Different ferric dosing strategies could result in different control mechanisms of sulfide and methane production in sediments of gravity sewers

Accepted Manuscript Different ferric dosing strategies could result in different control mechanisms of sulfide and methane production in sediments of gravity sewers Juanjuan Cao, Liang Zhang, Jiaying Hong, Jianliang Sun, Feng Jiang PII:

S0043-1354(19)30688-8

DOI:

https://doi.org/10.1016/j.watres.2019.114914

Article Number: 114914 Reference:

WR 114914

To appear in:

Water Research

Received Date: 13 May 2019 Revised Date:

15 July 2019

Accepted Date: 23 July 2019

Please cite this article as: Cao, J., Zhang, L., Hong, J., Sun, J., Jiang, F., Different ferric dosing strategies could result in different control mechanisms of sulfide and methane production in sediments of gravity sewers, Water Research (2019), doi: https://doi.org/10.1016/j.watres.2019.114914. This is a PDF file of an unedited manuscript that has been accepted for publication. As a service to our customers we are providing this early version of the manuscript. The manuscript will undergo copyediting, typesetting, and review of the resulting proof before it is published in its final form. 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.

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ACCEPTED MANUSCRIPT 1

Different ferric dosing strategies could result in different control

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mechanisms of sulfide and methane production in sediments of gravity

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sewers

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Juanjuan Caoa,1, Liang Zhangb,1, Jiaying Honga, Jianliang Suna, Feng Jianga,c,*

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a

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Treatment, School of Chemistry & Environment, South China Normal University, Guangzhou, China

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b

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Nanyang Technological University, 50 Nanyang Avenue, Singapore, 639798, Singapore

Guangdong Provincial Engineering Technology Research Center for Wastewater Management and

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Advanced Environmental Biotechnology Centre, Nanyang Environment & Water Research Institute,

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School of Environmental Science and Engineering, Sun Yat-sen University, Guangzhou, China

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*Corresponding author:

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Dr. Feng Jiang

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Email: [email protected] / [email protected]

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1: Juanjuan Cao and Liang Zhang contributed equally to this work.

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Abstract

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Ferric salt dosing is widely used to mitigate sulfide and methane emissions from sewers. In

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gravity sewers with sediments, responses of sulfate-reducing bacteria (SRB) and

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methanogenic archaea (MA) residing in different zones to Fe3+ dosing strategies still

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remain unknown. In this study, we investigated the changes in behavior of SRB and MA in

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different depths of sewer sediment using laboratory-scale sewer sediment reactors with

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different Fe3+ dosing strategies (different instant dosages and frequencies). All Fe3+ dosing

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strategies examined efficiently suppressed sulfide concentration for a short time, but the

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control mechanisms were different. When a low-dosage, high-frequency Fe3+ dosing

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strategy was employed, Fe3+ could not penetrate into the sewer sediment, therefore, the

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abundances of SRB and MA in all zones of sewer sediment did not change substantially. As

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a result, the active sulfide-producing and methane-producing zones kept unchanged.

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Sulfide was controlled mainly via chemical sulfide oxidation and precipitation, and

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methane formation was not influenced. In contrast, when a high-dosage, low-frequency

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Fe3+ dosing strategy was used, the SRB activity in the upper layer of the sewer sediment

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was nearly fully suppressed according to the down moving zones of sulfide production

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(from 0-5 mm to 20-25 mm) and lower sulfate reduction, in which sulfate reduction

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decreased by 56% in the long-term trial. The generated sulfide was further removed via

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chemical sulfide oxidation and precipitation. This strategy also significantly suppressed MA

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activity (21% reduction in methane production). However, considering long-term

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satisfactory control of sulfide and methane production, low operational cost and less

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ACCEPTED MANUSCRIPT sediments deposited in gravity sewers, a low-dosage, high-frequency Fe3+ dosing strategy

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would be a more cost-effective solution for sulfide control in gravity sewers with thin

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(<20mm) or thick (>20 mm) sediments if methane mitigation does not need to be taken

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into account.

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Keywords: Sulfate-reducing bacteria (SRB), Methanogenic archaea (MA), Sulfide and

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methane control, Microelectrode analysis, Stratified microbial structure

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1. Introduction

Sewer networks are used for transporting wastewater from urban settlements to

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wastewater treatment plants (WWTPs). Under anaerobic conditions, sufficient organic

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compounds and sulfates in the wastewater favor the growth of sulfate-reducing bacteria

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(SRB) and methanogenic archaea (MA) in sewer systems. SRB can reduce sulfate to sulfide,

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which is responsible for noxious sewer odors and corrosion (Pikaar et al., 2014). MA can

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generate methane (CH4), which is also an issue, as CH4 has a low explosive limit and is a

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potent greenhouse gas with 21 times higher global warming potential than carbon dioxide

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(Jiang et al., 2013b, Pikaar et al., 2014). In large cities, this problem is worse because of a

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longer hydraulic retention time (HRT) in the expanded sewer networks.

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To minimize the emission of sulfide and methane, iron salts (including FeSO4, FeCl2,

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FeCl3, Fe(NO3)3, and FeSO4) have been extensively used in sewers (Zhang et al., 2009,

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Ganigue et al., 2011). Ferric ions (Fe3+) chemically oxidize sulfide to elemental sulfur as the 3

ACCEPTED MANUSCRIPT ferric ions reduce to Fe2+, which can subsequently precipitate sulfide to generate FeS

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(equations 1 and 2) (Nielsen et al., 2005, Firer et al., 2008). However, most previous

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studies on the effect of ferric salt dosing on sulfide and methane formation were

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performed in sewer biofilms (Zhang et al., 2009, 2010, Ganigué et al., 2018), in which the

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addition of ferric chloride was found to significantly reduce sulfide and methane

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production in the anaerobic sewer biofilm of rising mains by more than 50%.

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2 F e 3 + + HS − → 2 Fe 2 + + S 0 ↓ + H +

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F e 2 + + HS − → FeS ↓ + H +

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In gravity sewers, sediment is considered to be biologically active (Schmitt and

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Seyfried, 1992). The sulfide-producing capacity of sewer sediment is significantly greater

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than that of sewer biofilms (generally present in rising mains) (Schmitt and Seyfried, 1992,

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Liu et al., 2015b), indicating that the sulfide and methane production from sewer

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sediment need to be accounted for when controlling their emission. However, the effects

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of Fe3+ addition on controlling sulfide and methane production and the sulfidogenic and

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methanogenic activities in sewer sediments have been poorly documented.

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The higher thickness of sewer sediment than sewer biofilm (centimeters vs.

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hundreds of micrometers) limits substrate diffusion into the deep layers of sediment. Due

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to the diffusion limits of sulfate, dissolved oxygen, sulfide, soluble chemical oxygen

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demand (sCOD), and other substrates (e.g. Fe3+), sewer sediments are reported to have a

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stratified microbial structure and activity of sulfidogenic and methanogenic communities

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(Liu et al., 2015c). When dosing a gravity sewer with iron salt, the ferric and ferrous ions

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may not fully penetrate into sewer sediment, which may result in different effects on the

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ACCEPTED MANUSCRIPT inhibition of sulfidogenic and methanogenic activities compared to that in a sewer biofilm.

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This suggests that a higher dosage of ferric salt is necessary to effectively suppress

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sulfidogenic and methanogenic activities in sewer sediments. However, to minimize the

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operational cost, dosing frequency should be reduced with increased iron dosage.

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Furthermore, when a high dose of ferric salts was used, the excess of Fe3+ under neutral

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pH conditions may quickly convert into amorphous ferric oxyhydroxide, Fe(OH)3 (Davydov

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et al., 1998) and settle on the sediment surface. The formed Fe(OH)3 layer may hinder the

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diffusion of Fe3+ into sewer sediments to influence sulfidogenic and methanogenic

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bacteria. Even if the Fe3+ could penetrate into the sediment, the sulfidogenic and

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methanogenic bacteria residing in different zones may exhibit different responses due to

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the uneven Fe3+ concentrations in the different zones of a sewer sediment. To our

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knowledge, no previous study has focused on the effects of different ferric dosing

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strategies on the spatial dynamics of sulfidogenic and methanogenic activities in sewers

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with sediments, which highlights the need for a direct comparison in a single experiment.

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Therefore, filling the aforementioned knowledge gaps is an important

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prerequisite in the ongoing quest to improve the effectiveness of sulfide and methane

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control in the sediments of sewers receiving ferric salt dosing. To this end, three parallel

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sewer sediment reactors were constructed in this study to mimic gravity sewer conditions

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and were continuously operated for 212 days. Different ferric chloride dosing strategies

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(details shown in Section 2.1) were employed during the long-term operation of the sewer

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sediment reactors. We aimed to: (a) investigate the effectiveness of ferric salt dosing for

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controlling sulfide and methane production in gravity sewers with sediments under

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ACCEPTED MANUSCRIPT different dosing strategies; (b) profile and analyze the inhibitory effect of ferric salt

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addition on sulfidogenic and methanogenic bacteria in different zones of the sewer

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sediment; (c) characterize the control mechanisms of sulfide and methane production

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under different Fe3+ dosing strategies.

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2. Material and methods

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2.1. Lab-scale sewer reactor setup and operation

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The three identical laboratory-scale reactors were constructed to mimic the

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conditions in gravity sewers (Fig. 1). Each reactor had a total volume of 4.6 L and a total

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height of 30 cm. The inner diameter and outlet height of each reactor was 14 cm and 27.9

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cm, respectively, resulting in a working volume of 4.3 L. In order to conveniently

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performing the following microelectrode analysis and collecting sludge samples at

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different depths of sewer sediments, an additional cylinder (12 cm diameter and 6 cm

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height) was placed at the bottom of each reactor to serve as the sediment carrier. The

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seeding sludge was collected from a sewage sewer located in Guangzhou, China, and 600

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mL of the seeding sludge was added to each reactor to obtain a sediment depth of 40 mm.

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A mechanical overhead stirrer with a speed of 40 rpm/min was positioned at 16 cm to

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achieve aeration in the reactor. As such, hydraulic shear stress conditions mimicking those

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in real gravity sewers can be obtained (Liu et al., 2015c).

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To avoid fluctuations in sulfide production owing to varied organic and sulfate

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concentrations in the influent wastewater, synthetic domestic wastewater was used in this

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study and was prepared every 12 h following previously described methods (Jiang et al., 6

ACCEPTED MANUSCRIPT 2013a, Zhang et al., 2018a). The synthetic wastewater consisted of 244 mg/L glucose, 122

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mg/L yeast extract, 326 mg/L sodium acetate, 177.5 mg/L Na2SO4, 9 mg/L KH2PO4, 24 mg/

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L K2HPO4, and 2.5 mL of a concentrated trace element solution (Table S1) to obtain 534 ±

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53 mg/L COD and 40 ± 5 mg S/L SO42-. The synthetic wastewater was stored at 4 °C before

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being fed into the three reactors. The influent flow rate of each reactor was kept at 11.9

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ml/min with an HRT of 6h. The reactors were covered with aluminum foil to avoid light

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exposure and prevent algae growth and kept at 25 °C using a water bath.

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The three sewer sediment reactors were continuously operated for 212 days,

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during which they were run more than 60 days before ferric chloride addition in order to

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cultivate mature intact sewer sediment and to achieve pseudo steady-state conditions. To

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prevent ferric hydroxide precipitation, a ferric ion stock solution was prepared by

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dissolving reagent-grade ferric chloride (FeCl3·6H2O) in deoxygenated water containing

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0.01 M hydrochloric acid (HCl). During the operational period, three different dosing

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strategies with different instant dosages and frequencies were applied in reactor 1 (R1) or

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reactor 3 (R3) to study the effects of different Fe3+ dosing strategies on sulfide and

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methane production and the functional microbial communities, such as low Fe3+ dosage

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with high frequency (Strategy 1), medium Fe3+ dosage with medium frequency (Strategy 2),

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and high Fe3+ dosage with low frequency (Strategy 3, see Table 1). Strategy 2 (medium

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dosage with medium frequency) was used as a transitional stage from Strategy 1 to

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Strategy 3 in the R3 operation. Reactor 2 (R2) without Fe3+ addition throughout the entire

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experiment was used a control. In R1, only Strategy 1 was employed, and this strategy was

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used during days 84-141 and days 176-207. In R3, strategies 1, 2, and 3 were applied

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ACCEPTED MANUSCRIPT during days 64-76 and days 85-94, days 113-133 and days 176-207, respectively (Table 1).

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In addition, each Fe3+ dosing event was completed within a very short time (3 min), the

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effect of change in HRT on sulfide production was not considered in this study.

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2.2. Microelectrode analysis

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At the end of the experiment, strategies 1 and 3 were used again in the R1 and R3,

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respectively. After one day, the steady state profiles of hydrogen sulfide (molecular H2S),

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pH, and oxidation reduction potential (ORP) in the sewer sediment were measured using

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microelectrode sensors (Unisense A/S, Denmark) after calibration following the

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manufacturer’s instructions. All the diameters of H2S, pH, and ORP tips were 25 μm.

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According to the hydrogen sulfide and pH profiles, the total dissolved sulfide concentration

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can be calculated based on the dissociation equation described by Kühl and Jørgensen

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(1992). ORP profiles can be utilized to determine the redox conditions in the sewer

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sediment. Microelectrodes were mounted on a manual manipulator and positioned on the

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surface of the sediment. The H2S, pH, and ORP gradients through the sediment (0-30 mm

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depth) were obtained by moving the microelectrodes in increments of 0.5 mm. The

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sulfide-producing activity within the sediment was calculated using the diffusion reaction

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model developed by Sun et al. (2014).

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2.3. Sludge sampling and analysis

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After the microelectrode analysis (Section 2.2), the cylinders holding sediments were

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taken out from the sewer reactors. The sludge samples from each cylinder were

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ACCEPTED MANUSCRIPT transferred to respective self-made layered samplers (shown in Fig. S1) by vertically

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pressing the samplers from the surface to go through the whole depth of the sediments,

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and then were frozen at -20 °C for one day. The frozen sludge samples could be easily

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separated from the samplers when the samplers with frozen sludge samples were placed

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at room temperature (~25 °C) for approximately one minute. Subsequently, the frozen

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sludge samples were cryosectioned into three sections to represent the upper (0-13.3

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mm), middle (13.4-26.6 mm), and bottom layer (26.7-40 mm) in each sewer reactor. These

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stratified sludge samples were used for microbial community analysis.

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2.4 Chemical analysis

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The total organic carbon (TOC), sulfate, thiosulfate, methane, and iron (Fe2+ and

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Fe3+) concentrations were measured after filtration (0.22 μm pore size). TOC was analyzed

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using a TOC analyzer (Shimadzu TOC-5000A). COD was calculated by a theoretical ratio of

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2.67 g COD/g TOC to avoid the effect of sulfide on COD analysis (Liang et al., 2016). Total

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dissolved sulfide (H2S, HS- and S2-) was determined using the Methylene Blue Method

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(APHA, 2005). Sulfate and thiosulfate were analyzed using an ion chromatograph (ICS-900).

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Concentrations of iron (Fe2+ and Fe3+) were determined via a 1,10-phenanthroline

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spectrophotometry method at 512 nm (APHA, 2005). pH was measured using a pH meter

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(HQ40D, HACH). Elemental sulfur in sludge samples was extracted according to the

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method developed by McGuire and Hamers (2000) and quantified with a

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high-performance liquid chromatograph (HPLC, Shimadzu LC-16, Japan) equipped with a

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Kromasil column (C18, 5μ, 100 Å) and a UV detector operating at 254 nm.

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ACCEPTED MANUSCRIPT Water samples were collected from the reactors for CH4 analysis using a 3-ml plastic

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syringe and transferred into pre-evacuated Exetainer tubes (Labco, Wycombe, UK) using

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syringe needles. The tubes were mixed overnight in a shaker to bring the gas and liquid

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phases into equilibrium. Most of the methane (w 97% at 25 °C) was transferred to the gas

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phase during this process. Methane concentrations in the gas phase of the tubes were

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measured by a Gas Chromatograph (Agilent GC-2014) equipped with a thermal

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conductivity detector (TCD) (Alberto et al., 2000). Then, the methane concentration in the

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water samples can be calculated based on mass balance and Henry's law (Guisasola et al.,

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2008).

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2.4. DNA extraction, PCR amplification, Illumina Miseq sequencing, and bioinformatic

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analysis

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After the microelectrode analysis, sludge samples were collected from the upper,

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middle, and bottom layers of each reactor to analyze the microbial community diversities

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by Illumina Miseq sequencing. The total genomic DNA was then extracted from the

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condensed sludge with a FastDNA Soil Kit (MP Biomedicals, CA, USA) following the

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manufacturer's instructions. The DNA extracted from the aforementioned samples was

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subjected to Illumina Miseq sequencing for bacterial and archaeal community analysis.

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The

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(5’-GTGCCAGCMGCCGCGGTAA-3’)

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(5’-GGACTACHVGGGTWTCTAAT-3’) to target the hypervariable V4 region of the bacteria

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domain (Kozich et al., 2013). The 16S rRNA was amplified with the forward primer 1106F

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1378R

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(5’-TGTGCAAGGAGCAGGGAC-3’) to target the hypervariable V9 region of the archaea

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domain (Watanabe et al., 2007). The Illumina Miseq sequencing service was provided by

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BIG (Shenzhen, China). The obtained paired-end raw 16S rRNA gene sequences were

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assembled and filtered in Mothur (v1.31.2) (Kozich et al., 2013). The aligned sequences

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were checked for chimeras using USEARCH (v7.0.1090) in QIIME (v4.2.40) and classified

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into operational taxonomic units (OTUs) with a 97% similarity cutoff (Caporaso et al.,

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2010). Taxonomy was assigned by Ribosomal Database Project (RDP) Classifier (v 2.2) with

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a confidence threshold of 0.6. The alpha diversity indices (Observed species, Chao1, ACE

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index, Shannon estimator and Simpson estimator) and the Unweighted UniFrac distances

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for Principal coordinates analysis (PCoA) were calculated following the QIIME pipeline

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(Caporaso et al., 2010). A PCoA based on the unweighted UniFrac distance matrix was

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performed to determine the beta diversity of the bacterial community (Lozupone et al.,

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2011). Heatmap analysis was conducted using the pheatmap package in R (v3.1.1,

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http://www.r-project.org/).

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3. Results

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3.1. Sulfide, sulfate, and methane profiles in the sewer sediment reactors

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The three reactors (R1, R2, and R3) were continuously operated for at least 60 days

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before Fe3+ addition, during which all reactors achieved stable and similar performance in

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terms of sulfide production (Fig. 2). When low Fe3+ dosage with high frequency (Strategy 1)

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was employed in R1, the sulfide concentration was reduced by 94% on average compared 11

ACCEPTED MANUSCRIPT to that in R2, decreasing from 22.0 ± 4.7 mg S/L to 1.4 ± 0.8 mg S/L (Fig. 2a). Interestingly,

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the concentrations of the effluent sulfate and COD in R1 were similar to those in R2 (Fig.

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3a and S2a), indicating that sulfate reduction was not influenced during the dosing period

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in R1. Meanwhile, the sulfide production rebounded to the original level after only one

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day since the low-dosage, high-frequency Fe3+ dosing strategy (Strategy 1) was halted. The

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present results indicate that low-dosage, high-frequency Fe3+ dosing had negligible

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inhibitory effects on sulfide-producing activity.

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In R3, three different Fe3+ dosing strategies were tested (Table 1 and Fig. 2b). Fe3+

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dosing with Strategy 1 resulted in a similar trend to that in R1, in which the sulfide

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concentration was significantly reduced to 1.3 ± 1.0 mg S/L, but quickly rebounded to

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normal levels after stopping the Fe3+ dosing. When higher Fe3+ dosages with lower dosing

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frequencies (Strategies 2 and 3) were employed (Table 1), however, satisfactory sulfide

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control was not achieved; these strategies only achieved effective sulfide control for 1-2

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days and 3-4 days, respectively, for each dosing event (Fig. 2b). Notably, compared to

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Strategy 1, the effluent concentrations of sulfate, COD, and iron (Fe2+ and Fe3+) significantly

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increased during the period when Strategies 2 and 3 were employed, among which sulfate

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reduction decreased by approximately 56% (Figs. 3a, S2b and S3b), indicating that the

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sulfate-reducing activity in R3 was inhibited by high Fe3+ dosage. In addition, the effects of

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different Fe3+ dosing strategies on CH4 production in the three reactors were also

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investigated. We observed that low and medium dosage of Fe3+ with high and medium

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frequencies (Strategy 1 and 2) did not influence the CH4 production, producing 0.42 ± 0.05

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mg/L CH4. As for the high-dosage, low-frequency Fe3+ dosing strategy (Strategy 3), this

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ACCEPTED MANUSCRIPT strategy slightly reduced CH4 reduction by 21% on average from 0.42 ± 0.05 mg/L to 0.33 ±

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0.06 mg/L (Fig. 3b).

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3.2. Microsensor analysis of Sulfide and ORP depth profiles

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Total dissolved sulfide concentration, pH, and ORP values were measured in the

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sewer sediment at the end of the experiment. Due to limitations of the manual

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manipulator (maximum depth of 30 mm), we could not measure the sulfide concentration

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and ORP value within the bottom 10 mm of the sewer reactors. However, we can expect

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that the varying trends of total dissolved sulfide concentration, pH, and ORP values in the

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depth of 25-30 mm (Fig. 4) could be used to predict the trends in the bottom 10 mm. In R2

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without Fe3+ addition, sulfide was continuously generated throughout the sewer sediment,

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producing 19.6 ± 2.1 mg S/L sulfide on average (Fig. 4a). In R1 with low Fe3+ dosage, the

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sulfide-producing zone (0-5 mm) was not subjected to substantial changes compared to

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the corresponding zones in the R2 sediment, but sulfide accumulation was less than 0.5 ±

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0.1 mg S/L throughout the sediment. Note that the zone in the sewer sediments where

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sulfide sharply increased was determined as sulfide-producing zone in this study (Liu et al.,

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2015b, Liu et al., 2015c). On the contrary, in R3 with high Fe3+ dosage, sulfide was not

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observed in the top 20 mm of the sediment, but it rebounded to a maximum of 17.6 mg

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S/L in the deeper layer of the sewer sediment (> 20 mm). The present result indicates that

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high Fe3+ dosage changed the sulfide-producing zone of the sewer sediment in the surface

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layer but not in the deep layer of sewer sediment.

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Regarding the ORP profile (Fig. 4b), although the low-dosage, high-frequency Fe3+ 13

ACCEPTED MANUSCRIPT dosing strategy was applied in R1, the ORP levels in R1 sediment were similar to those in

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the R2 sediment. When high Fe3+ dosage was applied into R3 with low frequency (Strategy

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3), the ORP value in the top 8 mm of the sediment increased to approximately 700 mv,

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followed by a gradual decrease to -206 mv at 30 mm with the increased sediment depth,

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but was always higher than the values in R1 and R2. This may be due to the limited

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diffusion capacity of Fe3+ in R3, as shown in Fig. 5. The results suggest that the added Fe3+

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diffused into the R3 sediment to increase ORP levels and suppressed sulfate reduction

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when a high Fe3+ dosage was applied. When a low Fe3+ dosage was applied in R1, Fe3+

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barely diffused into the R1 sediment, but effectively limited the sulfide production in the

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sediment. It can be inferred that when Strategy 1 was applied in R1, Fe3+ was first reduced

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to Fe2+ on the sediment surface and then diffused into the deeper layers where it lowered

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the sulfide concentration by forming FeS. This explains why we did not observe a

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substantial change in sulfide-producing zones compared to the corresponding zones in R2.

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3.3. Microbial community analysis

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The sludge samples taken from the upper, middle, and bottom layers of the

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sediments in three reactors were analyzed to investigate the effects of Fe3+ addition on the

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metabolically active bacterial and archaeal communities. The α-diversity estimators show

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that all sludge samples had similar bacterial diversities except for a slightly lower diversity

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in the sludge sample in the upper layer of R3 (Table S2), revealing that high Fe3+ dosage

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reduced the bacterial community diversity in the upper layer of R3. PCoA based on

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unweighted UniFrac shows the β-diversity of the bacterial community on different layers

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of the three reactors, showing that the bacterial community in the upper, middle, and

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ACCEPTED MANUSCRIPT bottom layers of R3 sediment were separated from those in the R1 and R2 sediments (Fig.

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6 and Fig. S4). The present result indicates that only high Fe3+ dosage significantly shaped

306

the bacterial community in the sewer sediment reactor. As for the archaeal communities

307

in the three sewer reactors, the total number of OTUs and the diversity of the archaeal

308

communities in the three reactors were similar (Table S3). In addition, PCoA showed that

309

there was not a clear separation between the three reactors (Fig. S5). Thus, both low and

310

high Fe3+ dosages did not substantially change the archaeal abundance and diversity in the

311

sewer sediments.

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The dominant phyla in the three reactors were the same, such as Bacteroidetes,

313

Chloroflexi, Firmicutes, Proteobacteria, Spirochaetes and Synergistetes. In the R2 without

314

Fe3+ addition, the total relative abundance of the six phyla were 90.1%, 83.9% and 84.0%

315

in the three layers in sequence, respectively (Fig. S6). In the R1 with low Fe3+ dosage, the

316

total abundance of the predominant phyla was the same with the corresponding layers in

317

the R2 sediment. In the R3 with high Fe3+ dosage, although the predominant phyla were

318

same with those in the R2, the relative abundances of several phyla were significantly

319

changed. For instance, compared to the relative abundances of Bacteroidetes (10.2%),

320

Chloroflexi (10.1%) and Synergistetes (14.3%) in the upper layer of R2 sediment, their

321

relative abundances in the upper layer of R3 sediment changed to 33.6% and 4.8% and

322

1.4%, respectively (Fig. S6). The similar trend about the three phyla was also observed in

323

the middle layers of R3 sediment. Moreover, in the middle layer of R3 sediment, the

324

relative abundance of Proteobacteria was increased to 27.1% compared to 19.0% that in

325

R2 sediment. In the bottom layer of R3 sediment, the relative abundances of the six phyla

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ACCEPTED MANUSCRIPT were similar with those in the R2 sediment expect for an increase abundance of Firmicutes

327

(13.2% in the R3 vs. 4.2% in the R2). As for archaeal community, Euryarchaeota

328

predominated in all the layers of the three sewer sediment reactors, accounting for more

329

than 99% (Fig. S7a). However, the low and high Fe3+ dosages had a minor effect on

330

archaeal community at the phylum level (Fig. S7a).

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At the genus level, the bacterial community also exhibited differences among the

332

three reactors. As for the fermentation-related bacteria, Lactococcus, Paludibacter, T78,

333

and Trichococcus were the most commonly identified fermentative genera in the three

334

reactors (Fig. 7), which often found in sewer systems (Mohanakrishnan et al., 2011, Liu et

335

al., 2015a, Liang et al., 2016). These microbes can decompose complex organics into

336

simple compounds for SRB growth and respiration. With the addition of either low or high

337

Fe3+ dosage, the total abundance of fermentative genera was increased in R1 and R3,

338

especially in the upper layer of R3, in which the relative abundances of Lactococcus (5.6%)

339

and Trichococcus (5.8%) in the upper layer of R3 sediment were significantly higher than

340

those in R2 sediment (0.1% and 0.4%). Baek et al. (2015) also observed that the

341

abundance of Trichococcus in the anaerobic digestion system with the supplementation of

342

(semi)conductive ferric oxides. The present result suggests that the presence of Fe3+

343

provided a favorable environment for fermentative bacteria to flourish in the sewer

344

sediment reactors.

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Ferric salt dosing also significantly shaped the sulfidogenic communities in

346

sediments, as shown in Figs. 7 and 8. Six main sulfate-reducing genera were identified in

347

the

sewer

sediments,

namely

Dehalobacterium, 16

Desulfobacter,

Desulfobulbus,

ACCEPTED MANUSCRIPT Desulfococcus, Desulfomicrobium, and Desulfovibrio (Fig. 7), which have also been

349

reported to inhabit sewer systems (Sun et al., 2014, Liu et al., 2015a). The total relative

350

abundance of SRB in each layer of sediment from the three reactors are presented in Fig.8,

351

suggesting that the distributions of SRB were similar between R1 sediment and R2

352

sediment, which was significantly different from that in R3 sediment. In detail, in R2

353

sediment without Fe3+ addition, most of the SRB were located in the upper layer (Fig. 8),

354

which was attributed to the limited diffusion capacity of organic compounds. The total

355

percentages of the six sulfate-reducing genera in the upper, middle, and bottom layers was

356

15.5%, 7.3%, and 4.1%, respectively (Fig. 8). In R1 with low Fe3+ dosage, we observed a

357

similar SRB abundance with that in R2 sediment, indicating that low Fe3+ dosage generally

358

had a very weak impact on SRB abundance and distribution. However, in R3 with high Fe3+

359

dosage, the SRB abundance in the upper layer significantly decreased to 1.3%, in which

360

the relative abundance of Dehalobacterium, Desulfobacter, Desulfobulbus, Desulfococcus,

361

Desulfomicrobium, and Desulfovibrio were 0.03%, 0.2%, 0.3%, 0.01%, 0.4 % and 0.4%,

362

respectively, compared to 0.4%, 7.2%, 1.5%, 0.5%, 4.5%, 1.9 in order in the upper layer of

363

R2 sediment. The highest SRB abundance in R3 sediment appeared in the middle layer

364

(10%), which was even higher than that in R2 (7.3%). These results suggest that the growth

365

of SRB in sewer sediment was inhibited by high Fe3+ dosage in upper layer but promoted in

366

the middle layer. In the bottom layer, the SRB abundance in R3 (5.3%) was similar to that in

367

R2.

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Fe3+ addition may also influence iron-reducing bacteria (IRB) activity. Therefore,

369

IRB was also analyzed in this study. Clostridium, Geobacter, and Sulfurospirillum as well as 17

ACCEPTED MANUSCRIPT some sulfate-reducing genera (Desulfobacter, Desulfobulbus, and Desulfovibrio) were the

371

most commonly identified iron-reducing genera in the three sewer reactors (Fredrickson

372

and Gorby, 1996, Herrera and Videla, 2009). IRB also exhibited spatial distribution in the

373

sewer sediments (Fig. 8). In R2 without Fe3+ addition, the detection of IRB in the upper,

374

middle, and bottom layers of the sediment was mainly attributed to the presence of

375

Desulfobacter, Desulfobulbus, and Desulfovibrio, accounting for 12.2%, 5.4%, and 1.9%,

376

respectively (Fig. 7). In R1 with low Fe3+ dosage, the relative abundance of IRB was similar

377

with those in R2 except for a slightly lower IRB abundance in the middle layer of R1

378

sediment, which was due to a slight decrease in Desulfobacter abundance (from 2.0% to

379

0.2%). The present result indicates that, in general, low Fe3+ dosage had a negligible effect

380

on IRB abundance. In R3 with the high Fe3+ dosage, the relative abundance of IRB in the

381

upper layer of R3 sediment was significantly lower than those in the R2 sediment, which

382

resulted from a significant decrease in relative abundances of Desulfobacter,

383

Desulfobulbus, and Desulfovibrio in the upper layer of R3 sediment compared to that in

384

the R2 sediment (Fig. 7). However, the relative abundance of Sulfurospirillum in the upper

385

layer of R3 sediment was significantly higher than that in the upper layer of R2 sediment

386

(4.9% vs. 1.2%), indicating that Desulfobacter, Desulfobulbus, and Desulfovibrio cannot

387

tolerate the toxic effect of high Fe3+ dosage but Sulfurospirillum can. In the middle and

388

bottom layers of R3 sediment, IRB abundance dramatically increased (12.0 and 6.9%,

389

respectively) (Fig. 7). This increase may be attributed to the diffusion of Fe3+ into the deep

390

layer in R3 sediment.

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ACCEPTED MANUSCRIPT 391

As for the genus-level archaeal community, Candidatus Methanoregula,

392

Methanobacterium,

393

Methanosaeta, Methanosarcina, and Methanospirillum were the identified methanogenic

394

genera (Fig. S7b), among which Methanosaeta with a relative abundance of > 82% was the

395

predominant methanogenic genus in all layers of the three reactors (Fig. S7b). The genus

396

Methanosaeta uses acetate instead of hydrogen as a substrate (Sun et al., 2014) and is

397

mainly responsible for methane emission in sewers under anaerobic conditions (Rotaru et

398

al., 2014). With the low Fe3+ dosage in R1, the abundances of methanogenic genera were

399

similar with those in R2 sediments. With high Fe3+ dosage in R3, the total relative

400

abundances of the methanogenic genera in the middle and bottom layers of R3 sediments

401

were similar with those in R2 sediments. However, in the upper layer of R3 sediment, the

402

total relative abundance of methanogenic genera reduced by 9% compared to that in R2

403

sediment (90.1% vs. 98.8%) (Fig. 8), in which the relative abundance of Methanosaeta

404

decreased to 82.7% compared to 96.8% in the upper layer of R2 sediment. It should be

405

noted that the very high relative abundance of MA was attributed to archaeal and

406

bacterial communities were separately analyzed via Illumina Miseq sequencing, and their

407

abundances were calculated based on the total number of sequences of archaeal

408

community rather than the total number of sequences of archaeal and bacterial

409

communities. The present results are consistent with the similar methane production

410

findings between R1 and R2. They also support that methane production reduced by

411

approximately 20% in R3 with the high Fe3+ dosage (Fig. 3b). Therefore, high Fe3+ dosage

412

only partially suppressed methanogenic activity in the upper layer of R3 sediment.

Methanomassiliicoccus,

Methanomethylovorans,

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Methanolinea,

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ACCEPTED MANUSCRIPT 413

4. Discussion

415

4.1. Mechanisms of sulfide and methane control in sewer sediment reactors receiving

416

iron dosing with different strategies

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Three different ferric chloride dosing strategies, low, medium, and high Fe3+ dosage,

418

with different dosing frequencies were investigated in this study. With all tested strategies,

419

Fe3+ dosing efficiently controlled sulfide accumulation in sewer reactors over time, which is

420

consistent with the findings by Zhang et al. (2009), Gutierrez et al. (2010), and Lin et al.

421

(2017), who found that ferric salts could significantly reduce sulfide accumulation in sewer

422

systems. However, ferric dosing significantly suppressed methane production by more

423

than 50% (Zhang et al., 2009, 2010, Ganigué et al., 2018), which was higher than that

424

observed in in this study (21% on average). It may be attributed to the difference between

425

sewer sediment and sewer biofilm. Sewer sediment is thicker than sewer biofilm

426

(centimeters vs. hundreds of micrometers) (Liu et al., 2015c), and MA mainly inhabit in the

427

deeper layer of sewer sediment. As such, MA in sewer sediments may be subjected to less

428

effect of ferric dosing than that in sewer biofilms. Additionally, the present results showed

429

that the control mechanisms of different iron dosing strategies against sulfide and

430

methane emission in the reactors differed. It should be noted that medium Fe3+ dosage

431

with medium frequency was only considered as a transitional stage from low to high

432

dosage. Therefore, we only discussed the effects of the other two Fe3+ dosing strategies on

433

sulfidogenic and methanogenic activities in the sewer sediments.

434

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When the low-dosage, high-frequency Fe3+ dosing strategy (Strategy 1) was used 20

ACCEPTED MANUSCRIPT in R1 and R3, sulfide concentration in bulk water was suppressed by 94% on average (Fig.

436

2), and methane production was not affected. In this strategy, the molar ratio of Fe:S was

437

1.3:1 based on the average Fe3+ concertation (49 mg/L) and sulfide concentration (22.0 ±

438

4.7 mg S/L) in the absence of Fe3+, which was close to the minimal molar ratio of Fe:S

439

(0.9:1) proposed by Firer et al. (2008). This low Fe3+ dosage with high-frequency did not

440

cause an obvious change in either the sulfate-reducing or methane-producing activity (Fig.

441

3), although many previous studies reported the inhibitory effects of Fe3+ on SRB (Lovley

442

and Phillips, 1987, Utgikar et al., 2001) and MA (Van Bodegom et al., 2004, Reiche et al.,

443

2008). A previous study also demonstrated that sulfide removal from a sewer was mainly

444

attributed to physicochemical reactions between sulfide and Fe3+ rather than inhibiting the

445

sulfate-reducing activity of sewer biofilm when iron-rich drinking water treatment sludge

446

was dosed (Fe3+ was the main form of iron) (Sun et al., 2015). Our microbial analysis

447

further confirmed this conclusion, in which the spatial distribution and abundances of SRB

448

and MA communities in R1 were similar to those in R2 without Fe3+ addition (Figs. 7 and 8).

449

The results suggest that sulfide control by Fe3+ dosing with Strategy 1 was mainly

450

attributed to chemical sulfide oxidation and precipitation in bulk water (Equations 1 and 2),

451

which is partially supported by the high amount of elemental sulfur deposited in the upper

452

layer of R1 sediment (Fig. S8). Additionally, the sulfide production zone (0-5 mm) did not

453

have a substantial change, which is consistent with the findings observed by Liu et al.

454

(2015c) and Liu et al. (2016), in which they found that the sulfidogenic activity mainly

455

maintained in the surface layer as substrates (e.g. COD and sulfate etc.) could only

456

penetrate a few millimeters into the sewer sediments.

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21

ACCEPTED MANUSCRIPT When the high-dosage, low-frequency Fe3+ dosing strategy (Strategy 3) was

458

employed in R3, the applied molar ratio of Fe:S was 1.5:1, which was similar to that in

459

Strategy 1. The SRB activity only in the upper layer of sewer sediment was nearly

460

completely suppressed. The activity of MA was also significantly suppressed in the upper

461

layer of sewer sediment. The inhibitory effects on SRB and MA could be resulted from the

462

toxicity of Fe3+ or the high ORP values induced by Fe3+. Compared to Strategy 1, the high

463

dosage with low frequency in Strategy 3 (7163.8 mg Fe3+ every 7 days vs. 210.7 mg Fe3+

464

every 6 hours) ensured more ferric salt precipitation into the sewer sediment in the R3 (Fig.

465

5), and more Fe3+ could diffuse into the deep sediment layers to increase the ORP level (Fig.

466

4b). However, when deeper than 20mm, the ORP values were only slightly higher than

467

that in R2, indicating that a little amount of Fe3+ diffused into the deep layer (>20 mm).

468

The limited diffusion of Fe3+ ions into sewer sediment in R3 could be probably attributed

469

to Fe3+ ion dissociation. This is because dissociated Fe3+ ions can be easily converted into

470

insoluble amorphous Fe(OH)3 in the upper layer (Davydov et al., 1998). Moreover, the

471

amount of Fe3+ could lower than the threshold that may cause inhibitory effects on SRB

472

activity. Utgikar et al. (2002) also found that the effect of heavy metals can be stimulatory

473

at a low concentration and toxic/inhibitory at a high concentration. Thus, we observed

474

that the main sulfide production zone in R3 sewer sediment shifted from 0-5 mm to 20-25

475

mm in depth, which was supported by the highest abundance of SRB present in the middle

476

layer of R3. Moreover, sulfide production sharply increased in the deep layer (Fig. 4a).

477

Meanwhile, the sulfide produced in the deeper layer could be oxidized to sulfur by the

478

presence of Fe3+, leading to a significant accumulation of sulfur in the deep layer of R3

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ACCEPTED MANUSCRIPT sediment compared to that of R1 sediment (Fig. S8), which in return contributed sulfide

480

production via sulfur reduction (Qiu et al., 2017, Zhang et al., 2018b, Zhang et al., 2018c).

481

The present results are consistent with the abundances of SRB and MA in the deep layer of

482

sewer sediment. The SRB and MA abundances in middle and bottom layers both increased

483

(Fig. 8). The increase in abundances of SRB and MA in the deep layer may be due to the

484

fact that the inhibited SRB and MA in the upper layer consumed fewer organic compounds,

485

resulting in greater diffusion of organic compounds into the deep sediment layer to

486

support the growth and activities of SRB and MA.

487

4.2. Practical Implications

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Iron salts are very common chemicals for control of sulfide and methane production

489

in sewers (Ganigue et al., 2011). In general, dissolved sulfide concentration is difficult to

490

decrease to a very low level (0.1-0.2 mg S/L) and excessive iron salts are generally needed

491

to achieve complete sulfide control (Firer et al., 2008, Zhang et al., 2008). The higher ratios

492

would be ascribed to low pH, low sulfide concentration targeted (Firer et al., 2008). Zhang

493

et al. (2010) observed that a much higher Fe/S molar ratios (3.1-66.1) were used to

494

effectively control sulfide concentration in laboratory-scale rising main sewers. In this

495

study, two similar and plausible Fe/S ratios (1.3:1 and 1.5:1), which were greater than the

496

minimal Fe/S ratio (Fe/S=1.3:1), with different frequencies were tested to provide general

497

guidelines.

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498

Under the premise of similar Fe/S molar ratios, although a low-dosage,

499

high-frequencies ferric dosing strategy (Strategy 1) could not inhibit SRB and MA activities

500

in sewer sediments, but it would be a cost-effective solution for sulfide control with both 23

ACCEPTED MANUSCRIPT thin and thick sewer sediments if there is no specific demand in control of methane

502

generation. First, dosing ferric salts at an interval of 6 h was demonstrated to efficiently

503

control sulfide emission. Previous studies also found that a frequency of pumping events

504

within several hours efficiently reduced sulfide emission from sewers (Zhang et al., 2009,

505

2010, Zhang et al., 2011, Ganigué et al., 2018). Second, ferric salts can be precisely dosed

506

based on the sulfide production as sulfide was removed only via chemical oxidation and

507

precipitation. Third, adding ferric salt in sewers would facilitate the phosphate removal

508

and sludge dewatering in a downstream WWTP (Gutierrez et al., 2010, Rebosura Jr et al.,

509

2018), if the dosed iron salt could run off from the sewers into the WWTP. Thus, this

510

strategy (Strategy 1) could also bring great benefits to the downstream WWTPs as

511

approximately 98% of the Fe3+ dosed with Strategy 1 ran out from the sewer reactor (R1)

512

based on the iron balance analysis (see Fig. 5 and S9). Additionally, since the majority of

513

iron would finally entered into the downstream WWTPs, the formed sediments deposited

514

in sewers from FeS precipitates would be neglected.

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In contrast, a high-dosage, low-frequency ferric dosing strategy (Strategy 3) may

516

not be viable in gravity sewers. On one hand, in gravity sewers with thin sediments (depth

517

< 20 mm), this strategy could efficiently inhibit SRB and MA activities, but it could not

518

achieve satisfactory control effectiveness of sulfide and methane during a long-term

519

period (Fig. 2 and 3). On the anther hand, in gravity sewers with thick sediments

520

(depth >20 mm), this strategy could not efficiently inhibit SRB and MA activities in the

521

deep layer, indicating that sulfide and methane could not be effectively controlled.

522

Moreover, a large proportion of the Fe3+ dosed with Strategy 3 was deposited in the

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24

ACCEPTED MANUSCRIPT sediment in R3 (approximately 49%) (see Fig. 5 and S9). Necessary measures need to be

524

periodically taken to remove the sediments from sewers, such as hydraulic flushing with a

525

high flow rate (Liang et al., 2019). As such, the operational cost would further increase.

526

Additionally, the high-instant dosage, low-frequency ferric dosing strategy could reduce

527

COD loss to some extent (Fig. S2) during wastewater transportation in gravity sewers due

528

to the inhibited SRB and MA. However, considering the unsatisfactory long-term control

529

effectiveness of sulfide and methane, the benefit of minimizing COD loss in gravity sewers

530

would not be attractive.

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531

5. Conclusions

533

The effects of different ferric dosing strategies on sulfidogenic and methanogenic activities

534

in sewer sediments were investigated via laboratory-scale sewer sediments reactors. The

535

main findings are shown as following:

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Different ferric dosing strategies effectively significantly reduced sulfide levels in

537

the sewage (> 90%), but these strategies had different impacts on SRB activity. A

538 539

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low-dosage, high-frequency ferric dosing strategy did not inhibit SRB activity and did not change sulfide-producing zone of a sewer sediment as Fe3+ could not

540

penetrate into the sediment. Sulfide was controlled only via chemical oxidation

541

and precipitation. In contrast, a high-dosage, low-frequency ferric dosing

542

strategy nearly completely SRB activity based on the fact that sulfate reduction

25

ACCEPTED MANUSCRIPT was decreased and sulfide-producing zone moved to deeper layers. Thus, sulfide

544

was removed through inhibition of SRB activity and chemical oxidation and

545

precipitation.

546

A low-dosage, high-frequency ferric dosing strategy had negligible effects on MA

547

activity and methane production. A high-dosage, low-frequency ferric dosing

548

strategy could only partially suppress MA activity and reduce methane

549

production (21% reduction on average).

550

When the molar Fe/s ratio is certain, a low-dosage, high-frequency Fe3+ dosing

551

strategy would be more cost-efficient for sulfide control in gravity sewers in

552

practice.

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Therefore, a deep insight into the effects of different ferric dosing strategies on the control

554

of sulfide and methane production obtained in this study could provide useful information

555

for optimizing odor control strategies in gravity sewers receiving ferric salts.

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Acknowledgements

558

The authors acknowledge support from the National Natural Science Foundation of China

559

(51638005), the Guangdong Provincial Science and Technology Planning Project

560

(2017B050504003).

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Table 1. The three different Fe dosing strategies used in R1, R2 and R3. 3+

Operational period (d)

3+

Instant

Intervals

dosing

dosage

between two

dosage

ratio (mol

strategy

(mg Fe)

dosages

(mg Fe/L)*

Fe/mol S)

R1

R2

84-141 and Strategy 1

Average Fe/S

#

R3 64-76 and

-

210.7

176-207

Average Fe

85-94

6h

-

-

113-133

1684.7

3d

Strategy 3

-

-

176-207

7163.8

7d

49.0

1.3

32.6

0.8

59.5

1.5

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Strategy 2

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Fe

“–“ represents the strategy was or was not performed in the reactors, respectively. 3+

*: the average Fe dosage was equal to the result of using the instant dosage divided by total volume

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of wastewater passing through the sewers during an interval between two dosages.

#: the sulfide concentration in R2 (a control reactor) to calculate the average Fe/S molar ratios in the 3+

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three different Fe dosing strategies.

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Fig. 1. Simplified diagram of the laboratory-scale sewer sediment reactors.

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Fig. 2. Comparisons of the effluent sulfide concentration (a) between R1 and R2 and

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(b) between R2 and R3 during the long-term operation. Strategy 1 represents low Fe3+ dosage with high frequency; Strategy 2 represents medium Fe3+ dosage with

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medium frequency and Strategy 3 represents high Fe3+ dosage with low frequency. Strategy 1 was used in R1 and R3. Strategy 2 and 3 were used in R3.

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Fig. 3. (a) The influent and effluent sulfate (a) and CH4 (b) concentration in R1, R2 and

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R3 during the long-term operation. Strategy 1 represents low Fe3+ dosage with high frequency; Strategy 2 represents medium Fe3+ dosage with medium frequency and

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Strategy 3 represents high Fe3+ dosage with low frequency. Strategy 1 was used in R1 and R3. Strategy 2 and 3 were used in R3. The strategy 1 used in R1 is not marked in this figure, but the strategies employed in R3 are marked.

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Fig. 4. Profiles of (a) measured total dissolved sulfide and (b) ORP in the sewer

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sediments of R1, R2 and R3.

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Fig. 5. The pictures of the sewer sediments in R1, R2 and R3 at the end of experiment.

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Note that the top 20mm represents the liquid phase.

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Fig. 6. Principal coordinates analysis (PCoA) of the bacteria in the sludge samples. The PCoA plot was based on the unweighted UniFrac distance matrix and relative

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abundances of bacterial OTUs were used as the dataset.

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Fig. 7. Heatmap showing the relative abundances and distributions of genus-level bacterial community in the different layers of the three sewer sediment reactors. The genera with <0.5% of relative abundance in all the sludge samples are not shown

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in this figure. The color code indicates the relative abundance, ranging from light purple (low abundance) to green (high abundance). (For interpretation of the references to color in this figure legend, the reader is referred to the web version of this article.)

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Fig. 8. The total relative abundance of sulfate-reducing bacteria (SRB), methanogenic archaea (MA) and iron-reducing bacteria (IRB) in the upper, middle and bottom

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layers in the three sediment reactors. (For interpretation of the references to color in

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this figure legend, the reader is referred to the web version of this article.)

ACCEPTED MANUSCRIPT Sulfide control by low Fe3+ dosage was mainly due to chemical sulfide oxidation and precipitation High Fe3+ dosage completely suppressed SRB activity only in the surface layer of

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sediment High Fe3+ dosage altered active sulfide-producing zone in the sediment

Only high Fe3+ dosage partially suppressed methanogenic activity and reduced

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methane production

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A low-dosage, high-frequency Fe3+ dosing strategy would be preferentially used

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in gravity sewers

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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.

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☐The authors declare the following financial interests/personal relationships which may be considered as potential competing interests: