categories, the environmental

455

impact to ecosystem quality is of least importance. This could be attributable to the

456

relatively low leachability of sewage sludge biochar, resulting in low heavy metal

19

457

pollution to ecosystems.

458 459

Figure 6

460 461

The analysis of impact hotspot was carried out to discover the primary source of

462

environmental impact caused by sewage sludge biochar during its life cycle. It was

463

assumed that there are three stages in the life cycle of sewage sludge biochar used as a

464

catalyst for the Fenton oxidation of TCE: 1) thermal drying to produce dry sewage

465

sludge, 2) microwave pyrolysis to produce sewage sludge biochar, and 3) land

466

application to degrade the TCE contamination in groundwater. The impact hotspot

467

analysis result is shown in Fig. 7. It can be seen that the primary source of

468

environmental impact is the stage of microwave pyrolysis, taking approximately

469

86–90%. The impact source of secondary importance is the stage of land application

470

with the share of approximately 9–13%. The impact source of thermal drying stage only

471

takes approximately 0.8–0.9%. To find out the key impact source of microwave

472

pyrolysis of sewage sludge, the impact hotspot of the process was also analyzed, as

473

shown in Fig. A.6. The result shows that approximately 88% of environmental impact

474

comes from the consumption of electricity. High energy usage is one of the crucial

475

disadvantages to limit the practicality and applicability of microwave pyrolysis (Huang

476

et al., 2016). If the microwave heating technology can be improved to reduce the input

477

energy required, the environmental impact caused by microwave pyrolysis will be

478

largely decreased to make it more feasible for the production of sewage sludge biochar

479

used as a catalyst for heterogeneous Fenton oxidation of TCE. Besides, it would be

480

necessary to evaluate the risk of oxygenated organic compounds formation (Makoś et

20

481

al., 2019).

482 483

Figure 7

484 485

In addition to environmental impact, the costs of producing sewage sludge biochar

486

using microwave pyrolysis and removing TCE using Fenton oxidation are of great

487

importance as well. According to the inventory data as aforementioned, to remove 1

488

mg/L of TCE from 1 m3 of contaminated groundwater, it would need approximately 19

489

U.S. dollars for the chemicals. Besides, to obtain the required amount of sewage sludge

490

biochar (approximately 9.2 g) produce by microwave pyrolysis at 300 W of microwave

491

power level for 30 min of processing time, it would need approximately from 142 to

492

710 US dollars for the electricity used for the overall microwave heating system

493

(assuming electricity rate at 0.1–0.5 U.S. dollar per kWh of electricity). Therefore, to

494

remove TCE from groundwater by using Fenton oxidation and microwave pyrolysis

495

techniques, the cost of energy is much higher than that of chemicals. Although

496

microwave pyrolysis can save energy, cost and time in comparison with conventional

497

pyrolysis (Li et al., 2016), it is inevitable that thermal treatment of biomass requires a

498

large amount of input energy. To overcome this disadvantage, it would need to apply

499

physical, chemical, biological, or combined processes for the pretreatment of sewage

500

sludge, to make it more suitable for microwave pyrolysis which is carried out at lower

501

microwave power level and for shorter processing time.

502 503

4. Conclusions

504

21

505

Sewage sludge biochar can be efficiently produced by using microwave pyrolysis

506

at relatively low microwave power levels. The highest TCE removal efficiency was

507

approximately 83%, obtained by using 300 W biochar at pH 3.1. Compared with 200 W

508

biochar, 300 and 400 W biochars can provide better catalytic effect, possibly due to

509

their higher iron contents and specific surface areas. TCE removal efficiency was in

510

reverse proportion to pH value, which matches the requirement of acidic condition for

511

Fenton oxidation. When using 200 W biochar at pH 4.8 and 6.8, TCE removal

512

efficiencies were lower than those of non-biochar blank. This may be attributable to

513

chlorinated organic compounds on the sewage sludge biochar which were released as or

514

converted into TCE, resulting in lower removal efficiency. The experimental results of

515

non-H2O2 blank indicate that part of the TCE removal may come from the adsorption

516

onto sewage sludge biochar. The reactivity of sewage sludge biochar can be maintained

517

after several uses. The durability of sewage sludge biochar provides an advantage for

518

using as a permeable reactive barrier to remediate groundwater pollution.

519 520

The hazardous heavy metals, including chromium, copper, nickel, lead, and zinc

521

were found in the leachate generated from the TCLP test of sewage sludge biochar. The

522

concentrations of the heavy metals except lead were lower than the standards for

523

non-drinking water use but higher than those for drinking water use. Zinc possessed the

524

highest leachability which was directly proportional to microwave power level. Copper,

525

iron, and zinc were found in the reaction solutions of heterogeneous Fenton oxidation.

526

The concentrations of iron and zinc increased with decreasing pH value in most cases,

527

so the effect of pH value on metal leachability is significant. The lower metal

528

concentrations at higher pH values may be owing to the formation of hydroxides or the

22

529

adsorption phenomenon of metal ions onto the biochar.

530 531

Because of the lowest reactivity of 200 W biochar, its dosage required for the

532

heterogeneous Fenton oxidation of TCE is highest, resulting in the highest

533

environmental impact. The 300 W biochar causes the lowest environmental impact, so it

534

could be most environmentally-friendly. Among the four endpoint impact categories,

535

human health is the highest concern because of the relatively high environmental

536

impacts in the midpoint categories of respiratory inorganics and carcinogens. The

537

environmental impact to ecosystem quality is of least importance. This is because of the

538

low leachability of sewage sludge biochar, resulting in low heavy metal pollution to

539

ecosystems. The analysis of impact hotspot shows that the primary source of

540

environmental impact is the microwave pyrolysis because of high energy usage. If the

541

input energy required for microwave heating can be lowered, its environmental impact

542

will be largely decreased to make it more feasible for the production of sewage sludge

543

biochar.

544 545

Acknowledgments

546 547

This work was financially supported by the NTU Research Center for Future Earth,

548

from The Featured Areas Research Center Program, within the framework of the Higher

549

Education Sprout Project, by the Ministry of Education (MOE) in Taiwan

550

(108L901003), and the Ministry of Science and Technology (MOST) of Taiwan (No.

551

107-2621-M-002-005).

552

23

553

Appendix A. Supplementary data

554 555 556

Supplementary data associated with this article can be found, in the online version, at http://

557 558

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687 688

29

689

Figure captions

690 691

Fig. 1. System boundary of sewage sludge recycling for TCE degradation.

692

Fig. 2. Mercury and THCs emissions from microwave pyrolysis of sewage sludge.

693

Fig. 3. SEM images of sewage sludge after microwave pyrolysis at (a) 200 W, (b) 300

694

W, and (c) 400 W.

695

Fig. 4. Fenton oxidation of TCE at different pH values.

696

Fig. 5. Metal concentrations in reaction solutions under different conditions.

697

Fig. 6. End-point environmental impact in the life cycle of sewage sludge biochar.

698

Fig. 7. Impact hotspot analysis of sewage sludge biochar.

699 700

30

701

Table legends

702 703

Table 1 Items and sources of inventory data.

704

Table 2 Results of microwave pyrolysis of sewage sludge.

705

Table 3 Characteristics of raw and pyrolyzed sewage sludge.

706

Table 4 Metal compositions of raw and pyrolyzed sewage sludge.

707

Table 5 TCLP results of sewage sludge biochar.

708

31

Table 1 Items and sources of inventory data. Process

Category

Thermal Drying

Energy input

Item

Data Source

Electricity

Operational data from the Dihua Sewage Treatment Plant

Digestion gas DC power supply Energy input Microwave generator Acetone Material input

Microwave Pyrolysis

Experimental results

Nitric acid Dried sewage sludge

Mercury

Difference between mercury contents of raw and pyrolyzed sewage sludge

THCs

Detected by GC-FID

Emission to air

Hydrogen peroxide Material input

Experimental results Sewage sludge biochar Copper Lead

TCE degradation Emission to water

Zinc TCLP results Iron Chromium Nickel

Table 2 Results of microwave pyrolysis of sewage sludge. Microwave power level (W)

Maximum temperature (°C)

Solid yield (%)

200±4.4

192.6±20.4

63.3±2.4

300±8.4

411.9±3.9

37.9±0.9

400±7.5

429.8±9.0

35.2±0.4

Table 3 Characteristics of raw and pyrolyzed sewage sludge.

Moisture (wt%)

Volatile matter (wt%)

Fixed carbon (wt%)

Ash (wt%)

Specific surface area (m2/g)

PZC

Raw

9.19±0.23

56.21±0.68

9.06±0.68

25.54±0.52

200W

3.33±0.13

39.05±2.04

21.25±1.09

36.37±1.08

0.879±0.10

3.86

300W

4.29±0.58

11.52±0.33

25.52±1.67

58.67±2.11

37.36±3.23

5.46

400W

3.51±0.05

11.76±0.17

20.78±0.38

63.96±0.50

33.07±7.22

5.31

Table 4 Metal compositions of raw and pyrolyzed sewage sludge. Cd

Cu

Fe

Ni

Pb

Zn

As

Cr

Hg

Raw

N.D.

0.188

15.512

N.D.

N.D.

1.225

N.D.

0.032

0.299

200W

0.016

0.431

25.406

0.102

0.120

2.455

N.D.

0.028

0.201

300W

0.021

0.628

36.269

0.163

0.235

3.308

N.D.

0.048

N.D.

400W

0.011

0.623

42.331

0.101

0.323

3.554

N.D.

0.033

N.D.

* Unit = mg/g. ** N.D.: not detected.

Table 5 TCLP results of sewage sludge biochar.

Item

200W

300W

400W

Groundwater pollution control standard Non-drinking water use

Drinking water use

As

N.D.

N.D.

N.D.

0.5

0.05

Cd

N.D.

N.D.

N.D.

0.05

0.005

Cr

0.25

0.22

0.17

0.5

0.05

Cu

2.88

0.95

1.97

10

1.0

Ni

0.11

0.29

0.41

1.0

0.1

Pb

0.60

0.64

0.62

0.5

0.05

Hg

N.D.

N.D.

N.D.

0.02

0.002

Zn

3.19

25.67

32.21

50

5.0

Fe

3.43

1.65

1.87

–

–

* Unit = mg/L. ** N.D.: not detected.

Fig. 1. System boundary of sewage sludge recycling for TCE degradation.

Fig. 2. Mercury and THCs emissions from microwave pyrolysis of sewage sludge.

Fig. 3. SEM images of sewage sludge after microwave pyrolysis at (a) 200 W, (b) 300 W, and (c) 400 W.

Fig. 4. Fenton oxidation of TCE at different pH values.

Fig. 5. Metal concentrations in reaction solutions under different conditions.

Fig. 6. End-point environmental impact in the life cycle of sewage sludge biochar.

Fig. 7. Impact hotspot analysis of sewage sludge biochar.

Highlights:

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Heterogeneous Fenton oxidation of TCE over sewage sludge biochar was studied.

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Sewage sludge biochar was produced by microwave pyrolysis.

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Biochar produced at 300 W provided the highest TCE removal efficiency.

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Environmental impact was evaluated by life cycle assessment.

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Human health is the highest concern whereas ecosystem quality is the least.

Declaration of interests ☑ The authors declare that they have no known competing financial interests or personal relationships that could have appeared to influence the work reported in this paper. ☐ The authors declare the following financial interests/personal relationships which may be considered as potential competing interests: