Sustainable Strategy for Enhancing Anaerobic Digestion of Waste

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A sustainable strategy for enhancing anaerobic digestion of waste activated sludge: Driving dissimilatory iron reduction with Fenton sludge Mingwei Wang, Zhiqiang Zhao, and Yaobin Zhang ACS Sustainable Chem. Eng., Just Accepted Manuscript • DOI: 10.1021/ acssuschemeng.7b03637 • Publication Date (Web): 18 Dec 2017 Downloaded from http://pubs.acs.org on January 3, 2018

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ACS Sustainable Chemistry & Engineering

1

Author list

2

Dr. Mingwei Wang

3

E-mail address: [email protected]

4

Dr. Zhiqiang Zhao

5


6

Prof. Yaobin Zhang

7


First author

Corresponding author

8 9

Affiliations: Key Laboratory of Industrial Ecology and Environmental Engineering

10

(Dalian University of Technology), Ministry of Education, School of Environmental

11

Science and Technology, Dalian University of Technology, Dalian 116024, China.

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Address: Key Laboratory of Industrial Ecology and Environmental Engineering

13

(Dalian University of Technology), Ministry of Education, School of Environmental

14

Science and Technology, Dalian University of Technology, No.2 Linggong Road,

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Ganjingzi District, Dalian City, Liaoning Province.

ACS Paragon Plus Environment

ACS Sustainable Chemistry & Engineering 1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 25 26 27 28 29 30 31 32 33 34 35 36 37 38 39 40 41 42 43 44 45 46 47 48 49 50 51 52 53 54 55 56 57 58 59 60

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A sustainable strategy for enhancing anaerobic digestion of waste activated sludge: Driving dissimilatory iron reduction with Fenton sludge

20 21

Authors:

22

Mingwei Wang, Zhiqiang Zhao, Yaobin Zhang*

23 24

Affiliations:

25

Key Laboratory of Industrial Ecology and Environmental Engineering (Dalian

26

University of Technology), Ministry of Education, School of Environmental Science

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and Technology, Dalian University of Technology, Dalian 116024, China.

28 29 30 31

* Correspondence: Tel: +86 411 8470 6263, Fax: +86 411 8470 6263;

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Abstract

35

Fenton process has been extensively applied for treatment of refractory organic

36

pollutants. While the potentially hazardous iron-containing sludge generated from the

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Fenton process requires proper treatment and disposal, due to its high Fe contents and

38

toxic organic matters involved. Considering that Fe(III) oxides exhibits an ideal

39

potential for enhancing anaerobic digestion (AD), in this study Fenton sludge with a

40

high-abundance of Fe(III) was introduced in AD of wasted activated sludge (WAS)

41

with the aims to improve the sludge digestion as well as to remove the organic matters

42

in Fenton sludge. Results showed that methane production and sludge reduction of

43

WAS were significantly improved, and the organic matters contained in Fenton sludge

44

was removed by 70.0%. Meanwhile, nearly a half of in Fenton sludge was converted

45

to Fe2+ via dissimilatory iron reduction during the digestion, in agreement with

46

microbial community analysis. The study suggests a Fe recycling between AD and

47

Fenton process that Fenton sludge can be used as an iron source to enhance AD,

48

during which most of harmful organic matters in Fenton sludge was removed and

49

Fe(II) generated can be serve as a reactant again for a new Fenton reaction.

50

Keywords: Fenton sludge; Dissimilatory iron reduction; Anaerobic digestion (AD);

51

Wasted activated sludge (WAS)



52

Introduction

53

Fenton and Fenton-like processes have been extensively applied for the treatment of

54

refractory wastewaters1, 2. This advanced oxidation technology utilizes hydroxyl

55

radical (•OH) produced from the catalyzing reaction between Fe2+ and hydrogen

56

peroxide under pH of 3-4 to oxidize refractory organics. During the processes,

57

however a mass of iron-containing sludge is generated when the pH of effluent is

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adjusted to neutral. This Fenton sludge contains lots of organic pollutants, heavy

59

metals and other harmful matters, thus having been listed as hazardous wastes in some

60

countries, which leads to high sludge disposal cost and requires to be treated carefully

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prior to discharge into environment3, 4.

62 63

Traditional methods such as combustion and cement stabilization cannot eliminate the

64

environmental risks of Fenton sludge and possibly result in pollutant transfer.

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However, iron making up as high as 20-40% (dry weight) of Fenton sludge mainly

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exists as a form of insoluble ferric iron5. Reusing the iron in Fenton sludge can be an

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attractive way to dispose this residue. In previous studies, Fenton sludge had been

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reused as an iron source for synthesizing coagulant6 or Fe-based heterogeneous

69

catalysts5, which however needed extra chemicals or energy input. Developing new

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ways to cost-effectively dispose Fenton sludge is highly desired to widely employ this

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advanced oxidative process in refractory wastes treatment.

72 73

Iron reducing bacteria (IRB) utilizes insoluble Fe(III) as terminal electron acceptor to


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gain energy from oxidation of organic compounds, are commonly present in anaerobic

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environments. They drive the microbial dissimilatory iron reduction proceeding and

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play a key role in iron cycling7-9. Interestingly, (semi) conductive Fe(III)/Fe(III)-Fe(II)

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oxides exhibit a positive effects on AD in various settings10-12. The potential

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mechanism involved is regarded that Fe(III)/Fe(III)-Fe(II) oxides enrich the IRB

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microorganisms that are capable of utilizing a variety of substrates and participating

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in the decomposition of complex organic matters via the dissimilatory iron reduction10.

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Many researchers investigated the effect of different iron oxides, i.e., hematite13, ferric

82

oxyhydroxide10 and magnetite11, 14, on the anaerobic treatment and observed a positive

83

effect of methane production or organic removal rate. Besides, natural iron minerals

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(hematite or magnetite) have also been confirmed to increase methane production and

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organic decomposition in paddy soils15 or in anaerobic sludge digesters16. In our

86

previous study, it was observed that adding ferric iron had positive effects on methane

87

production from pretreated sewage sludge17.

88 89

Further studies demonstrated that addition of conductive iron oxides such as

90

magnetite and hematite may also serve as the electrical conduits to facilitate direct

91

interspecies electron transfer (DIET) that is considered as an alternative to

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interspecies hydrogen/formate transfer (IHT/IFT) which could accelerate the

93

syntrophic conversion of alcohols and volatile fatty acids (VFAs) to methane15. And it

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was detected a high-abundance Syntrophomonadaceae known to proceed the

95

syntrophic conversion of VFAs to methane with the hydrogen-utilizing methanogens



96

via IHT.

97 98

Although different types of iron oxides had been used to improve the anaerobic

99

digestion performance, however, to the best of our knowledge, there had been no

100

studies to investigate iron oxides in Fenton sludge for improving methanogenesis and

101

sludge reduction during the sludge anaerobic digestion process. And there was no

102

study to treat Fenton sludge combined with biological process. Based on the above

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consideration, iron-containing Fenton sludge was used as a ferric iron source for

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enrichment of IRB microorganisms to accelerate anaerobic digestion of wasted

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activated sludge (WAS). In this study, Fenton sludge was dosed into anaerobic

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digesters treating WAS to investigate (1) the removal of organic matters of Fenton

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sludge (2) effects of Fenton sludge on anaerobic digestion, and (3) generation of

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ferrous iron from dissimilatory iron reduction. We expect to offer a sustainable

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strategy for enhancing anaerobic digestion of WAS as well as to dispose Fenton

110

sludge environmentally friendly.

111 112

Experimental Section

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Preparation of Fenton sludge powder.

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Iron-containing sludge was derived from a Fenton process that treated landfill

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leachate. Molar ratio of Fe2+ to H2O2 was 1:3, using 30 mmol/L FeSO4•7H2O as Fe2+

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source. After 24h static settlement, the Fenton sludge was filtered then dried at 105 ℃

117

for 4 h. Afterwards, it was pulverized into powder for use.


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118 119

Substrates and inoculum.

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WAS collected from a wastewater treatment plant (Dalian, China) was used as

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substrates for this study. Prior to the experiments, the solid content of WAS was

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diluted to about 6% with deionized water. The seed sludge was collected from an

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anaerobic digester of a waste sludge treatment plant of Dalian (China) with a

124

concentration of volatile solids (VS) about 35 g/L.

125 126

Batch experiments.

127

Two batch experiments were conducted in this study. The first experiment was to

128

investigate whether organic matters involved in Fenton sludge could be removed in

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anaerobic digestion. Before inoculation, the seed sludge was washed for three times

130

with 0.1 M PBS to remove organics from sludge as much as possible then digested for

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7 days till biogas production ceased. Five 120mL serum bottles were used for

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anaerobic digestion. The first four bottles were inoculated with 10 mL seed sludge

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taken from abovementioned. Then 0, 0.5, 1.0, 2.0 g Fenton sludge were added to the

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four groups labeled with F0, F0.5, F1.0, F2.0, respectively. The fifth group was added

135

with only 0.5 g Fenton sludge but no seed sludge (labeled with Fonly) to clarify

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whether the Fenton sludge would be degraded by itself. Then all groups were diluted

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into 20mL using deionized water. The trace elements were added according to the

138

reference18. Upon digestion, all the bottles were sealed with Teflon-faced butyl rubber

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stoppers and then flushed with N2 for 0.5 h in the headspace. The digestion was



140

operated for 16 days.

141 142

The second experiment was operated in another four 250 mL serum bottles to clarify

143

whether Fenton sludge could enhance the efficiency of anaerobic digestion. The WAS

144

was mixed with the seed sludge at a ratio of 9:1. The main characteristics of the mixed

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substrates are listed in Table 1. A mixture (200 mL) of WAS and seed sludge was

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incubated in each serum bottle. Then 0, 0.5, 1.0, 2.0 g Fenton sludge was respectively

147

added in these four bottles (R0, R0.5, R1.0, R2.0). Before the digestion, the oxygen of

148

the headspace and sludge of the bottles was removed via nitrogen gas aeration for 0.5

149

h. Afterwards the bottles were sealed by a cap which was drilled two holes to connect

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with a biogas sampling bag and a liquid sampling pipe. The digestion was operated

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for 24 days. During the digestion, the biogas produced from each bottle was collected

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into gasbag for analysis. 2 mL sludge was taken out every two day to measure

153

short-chain fatty acids.

154 155

All experiments were operated in the dark at 37±2 °C in an air-bath shaker (120 rpm)

156

and replicated in triplicate.

157 158

Chemical Analysis.

159

A scanning electron microscope (SEM, S4800, Hitachi, Japan) equipped with an

160

energy dispersive spectrometer (EDS) system was used to describe morphology

161

features of Fenton sludge. The elements of Fenton sludge containing C, H and N were


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analyzed by elemental analyzer (Vario EL, Elment, Germany). The main phase and

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crystalline properties of Fenton sludge was characterized using X-ray diffraction

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(XRD, Empyrean, PANalytical, Netherlands) and element chemical states were

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further analyzed by a X-ray photoelectron spectroscopy (XPS, ESCALABTM 250Xi,

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Thermofisher, America). At the beginning and end of the experiment, total chemical

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oxygen demand (TCOD), polysaccharide, protein, total solid (TS), volatile solid (VS)

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were measured. TS, VS, TCOD and SCOD were determined according to Standard

169

Methods for the Examination of Water and Wastewater (APHA, 1998). Proteins were

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analyzed with Lowry’s method using bovine serum albumin as a standard solution19.

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Polysaccharide was measured with phenol–sulfuric acid method using glucose as a

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standard solution20. The volume of biogas collected by the gas sampling bag was

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measured by a syringe. The CH4 and CO2 proportion of biogas were measured using a

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gas chromatograph with a thermal conductivity detector (TCD) (Tianmei,

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GC-7900P/TCD, China)21. Short-chain VFAs (including acetate, propionate butyrate

176

and valerate) were analyzed using another gas chromatograph with a flame ionization

177

detector (FID) (Tianmei, GC-7900P/FID, China) every two days. The analytic

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methods of gas chromatograph were according to the report by Jiang et al22. The total

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iron content of the Fenton sludge was determined using Inductively Coupled Plasma

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(ICP, Optima2000DV, perkinelmer, America). Fe2+ and total iron were analyzed by an

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adaptation of the ferrozine technique23. The oxidation reduction potential (ORP) was

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measured by an ORP combination class-body redox electrode (Sartorius PY-R01,

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Germany).



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184 185

DNA

extraction,

PCR

amplification

and

high-throughput

16S

rRNA

186

pyrosequencing.

187

Microbial community structure of initial seed sludge and digestion sludge of control

188

reactor (R0) and Fenton sludge reactor (R2.0) on day 8 (middle stage) and on day 24

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(final stage) were analyzed via high-throughput 16S rRNA pyrosequencing.

190 191

The detailed methods of DNA extraction, PCR24 and sequencing25 are provided in the

192

Supporting Information.

193 194

Results and Discussion

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Characterization of Fenton sludge powder.

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The obtained powder Fenton sludge appeared reddish brown color, in accordance with

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the high content of Fe in the Fenton sludge (26.8wt%, Fig. S1(b)). As shown in Fig.

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S1(a), SEM analysis demonstrated that the dried sample appeared irregular brick-like

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particles with a size less than 500 nm. The EDS analysis revealed that the contents of

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C, H and O in Fenton sludge was 20.8%, 1.6%, 36.3%, respectively. The high

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contents of these three elements were more likely linked to organic matters involved

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in the Fenton sludge. The organics of Fenton sludge primarily were resulted from

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landfill leachate that had been treated by Fenton process. Accordingly, the VS of

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Fenton sludge accounted for 24.85%.

205


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The structure and fractionation of ferric oxides are influencing factors that cannot be

207

neglected to drive iron reduction, because iron reducers have significantly different

208

capacities to transform ferric iron minerals with varied crystallinity, solubility and

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electrical potential. Many researchers had studied the effects of adding different ferric

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iron compounds in amorphous as well as crystalline forms, conductive as well as

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semi-conductive on anaerobic digestion in various settings26-28. To identify its phase

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and crystalline properties, Fenton sludge was characterized by X-ray diffraction and

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the pattern is presented in Fig. 1(a). The peaks (labeled with H) could be readily

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indexed a rhombohedral cell of α-Fe2O3 (hematite, space group: R3c) which was

215

consistent with the values given in the standard card (JCPDS, no.2-919). The

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unknown peaks (labeled with U) indicated that these were a certain impurities mixed

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in the flocculation of Fenton reagent. Chemical bonding states of the Fenton sludge

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were further analyzed by XPS. The main elements of Fenton sludge, such as C, O, Na,

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S, Fe, all had response in the XPS scanning (Fig. 1(b)). The Fe 2p XPS spectra (Fig.

220

1(c)) of the sample exhibit two peaks at 724.6 and 711.2 eV, corresponding to the Fe

221

2p1/2 and Fe 2p3/2 spin–orbit peaks of Fe2O3 (Fig. 3b). Moreover, a satellite peak at

222

718.9 eV (indicated by a circle), which is the characteristic of Fe2O329.

223 224

Removal of organics matters of Fenton sludge during the anaerobic digestion.

225

To investigate whether the organics involved in Fenton sludge could be removed

226

during anaerobic digestion, the Fenton sludge mixed with the seed sludge was added

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in a digester with no WAS added. The initial TCOD were 21931±2054, 27196±1195



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, 33513±1445, 43540±845 mg/L (Fig. 2) under the Fenton sludge dosage of 0, 0.5,

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1.0, 2.0 g (F0, F0.5, F1.0, F2.0) respectively, indicating that the seed sludge contributed a

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background TCOD of about 20000mg/L and the TCOD significantly increased with

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increase of Fenton sludge. After 16 days digestion, TCOD of the four groups were

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19406±719, 21063±779, 24021±1407, 28962±781 mg/L in F0, F0.5, F1.0, F2.0,

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respectively. Namely, the TCOD removal of the four groups was 11.5%, 22.6%,

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28.3%, 33.5%, respectively. Remarkably, the TCOD removal of the 0 g Fenton sludge

235

group (F0) was resulted from the endogenous respiration of seed sludge itself, which

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was a background value that should be subtracted when assessing the decomposition

237

of the Fenton sludge during the digestion. Considering this, after 16 days anaerobic

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digestion, the removal of organic matters in Fenton sludge were calculated as 70.0%,

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67.6%, 58.4% for adding 0.5, 1.0, 2.0 g Fenton sludge, respectively. A seed

240

sludge-free digestion test (Fonly) showed almost no COD removal in Fenton sludge,

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meaning that Fenton sludge could not be degraded without the seed sludge.

242 243

After 16 days, the accumulative CH4 production was 8.0±0.9, 12.7±3.6, 9.4±1.4,

244

8.4±1.9 mL for F0, F0.5, F1.0, F2.0 (Fig. 3). It indicated that CH4 production of each

245

Fenton sludge addition had no significant discrepancies. Comparatively, the

246

accumulative CO2 production was 2.5±0.4, 16.5±1.5, 20.2±0.8, 44.8±2.4 mL in

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the four groups, respectively, which significantly increased by addition of Fenton

248

sludge. Organics in the Fenton sludge provided electron donors available for

249

anaerobic digestion including methanogenesis, thereby increasing methane production.


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250

On the other hand, Fe(III) could serve as an electron acceptor to compete with

251

methanogens for common electron donors9. In other words, dissimilatory iron

252

reduction was an alternative to anaerobic respiration. Therefore, only slight increases

253

of methane production were obtained with Fenton sludge supplemented. Previous

254

studies demonstrated iron reducing bacteria could completely oxidize multi-carbon

255

compounds to carbon dioxide30. Thus CO2 could be produced from both

256

methanogenesis and dissimilatory iron reduction, and then production of CO2

257

increased obviously. The results suggested that the organic matters in the Fenton

258

sludge could serve as the electron donors for the anaerobic metabolism.

259

Correspondingly, organics were removed when Fenton sludge was added into the

260

anaerobic digestion.

261 262

Enhancing anaerobic digestion by adding Fenton sludge.

263

Anaerobic digestion of WAS is an efficient and sustainable technology to stabilize

264

sludge by means of sludge reduction and methane production simultaneously31. WAS

265

composed of high-strength macromolecule organics, such as polysaccharide and

266

protein, and their decomposition is quite slow, which results in the low fermentation

267

efficiency and long retention time32, 33. Previous study observed the increased methane

268

production from anaerobic digestion of WAS with adding ferric oxides10, 17. Therefore,

269

Fenton sludge was added in anaerobic digesters to investigate whether Fenton sludge

270

could be used as a ferric source to enhance anaerobic digestion of sludge.

271



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272

Methane yield is a valuable by-product of anaerobic digestion and an important index

273

to assess performance of anaerobic sludge digestion. After 24 days anaerobic

274

digestion, the accumulative methane production was 2153±194， 2423±10, 2531±

275

66, and 2604±87 mL under the Fenton sludge dosage of 0, 0.5, 1.0, 2.0 g, as shown

276

in (Fig. 4). Namely, the methane yield was 223, 251, 262, 270 mL/g-VS, respectively.

277

The methane production rate of the anaerobic sludge digestion plant of Dalian (China)

278

that the seed sludge was collected in this work were 210-240 mL/g VS methane

279

production which was consistent with our results. Compared with the control reactor,

280

the accumulative methane production of the Fenton sludge reactors increased 12.6%

281

for R0.5, 17.6% for R1.0 and 21.0% for R2.0. Especially, more increment of methane

282

production occurred in the initial stage. The methane production reached 300±27,

283

415±46, 423±39, 536±73 mL in R0, R0.5, R1.0 and R2.0 in the initial 8 days,

284

respectively, meaning that the methane production of these three reactors increased by

285

38.3%, 41.1%, and 78.8% compared with the control group (R0). The first few days of

286

anaerobic digestion is generally associated with the hydrolysis of sludge, a

287

rate-limiting stage of anaerobic digestion. The improvement of methane production

288

during anaerobic digestion of WAS especially in the initial 8 days indicated that the

289

addition of Fenton sludge enhanced sludge hydrolysis to produce small-molecule

290

substrates

291

Fe(III)-reducing microorganisms enriched by ferric oxides were capable of utilizing a

292

variety of substrates to participate in the decomposition of complex organics via the

293

dissimilatory iron reduction. Fenton sludge dosed in this work seemingly appeared a

available

for

methanogenesis.

Baek10


and

Zhang17

suggested

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294

similar effect.

295 296

TS, VS and TCOD removal are other main parameters to measure the efficiency of

297

anaerobic digestion and sludge reduction. The changes in VS, TS and total COD of

298

sludge before and after the digestion with the addition of Fenton sludge are shown in

299

Fig. 5(a). The addition of Fenton sludge promoted the TCOD removal compared with

300

the control group. The remained COD of the sludge after digestion was 37097±1128,

301

34410±377, 33043±76 and 30797±768 mg/L in R0, R0.5, R1.0, R2.0, respectively.

302

Accordingly, the TCOD removal was 42.7%, 46.9%, 49.0% and 52.5%, respectively.

303

If subtracting the COD from the Fenton sludge added, the TCOD removal ratio of the

304

WAS were 42.7%, 47.3%, 49.9%, 54.1%. The addition of Fenton sludge composing

305

of 75% inorganic substances and 25% organic substances inevitably increased the TS

306

of the reactors. The VS and TS (VS / TS) were 24.6 / 34.4, 24.2 / 33.3, 23.8 / 31.6 and

307

23.0 / 28.5 g/L, respectively in R0, R0.5, R1.0 and R2.0 group after subtracting the initial

308

VS and TS resulted from Fenton sludge. The VS and TS removal ratio were 49.0% /

309

45.4% for R0, 50.4% / 47.6% for R0.5, 51.7% / 49.9% for R1.0 and 54.4% / 54.8% for

310

R2.0. It meant that the addition of 2.0 g Fenton sludge increased the removal of TCOD,

311

VS and TS by 11.4%, 5.4% and 9.4%, respectively.

312 313

Polysaccharide and protein are two main organic components of WAS. During the

314

anaerobic digestion, those substrates were finally mineralized into methane and

315

carbon dioxide, accompanied with the sludge reduction. As shown in Fig 5(b). After



316

24 days digestion, the total polysaccharide in R0, R0.5, R1.0 and R2.0 was 464±12, 443

317

±10, 411±19 and 410±14 mg/L, while the total protein was 545±19, 483±4, 454

318

±4 and 426±10 mg/L. It meant that the removal efficiency of polysaccharide was

319

38.9%, 41.6%, 45.9% and 46.0%, respectively. And the removal efficiency of protein

320

was 33.7%, 41.3%, 44.8% and 48.2%, respectively. The enhancement of organic

321

matters reduction may be attributed mainly to the activity of the enzymes associated

322

with hydrolysis acidification which was observed to significantly increase with the

323

addition of Fe in previous study34.

324 325

Fig. 6 demonstrates the change of VFAs (acetate, propionate, butyrate, valerate) and

326

pH every two days during the 20-day experiment. With increase of Fenton sludge

327

from 0 to 0.5, 1.0 and 2.0 g, the concentration of VFAs increased from 3118 to 3860,

328

4010, and 4081mg/L in the initial two days. It suggested that the addition of Fenton

329

sludge promoted the hydrolysis of WAS to produce more simple organics. Then the

330

total VFAs of all reactors accumulated rapidly and peaked on day 4. Afterwards,

331

VFAs in all reactors declined significantly. On the end of the fermentation, the

332

concentration of VFAs were 1122±25, 820±147, 705±116, 387±94 mg/L in R0,

333

R0.5, R1.0 and R2.0, respectively, indicating that Fenton sludge also accelerated the

334

VFAs decomposition. In other words, Fenton sludge promoted both VFAs production

335

and consumption during the anaerobic digestion, which was in agreement with the

336

methane production (Fig. 4) as well as pH profiles (Fig. 7). Complex substrates are

337

oxidized to organic acids with anaerobic fermentative bacteria, followed by


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338

consumption of fermentation products by syntrophic acetogen bacteria and

339

methanogens, which can eliminate feedback inhibition of fermentation35. Conversion

340

of complex substrates (proteins and carbohydrates) to simples such as short-chain

341

fatty acids or hydrogen is the business of anaerobic fermentative bacteria, which

342

however are low performance in the initial stage of sludge digestion35, 36. Results in

343

this study suggested that Fenton sludge could significantly accelerate the

344

decomposition of complex substrates such as proteins and polysaccharide (Fig. 5)

345

contained in WAS, as well as the consumption of VFAs. As a result, anaerobic

346

digestion proceeded smoothly. From Fig. 7, the pH decreased to 6.57±0.15 for R0,

347

6.85±0.09 for R0.5, 6.99±0.06 for R1.0 and 7.05±0.02 for R2.0 in the initial stage,

348

and then rose to 7.83±0.10, 8.07±0.04, 8.10±0.03, and 8.13±0.01 in the next few

349

days. Interestingly, the pH of Fenton groups (R0.5, R1.0, R2.0) did not decrease to more

350

acidic levels with producing more VFAs in the initial days. On the contrary, the pH

351

rose along with the increase of Fenton sludge dosage. It is well recognized that

352

methanogens are sensitive to cultivation environment and pH is the critical factor for

353

achieving sustainable fermentation. The results suggested the addition of Fenton

354

sludge enhanced fermentation by attenuating acidification through reduction of iron

355

oxides Consistently, Dong13 et al found that hematite enhanced the consumption of

356

electron equivalents from organic substrates, effectively consumed acid produced by

357

fermentation. That was likely another reason for the more increment of methane

358

production of Fenton sludge groups in the initial stage.

359



360

Propionate is one of the main factors limiting the efficiency of AD because of the

361

relatively slow syntrophic metabolism (∆G = +76.1 kJ/mol)37. It was observed a

362

high-composition propionate (20.4-89.9%) in the control reactor (R0) during the

363

entirely anaerobic fermentation process (Fig. S2). With the increase of dosage of

364

Fenton sludge from 0 to 0.5, 1.0 and 2.0 g, the accumulation of propionate eased

365

gradually. The average ratio of propionate of VFAs of R0, R0.5, R1.0, and R2.0 during

366

the anaerobic digestion was 45.6%, 41.9%, 32.7%, and 24.5% respectively (Fig. S2).

367

Instead, the ratio of acetate of these four groups increased from 20.9% to 25.8%, 34.7%

368

and 46.3% (Fig. S2). This result suggested that the Fenton sludge could enhance

369

decomposition of propionate, in agreement with Zhang17 et al, who used rusty iron

370

scraps to enhance anaerobic digestion of WAS.

371 372

Dissimilatory iron reduction is an energetically favorable process to oxidize organics

373

compounds including VFAs and complicated contaminants such as aromatic

374

hydrocarbons, halogenated solvents and chlorinated benzenes. That can be a reason

375

for the higher removal of TCOD, VS, polysaccharide and protein. From the

376

perspective of thermodynamics analysis, Fe(III) reduction(-1410 KJ/mol, hematite)

377

gained more free energy than methanogenesis (-31~-185.5 KJ/mol, based on different

378

types of methanogenesis)38. In other words, dissimilatory iron reduction is more

379

favorable in thermodynamics than methanogenesis. Under low organic concentration,

380

the competition between Fe(III) reduction and methanogenesis may limit methane

381

production39, 40, just like the results of Fig. 3. While dissimilatory iron reduction


Page 18 of 45

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382

occurring under abundant organic conditions such as anaerobic digestion of WAS, the

383

methanogenesis was no longer restricted due to the abundant electron donors.

384

Reversely, dissimilatory iron reduction was an electron sink to couple with organics

385

oxidation, and the dissimilatory iron reduction had priority over methanogenesis in

386

the competition for electrons.

387 388

Migration and transformation of Fenton sludge.

389

Dissimilatory iron reduction was the reason for the higher performance of the

390

anaerobic sludge digestion with the addition of Fenton sludge. There were only a little

391

amount of iron released to liquid in all reactors after 24 days anaerobic digestion, and

392

almost all were ferrous due to the anaerobic environment (Fig. S3). Most of iron was

393

bonded with WAS in the sludge phase, because of the flocculation of ferrous and

394

ferric iron with macromolecular organics in the sludge. The total ferrous in sludge and

395

aqueous phases of R0, R0.5, R1.0 and R2.0 were 259±14, 508±13, 668±29, and 804

396

±23 mg/L (Fig. S3). As reported by other researchers, the dissolved Fe2+ might

397

decrease ORP of the anaerobic digester which was beneficial for acetate production

398

and reducing the propionate accumulation12, 41. In this study, the ORP was decreased

399

from -375±32 to -405±6, -417±23 and -436±12 mV with the dosage of Fenton

400

sludge from 0 to 0.5, 1.0 and 2.0g, respectively. This result was corresponding to the

401

VFA composition, namely the lower OPR decreased the accumulation of propionate.

402

Moreover, the activities of several key enzymes associated with hydrolysis and

403

acidification might be enhanced in the presence of iron32. There was a little amount of



404

iron in the control group (R0) after the digestion, likely a result from Fe-related

405

coagulants used during the sludge dewatering in the Wastewater Treatment Plant.

406

Subtracting the ferrous and ferric iron coming from the WAS, 44.13% for R0.5, 36.39%

407

for R1.0, and 24.18% for R2.0 ferric iron in Fenton sludge were converted into ferrous,

408

indicating dissimilatory iron reduction had been occurring during the digestion.

409 410

Microbial community analysis.

411

Microbial community structure of initial seed sludge and digestion sludge on the

412

middle (M) and final (F) stage were analyzed to gain insight into the microbial factors

413

linked to the performances (Fig. 8). Methanothrix was the dominant methanogens in

414

both control reactor and Fenton sludge reactor during the entirely anaerobic digestion

415

indicating that the aceticlastic pathway was probably the main route for

416

methanogenesis in all reactors42. They accounted for about 75.6% of the communities

417

in the seed sludge. But on middle stage which was considered as hydrolytic

418

acidogenesis stage, their relative abundance decreased to 67.4% of Fenton (M), while

419

there was no obvious change in Control (M). The decrease of Methanothrix was

420

accompanied by a significant increase in Methanospirillum species, the well-known

421

hydrogen-utilizing methanogens in many traditional anaerobic digesters43. They

422

accounted for only 2.2% of the communities in the seed sludge and increased to more

423

than 18.5% in Fenton (M), which was 10% higher than that in Control (M). It was

424

well recognized that the stage of hydrolytic fermentation is an acid accumulation

425

process. The significant increase in the abundance of Methanospirillum species might


Page 20 of 45

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426

be ascribed to the Fenton sludge that accelerated the decomposition of complex

427

organics to simples with the release of a large amount of hydrogen that stimulated the

428

microbial communities to enrich Methanospirillum species. Consistently, there was a

429

faster accumulation of VFAs in Fenton sludge reactors (Fig. 6). Also, the significant

430

methane production especially in the initial 8 days with the Fenton sludge

431

supplemented (Fig. 4) might due to the increased Methanospirillum species for they

432

are hydrogen-utilizing methanogens capable of maintaining the balance of hydrogen

433

partial pressure of anaerobic system. On the final of the digestion, most organic

434

matters were consumed and there were almost no VFAs residual. In contrast, it had a

435

remarkable accumulation of propionate in control reactors. Accordingly, the relative

436

abundance of Methanospirillum species of Control (F) was 37.2% higher that of

437

Fenton (F). Methanothrix species accounted for about 51.7% of the communities for

438

control (F) and 57.5% of that for Fenton (F), because of the Fenton sludge reactor

439

maintain a much higher ratio of acetate than the control reactor. Besides the ability of

440

generating methane from acetate cleavage, Methanothrix species in the aggregates

441

have a complete complement of genes for the enzymes necessary for the reduction of

442

carbon to methane. Rotaru42 suggested that Geobacter with highly expressed genes

443

for extracellular electron transfer via electrically conductive pili and Methanothrix

444

species could exchange electrons via direct interspecies electron transfer (DIET).

445

These imply the potential of Methanothrix to directly accept electrons and participate

446

in the electric syntrophy.

447



Page 22 of 45

448

The bacterial communities provided a further insight into the performances of

449

hydrolysis and acidogenesis in the presence of Fenton sludge. On middle stage, the

450

dominant bacterial detected of Fenton sludge reactor were Clostridium species,

451

accounted for about 7.3% of the communities, that were well-known as the

452

Fe(III)-reducing genus44, 45, while accounted for less than 2.0% of control reactor and

453

initial seed sludge. Clostridium species had the type IV pili for extracellular electron

454

transfer to the insoluble Fe(III) oxides with the reduction of Fe(III) to Fe(II). Together

455

with the higher concentration of Fe2+ detected in Fenton sludge reactors (Fig. S3), it

456

was suggested that Fenton sludge supplemented could enrich the Fe(III)-reducing

457

microorganisms via the dissimilatory iron reduction. As a result, the performances of

458

Fenton sludge reactors especially significantly improved in the hydrolytic

459

fermentation. It suggested the Fenton sludge supplemented to the anaerobic digestion

460

of WAS might facilitate the decomposition of complex organics, such as the

461

carbohydrates and proteins which were main compositions of WAS (Fig.5). The

462

reason should be ascribed to the enriched Fe(III)-reducing microorganisms that

463

participated in the conversion of complex organics to simples. The reduction potential

464

of α-Fe2O3/Fe2+ was -0.287 V (pH =7) which was significantly lower than that of the

465

chelated

466

Fe(III)-citrate/Fe(II)-citrate (+0.385 V)46. Chelated Fe(III) is on the favorable end of

467

the spectrum; however, neither Geobacter nor Shewanella extracts the maximum

468

energy available from chelated Fe(III), as evidenced by their poor growth yields. It

469

therefore seems that IRB has adapted to use low-potential substrates [e.g. Fe(III)

Fe,

such

as

Fe(III)-NTA/Fe(II)-NTA


(+0.372

V)

and

Page 23 of 45 1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 25 26 27 28 29 30 31 32 33 34 35 36 37 38 39 40 41 42 43 44 45 46 47 48 49 50 51 52 53 54 55 56 57 58 59 60


470

minerals] rather than maintaining an electron transport chain that would allow it to

471

extract greater energy from higher potential acceptors [e.g. Fe(III)-citrate]46.

472

Therefore, the insoluble Fe(III)/Fe(III)-Fe(II) oxides with a lower or more negative

473

reduction potential, such as hematite (the main form in Fenton sludge after dried in

474

this study), could support the growth of the Fe(III)-reducing genus and enrich them

475

better. Syntrophomonas belong to Syntrophomonadaceae which are the family related

476

to the syntrophic VFA-oxidizing bacteria not only took part in long-chain fatty acids

477

degradation47-50, but also could form a syntrophic metabolism with methanogens to

478

promote methane production. They accounted for about 2.4% the communities in the

479

seed sludge and increased to 5.0% on the middle stage and 20.0% on final stage in

480

Fenton sludge reactor, while there were 2.2% and 17.3% of that without Fenton

481

sludge supplemented. Their proliferation during the anaerobic digestion was

482

consistent with the change of VFAs in reactor (Fig. 6). For the initial few days, the

483

Fenton sludge reactor had a higher acidification efficiency to produce sufficient VFAs

484

to enrich Syntrophomonas and they even turned into the dominant genus in Fenton

485

sludge reactor on the final stage of the anaerobic digestion. As a result, the Fenton

486

sludge reactors had a more rapid metabolism of VFAs and there were almost no

487

propionate and other VFAs residues. While there were an obvious accumulation of

488

propionate in the control reactor (Fig.S2). Further studies51 detected Syntrophomonas

489

in the biofilms or anodic biofilms, and indicated that they were able to contribute to

490

butyrate degradation and electricity generation. That implied Syntrophomonas species

491

were potential to participate into DIET for sludge decomposition and methane



Page 24 of 45

492

production in the Fenton sludge supplemented reactors. While genomic analyses have

493

so far not been able to identify the mechanism of extracellular electron transfer in

494

Syntrophomonas.

495 496

To the best of our knowledge, this is the first time to introduce Fenton sludge as an

497

iron source to enhance the anaerobic digestion and compared the changes of microbial

498

community structures in different stages during the digestion. For the hydrolytic

499

acidification stage, the addition of Fenton sludge was more likely to play a role as

500

insoluble Fe(III) oxides which made it possible for dissimilatory iron reduction take

501

place. As a result, IRB such as Clostridium species were enriched following the

502

decompostion of complex organic matters and then more small-molecule substrates

503

were

504

methanogenesis stage, Fe(II) was generated from the Fenton sludge via the

505

dissimilatory iron reduction, the dissolved Fe(II) might decrease ORP which made it

506

beneficial for acetate production and reducing the propionate accumulation. Moreover,

507

Syntrophomonas became the dominant genus in the Fenton sludge reactor on the end

508

of the anaerobic digestion that accelerated the metabolism of VFAs, which could

509

eliminate feedback inhibition of fermentation and maintained the anaerobic digestion

510

proceeding steadily. The microbial community analysis along with other results

511

parameters suggested the addition of Fenton sludge had significantly improved

512

methane production and organic removal of WAS.

produced

available

for

methanogenesis.

On

513


the

acetogenesis

and

Page 25 of 45 1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 25 26 27 28 29 30 31 32 33 34 35 36 37 38 39 40 41 42 43 44 45 46 47 48 49 50 51 52 53 54 55 56 57 58 59 60


514

Implications.

515

In most cases, Fenton sludge was regarded as a hazardous waste, which cannot be

516

efficiently disposed by traditional methods such as stabilization, solidification, landfill

517

process or combustion1. Using Fenton sludge as an iron resource to enhance anaerobic

518

digestion was an alternative method to reuse this iron-containing sludge, in which

519

most organic matters involved in the Fenton sludge were also removed to decrease its

520

environmental risks. Moreover, Fe(II) generated from the dissimilatory iron reduction

521

accounted for 44.1% (508 mg/L) of total Fe content of 0.5 g Fenton sludge after

522

anaerobic digestion. When adjusting the pH of the digested sludge to 5.33, almost a

523

half of Fe(II) was dissolved from the sludge (Fig. S4). It meant that the Fe2+ produced

524

from the dissimilatory iron reduction could be reused as Fenton reagent without

525

chemical reductant or electrochemical assistance. Thus a recycling process can be

526

established based on the present study as following steps: (Fig. 9) (1) a given ratio of

527

Fenton sludge is added into WAS for anaerobic digestion. The anaerobic digestion is

528

enhanced with the principle of dissimilatory iron reduction. Meanwhile, organics

529

involved in the Fenton sludge are decomposed during the anaerobic digestion. Also,

530

ferric iron in the Fenton sludge is reduced to ferrous iron due to dissimilatory iron

531

reduction (2) pH of fermentation liquid is adjusted to a weak acidic pH to release

532

ferrous iron from the sludge phase into the aqueous phase, which can be used as

533

Fenton reagent. (3) Wastewater such as landfill leachate is mixed with the Fe(II)

534

obtained in step (2) and hydrogen peroxide at a certain proportion to proceed Fenton

535

reaction. (4) After Fenton process, the pH is adjusted into neutral, and the Fenton



536

sludge settlement is recycled into the anaerobic digester.

537

Conclusion

538

Fenton sludge with high iron content could be a potential resource for enhancing

539

anaerobic digestion of WAS. The methane production increased by 20.95% and the

540

sludge reduction ratio increased by 9.4% with Fenton sludge supplemented after 24

541

days operation. The addition of Fenton sludge accelerated the hydrolysis of

542

polysaccharide and protein to produce VFAs available for methanogenesis. The

543

potential mechanism involved was considered that Fe(III) oxides in Fenton sludge

544

could enrich the IRB such as Clostridium that are capable of utilizing a variety of

545

substrates and participating in the decomposition of complex organic matters via the

546

dissimilatory iron reduction. Also, the conversion of propionate to acetate was

547

enhanced with Fenton sludge supplemented. Moreover, organic matters in Fenton

548

sludge could be removed by 70% during anaerobic digestion. Using Fenton sludge as

549

an iron source in anaerobic digestion of WAS can both cut down the cost of disposal

550

of Fenton sludge and obtain a more efficient anaerobic digestion, which can be a

551

sustainable strategy for enhancing anaerobic digestion and a new way to solve the

552

environmental problem of Fenton sludge.

553


Page 26 of 45

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554

Associated Content

555

Supporting Information Available.

556

Details about method for DNA extraction, PCR amplification and high-throughput

557

16S rRNA pyrosequencing, scanning electron microscope image and energy

558

dispersive spectrometer spectrum of powder Fenton sludge (Figure S1), propionate

559

and acetate ratio of volatile fatty acids (Figure S2) and experimental data about

560

dissolution of iron are available(Figure S3 and Figure S4) in the Supporting

561

Information.

562 563

Conflict of interest statement

564

The authors declare that the research was conducted in the absence of any commercial

565

or financial relationships that could be construed as a potential conflict of interest.

566 567

Acknowledgments

568

The authors acknowledge the financial support from the National Natural Scientific

569

Foundation of China (21777016).

570 571



572 573

References:

574

1. Bautista, P.; Mohedano, A. F.; Casas, J. A.; Zazo, J. A.; Rodriguez, J. J. An

575

overview of the application of Fenton oxidation to industrial wastewaters treatment. J.

576

Chem. Technol. Biotechnol. 2008, 83, (10), 1323-1338.

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2. Badawy, M. I.; Ali, M. E. M. Fenton's peroxidation and coagulation processes for

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3. Ma, X.; Xia, H. Coagulation combined with fenton process for the treatment of

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hydrogen production from glucose at low temperature in microbial electrolysis cells.

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Bioresource Technol. 2012, 124, 68-76.

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37. Schmidt, J. E.; Ahring, B. K. Effects of hydrogen and formate on the degradation

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of propionate and butyrate in thermophilic granules from an upflow anaerobic sludge

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blanket reactor. Appl. Environ. Microbiol. 1993, 59, (8), 2546-2551.


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38. Zhang, L.; Keller, J.; Yuan, Z. Inhibition of sulfate-reducing and methanogenic

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activities of anaerobic sewer biofilms by ferric iron dosing. Water Res. 2009, 43, (17),

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4123-4132.

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39. Lueders, T.; Friedrich, M. W. Effects of amendment with ferrihydrite and gypsum

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on the structure and activity of methanogenic populations in rice field soil. Appl.

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Environ. Microbiol. 2002, 68, (5), 2484-2494.

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40. Bodegom, P. M.; Scholten, J. C. M.; Stams, A. J. M. Direct inhibition of

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methanogenesis by ferric iron. FEMS Microbiol. Ecol. 2004, 49, (2), 261-268.

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41. Peng, S.; Xue, J.; Shi, C.; Wang, J.; Chen, T.; Yue, Z. Iron-enhanced anaerobic

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digestion of cyanobacterial biomass from Lake Chao. Fuel. 2014, 117, (A), 1-4.

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42. Rotaru, A.; Shrestha, P. M.; Liu, F.; Shrestha, M.; Shrestha, D.; Embree, M.;

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Zengler, K.; Wardman, C.; Nevin, K. P.; Lovley, D. R. A new model for electron flow

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during anaerobic digestion: Direct interspecies electron transfer to Methanosaeta for

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the reduction of carbon dioxide to methane. Energy Environ. Sci. 2014, 7, (1),

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43. de Bok, F.; Luijten, M.; Stams, A. Biochemical evidence for formate transfer in

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syntrophic propionate-oxidizing cocultures of Syntrophobacter fumaroxidans and

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Methanospirillum hungatei. Appl. Environ. Microbiol. 2002, 68, (9), 4247-4252.

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44. Park, H. S.; Kim, B. H.; Kim, H. S.; Kim, H. J.; Kim, G. T.; Kim, M.; Chang, I.

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S.; Park, Y. K.; Chang, H. I. A novel electrochemically active and Fe(III)-reducing

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bacterium phylogenetically related to Clostridium butyricum isolated from a microbial

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fuel cell. Anaerobe. 2001, 7, (6), 297-306.



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45. Dobbin, P. S.; Carter, J. P.; San Juan, C.; von Hobe, M.; Powell, A. K.;

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Richardson, D. J. Dissimilatory Fe(III) reduction by Clostridium beijerinckii isolated

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from freshwater sediment using Fe(III) maltol enrichment. FEMS Microbiol. Lett.

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1999, 176, (1), 131-138.

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46. Bird, L. J.; Bonnefoy, V.; Newman, D. K. Bioenergetic challenges of microbial

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iron metabolisms. Trends Microbiol. 2011, 19, (7SI), 330-340.

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47. Zhang, C. Syntrophomonas curvata sp. Nov., An anaerobe that degrades fatty

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acids in co-culture with methanogens. Int. J. Syst. Evol. Microbiol. 2004, 54, (3),

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969-973.

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48. Wu, C.; Liu, X.; Dong, X. Syntrophomonas erecta subsp. Sporosyntropha subsp.

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Nov., A spore-forming bacterium that degrades short chain fatty acids in co-culture

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with methanogens. Syst. Appl. Microbiol. 2006, 29, (6), 457-462.

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49. Hatamoto, M.; Imachi, H.; Fukayo, S.; Ohashi, A.; Harada, H. Syntrophomonas

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palmitatica sp. Nov., An anaerobic, syntrophic, long-chain fatty-acid-oxidizing

719

bacterium isolated from methanogenic sludge. Int. J. Syst. Evol. Microbiol. 2007, 57,

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(9), 2137-2142.

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50. McInerney, M. J.; Bryant, M. P.; Hespell, R. B.; Costerton, J. W.

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Syntrophomonas wolfei gen. Nov. Sp. Nov., An Anaerobic, Syntrophic, Fatty

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51. Zhao, Z.; Zhang, Y.; Quan, X.; Zhao, H. Evaluation on direct interspecies

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electron transfer in anaerobic sludge digestion of microbial electrolysis cell.

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Bioresource Technol. 2016, 200, 235-244.


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727

Tables & Graphics

728

Table 1. The main characteristics of the mixture. a Parameters

Initial value

TS b

63.0±0.10 g/L

VS c

48.3±0.11 g/L

pH

7.44

TCOD d

64.8±2.05 g/L

SCOD e

5.0±0.4 g/L

Total Polysaccharide

759.4±63.63 mg/L

Soluble Polysaccharide

197.9±20.11 mg/L

Total protein

822.0±73.00 mg/L

Soluble protein

422.7±18.19 mg/L

729

a

Average data and standard deviation obtained from three tests.

730

b

TS: total solids.

731

c

VS: volatile solids.

732

d

TCOD: total chemical oxygen demand.

733

e

SCOD: soluble chemical oxygen demand



734

735

736 737

Figure 1. (a) XRD patterns of the powder Fenton sludge. The symbols correspond to

738

the phases: H-hematite; U-unknown. (b) XPS spectra of Fenton sludge powder of

739

wide scan and (c) Fe 2p spectra of Fenton sludge powder


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740

741 742

Figure 2. TCOD change of five reactors and COD removal rate of Fenton sludge



743 744

Figure 3. Accumulative biogas production during anaerobic digestion in different

745

groups (a) accumulative methane production, (b) accumulative carbon dioxide

746

production


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747

748 749

Figure 4. Accumulative methane production during 24 days anaerobic digestion



750

751 752

Figure 5. Organic matter removal during the 24 days anaerobic digestion. (a) TS, VS

753

and TCOD of raw sludge and final stage of different reactors. (b) Polysaccharide and

754

protein of raw sludge and final stage of different reactors


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755

756

757 758

Figure 6. The change of VFA of reactors with (a) 0 g Fenton sludge, (b) 0.5 g Fenton

759

sludge, (c) 1.0 g Fenton sludge, (d) 2.0 g Fenton sludge



760

761 762

Figure 7. The change of pH of four reactors during the 24 days digestion


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763

764

765 766

Figure 8. (a) Archaeal and (b) bacterial communities of seed sludge and digestion

767

sludge.



768

769 770

Figure 9. Recycle of Fenton sludge in anaerobic digestion based on dissimilatory iron

771

reduction


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Synopsis: A sustainable strategy for disposing Fenton sludge environmentally

773

friendly as well as to enhance anaerobic digestion of WAS

774

TOC graphic

775 776 777 778 779 780

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Sustainable Strategy for Enhancing Anaerobic Digestion of Waste

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