Effects of dosing iron- and alum-containing waterworks sludge on sulfide and phosphate removal in a pilot sewer

Journal Pre-proofs ‘Effects of dosing iron- and alum-containing waterworks sludge on sulfide and phosphate removal in a pilot sewer Sohan Shrestha, Jagadeeshkumar Kulandaivelu, Keshab Sharma, Guangming Jiang, Zhiguo Yuan PII: DOI: Reference:

S1385-8947(20)30064-4 https://doi.org/10.1016/j.cej.2020.124073 CEJ 124073

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Chemical Engineering Journal

Received Date: Revised Date: Accepted Date:

10 October 2019 4 January 2020 8 January 2020

Please cite this article as: S. Shrestha, J. Kulandaivelu, K. Sharma, G. Jiang, Z. Yuan, ‘Effects of dosing iron- and alum-containing waterworks sludge on sulfide and phosphate removal in a pilot sewer, Chemical Engineering Journal (2020), doi: https://doi.org/10.1016/j.cej.2020.124073

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Effects of dosing iron- and alum-containing waterworks sludge on sulfide and phosphate removal in a pilot sewer Sohan Shrestha, Jagadeeshkumar Kulandaivelu, Keshab Sharma, Guangming Jiang, Zhiguo Yuan* Advanced Water Management Centre (AWMC), The University of Queensland, St. Lucia, QLD 4072, Australia *Corresponding

author: [email protected]

Abstract Reusing waterworks aluminium (Al)- or iron (Fe)-sludge instead of chemical coagulants such as Al- or Fe-salts, is a credible solution for sulfide and phosphate control in sewer bulk phase for environmental and economic reasons. We comprehensively evaluated and compared the effects of direct dosing of waterworks Fe-/Al-sludge in pilot-scale sewer rising mains, particularly focusing on removal of sulfides and phosphate and underlying possible mechanisms. Changes in other sewage characteristics were also examined. Waterworks Fesludge dosing was effective for sulfide removal at a ratio of 0.290.06 mg S/mg Fe, but exhibited limited effect on phosphate removal. Likewise, Al-sludge was effective for phosphate removal at ratio of 0.290.01 mg P/mg Al, but with limited effect on sulfide removal. The mixing of the sludge stream with raw wastewater, i.e. dilution effect, was primarily responsible for observed reduction in soluble chemical oxygen demand (sCOD) concentrations, under both Fe-/Al-sludge dosing. The Fe-/Al-sludge dosing did not cause any increment in dissolved methane (CH4) and nitrous oxide (N2O) formation, nor release of other metals. Combined spectroscopic, spectrometric, and microscopic analyses suggest a precipitation reaction between sulfide and ferric ions in Fe-sludge, is likely to be the dominant mechanism for sulfide 1

removal when dosing Fe-sludge. In terms of phosphate removal with Al-sludge dosing, ligandexchange processes between surface hydroxyl (-OH) groups and PO43- ions, favouring the formation of both inner- and outer-sphere surface Al-phosphate complexes, appears to be the dominant mechanism. These findings showed the potential multiple benefits of dosing waterworks Fe-/Al-sludge in sewers. Further system-wide, long-term studies including a comprehensive cost-benefit analysis are warranted for optimisation. Keywords: Reuse; waterworks Al-sludge; waterworks Fe-sludge; sulfide; phosphate; pilotscale sewer

1.

Introduction

Excess release of phosphorus (P) to freshwater bodies due to anthropogenic activity can cause eutrophication [1] or toxic algal blooms [2], which can negatively impact aquatic organisms and human health. Consequently, water utilities have been subject to stringent P discharge standards in recent years. Achieving targeted P removal has been a common practice in fullscale wastewater treatment plant (WWTP), but the same approach has not been applied to sewers. Sewage is also considered one of the main contributing sources of P reaching freshwater bodies, for example, through sewer overflow [3]. Likewise, sulfide build-up in sewer networks is also a serious issue causing health hazards, nuisance odour, and in-sewer corrosion [4]. Inorganic salts, particularly iron (Fe)-salts, are commonly used for controlling sulfide in sewers [5]. However, effective sulfide control demands continuous Fe-salt dosing, which incurs high operational costs. The global cost of inorganic coagulants used for water/wastewater treatment in 2018 was $1.37 billion, and this is predicted to reach $1.84 billion by 2023 [6]. Considering the enormous cost of chemical dosing, seeking low costs alternatives is a major imperative. In this context, the reuse of cost-effective materials such as waterworks aluminium-rich or iron-rich sludge (denoted as ‘Al- or Fe-sludge’ hereafter) as

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alternatives to chemical coagulants (Al-, Fe-salts), is an appealing solution for both sulfide and phosphorus removal in sewers.

Waterworks Al- or Fe-sludge, are inevitable by-products in water treatment plants (WTPs), when Al- or Fe-salt is used as primary coagulant. Babatunde and Zhao [7] reported that the average production of WTP sludge globally exceeds 3.65106 dry tonnes per annum. Annual production of WTP sludge by an Australian water utility varies between 150 to 43,500 dry tonnes per annum [8, 9]. The production of WTP sludge is expected to increase over the years, considering the growing demand for potable water by a rapidly increasing global population. Hence, sustainable management of such enormous volumes WTP sludge is essential.

Traditionally waterworks sludge has been dried and stockpiled on-site for years, however it is becoming common to dispose of WTP sludge in sewers, lagoons, and landfill sites [9, 10]. The disposal options for WTP sludge, vary depending on country, local conditions and size of WTP operation [11]. In Australia, end use or disposal of WTP sludge varies in each State [9]. Of particular interest is disposal of WTP sludge to sewers, which is more common in Europe, the U.K. and the U.S.A. [12, 13]. There are many instances of sewer disposal of waterworks Alsludge being implemented in Australia [9]. In doing so, WTP sludge could be treated together with sewage sludge in the downstream WWTP, offering potentially wide-ranging economic benefits [10].

There has been growing interest in the direct reuse of waterworks Al- and Fe-sludge in sewers in recent years, but the impacts of such practice are not well understood. Studies on the feasibility of reusing Al-sludge for phosphate [14-17] or particulate pollutants removal [18], were primarily based on bench-scale experiments, often using synthetic wastewater. Studies 3

reusing Fe-sludge for phosphate removal [16, 19] and sulfide removal [20] also adopted a similar approach. In terms of full-scale application, Edwards et al. [21] showed a substantial reduction in sulfides in digester gas of a receiving WWTP on direct application of WTP Fesludge to real sewers, but the efficacy of Al-sludge on sulfide removal in the digester gas was marginal. The influence of Fe-sludge dosing on control of sulfide in sewers was not investigated. To our best knowledge, there have been no comprehensive studies to date on the application of waterworks Al- and Fe-sludge to sewers at pilot- or full-scale and their respective roles in sulfide and phosphate removal. In addition, the impacts on other sewage characteristics, such as particulate pollutants, dissolved methane, and nitrous oxide, are not fully understood. Successful reuse of Al- and Fe-sludge instead of chemical coagulants would have a major impact on both sustainable WTP sludge management and urban wastewater management. Undoubtedly, reuse of waterworks sludge will deliver both economic and environmental benefits for water utilities, by reducing the burden of WTP sludge treatment/disposal, and reducing chemical consumption in urban water system.

In this study, we aim to investigate the impact of reusing waterworks Al- and Fe-sludge on sulfide, phosphate, and other major sewage characteristics. Two pilot-scale rising main (RM) sewers are used in this study, one as an experimental system, and the other as a control. For this, two sets of experiments were conducted, reusing Fe-sludge and Al-sludge, respectively. These sludges were sourced from two different WTPs. Batch tests were also undertaken to assess impacts of Al:P dosage ratio and suspended solids on phosphate removal in sewage.

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2. Material and methods 2.1 Pilot rising mains set-up and operation This study was conducted on two pilot rising main (RM) sewers, located at the Luggage Point WWTP (Queensland, Australia). Each of the two RM pipes has an internal diameter of 0.1 m (A/V = 4.0/0.1 m = 40 m−1) and a length of 300 m. One of the sewer pipes was operated as ‘control’ and the other as ‘experimental’ line. These pilot sewers were supplied with wastewater, directed from the inlet of the Luggage Point WWTP to a storage tank (Tank 1) located adjacent to the pilot sewer systems (see Figs. 1a-b). Three sampling ports were installed in both sewer lines at the locations of 15 m, 105 m, and 210 m for manual sampling. The wastewater was pumped into the RM systems using a SHE50-16075 centrifugal 3-phase pump (7.5 Kilowatt, Lowara Pumps, Australia). The pump was equipped with Hydrovar variable frequency drive for flow control, and was fitted with an inline magnetic resonance flow meter (IFM SM2000) to maintain the designed flow rate.

Fig. 1. (a) Schematic of the pilot rising main with an internal diameter of 100 mm and a length of 300 m. The experimental line is shown here. The control line was identical but without the sludge dosing mechanism. Temperature and pH were monitored online by pumping sewage out at 45 m then returning it at 75 m using Masterflex Peristaltic Pumps (Cole-Parmer, USA) 5

maintaining 3,400 mL/min at 600 rpm; (b) pictorial representation of real rising main pipes (‘outer pipe’ is designated as experimental line while ‘inner pipe’ as control line), where influent wastewater enters from the bottom pipe and exits via the top

2.2 Fe- and Al-sludge dosing to pilot sewer Waterworks Fe- and Al-sludges obtained from two different sources were used herein. The Fesludge was obtained from the Gold Coast Desalination Plant located at Tugun, Queensland, Australia. The Al-sludge was obtained from Mount Crosby Westbank Water Treatment Plant (Queensland, Australia). The Fe- and Al-sludges were obtained in slurry and dewatered sludge cake forms, respectively (Fig. S1) (Supplementary Information, SI).

Prior to feeding into the system, the Fe-sludge was mixed and diluted with subnatant of the Dissolved Air Floatation (DAF) unit of Luggage Point WWTP. Such mixing was done in advance of feeding raw wastewater to the sewer lines. Dilution of the sludge was undertaken herein to make the sludge ‘pumpable’ into sewer pipe. The targeted total solids (TS) elevation of 1.0 g/L inside the sewer pipes was selected based on the local trade waste sewer acceptance guidelines. The characteristics of both sludges, DAF subnatant used for dilution, and the sewer feed wastewater are given in Table 1.

Both sewer pipes were at pseudo-steady state steady state prior to sludge dosing experiment. At the beginning of the experiment, both sewer pipes were replenished with fresh raw wastewater by running the feed pumps continuously for 15 min. Once the RM feed pumps were turned off, the re-circulation pumps (Fig. 1-a) were turned on, which re-circulated wastewater, at a flow rate of 110 L/min, i.e. 0.22 m/s flow velocity, from 240 m and back to the start of the pipe (at 4 m) in both sewer lines, with an aim to keep the solids in suspension during the

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experiment. The waterworks Fe- (Experiment 1) or Al-sludge (Experiment 2) in Tank 2 was then dosed into the experimental sewer line at flow rate of 12 L/min for 20 min (total dosage = 240 L). This flow rate was maintained by a peristaltic pump (L/S Precision Modular Drive, Cole Parmer, United States). The submersible pump was used in the sludge feed tank (Tank 2) to achieve proper mixing of diluted Fe-/Al-sludge during injection of sludge mixed liquor to the experimental line. The dosing amount of Fe-/Al-sludge was optimised to achieve the targeted TS concentration. The re-circulation mode was maintained for 6 hr, within the typical range of hydraulic retention time for real sewer pipes.

To assess the impacts of Fe-/Al-sludge dosing on dissolved sulfide, phosphate, and other sewage characteristics, wastewater quality parameters were monitored through grab sampling. However, the sediment depth or composition of settled materials in the sewer pipes were not measured during both Fe-/Al-sludge dosing tests. Grab samples were taken before (pre-dosing samples, t0 = 0 min) and after sludge dosing at pre-designated time intervals over 6 hr (t20 = 20 min, t40 = 40 min, t60 = 60 min, t120 = 120 min, t180 = 180 min, t240 = 240 min, t300 = 300 min, t360 = 360 min). The grab samples were taken simultaneously at three different locations (15 m, 105 m, and 210 m) in both sewer pipes (Fig. 1-a). Considering the equilibrium conditions maintained throughout the pipe, sampling at three locations was considered triplicate (n=3). Wastewater samples were analysed, using the methods outlined in Section 2.4, for total and volatile solids, total and soluble Fe/Al, total and soluble chemical oxygen demand (COD), dissolved sulfur species, and phosphate (PO43-) concentrations. In addition, pH and temperature were monitored online in both sewer lines (Fig. 1-a).

For the Fe-sludge dosing test, samples were taken from the feed tank (Tank 2) (during the sludge feeding period) and the experimental line at the end of the test for characterisation using

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X-ray Diffraction (XRD), Scanning Electron Microscopy coupled with Energy Dispersive XRay (SEM-EDX), and Attenuated Total Reflectance-Fourier Transform Infrared (ATR–FTIR) spectroscopy. These analyses were carried out to understand the possible mechanisms behind dissolved sulfide and phosphate removal when reusing Fe-sludge. Details of these analyses are provided in Section 2.4.

2.3 Batch tests to investigate effects of Al:P dosage ratio and suspended solids on phosphate removal Batch tests were carried out on unfiltered and filtered sewage, at various Al:P molar ratios, to evaluate the effects of Al:P dosage ratio and suspended solids on phosphate removal. This was required as Al-sludge was dosed in excess during the pilot study. All tests were carried out in borosilicate glass reactors (Fig. S2). The filtered sewage was prepared using glass microfiber filters [CAT No. 1822-047, Grade GF/CTM (1.2 m). Al-sludge was dosed into the reactor at various dosing rates, based on the molar ratio (Al:P) of the total Al concentration versus total dissolved P concentration in the sewage. Both pH and DO were measured online. Prior to each test, the reactor was filled with fresh sewage (no headspace). Each test was performed in duplicate (n=2) at pH 7.50.1, with DO 0.70.1 mg/L, and at ambient temperature of 242.0 C. An overview of the experimental design is presented in Table 2.

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Table 1. Characteristics of feed wastewater (raw sewage), DAF subnatant (used as diluent), Fe-/Al-sludge including diluted Fe-/Al-sludge (n = number of measurements as indicated in parentheses; if n is not stated, a single measurement was taken) Fe- and Samples Alsludge dosing tests Feed wastewater

Fe-sludge dosing

DAF subnatant

Original Fe-sludge

Sample Type

NH4-N (mg N/L)

PO4-P S(-II) (mg P/L) (mg S/L)

SO42(mg S/L)

tCOD (mg/L)

(UF) (F)

N.D.

(UF)

16.7 (n=2) N.D.

884.0 (n=2) 0.01 (n=2)

5.2 (n=2) N.D.

140.5 (n=2) 75.6 (n=2)

3126.787. 8 (n=3)

1.00.6 (n=6) N.D. (n=6)

2.40.3 (n=6) 0.30 (n=6)

6.00.1 (n=6) 4.50.1 (n=6)

28.30.4 (n=6) 26.10.5 (n=6)

388.86.8 (n=6)

N.D.

0.5

1.2

37.4

122.11.4 mg/g DS (n=3)

4.30 mg/g DS (n=3)

1.210 mg/g DS (n=3)

4.30.1 mg/g DS (n=3)

(UF)

DAF subnatant

(UF)

Original Al-sludge

(F) Dry solids (DS)

1.80.1 (n=6) 0.20 (n=6) N.D.

5.50.1 (n=6) 4.10 (n=6) 1.2

37.90.2 (n=6) 36.50.4 (n=6)

sCOD (mg/L)

Solid contents TSS VSS TS (mg/L) (mg/L) (g/L)

0.20.1 (n=6) N.D. (n=6)

(F)

Al-sludge dosing

S(T), S(sol.) (mg S/L)

Unfiltered (UF) Filtered (F)**

(F) Feed wastewater

Parameters* and respective values Al(T), Fe(T), P(T), P(sol.) (mg P/L) Al(sol.) Fe(sol.) (mg Al/L) (mg Fe/L)

393.714.3 (n=6) 39.80.2 (n=6)

4.60 (n=6)

16.00.3 (n=6)

25.50.3 (n=6)

N.D.

1.7

0.04

37.0

45.4

229.36. 5 (n=6)

179.2 140.0 1.70.0 0.50.1 20.1 11.8 (n=6) (n=6) (n=6) (n=6)

206.7

41.0 (n=2)

VS (g/L)

100.0 1.6

0.3

29.0 (n=2) 52.00.0 10.23.0 1 (n=3) (n=3) 1501.75 5.9 (n=3)

42.80.4 (n=6)

5.30.05 (n=6)

24.50.2 (n=6)

9.30.2 (n=6)

267.24. 2 (n=6) 25.0 (n=2)

2.63

0.13

0.04

193.6 167.2 1.00.01 0.40.04 3.0 3.0 (n=6) (n=6) (n=6) (n=6)

1.35

0.19

37.8

0.20.002 13.70.2 8.50.1 %TS %VS g/g DS (n=3) (n=3) (n=3) * Al(T), Fe(T), P(T), S(T) = total (dissolved + particulate) Al, Fe, P, and S concentrations (unfiltered samples); Al(sol.), Fe(sol.), P(sol.), and S(sol.) = total dissolved Al, Fe, P, and S concentrations (filtered samples) **samples were filtered using 0.22 m syringe PES filter, Millex total dissolved sulfide = S(-II) N.D. = non-detectable (<0.01), below the instrument limit of quantification (LOQ) DS = dry solids, such that moisture content (%) = 86.30.2

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During each test, total Al (Al(T)), total dissolved Al (Al(sol.)), and phosphate (PO43-) concentrations were monitored and analysed using methods described in Section 2.4. Samples were taken before- (t0 = 0 min) and after Al-sludge dosing at pre-designated time intervals (t10 = 10 min, t30 = 30 min, t60 = 60 min, t90 = 90 min, t120 = 120 min). Samples taken at t0 = 0 min (before Al-sludge dosing) and t120 = 120 min (following end of test) during the batch tests 1, 3, and 4 (using unfiltered sewage), were further characterised using ATR–FTIR spectroscopy including 27Al and 31P solid-state Nuclear Magnetic Resonance spectroscopy (see Section 2.4). The samples taken t0 = 0 min were also characterised using XRD analysis. These XRD and spectroscopic analyses were undertaken to determine the mechanism behind PO43- removal, when dosing Al-sludge.

Table 2. Overview of experimental design for batch tests to investigate effects of Al:P dosage ratio and suspended solids on phosphate removal. Here, test duration = 2 hr, number of replicates, n = 2 Batch tests Tests using unfiltered sewage

Test No. 1 2 3 4

Al:P molar ratio 1:1 1.5:1 2:1 3:1

Tests using filtered sewage

5 6

2:1 3:1

2.4 Analytical methods For measurement of dissolved sulfur species (sulfide, sulfate, sulfite, thiosulfate) concentrations, 1.5 mL sample was filtered (0.22 μm, Millipore, Millex GP) immediately after collection, and preserved with a 0.5 mL sulfide anti-oxidant buffer (SAOB) solution. Samples were then analysed using ion chromatography (IC) with a UV and conductivity detector (Dionex ICS-2000), as described elsewhere (Keller-Lehmann et al. 2006). Samples for 10

phosphate (PO43-) analysis were filtered (0.22 μm, Millipore, Millex GP) immediately after sampling, and then analysed using a Lachat QuickChem 8000 flow injection analyser (FIA) (Lachat instrument, Milwaukee, Wisconsin). Total Fe (Fe(T)), Al (Al(T)), P (P(T)), and S (S(T)) concentrations (dissolved + particulate) in unfiltered samples, and total dissolved Fe (Fe(sol.)), Al (Al(sol.)), P (P(sol.)), and S (S(sol.)) concentrations in filtered samples (0.22 μm, Millipore, Millex GP), were measured using Inductively Coupled Plasma Optical Emission Spectroscopy (ICP-OES) (Perkin Elmer Optima 7300DV, Waltham, MA, USA). Before ICP-OES analysis, samples were digested using 70% nitric acid (HNO3). Total and volatile solids (TS, VS) and their suspended solids fractions (TSS, VSS) were measured using standard methods [22]. Total and soluble chemical oxygen demand (tCOD and sCOD) were measured using COD cell test kits (Merck, range 25–1500 mg COD/L and 500–10000 mg COD/L). For sCOD determination, samples were filtered (0.22 μm, Millipore, Millex GP) prior to analysis. Similarly, total organic carbon TOC content was measured using a TOC analyzer (TOC-5000A, Shimadzu, Japan). Likewise, dissolved methane (CH4) and nitrous oxide (N2O) were measured using the previously described protocol [23, 24].

In the Fe-sludge dosing study, Fe-sludge samples collected before- and after dosing into the experimental sewer line were characterised by using XRD, SEM-EDX, and ATR-FTIR spectroscopy. In the Al-sludge dosing study, freshly collected sludge samples were characterised by XRD. In addition, unfiltered sewage samples obtained before- and after Al-sludge dosage at various Al:PO43- molar ratios, were analysed using both ATR-FTIR and Solid-State NMR spectroscopy. Details of XRD, SEM-EDX, ATR-FTIR and 27Al and 31P solid-state NMR analysis techniques are provided in Supplementary Information (SI-1).

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2.5 Statistical analysis Mean values, standard deviation (SD), and standard error of mean (SEM) were calculated for each parameter sampled at three locations (15 m, 105 m, and 210 m, n=3). Notably, 240 L of Fe-/Al-sludge was added to the 236 m (=240 m - 4 m) sewer pipe (volume = 1850 L), giving a dilution of 11.5%. This dilution effect was incorporated to all parameters in data analysis for the Fe-/Al-sludge dosing study (see Sections 3.1 and 3.3). Because of transition phase immediately post-dosing, data collected in the first 40 min were not considered for the data analysis. Further, student’s t-test with Welch’s correction and a 95% confidence interval (CI95%) was applied to determine whether the differences observed between mean values of respective parameters in control and experimental sewer lines, were statistically significant, based on the p-values (p<0.05). All statistical tests were undertaken using GraphPad Prism software (version 7.03).

3. Results and discussion 3.1 Effects of Fe-sludge dosing on sewage characteristics Figs. 2a-f show the changes in suspended solids, Fe, and Al concentrations with Fe-sludge dosing. Mean differences in TSS and VSS concentrations between experimental and control sewer lines were 71.8±6.9 mg TSS/L (Fig. 2a) and 25.6±2.9 mg VSS/L, respectively (Fig. 2b). A decreasing trend in TSS and VSS concentration in both sewer lines was observed, which suggests the settling of solids occurred in sewer pipes during recirculation. Fe(T) concentrations also decreased over time, displaying a similar trend (Fig. 2c). Mean Fe(T) concentrations in control and experimental sewer lines were 2.0±0.3 mg Fe/L and 22.0±4.0 mg Fe/L, respectively. However, mean Fe(sol.) concentrations were relatively low at <0.5 mg Fe/L in both control and experimental sewer pipes (Fig. 2d). Interestingly, the concentration-time profiles of Al(T) were similar to the Fe(T) trend (Fig. 2e). Mean Al(sol.) concentrations were relatively

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low at <0.1 mg Al/L in both sewer lines (Fig. 2f). In addition, respective values for pH (7.0±0.02) (Fig. S3-a) and temperature (25.6±0.4 C) (Fig. S3-b) were similar in both sewer lines during Fe-sludge dosing.

Figs. 3a-b show a decrease in sulfide, sulfite, and thiosulfate concentrations in the experimental line. Notably, the total dissolved sulfides level initially present in the sewage was unusually high 16.0±0.3 mg S/L because the raw sewage was sourced from the inlet of the treatment plant (i.e. end of the sewer network). Sulfide concentration decreased rapidly at the beginning of the test, followed by a relatively slow increase with time. Dilution could be contributing to the initial decrease in sulfide concentration as DAF subnatant does not contain sulfide (Table 1). In contrast, sulfide continued to rise slowly in the control line. The mean difference in sulfide concentration between the control and experimental sewer lines was 12.8±0.5 mg S/L (p<0.05), equivalent to 80% of the sulfide concentration in raw sewage (Table 1). Overall, sulfide concentration decreased by 57.61.3% in the experimental line (Fig. 3a). The total dissolved sulfide removed per total Fe added (as Fe-sludge) was 0.290.06 mg S/mg Fe (0.510.1 mol S/mol Fe). This implies that the physicochemical processes occurring in-sewer between Fesludge and sulfide are responsible for sulfide removal. The visual observation of blackcoloured sewage following Fe-sludge dosing (Fig. S1) suggests that FeS formed through a precipitation reaction between iron and sulfide. Sun et al. [20] observed sulfide removal in a laboratory sewer reactor following addition of Fe-sludge, but this is the first time that such a process has been demonstrated in a pilot sewer. The observed sulfide to Fe molar ratio of 0.51 is lower than the theoretical stoichiometry, which is 0.67 for reactions between ferric ions and sulfide and 1.0 for reactions between ferrous irons and sulfide [25]. The observed ratio is similar to that reported by Sun et al. [20] for a laboratory sewer reactor (0.50±0.02 mol S/mol Fe). 13

There was a sharp increase in sulfate in the experimental line following addition of Fe-sludge (Fig. 3b). This is largely due to a higher sulfate concentration in DAF subnatant as compared with raw sewage (Table 1). As described in Section 2.2, the DAF subnatant was used to dilute the Fe-sludge prior to its addition. This increase caused significantly higher sulfate concentrations in the experimental line than in control sewer line (Fig. 3b). The mean difference was 17.0±2.4 mg S/L (p<0.05). Sulfate was consumed in both the experimental and the control lines, as evidenced by the sulfide profiles (Fig. 3a). However, the sulfate consumption rate and sulfide production rate in the experimental lines appear to be slightly lower that these rates in the control line. This may be due to potentially inhibitory effects of Fe-sludge dosing on sulfate-reducing activity of anaerobic sewer biofilms. Lovley and Phillips [26] showed sulfate reduction in sediments was inhibited by 86-100% when dosing ferric (oxy)hydroxide (FeOOH). Interestingly, FeOOH is one of the predominant mineral precipitates found in the Fe-sludge used herein, as evidenced by XRD spectra (Fig. 4a). Bratby [27] also reported that waterworks Fe-sludge typically contains ferric hydroxides bound with organic or inorganic compounds. The mean difference in the sulfite concentrations between the control and experimental sewer lines was 0.4±0.05 mg S/L (p<0.05), while the mean difference in thiosulfate concentration was 1.0±0.1 mg S/L (p>0.05) (Fig. 3b).

Fig. 3c shows there was a sharp decrease in phosphate concentration in the experimental line following Fe-sludge addition. Again, this is largely due to the much lower phosphate concentration in the DAF subnatant as compared with raw sewage (Table 1). This caused a mean difference in phosphate concentrations between the two sewer lines of 0.8±0.1 mg P/L (p<0.05) (4.9±0.02 mg P/L in control and 4.0±0.1 mg P/L in experimental), equivalent to 21.8% of that in raw sewage (Table 1). This represents a decrease in PO4-P concentration by

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17.31.5% in the experimental sewer line. This marginal mean difference in phosphate concentrations can be attributed to the combination of dilution and Fe-induced removal, where a dilution effect holds notable contribution (i.e. 0.7 mg P/L). This implies that the Fe-P reaction contributed little to the decrease in phosphate concentrations in experimental line. Here, the observed decrease in sulfide with limited impact on phosphate removal with Fe-sludge addition corresponds to the prevalent pH (7.0±0.02) (Fig. S3-a). In retrospect, effectiveness of sulfide and phosphate precipitation with Fe is likely to be influenced by the prevalent wastewater pH conditions albeit the detailed underlying chemistry under in-sewer condition is not investigated in this study.

Figs. 3d-e show the variations in the COD concentrations (tCOD, sCOD) between both sewer lines. The mean difference in tCOD concentration between the experimental and control sewer lines was 32.0±7.0 mg COD/L (p>0.05), which can be primarily attributed to dilution. This is because tCOD concentrations were similar in the Fe-sludge (after dilution with DAF) and raw wastewater in the beginning of experiment, following Fe-sludge dosing (Fig. 3d). Earlier, Sun et al. [20] showed an increase in tCOD concentration by 40.0±4.0 mg/L with Fe-sludge addition. Such observed differences may be attributed to individual properties and the origin of respective waterworks Fe-sludge. The Fe-sludge used herein, sourced from the seawater desalination plant, has relatively low tCOD content (60.1±1.7 mg/g dry solids) as compared with the tCOD content of Fe-sludge (352.0±9.0 mg/g dry solids), which was sourced from freshwater treatment plant [20]. Without considering the dilution effect, the mean difference in sCOD concentration between the control and experimental sewer lines was 35.2±4.1 mg COD/L (p<0.05) (Fig. 3e), which is equivalent to 15.3% of that in raw sewage (Table 1). Dilution of the sludge stream with the raw wastewater could have caused the decrease in sCOD

15

(30.0 mg COD/L), as Fe-sludge has a lower sCOD concentration after dilution with DAF as compared with sCOD in raw wastewater.

Similarly, differences in mean concentrations of dissolved CH4 and N2O in experimental and control sewer lines were 1.2±2.0 mg COD/L (p>0.05) (Fig. 3f) and 0.01±0.1 gN/L (p>0.05) (Fig. S3-c), respectively. This implies addition of WTP Fe-sludge did not cause an increase in greenhouse gas, GHG formation in sewers. In addition, the total concentrations of other metals (As, Cd, Co, Cr, Ni, Pb, Se, Zn) except Mn (0.2 mg/L) in the both sewer lines were comparable, i.e. <0.01 mg/L. This confirms that dosing of Fe-sludge in the sewer did not increase the loading of other metals to the receiving WWTP.

Fig. 2. Changes in sewage characteristics in control [C] and experimental [Exp] sewer pipes after dosing Fe-sludge. The vertical dashed line represents the time at which Fe-sludge was dosed to the experimental pipe. Concentration-time profiles are presented for (a) TSS, (b) VSS, (c) Fe(T), (d) Fe(sol.), (e) Al(T), and (f) Al(sol.). Each data point corresponds to the mean value of

16

three sampling points (15 m, 105 m, 210 m) in sewer pipes. Error bars represent standard error of mean (SEM).

Fig. 3. Changes in sewage characteristics in control [C] and experimental [Exp] sewer pipes after dosing Fe-sludge. The vertical dashed lines represent the sewage characteristics observed at 20 min and 40 min in both sewer pipes. Concentration-time profiles are presented for (a) total dissolved sulfide S(-II), (b) sulfite, sulfate, and thiosulfate, (c) PO4-P, (d) tCOD, (e) sCOD, and (f) dissolved CH4. Each data point corresponds to the mean value of three sampling points (15 m, 105 m, 210 m) in sewer pipes (except for the dissolved CH4, which correspond to sampling point 15 m). Error bars represent SEM.

3.2 Mechanism of sulfide removal in sewer when dosing Fe-sludge Fig. 4a shows the chemical composition of a sample collected after Fe-sludge dosing (see Section 2.2). XRD analysis revealed that ferric sulfate, i.e. Fe2(SO4)3, and iron oxyhydroxides, i.e. FeOOH, were the predominant mineral precipitates in the sample. This was similar to the chemical composition observed in Fe-sludge before dosing (Fig. 4a). Ferric activator 17

[Fe2(SO4)3] (Fig. 4a) can form iron-sulfide precipitates (FeSx) in the sample under reductive conditions (see Section 2.1). At the same time, the presence of ferric compounds such as FeOOH can result in sulfidization reactions between FeOOH and H2S, forming FeSx (FeS, FeS2) as in Eqs. (1)-(2) [28, 29]. 2FeOOH + 3H2S  2FeS + S0+ 4H2O

(1)

FeS + H2S  FeS2 + H2

(2)

Such possible interactions are reflected by the appearance of broader and sharper peaks in samples taken after dosing with Fe-sludge as compared with those in Fe-sludge alone (Fig. 4a). Notably, a peak corresponding to iron-sulfide (FeSx) precipitates appeared (Fig. 4a). This implies the physicochemical reaction, i.e. precipitation between sulfide and ferric ion in Fesludge is likely to be the dominant mechanism for sulfide removal (Fig. 3a), and is expected to proceed as outlined in Eq. (1). This is further supported by EDX-determined elemental composition of the Fe-sludge sample after dosing (Fig. S4). For this, both atomic percentage and atomic ratio were calculated for iron and sulfide. The atomic ratio of 0.89  0.92 (Table S1) indicates that the precipitates as FeS. As shown in Fig. 3a, the sulfide concentration decreased rapidly at the beginning, followed by a relative slower decrease. Such trend may be attributed to a decrease in reactive sites, i.e. surface FeOOH [20, 27], likely caused by continuous precipitation reactions occurring in sewer.

The infrared (IR) spectra, acquired from 400 to 4000 cm-1 for Fe-sludge samples before- and after dosing (Fig. 4b), showed the major changes occurred at 800 to 1800 cm-1. In the IR spectra of the sample obtained after Fe-sludge dosing (Fig. 4b), the band previously seen at 871 cm1

disappeared (Fig. 4b). This band represents the primary amine functional group (R-NH2,

where R is alkyl group) and carbonate group coordinated with Fe, originally present in the Fesludge sample before dosing (see Fig. 4b). This R-NH2 functional group could be responsible 18

for binding sulfides to form R-NH3-HS or R-NH3S- [30]. This is accompanied by the appearance of a sharper peak at 1025 cm-1 (Fig. 4b), which corresponds to S-O covalent bond stretching of inorganic sulfates [31]. This suggests the precipitation of dissolved sulfides with FeOOH (bound with other organic or inorganic compounds) present on the Fe-sludge surface, is likely dominant physicochemical process contributing to sulfide removal (Fig. 3a). In a nutshell, the overall chemical reaction schemes outlining the possible interactions between wastewater and Fe-sludge in a sewer are depicted in Fig. 4c.

Fig. 4. (a) X-ray diffraction pattern and (b) IR spectra of Fe-sludge samples, before and after Fe-sludge dosing into the pilot sewer, i.e. a mixture of raw sewage and Fe-sludge solids. Sample used for XRD analysis was collected at the end of each test; (c) diagrammatic representation outlining the possible chemical reactions between raw wastewater and Fe-sludge in sewer.

19

3.3 Effects of Al-sludge dosing on sewage characteristics Figs. 5a-f show the changes in suspended solids and Al, Fe concentrations with Al-sludge dosing. The mean difference in TSS concentration between experimental and control sewer lines was 509.4±23.1 mg TSS/L (Fig. 5a); while, the mean difference in VSS concentration was 139.0±2.5 mg VSS/L (Fig. 5b). Similar to Fe-sludge dosing, we observed decreasing TSS concentration in both sewer lines with time, implying that solids may be settling in the sewer pipes during recirculation. TSS concentration is likely to influence the Al(T) or Fe(T) concentrations in the sewer bulk phase, as previously observed in Figs. 2c-d. However, this was not the case as shown by the concentration-time profiles of Al (Al(T), Al(sol.),) and Fe (Fe(T), Fe(sol.)) in Figs. 5c-f. Neither Al(T) nor Fe(T) concentrations showed decreasing trends with time. The mean Al(T) concentrations in control and experimental sewer lines were 0.5±0.05 mg Al/L and 46±3.0 mg Al/L, respectively (Fig. 5c). However, mean Al(sol.) concentrations were insignificant (<0.5 mg Al/L) in both the sewer lines (Fig. 5d). Mean Fe(T) concentrations in control and experimental lines were similar at 2.0±0.2 mg Fe/L and 3.0±0.2 mg Fe/L, respectively (Fig. 5e); whereas mean Fe(sol.) concentrations were 0.4±0.06 mg Fe/L and 0.05±0.03 mg Fe/L, respectively (Fig. 5f). In addition, both pH (7.0±0.0) (Fig. S5-a) and temperature (25.0±0.1 C) (Fig. S5-b) were comparable in the two sewer lines during Al-sludge dosing.

Fig. 6a shows the changes observed in phosphate (PO4-P) concentrations in both sewer lines. Considering the dilution effect, the mean difference in PO4-P concentrations between the two sewer lines was 5.3±0.03 mg P/L (p<0.05). This represents a relative decrease in PO4-P concentration by 97.50.2% in the experimental line. Here, the dilution would cause a decrease in PO4-P concentrations by 0.6 mg P/L, which accounts for a small part of the observed difference. Here, additional analysis was undertaken to investigate the stoichiometry and the 20

phosphate removal mechanisms (Sections 3.4 - 3.5). The total PO43- removed per total Al added (as Al-sludge) was 0.340.02 mgP/mgAl. This measured ratio is below the theoretical Al:PO43ratio of 0.87:1. This is likely because Al was added in excess, as evidenced by the almost complete removal of PO43-. Such PO43- removal in sewers upon using waterworks sludge may impact the P recovery at the receiving WWTP. This requires the understanding of Al and P interactions and characterization of Al-P mineral precipitates in WWTP, which would facilitate in P recovery using appropriate separation techniques. In this context, Prot et al. [32] reported P can be recovered using magnetic separation process from digested sludge, rich in Fe-P mineral precipitates (e.g. vivianite). However, there has been no report that Al-bound P can be recovered. Many studies showed the efficacy of waterworks Al-sludge in phosphorus removal in wastewater treatment, as reviewed by Babatunde and Zhao [7], also the wide variety of end uses or disposal practices for waterworks Al-sludge had been reported [9]. In Australia, disposal to landfill is the most common adopted practice for much of the waterworks Al-sludge [9]. However, there is little existing information about the potential secondary benefits of Alsludge addition to sewers, in particular phosphate removal.

Figs. 6b-c depicts the changes observed in sewage dissolved sulfur species concentrations following Al-sludge dosing. Considering the dilution effect, the mean difference in sulfide concentration between the two sewer lines was 1.8±0.4 mg S/L (p<0.05) (Fig. 6b), equivalent to 7.3% of sulfide in raw sewage (Table 1). This equates to a small decrease in sulfide concentration (6.31.4%) in the experimental line. Here, the dilution would cause a decrease in sulfide concentrations by 2.8 mg P/L, which implies that dilution was the cause of the observed mean difference. Sulfide removal by precipitation with Fe or Al is unlikely, as the Al-sludge addition only increased total Fe concentration by 1.0 mg Fe/L (Fig. 5e), and Al is not known to precipitate with sulfide. Studies had reported that humic substances or organic 21

matter in sludge could mediate sulfide removal to some extent via oxidation, owing to high redox potential [21, 33]. Other possible in-sewer processes contributing to sulfide removal include the formation of organic sulfur compounds and sulfide adsorption onto sludge [30, 33]. Fig. 6c shows a relatively higher sulfate concentration in the experimental line. Considering the dilution effect, the mean difference in sulfate concentrations between the two sewer lines was 4.7±0.8 mg S/L (p<0.05). This difference can be primarily attributed to a relatively higher sulfate concentration in DAF subnatant as compared with raw wastewater (Table 1), as evidenced by a sharp increase in sulfate in experimental line at the beginning of the experiment. Theoretically, the dilution would cause an increase in sulfate concentration by 4.3 mg S/L. Sulfate was consumed in both the sewer lines, with concomitant sulfide production (Fig. 6b). Likewise, the mean difference in sulfite concentration between the two sewer lines was comparable (0.10.01 mg S/L) (p>0.05), while the mean difference in thiosulfate concentration was 0.7±0.1 mg S/L (p>0.05).

Figs. 6d-e show the variations in the COD concentrations (tCOD, sCOD) between the two sewer lines. The mean difference in tCOD between the two sewer lines was statistically significant (649.0±83.0 mg COD/L, p<0.05). Increased tCOD in the experimental sewer line than control line is likely due to release of organic matter from the Al-sludge. In contrast, Guan et al. [18] showed in sewage jar-tests that the tCOD concentration decreased by 15.0% with Al-sludge dosing at 18-20 mg Al/L. Without considering the dilution effect, the mean difference in sCOD concentration between the sewer lines was 33.9±6.9 mg COD/L (p<0.05) (Fig. 6e), equivalent to 12.7% sCOD in raw sewage (Table 1). This decrease in sCOD concentration is perhaps primarily caused by dilution as DAF subnatant does not contain much sCOD (Table 1). This is reflected by the sharp decrease in concentration observed in the first 20 min. Here, only the dilution would cause a decrease in sCOD concentration by 31.0 mg 22

COD/L, which accounts for the majority of the observed mean difference. In general, reduced soluble COD could impact the nitrogen removal process in receiving WWTP, especially when the WWTP incorporates a primary clarifier where suspended solids settle. Despite the increased tCOD concentrations in the experimental sewer line, there were no significant changes in respective dissolved CH4 (Fig. 6f) and N2O (Fig. S5-c) concentrations compared to the control line. The mean difference in dissolved CH4 between the two sewer lines was 0.35±2.7 mg COD/L, while the mean difference for N2O concentrations was 0.01±0.08 g N/L. This implies that discharge of WTP Al-sludge to sewers does not cause increased GHGs formation in sewers. In addition, the respective concentrations of metals (As, Cd, Co, Cr, Ni, Se, Zn) in both sewer lines were 0.01 mg/L, except for Mn (0.2 mgMn/L), Cu (0.1 mg/L) and Pb (0.1 mg/L). This confirms that dosing of Al-sludge does not increase the loading of other metals to the receiving WWTP.

Fig. 5. Changes in sewage characteristics in control [C] and experimental [Exp] sewer pipes after dosing Al-sludge. The vertical dashed line represents the time at which Al-sludge was dosed to the experimental pipe. Concentration-time profiles are presented for (a) TSS, (b) VSS, 23

(c) Al(T), (d) Al(sol.), (e) Fe(T), and (f) Fe(sol.). Each data point corresponds to the mean value of three sampling points (15 m, 105 m 210 m) in sewer pipes. Error bars represent SEM.

Fig. 6. Changes in sewage characteristics in control [C] and experimental [Exp] sewer pipes after dosing Al-sludge. The vertical dashed lines represent the sewage characteristics observed at 20 min and 40 min in both sewer pipes (after dosing of Al-sludge in experimental line). Concentration-time profiles are presented for (a) PO4-P, (b) total dissolved sulfide S(-II), (c) sulfite, sulfate, and thiosulfate contents, (d) tCOD, (e) sCOD, and (f) dissolved CH4. Each data point corresponds to the mean value of three sampling points (15 m, 105 m, 210 m) in sewer pipes (except for the dissolved CH4, which corresponds to sampling point 15 m). Error bars represent SEM.

3.4 Effects of Al:P dosage ratio and suspended solids on phosphate removal Figs. 7a and 7b show the influence of various Al:PO43- dose ratios on phosphate removal in sewage, unfiltered and filtered, respectively. Increasing addition of Al-sludge resulted in 24

increased PO43- removal in both cases. This is because Al-sludge predominantly consists of amorphous aluminium (hydr)oxides (as evidenced by the Al-sludge XRD pattern, Fig. S6), which provide active adsorption sites for available PO43- ions. At a higher Al-sludge dosing rate, more surface area and adsorption sites are available, resulting in higher PO43- removal (Figs. 7a-b). Akin Babatunde et al. [34] reported that the concentration of surface –OH groups increases via surface site density, with the increasing Al-sludge concentration. The possible mechanism governing PO43- removal is discussed in Section 3.5. On average, with unfiltered sewage the total PO43- removed per Al added (as Al-sludge) for Al:PO43- molar ratios of 1:1, 1.5:1, 2:1, and 3:1 was 0.30, 0.31, 0.26, and 0.28 mg P/mg Al, respectively (Fig. 7a, Table S2). For filtered sewage, the average total PO43- removed per Al added (as Al-sludge) for Al:P molar ratios of 2:1 and 3:1 was 0.34, and 0.28 mg P/mg Al, respectively (Fig. 7b, Table S2). Phosphate removal was comparable in unfiltered and filtered sewage when dosing Al-sludge of Al:PO43- molar ratio 2:1 and 3:1, respectively (Figs. 7a-b). This implies that PO43- removal in sewage was primarily due to added Al-sludge and it is unlikely that PO43- removal was due to interactions of added Al (as Al-sludge) with sewage suspended solids.

Fig. 7. Effect of various Al:PO43- dosage ratios on phosphate removal in (a) unfiltered and (b) filtered sewage. Here, each data point corresponds to mean value of duplicate tests (n=2)

25

3.5 Mechanisms of phosphate removal in sewers when dosing Al-sludge We conducted a separate set of bench-scale experiments (Supplementary Information, SI-2) to understand the characteristics and mechanisms of PO43- removal when dosing Al-sludge as observed in Fig. 6a and Fig. 7a. This experiment comprised two phases – first, the hydrolysis of added Al-sludge, and second, the subsequent adsorption test for PO43- removal. The results showed that the dominant pathway consisted of ligand-exchange resulting in formation of Alphosphate complexes (Fig. S7), which also suggests that surface-precipitation was contributing to PO43- removal. This finding prompted the further spectroscopy analyses to identify the Alcoordination number. For these analyses, bench-scale experiments were undertaken with sewage sampled before- and after dosing Al-sludge (see Section 3.4).

Fig. 8a shows the IR spectra of sewage samples, before and after Al-sludge (as Al:PO43- = 3:1) dosing. As previously stated, the formation of Al-phosphate complexes with Al-sludge dosing is reflected by particular absorbance bands in mid-infrared regions from 800-1200 cm-1. The peaks in this region are assigned to metal-orthophosphate complexes [35, 36]. Absorbance bands in this region can also be assigned to asymmetric stretching vibrations of the bridging PO2- (O=P–O−) and P–O–P, or asymmetric and symmetric stretching vibrations of the PO43ions [36]. Clearly, the relative peak intensity of Al-phosphate complexes is stronger in the sample taken after Al-sludge dosing (Fig. 8a), which implies that PO43- is chemically adsorbed on the Al-sludge surface.

Further, it is crucial to understand how the PO43- is associated with Al-OH on the sludge surface, whether by forming ‘inner-sphere’ or ‘outer-sphere’ surface complexes, or both. The occurrence of outer-sphere surface complexation resulting from PO43- ions exchanging with surface -OH groups, is evidenced by the ligand-exchange mechanism described in Fig. S7.

26

Evidence of inner-sphere complexation can also be gained from IR spectra (Fig. 8a). At 3200 cm-1 the IR spectra of the sample taken before Al-sludge dosing demonstrates strong OHstretching [37]. This stretching becomes more pronounced in IR spectra of the sample taken after Al-sludge dosing. In addition, an intense peak associated with the deformation vibrations of multi-centered hydroxyl groups of aluminium oxides (Al-OH) appeared at 1017-1018 cm-1 in the sample taken after Al-sludge dosing [38]. This implies the possible formation of innersphere surface complexes between PO43- ions and aluminium oxides (Fig. S6).

Figs. 8b and 8c show the 27Al NMR and 31P NMR spectra of sewage sampled before and after Al-sludge dosing, respectively. These

27Al

and

31P

NMR spectra further corroborate the

association of PO43- ions and aluminium oxides as outlined previously. The broad resonance peak observed in two regions of 27Al spectra, i.e.  -40 ppm to  30 ppm, and 40 ppm to 60 ppm in samples after Al-sludge dosing (Fig. 8b), resulting from a chemical shift δ(27Al) around 0 ppm, -5 ppm,  -25 ppm or 50 ppm [39, 40], is due to octahedral Al-coordination with a single PO43- or four PO43- and two hydroxyl/water ligands. This is further evidenced by the chemical shift (iso, Al) ppm results from 27Al spectra deconvolution presented in Table S3. The quantitative results (%) show that most Al was in octahedral co-ordination, which corroborates the dominance of Al-phosphate complexes, as previously deduced by the hydrolysis-adsorption experiment (see SI-2) and IR spectra (Fig. 8a). Notably, the resonance peaks’ intensity increased in samples with increasing aluminium loading as compared with raw sewage (Fig. 8b) at pH 7.3-7.6.

Likewise, 31P spectra shows the broadening of resonance peaks with a negative chemical shift around 0 ppm to -50 ppm in samples after Al-sludge dosing (Fig. 8c) as compared with raw sewage. This likely indicates the chemical adsorption of PO43- ions onto aluminium oxides on 27

the Al-sludge surface with the increasing Al:PO43- dosage ratio. Adsorption of PO3- on aluminium (hydr)oxides usually yields chemical shifts in the region 0 to -20 ppm [39, 41, 42] whereas aluminium phosphates yield chemical shifts from -10 to -30 ppm [42]. The 31P spectra observed under different Al:PO43- ratio (Fig. 8c) further corroborates the existence of outersphere complexes (= 10 to 0 ppm) and inner-sphere complexes (= 0 to -10 ppm), and also suggests the possible formation of surface precipitates (= -10 to -30 ppm) such as AlPO4.2H2O, Al3(OH)3(PO4)2 [42]. The possible formation of surface precipitates is supported by the deconvolution results of 31P spectra that most P-integral (%) experienced a chemical shift (iso, P) at

-21 ppm and -27 ppm (Table S4). Lookman et al. [43] attributed the resonance line at -27

ppm to crystalline AlPO4, while resonance lines at -22.2 ppm and -27.2 ppm were attributed to anhydrous AlPO4 [44]. The formation of surface precipitates in aqueous solution containing PO43- ions, is also likely be governed by the dissolution-reprecipitation process that can occur on Al-sludge addition.

Hence, we can assume that with Al-sludge dosing (enriched with aluminium (hydr)oxides), the predominant reaction will be ligand-exchange between the sludge surface -OH active functional group and H2PO4- or HPO4- (dominant in pH range 6.64-8.39). This in turn would result in the formation of both inner- and outer-sphere surface Al-phosphate complexes. This ligand-exchange substituting the surface -OH groups with PO43- ions could in turn slightly raise the pH of the solution, and also release other anions such as sulfate (SO42-) or humics during the process. In addition, surface-precipitates can form with a longer contact time.

28

Fig. 8. (a) IR spectra of sewage samples taken before and after batch experiment, using an Alsludge dosage molar ratio of Al:PO43- = 3:1. Here, star symbols are assigned to respective absorbance bands, where major changes were observed; (b)

27Al

spectra and (c)

31P

NMR

spectra of sewage samples – before and after Al-sludge dosing, respectively, with various Al:PO43- molar ratios (1:1, 2:1, 3:1), reacted for 2 hr with constant mixing at 25C and pH 7.3 - 7.6

3.6 Implications of dosing waterworks Al- or Fe-sludge in sewers This study evaluated the impacts of discharging waterworks Fe- or Al-sludge into pilot sewers. Neither Fe- nor Al-sludge caused an increase in GHG (dissolved CH4, N2O) production in sewers. No release of other metals was observed in both cases. Dilution caused by mixing of the Fe- or Al-sludge sludge stream with raw wastewater was found to be largely responsible for the observed reduction in sCOD concentrations.

Dosing of 22.1 mg Fe/L (as Fe-sludge) lead to a decrease in dissolved sulfide, while dosing of 45.7 mg Al/L (as Al-sludge) decreased phosphate concentrations. Based on waterworks Fesludge dosing to laboratory sewers, Sun et al. [20] reported that waterworks Fe-sludge was effective in controlling dissolved sulfide concentration, and also showed that phosphate removal could be an additional benefit. In contrast, in the current study dosing of Fe-sludge was effective for dissolved sulfide removal, while Al-sludge dosing was more effective for phosphate removal. Notably, our study was conducted in pilot sewers under more realistic 29

sewer conditions. The findings of the current study are in contrast to Edwards et al. [21], which reported Fe-sludge to be more effective than Al-sludge in phosphate removal. However, those results were primarily based on bench-scale experiments under controlled conditions, rather than Fe- or Al-sludge dosing in pilot sewers as in this study. The differences in the Fe-/Alsludges sources and the resultant varying sludge compositions could be another factor behind such contrasting result. Further details highlighting the novelty of this study as compared to the previous studies conducted to treat wastewater on applying Fe- or Al-sludge are provided in Table S5.

The results from the current study further confirm the benefits of applying Fe-sludge and Alsludge to full-scale sewers as a viable end use for sludge generated from WTPs. This could have significant environmental and economic benefits for water utilities in implementing a circular economy in urban water management via sustainable WTP sludge management. Instead of waterworks sludge being considered a waste product, it can be reused as a resource for pollutants or nutrients removal in an urban wastewater system.

Globally, average daily production of WTP sludge exceeds 10,000 tonnes [7]. The large volume of sludge incurs high disposal and storage costs. For instance, the total cost of waterworks sludge disposal is worth $33.47$44.62 million per annum in the Netherlands alone [45]. Pikaar et al. [4] reported that 56% of the surveyed 77 WTPs in Australia use alum as primary coagulant, and the case is similar in the U.S.A., the U.K., Canada, and China. The cost of Al-sludge disposal in landfill in Australia is $130$200/tonne excluding sludge transport which costs around $30$40/tonne [9]. The enormous costs associated with waterworks sludge disposal and handling demands a rethink of options such as sludge reuse or alternative disposal routes. 30

Disposal of waterworks sludge into sewers for management at the receiving WWTP enables system-wide cooperation, and has been identified as an option for beneficial reuse by utilities [7, 9]. The current study provides firm evidence of the multiple benefits observed in both sewers and the receiving WWTP [21], on discharge of waterworks Fe- or Al-sludge into sewers. A few Australian water utilities have adopted sewer disposal practices for Al-sludge, to reduce the phosphorus loads to receiving WWTPs [9]. Similarly, Filho et al. [46] also reported that long-term dosing of waterworks Al-sludge in sewers caused no adverse effects on removal efficiency of organics or nutrients in the receiving WWTP. Edwards et al. [21] estimated that savings of $20,500$96,500 per annum in a 17 MGD capacity treatment plant (excluding costs for coagulant storage and dosing facilities) could be made by reusing Fesludge in sewers.

Disposal of waterworks sludge into sewers is practical when a sewer main of the receiving WWTP is located in close proximity to the WTP [9]. However, in most cases the WTP is at a distant location in relation to the sewer main. In such cases, construction of a new sewer main would be required incurring additional infrastructure costs. In addition, disposal of waterworks sludge into sewers increases the solids concentration, potentially causing sludge sedimentation in sewers. Increased solids loading to the receiving treatment plant will also result from sewer discharge of waterworks sludge. Therefore, a holistic approach, which considers factors that affect both sewer and WWTP operation, must be adopted prior to full-scale sewer disposal of waterworks Fe- or Al-sludge. These factors include assessing: (i) impacts on sewer infrastructure such as sewer corrosion or blockages due to sludge sedimentation, and changes in sewage characteristics, (ii) current sludge handling capacity of receiving WWTPs, and (iii) impacts on treatment processes of receiving WWTPs, and final WWTP effluent quality 31

including quality of WWTP biosolids [9, 20, 47]. Hence, further long-term comprehensive study incorporating such holistic approach (or coupled with life cycle assessment) is essential for fully evaluating the impacts of waterworks sludge discharge in real-life full-scale sewer networks and receiving WWTPs for system-wide operation. In addition, a comprehensive comparative technological, environmental, and economic analysis is vital in maximizing the multiple benefits from waterworks sludge dosing to sewers. Besides, dosing of waterworkssludge directly at WWTP is also an option (e.g. P removal) which would eliminate the sludge sedimentation issue, but in-sewer sulfide control (as in case of Fe-sludge dosing) could not be achieved. However, there are also transport costs associated with this strategy.

4. Conclusion This study, for the first time, comprehensively investigated the effects of dosing waterworks Fe- and Al-sludge in pilot sewers, particularly focusing on removal of sulfides and phosphate, and associated mechanisms. Changes in other sewage characteristics were examined to understand the unintended consequences of waterworks sludge dosing. The key findings of this study are: 

Waterworks Fe-sludge was effective in removing dissolved sulfides while Al-sludge was effective for phosphate removal in sewer. Fe-sludge removes sulfide at a ratio of 0.290.06 mg S/mg Fe in sewage. Al-sludge removes phosphate at ratio of 0.290.01 mg P/mg Al in sewage (based on batch test results). Dosing of Fe- or Al-sludge in sewers caused no increase in GHG (dissolved CH4, N2O) formation, nor release of metals.



The physicochemical reaction (precipitation) between sulfide and ferric ions in Fe-sludge is likely to be the dominant mechanism for sulfide removal when dosing Fe-sludge in sewer.

32



Ligand-exchange occurring between Al-sludge surface -OH groups and PO43- ions is found to be the dominating mechanism for phosphate removal when dosing Al-sludge in sewer. Besides, the formation of surface-precipitates is also partly contributing to phosphate removal.



Long-term, system-wide studies are needed to comprehensively analyze the effects of dosing Fe-/Al-sludge in real and dynamic sewers, and potential for unintended consequences at the receiving WWTP.

Declaration of competing interest 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.

Acknowledgements Authors would like to thank the Australian Research Council (ARC), and industry partners including District of Columbia Water and Sewer Authority (DC Water), Queensland Urban Utilities (QUU), South East Queensland Water (Seqwater), Public Utilities Board - Singapore's National Water Agency (PUB), and Water Research Australia Limited (WaterRA) for their support through ARC Linkage Project LP140100386. Authors would also like to extend our gratitude to AWMC staff and students who supported and contributed to the project, in particular Dr Eloise Larsen for editing this manuscript. Finally, we would also like to thank Dr Ekaterina Strounina of Centre for Advanced Imaging (UQ) for NMR spectroscopy analysis.

Supplementary Information

References 33

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Declaration of interests

☒ All authors declare that they have no known competing financial interests or personal relationships that could have appeared to influence the work reported in this paper.

☐The authors declare the following financial interests/personal relationships which may be considered as potential competing interests: All authors would like to state that there are no such financial and personal relationships with any other people or organizations that could influence/create bias to this work. Also, all authors would like to declare there are no any such conflict of interests to declare

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41

Highlights 

Waterworks Fe-sludge is effective for dissolved sulfide control in sewers



Waterworks Al-sludge is effective for phosphate removal in sewers



Precipitation is the dominant mechanism for sulfide removal when dosing Fe-sludge



Ligand-exchange is the major pathway for phosphate removal when dosing Al-sludge

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