Feasibility of alternative sewage sludge treatment methods from a lifecycle assessment (LCA) perspective

Journal Pre-proof Feasibility of alternative sewage sludge treatment methods from a lifecycle assessment (LCA) perspective Soon Kay Teoh, Loretta Y. Li PII:

S0959-6526(19)34365-3

DOI:

https://doi.org/10.1016/j.jclepro.2019.119495

Reference:

JCLP 119495

To appear in:

Journal of Cleaner Production

Received Date: 26 June 2019 Revised Date:

8 October 2019

Accepted Date: 27 November 2019

Please cite this article as: Teoh SK, Li LY, Feasibility of alternative sewage sludge treatment methods from a lifecycle assessment (LCA) perspective, Journal of Cleaner Production (2019), doi: https:// doi.org/10.1016/j.jclepro.2019.119495. This is a PDF file of an article that has undergone enhancements after acceptance, such as the addition of a cover page and metadata, and formatting for readability, but it is not yet the definitive version of record. This version will undergo additional copyediting, typesetting and review before it is published in its final form, but we are providing this version to give early visibility of the article. Please note that, during the production process, errors may be discovered which could affect the content, and all legal disclaimers that apply to the journal pertain. © 2019 Published by Elsevier Ltd.

Feasibility of Alternative Sewage Sludge Treatment Methods from a Lifecycle Assessment (LCA) Perspective a

Soon Kay Teoh , Loretta Y. Li a

b,

*

National Environment Agency, 40 Scotts Road, Singapore 228231, Singapore

b

Department of Civil Engineering, University of British Columbia, 6250 Applied Science Lane,

Vancouver, B.C., V6T 1Z4, Canada *Corresponding author, email address: [email protected]

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Abstract Sewage sludge treatment and disposal are critical global issues, with concerns including sludge volume/weight, release of pollutants, and other environmental impacts. This study develops a semi-quantitative assessment methodology for selecting appropriate sludge treatment options on the basis of a lifecycle assessment approach. Various biological, chemical, thermal, and thermochemical sludge treatment methods described in the literature are reviewed and evaluated holistically by adopting the developed methodology to determine their comparative effectiveness in reducing sludge volume/weight and environmental impacts. Anaerobic digestion, pyrolysis, and supercritical water oxidation are found to be the best-performing treatment methods. They are not only more effective in reducing sludge volume/weight and pollutants but also have lower global warming and toxicity potential compared to most of the other methods reviewed. The potential for adverse environmental effects remains owing to the release of pollutants when the products of sludge treatment are utilised, e.g. as soil amendments or fuel. This necessitates further investigation to explore the toxicity impacts of a wider array of emerging pollutants from a lifecycle perspective as well as further development of sludge treatment methods to overcome the drawbacks of existing methods.1

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Abbreviations

AETP

aquatic ecotoxicity potential

ARDP

abiotic resource depletion potential

ASR

automotive shredder residue

BDE-209

decabromodiphenyl ether

COD

chemical oxygen demand

DEHP

diethylhexyl phthalate 1

Keywords: Sewage sludge treatment, lifecycle assessment (LCA), environmental pollutants, toxicity potentials, decision-making score

DEET

N,N-Diethyl-meta-toluamide/N,N-Diethyl-3-methylbenzamide

EP

eutrophication potential

ETP

ecotoxicity potential

GWP

global warming potential

HTP

human toxicity potential

LAS

linear alkyl-benzene sulphonates

LCA

lifecycle assessment

PAH

polycyclic aromatic hydrocarbon

PBDE

polybrominated diphenyl ether

PCB

polychlorinated biphenyl

PCDD/F

polychlorinated dibenzo-p-dioxin and dibenzofuran

PFOA

perfluorooctanoic acid

PFOS

perfluorooctanesulphonate/sulphonic acid

PPCP

pharmaceuticals and personal care products

RDF

refuse-derived fuel

SSRI

selective serotonin re-uptake inhibitors

SCWO

supercritical water oxidation

SWG

supercritical water gasification

TCLP

toxicity characteristic leaching procedure

TETP

terrestrial ecotoxicity potential

TP

toxicity potential

TSS

total suspended solids

TEQ

toxic equivalent

VOC

volatile organic compound

VS

volatile solids

VSS

volatile suspended solids

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1.

Introduction

Population growth, urbanisation, industrialisation, and inadequate wastewater management practices are resulting in critical unresolved challenges in municipal wastewater management. The increased adoption of secondary and tertiary wastewater treatment in recent decades has led to the emergence of two major challenges. First, the conventional activated sludge process, which is the most widely used biological process in secondary wastewater treatment (Wei et al., 2003), produces large amounts of excess sludge. Total sludge production in China showed an average annual growth of 13% from 2007 to 2013 (Yang et al., 2015). 6.25 million tonnes of dry solids were produced in China in 2013, while 13.8 million tonnes of dry sludge per year was captured during wastewater treatment in the United States (Seiple et al., 2017). The European Commission (2008) estimated that sewage sludge production in the European Union (EU) would amount to about 12 million tonnes of dry solids per year by 2020. The treatment and disposal of sludge is expensive, accounting for up to 60% of the total cost of wastewater treatment (Horan, 1990; Di Iaconi et al., 2017). The use of landfilling for final disposal has declined owing to the lack of available land, although ash from sludge incineration is generally destined for landfills because of its high heavy metal content (Wei et al., 2003). Similarly, land application of treated sludge has become increasingly restricted owing to environmental concerns regarding legacy or recalcitrant pollutants in the sludge (Aparicio et al., 2009; Alvarenga et al., 2017). These pollutants include heavy metals and organic pollutants, such as pharmaceuticals and personal care products, hormones, pesticides (Margot et al., 2015), persistent organic pollutants (POPs) (Hamid and Li, 2016), and emerging pollutants, e.g. phthalates and phenolics (Höhne and Püttmann, 2008; Gao and Wen, 2016). In the EU, the extent of use of landfilling as a sludge disposal method declined from 15% in 2005 to 7% in 2015 and that of agricultural application

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decreased from 43% in 2005 to 28% in 2015, while that of sludge incineration increased from 21% to 38% over the same period (Gutjahr and Müller-Schaper, 2018). Second, conventional secondary wastewater treatment processes are known to play a key role in the environmental cycling of pollutants (Hamid and Li, 2016). Although conventional wastewater treatment removes substantial proportions of volatile and biodegradable pollutants, and although it is not designed or optimised for such purposes (Clara et al., 2007; Mailler et al., 2014a; Mailler et al., 2015), hydrophilic or refractory organic compounds remain in the treated wastewater at ng/L–µg/L levels (Loos et al., 2013). A more severe problem may be encountered in the case of treated sewage sludge, in which pollutants with low aqueous solubility, high hydrophobicity, and limited biodegradability persist. For example, over 60%–90% of the total polybrominated diphenyl ethers (PBDEs) detected in influent wastewater accumulates in sewage sludge in wastewater treatment plants (North, 2004; Song et al., 2006; Deng et al., 2015). Table SI-1 provides a summary of the pollutants typically contained in sludge. Therefore, there is an urgent need for treatment methods that not only reduce sludge quantities but also remove, stabilise, or reduce the pollutants found in the sludge. The choice of such methods requires holistic assessment, often involving a difficult decision-making process that weighs competing concerns, such as effectiveness and environmental impact. The importance of considering the technical and environmental sustainability of solutions has been recognised (Moe and Gangarosa, 2009), along with other factors such as cost, especially in less developed economies. In their assessment of progress towards achieving the United Nations (UN) Millennium Development Goal (MDG) of doubling access to sanitation facilities, UNICEF and the World Health Organization (2015) reported that despite some progress, large disparities in

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access to such facilities remain between the rich and the poor. Lack of access to sanitation and wastewater treatment facilities intensifies the above-mentioned challenges. The main objective of this review is to develop a semi-quantitative methodology to evaluate various existing methods for managing municipal wastewater sludge and reducing disposal impacts. In particular, this study aims to identify sludge treatment methods and assess their feasibility in terms of effectiveness and environmental impact. Feasibility is assessed in terms of four factors, namely (i) effectiveness in reducing sludge volume/weight, (ii) effectiveness in reducing, removing, or stabilising pollutants, (iii) environmental impact based on lifecycle assessment (LCA) of global warming potential (GWP), and (iv) environmental impact based on LCA of toxicity potentials (TPs), including human toxicity, terrestrial ecotoxicity, and aquatic ecotoxicity potentials (HTP, TETP, AETP). Effectiveness in reducing sludge volume/weight was chosen as one of the factors to determine feasibility because of the high cost of sludge treatment and disposal (Horan, 1990; Di Iaconi et al., 2017). Effectiveness in reducing, removing, or stabilising pollutants was also chosen owing to the risk of environmental cycling of pollutants arising from sludge disposal (Hamid and Li, 2016). We have also included environmental impact based on LCA of GWP and TPs, as LCA has been extensively used to quantify the environmental impact of sludge treatment processes (Hospido et al., 2005) and GWP and TPs have attracted considerable attention in such studies on sludge treatment and disposal. Based on a review of the data compiled from the literature, an assessment methodology is developed and used to define and evaluate effectiveness and environmental impact. In summary, this review contributes towards improving our understanding of the relative feasibility of various sludge management methods. In addition, the proposed assessment method will be a useful decision-making tool for selecting appropriate options for implementation in a particular case.

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2.

Approach and Methodology

2.1

Feasibility Assessment Approach

Fig. 1 shows a diagrammatic representation of the approach adopted in this study. Previous research was broadly reviewed by collecting papers from the Web of Science website (http://apps.webofknowledge.com), as well as the ScienceDirect and Scopus databases. Multiple keywords related to the effectiveness and environmental impact of sludge treatment methods, such as “lifecycle assessment (LCA)”, “sludge treatment”, “sludge disposal”, “sludge volume reduction”, “pollutant reduction”, “environmental impact”, “global warming potential (GWP)”, and “toxicity potential (TP)”, were used as topic queries. A full list of keywords used is provided in Table SI-2. The collected papers were examined and included in the review if considered relevant. Papers citing or cited by each relevant paper collected using the topic queries were also included if they were considered appropriate. For the purposes of this review, sludge treatment methods were considered to consist of three steps (Fig. 2): (A) pre-treatment, (B) the actual treatment process, and (C) product end-use or disposal. Each step may incorporate one or more processes; in particular, the actual treatment step may incorporate one or more biological, chemical, thermal, or thermo-chemical treatment processes. Two important technical aims of the treatment process, in addition to those that are the focus of this review, are the elimination of pathogens remaining after the secondary treatment process and the recovery of materials or energy. Tertiary biological nutrient removal processes, which mainly aim to remove or recover nitrogen and phosphorus, can be incorporated before the steps shown in Fig. 2, but they are beyond the scope of this review, even though they may offer some benefits. Other possible pre-treatments, such as ozonation before anaerobic digestion

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(Carballa et al., 2011), are also beyond the scope of this review. The benefits of material or energy recovery may be reflected in reduced environmental impact under LCA. 2.2

Methodology for Assessing Effectiveness and Environmental Impact

Effectiveness is defined as the reported percentage reduction in sludge volume/weight, or pollutant mass or concentration, over the “conventional” process with which the treatment method was compared in the original study. Environmental impact is assessed via comparison with the results of LCA conducted by the original researchers. Environmental LCA is a tool for comprehensively evaluating the environmental impacts of processes, services, or goods (collectively termed as products) throughout their lifecycles (Hospido et al., 2005), and it has been extensively used to quantify the environmental impact of sludge treatment processes. Evaluation using LCA is important for elucidating the overall impact of sludge treatment and thus facilitating an examination of tradeoffs by considering lifecycle impacts. For example, a certain sludge treatment method may be effective in reducing or stabilising certain pollutants, but it may produce other pollutants. LCA compares such relative effects across multiple treatment methods by using a common unit of measurement and is thus useful for measuring the overall impact. As the numerical magnitudes of LCA impact potentials (or category indicators) are determined to a large extent by the methodologies and assumptions adopted and by the geographical context, they are first assessed within the context of each study before comparisons are drawn across studies. Guineé et al. (2001) identified several impact categories relevant to sludge treatment, such as GWP, TPs, abiotic resource depletion potential (ARDP), and eutrophication potential (EP). In our review, we assess only GWP and TPs, because they have attracted considerable attention in

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studies on the lifecycle environmental impacts of sludge treatment, possibly reflecting a greater overall concern with regard to these two factors. Further, the effectiveness of a sludge treatment method in reducing or stabilising pollutants is linked to the magnitude of its TPs. Although ARDP has been widely discussed in the scientific literature, it may be correlated with GWP if non-renewable energy use results in greenhouse gas emissions that contribute significantly to the overall GWP. Other potentials have not been discussed extensively. It should be noted that the studies reviewed in this paper assess TPs from a lifecycle perspective. TPs, measured in units of reference chemicals, are calculated indices based on both the inherent toxicity of substances and their potential doses, and are used to weight emissions inventoried as part of an LCA (Hertwich et al., 2001). The potential dose of a chemical can be calculated using a generic fate and exposure model, which determines its distribution in a model environment and accounts for different exposure routes, such as inhalation, ingestion, and dermal contact with water and soil (Hertwich et al., 2001). The LCA-based TPs discussed in this review are thus defined differently from toxicities derived through toxicological risk assessments, which are based on hazard quotients and cancer risks (Volosin and Cardwell, 2002). Most original studies presented their LCA findings of impact potentials in graphical form, as plots of un-normalised, normalised, or weighted values. When actual numerical values were not available, the values were estimated as accurately as possible from these plots. Owing to the difficulty in making meaningful comparisons across numerous normalisation or weighting methods, only un-normalised values were assessed and compared. As most studies investigated mid-point impacts (rather than end-point impacts), these impacts were considered. The majority of the original researchers selected one tonne of dry solids (t-DS) as the functional unit (FU). Un-

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normalised values are presented in the units of measurement used in the original studies; if possible, the units are converted into kg-CO2e/t-DS for GWP and kg-1,4-DCB-eq/t-DS for TPs. 2.3

Five-Point Scale Scoring System

This study developed a scoring system based on a semi-quantitative Likert-type five-point scale (Cartmell et al., 2006) to assign a relative numerical score to each of the four feasibility factors assessed. The scoring system enables decision-makers to holistically consider multiple feasibility concerns of the methods in a relatively simple manner. Methods that perform well are assigned positive scores, whereas those that perform poorly are assigned negative scores. The methods are scored against each feasibility factor on a scale of -2 (very negative) through 0 (neutral, or balance of negative and positive) to +2 (very positive). As the magnitudes of volume/weight reduction, pollutant reduction, and LCA impact potentials often cannot be meaningfully compared across studies, scoring is based on comparisons between different sludge treatment methods within each reviewed study. For the LCA impact potentials, it should be noted that scores are given on the basis of the magnitude of the net rather than the gross impact potentials; for example, a better score is given to a treatment method with compensating factors that reduce an otherwise high GWP or toxicity to a negative impact potential, than to another treatment method with low GWP or toxicity without any compensating factors. Semi-quantitative numerical scores for the two effectiveness factors and for GWP and TPs are given according to the matrix shown in Table 1. The semi-quantitative scores are then converted into a single overall score for each treatment method by considering the four feasibility factors as a whole. A single overall score further aids decision-makers in feasibility appraisal. The maximum potential overall score for any treatment

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method is +2; conversely, the minimum is -2. Decision-makers can use the same scoring system and tailor it to their needs by assigning their own weights according to what they consider important in their respective contexts. It should be noted that sensitivity analysis is beyond the scope of this paper, as only a few of the studies reviewed have included a thorough discussion on the sensitivity of their findings. Previous research has focused on the assessment of one or more feasibility concerns of a single sludge treatment method or combinations of sludge treatment methods in various geographical contexts, as well as on the basis of various modifications spurred by environmental regulatory requirements. In this paper, we assess and review the findings of 67 studies published between 2000 and 2018, identified through a literature search. Tables SI-3 and SI-4 summarise these studies, while Table SI-5 summarises the functional units, lifecycle stages, and system boundaries considered in each study. It should be noted that the sludge treatment methods in Tables SI-3 and SI-4 include only methods within the boundaries of the systems assessed by the studies, whereas the process steps that occur outside the system boundaries are excluded from the table. Evaluating and comparing the findings facilitates elucidation of the trends in the effectiveness and environmental impacts of the sludge treatment methods, as observed and analysed across studies. The following sections discuss these trends as well as the relative feasibility of the various sludge treatment methods.

3.

Results and Discussion

3.1

Effectiveness in Reducing Sludge Volume/Weight

Among biological treatment methods, anaerobic digestion is known to perform relatively well, resulting in substantial destruction of volatile solids (VS) and decrease in sludge dry weight. The

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typical range of values for VS destruction in mesophilic anaerobic digestion was reported to be 40%–50% (European Commission, 2001b). In a study to investigate the effect of adding crude glycerol on sludge digestion efficiency, it was found that, even without such addition, the sewage sludge dry weight decreased by around 20% within 6 days of reaction, while the VS destruction was around 11% (Kurahashi et al., 2017). Aerobic digestion and composting also achieve volume/weight reduction through moisture removal and partial conversion into gaseous products and heat. Salsabil et al. (2010) found that anaerobic digestion slightly out-performed aerobic digestion in terms of the total suspended solids (TSS) removal yield for all scenarios investigated (i.e. with or without pre-treatment). The TSS removal yield for aerobic digestion was 57%–76% while that for anaerobic digestion was 66%–86%. Pre-treatment steps before biological treatment, such as ultrasound treatment and ozonation, promote solubilisation and lysis, thereby enhancing VS reduction (Salsabil et al., 2010); however, these technologies are beyond the scope of this review. Co-digestion with high-organic-content waste can improve the activity of micro-organisms owing to the higher volatile-to-total solids ratio (Kurahashi et al., 2017), leading to greater VS destruction. In an investigation of the co-digestion of dewatered sewage sludge and food waste at various mixing ratios and solid retention times (SRT), Dai et al. (2013) found that an increase in the food waste ratio resulted in greater VS reduction. For example, at an SRT of 20 days, the VS reduction was 32.1 ± 1.1% for 100% dewatered sludge, but when food waste was added and mixed in at a sludge-to-food waste ratio of 2.4:1, the VS reduction improved to 45.5 ± 1.0%. At even higher food waste percentages (e.g. sludge-to-food waste ratio of 0.9:1), the VS reduction improved further (to 58.1 ± 0.8%).

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The main functions of lime stabilisation are the reduction of the microbial content of sludge and reduction of heavy metal availability (Wong and Selvam, 2006). Lime stabilisation is included in this review owing to its ability to stabilise heavy metal leaching. The addition of lime does not lead to volume/weight reduction; instead, it increases the overall volume/weight of the sludge (discounting moisture loss due to dewatering or drying). In terms of the typical amount of lime added to sewage sludge, the European Lime Association recommends addition of 50%–90% CaO per unit dry solids for 75 min to treat sludge at >55 °C and pH >12, or the addition of 20%– 40% CaO or equivalent Ca(OH)2 per unit dry solids for 3 months. A correspondingly large increase in weight can thus be expected. Furthermore, any high-pH leachate after landfill disposal can produce an adverse environmental impact, necessitating additional control steps to bring the pH within environmental regulation limits (e.g. B.C. Reg. 63/88 O.C. 268/88). In general, thermo-chemical treatment methods are among the most effective methods for reducing sewage sludge volume/weight, particularly for high-temperature treatment, such as incineration, pyrolysis, and gasification processes. Incineration was reported to reduce the volume of the sludge cake by up to 96% to stabilised ash (Vesilind and Ramsey, 1996). Pyrolysis, which is the process of thermal degradation in an inert atmosphere generally occurring at a temperature range of 300–900 °C (or even higher), reduced the volume of sewage sludge at 5.2 wt.% moisture by around 40%–50% to carbonaceous residues (Inguanzo et al., 2002). Hwang et al. (2007) found that the reductions of sewage sludge weight by pyrolysis and incineration were very similar, i.e. 63% and 62%, respectively. The principal stages of gasification, which converts carbonaceous content into combustible gas and ash in a net chemically reducing atmosphere, include drying, pyrolysis, oxidation, and reduction. Thus, while reliable estimates of volume

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reduction by gasification have not been uncovered in the literature review, they are expected to be similar to those achieved by pyrolysis. Hydrothermal carbonisation (~180–250 °C) and hydrothermal liquefaction (~250–400 °C) processes, which aim to recover solid carbonaceous fuel (i.e. hydrochar) and liquid bio-oil, respectively, appear to reduce sludge weight to a somewhat lower extent than incineration, pyrolysis, and gasification. Hydrothermal carbonisation was reported to recover around 60% of the input solid mass in the form of hydrochar (He et al., 2013). With the addition of various organic and inorganic additives at 10 wt.% to sewage sludge at a moisture content of 85 wt.%, the quantity of solid residues from hydrothermal liquefaction was around 12.8–22.6 wt.% of the total product weight (Qian et al., 2017). Sub/supercritical water gasification (≤400 °C and >400 °C respectively) appears to generate a lower average proportion of solid residue compared to that generated by hydrothermal carbonisation and hydrothermal liquefaction, but this output seems to vary considerably. Li et al. (2012) reported that the amount of solid residue obtained in sub/supercritical water gasification from completely dewatered sewage sludge (original moisture content estimated to be >80%) was around 68%–69% of the sludge dry weight (Li et al., 2012), while Zhang et al. (2010) reported solid residues of <30% of the original sludge dry weight by supercritical water gasification. Therefore, it should be noted that the amount of solid residue recovered from supercritical water gasification depends significantly on factors such as operating temperature and the physical and chemical characteristics of the sludge (Zhang et al., 2010). Both supercritical water oxidation (SCWO) and wet oxidation are expected to leave primarily inorganic residues, as the processes are reported to effectively oxidise organic matter primarily to carbon dioxide, water, and nitrogen (Svanström et al., 2004; Svanström et al., 2005; Houillon and Jolliet, 2005). The combined processes of supercritical water gasification and

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SCWO were reported to reduce the weight of sludge solids to 3.5% of the initial weight (Qian et al., 2015). While numerical values for sludge volume/weight reduction for other thermal and thermochemical methods have not been uncovered in the literature review, some inferences can be drawn with regard to the extent of volume/weight reduction. In sludge melting, sludge is heated to 1200–1500 °C; at such temperatures, organic matter is burnt and the remaining inorganic matter becomes a liquid, which solidifies into a glass-like slag upon cooling (Smith, 1992). As combustion temperatures are higher than incineration temperatures (leading to more complete combustion) and the slag is expected to be of higher density than incinerator ash, a greater volume reduction is attained than that in the case of incineration (Smith, 1992). Drying reduces the volume/weight contributed by the sludge moisture content. Reductions by as much as >85 wt.% dry solids may be required for certain applications, such as pre-treatment for pyrolysis or gasification (Spinosa at al., 2011) and land spreading (Lowe, 1995). However, because chemical conversion of solids to liquid or gaseous products does not occur to a significant extent, the extent of volume/weight reduction is expected to be less than that for thermo-chemical treatment methods. Table 2 summarises the effectiveness of sludge treatment methods in reducing sludge volume/weight, as well as in reducing, removing, or stabilising pollutants, and it suggests scores based on their comparative effectiveness. Incineration, pyrolysis, and gasification, which involve the complete removal of moisture and partial conversion of solids into gaseous and/or liquid products, lead to the greatest volume/weight reduction and have the highest score (+2). A score of +2 is given for the combined process of SCWO and wet oxidation owing to the significant volume/weight reduction reported (Qian et al., 2015). These methods are followed closely by

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hydrothermal carbonisation, liquefaction, and sub/supercritical water gasification, which are given the same score of +2. Digestion processes result in substantial destruction of the VS/TSS and decrease in the sludge dry weight, but they are given a slightly lower score (+1), as they do not perform as well. A score of -2 is given to lime stabilisation owing to the large resultant increase in weight. 3.2

Effectiveness in Reducing, Removing, or Stabilising Pollutants

Anaerobic digestion has been found to effectively degrade some pharmaceuticals and reduce the level of polychlorinated biphenyls (PCBs). However, there is some uncertainty regarding the fate of other organic pollutants, and anaerobic digestion is unable to biodegrade pollutants containing heavy metals. In particular, anaerobic digestion (both mesophilic and thermophilic) was found to be effective in reducing a range of pharmaceutical organics (including selective serotonin reuptake inhibitors (SSRIs) and oestrogens/endocrine disruptors), with an average reduction of around 30% (Malmborg and Magnér, 2015). Rosínska and Dąbrowska (2014) found that anaerobic digestion could effectively biodegrade both highly and less brominated PCB congeners in digested sludge products. The test sludge mixture was enriched with PCB congeners 28, 52, 101, 118, 138, 153, and 180 to initial concentrations of 151.0–375.3 µg kg-1 dry matter. After 21 days, the PCB congener concentrations decreased to 45.3–48.5 µg kg-1 dry matter (with concentrations in the control staying nearly the same). However, it should be noted that 21 days were required for this decrease; after 7 days of digestion, the PCB congener concentrations remained the same (150.2–374.1 µg kg-1 dry matter), while only partial degradation was attained after 14 days (to 67.9–231.8 µg kg-1 dry matter). This may pose problems if sufficient time is not allowed for less brominated congeners to degrade, as some of them (e.g. PCB-77, PCB-126, and PCB-169) are toxic (Ahlborg et al., 1994).

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On the other hand, the ability of anaerobic digestion to reduce other pollutants is less certain. Mailler et al. (2014b) investigated the fate of a wide range of pollutants in anaerobically digested sludge, including organotins, pesticides and herbicides, benzene-based products, volatile organic carbon compounds (VOCs), phenolics, diethylhexyl phthalate (DEHP), PBDEs, polycyclic aromatic hydrocarbons (PAHs), PCBs, and heavy metals. As heavy metals are not biodegradable or volatile, there was an increase in the heavy metal concentration in the final solid product compared to the input sludge, owing to the removal of the liquid matrix. The study reported that the biodegradation of most organotins was to the same extent as that of dry matter, indicated by little or no change in the pollutant concentration before and after sludge digestion. It was found that most alkylphenols, DEHP, and BDE-209 were removed to a greater extent than dry matter. Up to 42% of the original dry matter was removed, whereas up to 40%–95% of nonylphenols, nonylphenol monoethoxylate, nonylphenol diethoxylate, octylphenol, DEHP, and BDE-209 were removed, leading to lower concentrations in the final solid product; however, the authors noted that BDE-209 may be biodegraded to less brominated congeners. Subsequently, Mailler et al. (2017) investigated the fate of pharmaceuticals, hormones, perfluorinated acids, linear alkylbenzene sulphonate, alkylphenols, phthalates, PAHs, PCBs, and other pollutants after anaerobic digestion. They found that the concentrations of some pharmaceuticals (e.g. azithromycin, domperidone, lidocaine, sulphamethoxazole, tramadol) decreased from 40–130 µg kg-1 dry matter to as low as undetectable levels after digestion, while those of perfluorinated acids (PFOA, PFOS) decreased as well (from 316 µg kg-1 dry matter to 49 µg kg-1 dry matter for PFOS). However, the concentrations of some other pharmaceuticals, most hormones, linear alkylbenzene sulphonate, alkylphenols, PAHs, DEHP, and PCBs were found to increase after digestion. For example, the concentration of DEHP increased from 41,500

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µg kg-1 dry matter in raw sludge to 58,100 µg kg-1 dry matter in digested sludge, while those of nonylphenols, nonylphenol monoethoxylate, and diethoxylate increased from 940–1,720 µg kg-1 dry matter to 1,300–4,520 µg kg-1 dry matter. The authors attributed this increase in concentration to the greater removal of both dry matter and moisture (i.e. a decrease in the mass of the substrate). Considering the findings of the 2014 and 2017 studies together, the effect of anaerobic digestion on the concentrations of alkylphenols and DEHP seems to be somewhat uncertain. Poulsen and Bester (2010) reported that composting under thermophilic conditions could reduce the concentrations of some organic pollutants found in sewage sludge. They investigated 12 pollutants, including soaps and detergents, plasticisers (including DEHP), flame retardants, and other chemicals. The concentrations and masses of all 12 pollutants decreased during composting (seven of which were statistically significant), with mass reductions of 13%–89%. For example, the mass of DEHP (initial concentration of 31,000 ng/g dry matter) was reported to decrease by 84% over 24 days. However, the authors noted that the final concentration of DEHP was still significantly higher than the EU environmental standards. By comparison, constructed sludge treatment wetlands are generally less effective in reducing, removing, or stabilising pollutants. Uggetti et al. (2011, 2012) found that constructed wetlands did not reduce heavy metal concentrations. Plant uptake was shown to remove some pharmaceuticals and personal care products, such as ibuprofen and caffeine (Zhu and Chen, 2014), with reported removal efficiencies of >80%, whereas other pharmaceuticals were partially removed (e.g. DEET by 32.3%–78.4%, sulphamethoxazole by 33.6%–41.6%) or not removed significantly (e.g. carbamazepine and diclofenac sodium salt by <30%) (Zhu and Chen, 2014).

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Lime stabilisation was found to reduce the leaching of heavy metals from dewatered sewage sludge. The concentrations of Ni, Cu, and Zn in the leachate decreased from 0.55 mg/L, 2.42 mg/L, and 1.09 mg/L to 0.13 mg/L, 1.54 mg/L, and 0.01 mg/L, respectively, when the lime dosage was 10%, and to 0.09 mg/L, 1.04 mg/L, and 0.01 mg/L, respectively, when the dosage was increased to 20% (Liu et al., 2012). This represents reductions of 84%, 57%, and >99% for Ni, Cu, and Zn, respectively, at 20% dosage. Wong and Selvam (2006) found that composting sewage sludge mixed with sawdust and amended with lime at 0.63% dry wt. for 100 days reduced Cu, Mn, and Ni from 176, 141, and 64.0 mg kg-1 dry wt. to 166, 130, and 59.4 mg kg-1 dry wt., respectively, representing a reduction of around 6%–7%. They did not find reductions for Pb and Zn. When the lime dosage was increased to 1.63% dry wt., Cu, Mn, Ni, Pb, and Zn decreased by 6%–23%. Previously, Wong and Fang (2000) recommended that the lime dosage should be kept below 1% dry wt. if land application was intended as the end-use, as higher dosages were likely to inhibit microbial activity during composting if the pH was high. Further, liming was reported to be inappropriate for Cr- and Mo-polluted soils because of the high mobility of these metals in a neutral and weakly alkaline environment (Koptsik, 2014). Thus, liming as a method to stabilise heavy metals in sludge is not expected to be highly effective at dosages suitable for subsequent land application or universally for all heavy metals that may be present. During incineration, organic pollutants may undergo thermochemical conversion or destruction and be released primarily in stack gases. Several studies have shown that, in general, incineration does not necessarily lead to the net destruction of organic pollutants. Dai et al. (2014) found that the incineration of wet sewage sludge at different temperatures (700–950 °C) produced 2 to 13 times as much toxic polychlorinated dibenzo-p-dioxins and dibenzofurans (PCDD/Fs) in gaseous

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emissions as the amount originally present in the untreated sludge. When sewage sludge was coincinerated with various other waste types, such as automotive shredder residue (ASR) and refuse-derived fuel (RDF), it was observed that PCDD/Fs, dioxin-like PCBs, PCBs, and PAHs in the input waste were destroyed, whereas other PCDD/Fs, dioxin-like PCBs, PCBs, and PAHs were newly formed in the post-combustion zone (Van Caneghem et al., 2010; Van Caneghem et al., 2014), with the effect largely independent of the input concentrations. Overall, Van Caneghem et al. (2010) reported net destruction of dioxin-like PCBs, PCBs, and PAHs, but net formation of PCDD/Fs for a mixture of 70% RDF and 30% sludge, with input/output mass ratios of 5–14, 1,200–3,900, and 70–110 for dioxin-like PCBs, PCBs, and PAHs, respectively, and 0.03–0.1 for PCDD/Fs. For a mixture of 25% ASR, 25% RDF, and 50% sludge, net destruction of dioxin-like PCBs, PCBs, and PAHs (input/output mass ratios of 150–380, 4,900–6,900, and 1,000–8,200, respectively) was also reported, while the net change in the mass of PCDD/Fs was small (input/output mass ratio 0.95–3.35) (Van Caneghem et al., 2010). Jin et al. (2017) found that co-incineration of sludge and other waste with coal in cement kilns led to a net reduction of 70.4%–97.5% of PCBs in the flue gas compared to the input mass. Thus, the net destruction efficiency of organic pollutants by incineration appears to be high for dioxin-like PCBs, PCBs, and PAHs, but low for PCDD/Fs. Therefore, flue gas cleaning (or the use of inhibitors for dioxins (Zhan et al., 2016)) is necessary to further reduce organic pollutant emissions (Werther and Ogada, 1999). In the case of heavy metals and metalloids, which are not thermochemically destroyed during incineration, fly ash and bottom ash are the final sinks (Santos et al., 2013; Weibel et al., 2017), which complicates their recycling or disposal. Hoffman et al. (2016) found that sludge pyrolysis significantly reduces oestrogenicity by up to 95% in oestradiol equivalent at pyrolysis temperatures >400 °C. This reduction was considered

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to be due to the volatilisation of most oestrogens at such temperatures, followed either by partitioning to py-oil or py-gas, or thermal decomposition. Less than 5 wt.% of 17 toxic PCDD/Fs originally present in untreated sewage sludge was reported to survive pyrolysis at temperatures of 400–600 °C, explained by distillation and dechlorination effects (Dai et al., 2014). The potential for leaching of heavy metals from biochar (the solid residue from the pyrolysis process) produced at a pyrolysis temperature of 500 °C was small, with the leaching ratios of Cd, Cr, Pb, Zn, and Cu in the range of <0.01–0.1, defined as the ratio of the leaching concentration to the metal content (Hwang et al., 2007). Lu et al. (2016) further found that the leaching rate from biochar, defined as the ratio of the leachable heavy metal to the total content of the heavy metal, was reduced after pyrolysis compared to the input sludge. For example, at a pyrolysis temperature of 500 °C and leachate pH of 5, the leaching rate decreased from 0.43%– 88.87% to 0.06%–13.24% for a range of metals including Pb, Zn, Ni, Cd, As, Cu, and Cr (even though the retention rate, defined as the ratio of heavy metal quantities in biochar to that in sludge, exceeded 80%). On the other hand, Leng et al. (2015) found that 5%–20% of metals including Cu, Zn, Pb, Cd, Cr, Ni, V, Mn, Ba, Co, Ti, Sn, As, and Hg could be distributed in biooil, while Yuan et al. (2015) found that metals such as Zn, Ni, and Cd were at risk of exchange and leaching from bio-oil, suggesting that bio-oils produced by pyrolysis from metal-rich biomass such as sewage sludge should be pre-treated or upgraded before utilisation. There are also mixed results regarding the effectiveness of gasification in reducing, removing, or stabilising pollutants. Heavy metals mainly accumulate in the final carbonaceous residue. However, Marrero et al. (2004) demonstrated that low percentages of metals (ranging from 13.1 ± 14.1% for Cd to 61.2 ± 3.8% for As) were retained in the char from the gasifier after leaching with 50% nitric acid.

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The leaching rates of Cu, Cd, Ni, and Zn from liquefaction residues were suppressed after hydrothermal liquefaction compared with untreated sewage sludge, from 2.04%–7.31% in the untreated sludge to 0.14%–2.30% in the treated sludge (Huang et al., 2011). However, leachable Zn concentrations based on the toxicity characteristic leaching procedure (TCLP) were still above the US EPA threshold limit by 1.6–4.1 times. Leng et al. (2015) reported that heavy metals were distributed mainly into biochar, with around 5%–20% into bio-oil, when these products were obtained from sewage sludge liquefaction with ethanol or acetone; however, increasing the liquefaction temperature promotes distribution into bio-oil. Although heavy metals are distributed mainly into biochar, the significant amount of metals partitioned into bio-oil poses an environmental risk (Leng et al., 2015; Yuan et al., 2015). Using two risk assessment methods, Li et al. (2012) concluded that Cu, Zn, and Cd in solid residues obtained from supercritical water gasification of sludge pose a high risk in terms of eco-toxicity and bioavailability within soil, while the risks of Cr and Pb are minimised. With regard to organic pollutants, Xu et al. (2013) showed that PAHs were generated during supercritical water gasification and that a high reaction temperature, long reaction time, and low dry matter content favour the formation of mainly 4ring PAHs in the solid residue. They also noted that the total amount of PAHs in the solid residue met the Canadian soil quality standard for commercial use. Although our literature review did not uncover studies that investigated the destruction of organic pollutants by the wet oxidation or SCWO of municipal sewage sludge, the effectiveness of such destruction has been examined for other substrates. Catalytic wet air oxidation, which is somewhat similar to wet oxidation, has been assessed to determine its effectiveness in destroying refractory organic pollutants in industrial wastewater effluents. Wet oxidation is the aqueous oxidation of thickened sludge with oxygen at elevated temperature and pressure (e.g. 235 °C and

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40 bar) (Houillon and Jolliet, 2005). It transforms organic matter primarily into carbon dioxide and water vapour, destroying organic pollutants in the process and producing a mineral residue to be disposed of (Houillon and Jolliet, 2005). Catalytic wet air oxidation involves the use of catalysts, such as noble metals, metal oxides, and mixed oxides, to oxidise organic pollutants into biodegradable intermediates, carbon dioxide, water, and innocuous end products at elevated temperature (125–320 °C) and pressure (0.5–20 MPa) (Kim and Ihm, 2011). Using 5% CuO/95% activated carbon as a catalyst, a maximum chemical oxygen demand (COD) destruction of 89% was achieved for pulp and paper mill effluent treated by catalytic wet air oxidation at 443 K and 0.85 MPa (Garg et al., 2007). Williams and Onwudili (2006) assessed the effect of SCWO on organics in diesel fuel and waste landfill leachate. Organic species in diesel fuel spiked into the sand matrix at concentrations of 4–20 wt.% were decomposed at 96.6%–99.8%. Furthermore, a wide range of organics in waste landfill leachate were reported to be destroyed at >99.99%. Zou et al. (2013) studied the destruction of organics and the stabilisation of Cr, Cu, Pb, Zn, Ni, and Fe by SCWO of tannery sludge, which has high concentrations of organics (up to 54.2 wt.% dry matter) and chromium salts. The destruction efficiency of COD, measured as a surrogate for organic content, increased with the temperature, reaching ~95% at a process temperature of 500 °C and an oxygen-to-COD ratio of 3:1. As for heavy metals, Zou et al. (2013) suggested that these were concentrated in the solid ash residue owing to the poor solubility of inorganic compounds in supercritical water. Concentrations of Cr, for example, were found to lie in the range of 9.33–11.21 wt.% in ash compared to 4.71 wt.% dry matter basis in tannery sludge. A reduction in the leachability of Cr (from 11.41 mg L-1 to 0.12 mg L-1) was observed owing to SCWO at 400 °C at an oxidant ratio of 3:1. Similar trends were observed for Cu, Zn, and Fe.

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However, Ni experienced an increase in leachability, from 0.28 mg L-1 in raw sludge to 3.51 mg L-1 in the ash produced under the same oxidation conditions. Sludge melting has been shown to be a promising method for stabilising inorganic pollutants as well as avoiding the generation of additional organic pollutants during the process. Idris and Saed (2002) conducted leaching tests on melted ash from sludge incineration and showed that the quantities leached from the final product after melting treatment were extremely low compared to the standard limits, with the metal concentrations of As, Ba, Cd, Cr, Cu, Ni, and Pb ranging from undetectable amounts to 2.89 mg L-1 compared to the standard limits of 1.0–100.0 mg L-1. Hong et al. (2009) reported sludge melting to be advantageous over incineration, as dioxin production is reduced owing to crystallisation at high temperature. Table 2 shows the scores assigned for the comparative effectiveness of the treatment methods in reducing, removing, or stabilising pollutants. No sludge treatment method discussed in the literature is able to reduce, remove, or stabilise pollutants comprehensively. As such, the highest score awarded was +1 for methods that perform well. These include the following: (1) anaerobic digestion and composting, which effectively degrade at least some organic pollutants (but have no effect on heavy metals); (2) pyrolysis, which effectively degrades at least some organic pollutants and reduces heavy metal leaching, but some problems persist; (3) hydrothermal processes, which reduce heavy metal leaching and show limited organic pollutant destruction; and (4) wet oxidation, SCWO, and sludge melting, which show some effectiveness in destroying organics and reducing metal leachability. Methods that are moderately effective in reducing pollutant levels, such as constructed wetlands, incineration, and gasification, were assigned a score of 0.

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3.3

Global Warming Potential (GWP)

Two factors are observed to exert a clear influence on the GWP of sludge treatment methods: (1) end-use and final disposal methods of treated sludge, (2) overall or net energy/fuel consumption/substitution. Table 3 summarises the effects of these factors and suggests scores based on the methodology described in Table 1. Fig. SI-3 graphically summarises un-normalised GWPs of sludge treatment methods, as determined by the studies reviewed (unit conversion was performed as required). 3.3.1

Effects of End-use and Final Disposal

In general, the end-use and final disposal methods for biologically or chemically treated sludge differ from those of thermally or thermo-chemically treated sludge. Most scenarios for biologically or chemically treated sludge examined in studies considered land or agricultural application as the end-use, with only a small number considering landfilling or other end-uses (Table SI-6). Land or agricultural application of treated sludge affects the GWP in several ways. (1) As a soil amendment, treated sludge may reduce the GWP by offsetting the need for fertilisers and avoiding emissions related to their production, transport, spreading, and anaerobic degradation (i.e. biogeochemical emissions). (2) As a soil amendment, treated sludge may also reduce the GWP by sequestering soil carbon. (3) Land application of treated sludge may increase the GWP owing to the release of methane and nitrous oxide following anaerobic degradation in the soil (Johansson et al., 2008; Brown et al., 2010). The relative magnitudes of these factors determine the overall contribution to the GWP.

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Table SI-6 summarises how the above-mentioned factors were assessed in the studies, with detailed analyses of their effects. Most studies considered only two out of the three factors (i.e. carbon sequestration was not considered). There is some debate as to whether carbon sequestration by the addition of organic matter to soil should be included in LCA studies (Peters and Rowley, 2009), partly because the management of soils may not maintain the carbon store over the time scale used for GWP assessment (100 years). Consequently, soil carbon sequestration has been discounted by some studies. Emission factors vary widely across studies, with ranges of 0.02–6.3 kg CH4/t-DS and 0.00011–1.80 kg N2O/t-DS for methane and nitrous oxide emissions owing to the degradation of treated sludge in soil, and 50–328 kg CO2/t-DS avoided owing to fertiliser offset. Fig. 3 shows the effects of the main factors influencing the GWP arising from land application of treated sludge (fertiliser offset, methane and nitrous oxide emissions due to anaerobic degradation in soil, carbon sequestration), compared with the GWP of other factors not associated with land application. The values are plotted from the studies listed in Table SI-6, which have conducted such an assessment and published values for comparison. Fig. 3 shows that fertiliser offset, methane, and nitrous oxide emissions due to anaerobic degradation in soil, and carbon sequestration can all be significant factors contributing towards the overall GWP in land application end-uses. However, because the estimates vary across studies (as seen in Fig. 3), it is difficult to draw wider inferences as to how significant these factors may be in determining the overall GWP. We examine the findings of two studies that considered all three factors (Table SI-6). In the first of these studies, Brown et al. (2010) applied a greenhouse gas calculator tool (Biosolids Emissions Assessment Model, BEAM), developed for the Canadian Council of Ministers of the

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Environment (CCME) using data from peer-reviewed literature and municipalities, to nine scenarios in Canada. The modelling approach incorporated a critical assumption that nitrous oxide emissions from treated sludge applied on land are equivalent in magnitude to those from the synthetic fertilisers replaced. Thus, in the model, if treated sludge replaced fertilisers, credits from fertiliser offset and carbon sequestration accrued without any increase in the GWP owing to additional nitrous oxide emissions. If treated sludge did not replace fertilisers, only credits from carbon sequestration accrued, which was accompanied by an increase in the GWP owing to additional nitrous oxide emissions. Fig. 3 compares these two end-use possibilities (replacement vs. no replacement of fertilisers) for the scenario in British Columbia that involved anaerobic digestion, followed by land application of the treated product. In the second study that considered all three factors (Liu et al., 2013), the authors examined the effect of composting followed by land application of the treated product (which replaced the fertiliser) in one scenario. Fig. 3 shows the GWP for this scenario. Similar to Brown et al. (2010), Liu et al. (2013) excluded greenhouse gas emissions from sludge application in soil, noting that no significant differences were found between emissions from treated sludge and those from the replaced fertilisers. The authors noted that the overall GWP could be further decreased by 45% if carbon sequestration was considered, based on findings by other authors (including Peters and Rowley (2009) and Brown et al. (2010)). Some studies listed in Table SI-3 (Houillon and Jolliet (2005), Johansson et al. (2008), Carballa et al. (2011), Mills et al. (2014), Usapein and Chavalparit (2017)) did not adopt the same reasoning as Brown et al. (2010) and Liu et al. (2013) in assuming that nitrous oxide emissions from treated sludge applied on land are equivalent in magnitude to those from the synthetic fertilisers replaced. This deviation from the assumption made by Brown et al. (2010) and Liu et

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al. (2013) may significantly affect the GWP estimates, as can be seen in Fig. 3 for Johansson et al. (2008). Johansson et al. (2008) emphasised the benefits of SCWO in eliminating biogeochemical emissions. In this scenario, fertiliser avoidance accounted for nearly 10% of all GWP contributions, and together with the use of biogas in district heating, reduced the overall GWP to a negative value. Johansson et al. (2008) also noted considerable uncertainty in the magnitude of biogeochemical emissions from the land application of sludge treated by anaerobic digestion, as can be seen from the considerable differences between high and low estimates. Two studies listed in Table SI-6 (Peters and Rowley (2009), Mills et al. (2014)) were not plotted in Fig. 3, as their published findings did not provide related numerical details. The GWP of the application of sludge on land was not considered at all in two studies. Suh and Rousseaux (2002) investigated scenarios in France involving the anaerobic digestion, composting, lime stabilisation, or incineration of sludge, followed by either land application or landfilling. The effect of sludge degradation into landfill gas was considered, but the effect of degradation after land application was not considered. Xu et al. (2014) studied, within their system boundary, various anaerobic digestion configurations in China as well as their end-uses, including agricultural application, incineration, and landfilling. For the agricultural application end-use, the authors considered the toxicity of leaching of metals into the soil but not the GWP owing to anaerobic degradation of the sludge in soil. Some discrepancies in the methodologies adopted by the above-mentioned studies in determining the GWP are also apparent. The considered range of emissions due to land spreading differs across studies. Johansson et al. (2008) considered both methane and nitrous oxide emissions, while Houillon and Jolliet (2005), Brown et al. (2010), and Carballa et al. (2011) did not, with Houillon and Jolliet (2005) citing “a lack of available data”. This may not be

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significant if the same amounts of avoided emissions from fertiliser offset are netted off (Brown et al., 2010; Liu et al., 2013), but some uncertainty still remains as to whether artificial fertilisers and biosolids produce similar amounts of methane and nitrous oxide emissions, particularly with varying rates of biosolid application (Chiaradia et al., 2009). The choice of the system boundary can also lead to differences in the apparent significance of the factors associated with land spreading. Johansson et al. (2008) considered only processes after anaerobic digestion (e.g. transport, machine loading, storage, and spreading of treated sludge in one of their scenarios). Thus, potentially significant emissions from anaerobic digestion were not factored into their overall GWP comparison. On the other hand, Houillon and Jolliet (2005) included the liming process in their study, which produced a significant GWP owing to drying and liming, leading to a smaller percentage impact from fertiliser offset and biogeochemical emissions. Thus, the differences in the choice of system boundaries explain why findings cannot easily be compared across studies. In order to make sense of the findings, it is useful to first assess trends within a study examining multiple treatment options before further comparing them with trends in other studies. End-uses or final disposal methods for thermally/thermo-chemically treated sludge differ from those of chemically or biologically treated sludge. Of the studies reviewed, only drying/pasteurisation, SCWO, and wet oxidation processes led to land/agricultural application. This is not unusual, as drying/pasteurisation has been used primarily to remove pathogens from sludge, so that the product may be suitable for land/agricultural application (European Commission, 2001b). SCWO and wet oxidation produce inert outputs that can be safely spread on land without adverse effects, and no GWP was reported to be associated with land spreading (Svanström et al., 2004; Johansson et al., 2008). Residues from sludge incineration or co-

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incineration are usually landfilled. The studies surveyed in this review did not report GWP associated with such landfilling. Pyrolyzed/gasified sludge is incorporated into fuel products, leading to eventual fuel/energy substitution. Similarly, the contribution to GWP due to fuel use for drying and incineration dominated the study by Brown et al. (2010). The effect of fuel/energy consumption/substitution on GWP is explored in the following section. 3.3.2

Net Energy/Fuel Consumption/Substitution

Energy consumption by processes such as material transport, electricity use, or fuel use in sludge drying or combustion usually results in an increase in the GWP (e.g. if the energy replaced is non-renewable). Conversely, substituting energy generated from sludge treatment for use in the treatment process or elsewhere, or substituting sludge or sludge treatment products as fuel to offset the use of other fuels, usually results in a decrease in the GWP (e.g. if the fuel replaced is non-renewable). Fig. 4 compares the GWP with the net energy consumption for five studies that conducted such an analysis. For the same sludge treatment method (e.g. anaerobic digestion), Fig. 4 shows that the GWP can be either positive or negative depending on the choice of LCA methods, system boundaries, inventory analysis, and other parameters. Therefore, comparison of relative magnitudes is only meaningful within studies. Fig. 4 shows a clear trend of the GWP increasing with the net energy consumption across the studies. For example, in the study by Peters and Lundie (2001), lime stabilisation at the North Head plant in Sydney, Australia, resulted in a lower GWP compared to anaerobic digestion at the Bondi plant, but a higher GWP compared to anaerobic digestion at the Malabar plant. This was due to the use of biogas for power generation at Malabar (resulting in a negative GWP of around -200 kg CO2e/t-DS), but not at Bondi, where it was flared. Anaerobic digestion and drying as

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alternative process for North Head resulted in a lower GWP compared to lime stabilisation (by 45%) if the biogas from the digesters was used to power the dryers, but a higher GWP (by 10%) if natural gas was used instead. Thus, biogas generation and use are important in lowering the GWP of anaerobic digestion. In a study by Poulsen and Hansen (2003), alternative scenarios incorporating co-incineration in the Aalborg municipality, Denmark, were found to have the lowest GWP owing to the greatest substitution of energy and resources. In both pyrolysis scenarios examined by Cao and Pawłowski (2013), avoidance of greenhouse gas emissions due to bioenergy production (bio-oil in both scenarios; biogas in the scenario with anaerobic digestion), together with biochar substitution of the fertiliser, resulted in a net GWP offset. By assessing the findings of these studies as a whole, we can see that energy or fuel substitution plays an important role in reducing the GWP for energy-intensive thermal and thermo-chemical processes, such as drying, incineration, and pyrolysis, as well as for anaerobic digestion. For example, in the study by Cao and Pawłowski (2013), sludge pre-drying was the most energyconsumptive process, accounting for 53.1% and 81.8% of the total energy consumption for scenarios with and without anaerobic digestion, respectively. Anaerobic digestion, when incorporated, accounted for 34.8% of the total energy consumption. The pyrolysis process itself was less energy-consumptive than anaerobic digestion, accounting for 8.5% of the total energy consumption for the scenario incorporating anaerobic digestion. The contribution to the GWP also followed the same order: the drying operation was the largest GWP contributor, followed by anaerobic digestion and pyrolysis. The combined contribution of the other processes, such as dewatering and transport, accounted for <10% of the total GWP in both scenarios. The effect of transport on GWP was significant only for non-thermal/thermo-chemical processes, such as lime stabilisation and, at times, anaerobic digestion. For instance, in the study by Peters

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and Lundie (2001), transport of limed sludge by truck from North Head for a distance of 250 km was a significant contributor to both energy consumption and GWP, accounting for >50% of both total energy consumption and total GWP. By contrast, transport of dry sludge from North Head in the alternative scenario accounted for <20% of the total GWP, because of weight reduction due to a lower moisture content. 3.4

Toxicity Potentials (TPs)

LCAs of toxicity potentials arising from sludge treatment have, by and large, focused on the impacts of metallic contaminants in sludge as well as air emissions resulting from fuel consumption or combustion, or those avoided by fuel substitution. This may be a reflection of the accumulated body of scientific knowledge regarding the impact of metallic and air pollutants, as well as the focus of regulations. For instance, in the EU, the Sewage Sludge Directive 86/278/EEC (European Economic Community, 1986) sets limits for the amount of seven heavy metals. US EPA regulations (US EPA 40 CFR Part 503) also limit the presence of several heavy metals. However, the toxicity potentials of emerging pollutants have not been studied as thoroughly. This could be because the concentration levels of emerging pollutants in treated sludge have not been determined to the same extent as those of metallic pollutants, and studies have thus far focused mainly on determining the concentrations of emerging pollutants in treated sludge (e.g. Van Caneghem et al., 2010; Dai et al., 2014; Malmborg and Magnér, 2015). This has perhaps led to the difficulty in setting LCA characterisation factors as well as regulatory standards and criteria. Since the adoption of Directive 86/278/EEC, only a few member states (e.g. Austria, Denmark, and Germany) have set requirements for other contaminants (as well as implemented stricter limits for heavy metals), perhaps in response to growing scientific evidence regarding the wide range of pollutants found in sludge. Austria and Germany, for example, are

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among the few countries in Europe that have set limit values for organic compounds such as PCDDs, PCDFs, and PCBs in sludge (European Commission, 2001a), while Denmark has set limits on linear alkyl-benzene sulphonates (LAS), DEHP, nonylphenol, and PAHs. Thus far, US EPA rules and Environment Canada rules have not been promulgated for emerging pollutants such as PCDDs, PCDFs, and PCBs in untreated or treated sludge. The Canadian Council of Ministers of the Environment (CCME) notes that repeated application of untreated sludge could lead to an increase in PCDD/F concentrations in soil, owing to the presence of small amounts of these compounds in sludge and their persistence in soil, even though no guidelines have been developed thus far (CCME, 2002). Table SI-7 summarises the permissible limit values for a number of emerging pollutants adopted in some EU countries following the advent of Directive 86/278/EEC (European Commission, 2001a). With the growing awareness on emerging pollutants, the inadequacy of LCA studies in fully accounting for TPs has increasingly been acknowledged. Buonocore et al. (2018) noted that pathogens, pharmaceuticals, and personal care products (PPCPs) discharged into receiving water bodies were not characterised in LCA, leading to the underestimation of the impacts of TP. The authors also noted increasing attention towards more accurate characterisation models for toxicity-related impact categories, which should be integrated in LCA software and databases. In a review of recent scientific literature as part of this study, only two papers that assessed the impacts of the TP of emerging pollutants in sludge from a lifecycle perspective (Hospido et al., 2010; Harder et al., 2017) were identified. In the absence of LCAs, it becomes more challenging to compare the relative TPs of emerging pollutants vis-à-vis those of other pollutants. Notwithstanding the paucity of LCAs on the toxicity of emerging pollutants in sludge, this section reviews the findings of these studies, alongside the findings of Hospido et al. (2010) and

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Harder et al. (2017). Fig. 5 shows published values of TPs of different sludge treatment scenarios. Some trends were observed on reviewing the studies plotted in Fig. 5. First, in scenarios with agricultural application as the end-use or landfilling as the final disposal for treated sludge, trace metallic contaminants could be the largest contributors to both human and ecotoxicity potentials. Xu et al. (2014) compared several sludge treatment scenarios mainly involving anaerobic digestion in combination with other processes such as drying, with either landfilling or agricultural application as the end-use. The system boundaries included gravity thickening, anaerobic digestion, dewatering, drying, and end-use (landfill or agricultural application). The authors concluded that scenarios with agricultural application as the end-use had the highest contributions to terrestrial ecotoxicity because of the relatively high heavy metal emissions during the agricultural use stage. In these scenarios, virtually all (>90%) the terrestrial ecotoxicity was attributed to the end-use stage. Usapein and Chavalparit (2017) also found that using treated sludge as a fertiliser produced the greatest impact on terrestrial ecotoxicity, with heavy metals as the main contaminants, compared with other scenarios that considered landfilling as the final disposal, or cement clinker product as the end-use. On the other hand, the findings of some studies (e.g. Peters and Rowley (2009), Wang et al. (2013)) indicated that TPs due to fuel consumption or fuel substitution could dwarf the effect due to land application of treated sludge. Peters and Rowley (2009) compared the TETP and HTP for several scenarios, such as anaerobic digestion with agricultural application as end-use and drying with end-use as supplementary cement kiln fuel. The authors noted that “trace metallic contaminants delivered to agriculture contribute a minority of the indicator results [i.e., TPs] in scenarios with this end-use, whereas toxic inorganic chemicals and heavy metals emitted during the combustion of coal for electricity supply and fuel for transport are the most important

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contributors in all scenarios.” As a corollary, their findings suggested that the avoidance of coal combustion in the two scenarios that substituted treated sludge as fuel resulted in a net benefit in terms of the TPs. Wang et al. (2013) concluded that human health impacts (expressed as carcinogenic and non-carcinogenic impact potentials) for mono-incineration “were caused primarily by air emissions of particulates and nitrogen oxides, combined with the effluence of antimony and arsenic ions.” By comparison, the carbonisation scenario “had a beneficial effect on ecosystem quality because of the substitution of bio-coal for some of the coal used in the cofiring process.” It is noted that some studies did not consider pollutants remaining in treated sludge when determining the TPs. Peters and Lundie (2001) assessed the impact of air emissions, perhaps not surprisingly concluding that drying and dewatering (air emissions due to electricity consumption) and transportation (air emissions due to trucking) dominated the contributions to the HTP. Murray et al. (2008) considered the air emissions associated with burning biogas and the avoided air emissions due to fuel offset in deriving the environmental impacts. Both groups of authors did not consider the TPs of pollutants remaining in treated sludge applied on land. These two studies are not plotted in Fig. 5, as their published findings could not be easily translated to the plot. As mentioned previously, the literature review identified only two studies (Hospido et al. (2010), Harder et al., (2017)) that assessed the impacts of TP of emerging pollutants in sludge from a lifecycle perspective. The objective of the study by Hospido et al. (2010) was to investigate the impacts of emerging pollutants, such as PPCPs, in anaerobically digested sludge employed in agriculture by using an LCA methodology. The authors noted that LCA had been widely used for examining both sludge treatment and end-use options (e.g. Lundin et al. (2004), Houillon and Jolliet (2005), Cartmell et al. (2006), Murray et al. (2008), Peters and Rowley (2009)), but that

34

there were limited evaluations of the lifecycle environmental impact associated with emerging pollutants in wastewater or sludge. Further, Hospido et al. (2010) noted that four studies (Høibye et al., (2008), Muñoz et al. (2008), Wenzel et al. (2008), Muñoz et al. (2009)) had investigated the TPs due to emerging pollutants in wastewater, but either sludge treatment was not examined or incineration was assumed as the disposal pathway, and it was thus important to evaluate the environmental consequences of other sludge routes, such as land disposal. Hospido et al. (2010) determined that direct emissions of heavy metals into soil dominated HTP and TETP, accounting for more than two-thirds of the total impact in all scenarios studied. Their results also indicated that the impacts associated with the PPCPs present in sludge treated by anaerobic digestion were only 0.066%–0.079% of the total HTP and 0.008%–0.329% of the total TETP, depending on the treatment conditions, and thus by far less important than those of heavy metals. Nevertheless, it should be noted that the study found anaerobic digestion to reduce HTP and TETP by 55%–66% and 37%–99%, respectively. The objective of the study by Harder et al. (2017) was to estimate the HTP associated with metallic and organic contaminants released by the application of treated sludge from municipal wastewater treatment plants on agricultural land in Gothenburg, Sweden. The authors noted that the presence of organic contaminants in sewage sludge was generally predicted to not be of concern to either human health or the environment (VKM (2009), Diana et al. (2011), Sternbeck et al. (2013)). These three studies were all conducted under a risk assessment (RA) framework, whereby the risk was assessed by comparing the predicted intake of contaminants by certain organisms with the corresponding intake thresholds. Harder et al. (2017) noted that it was worth considering an additional aspect of the sewage sludge debate, i.e. “whether the impact on human health related to chemical contaminants, on the basis of a given human population as a whole, is

35

greater or lesser for one sewage sludge management option (e.g. application of sewage sludge to agricultural land) than for another (e.g. incineration of sewage sludge with subsequent phosphorus recovery).” Building on LCA studies using generic models such as USES-LCA (Hospido et al. (2010), Heimersson et al. (2014), Lane et al. (2015)) and USEtox (Yoshida et al. (2014)), Harder et al. (2017) found, by using the USEtox 2.0 LCA model, that the toxicity impacts were overwhelmingly due to four metals (chromium, mercury, lead, and zinc), and that they were greater than the impacts of organic pollutants by 3–4 orders of magnitude. Table 4 summarises the findings of this paper in terms of the net TPs and suggests scores based on the methodology in Table 1 for sludge treatment methods reviewed in this paper. The reasons for a certain score are described wherever they can be succinctly summarised in the table; otherwise, only the reference study is indicated. Separate scores are given for human toxicity, terrestrial ecotoxicity, and aquatic ecotoxicity potentials, because under the same conditions, a sludge treatment method may affect human toxicity and terrestrial and aquatic ecotoxicities to different degrees. The table shows that agricultural application of sludge in the scenarios studied led to a detrimental effect on the TPs, and fuel/energy substitution led to a beneficial effect on the TPs for energy-consumptive thermal and thermo-chemical methods that would otherwise have produced a detrimental impact.

4.

Overall Assessment and Limitations

For the overall assessment of various sludge treatment methods, their effectiveness and environmental impacts were holistically considered on the basis of the studies reviewed in this paper. The overall scores for each treatment method were computed by equally weighting the

36

individual scores for four feasibility factors. Methods to incorporate the relative importance of the feasibility factors and their mutual interactions, such as non-equal weighting methods, the use of utility functions, and multi-criteria decision analysis, are not considered in this review. These methods, though potentially useful in decision-making on system optimisation when considering multiple criteria or objectives (Azapagic and Clift, 1999; de Neuville, 1999), are beyond the scope of this study. Table 5 summarises the overall assessment of the sludge treatment methods by considering their effectiveness and environmental impacts together. In deriving the scores in Tables 3 to 5, we observed some trends evident from Figs. 3 to 5 and Fig. SI-1. Across all the studies reviewed, it can be seen that the GWP of sludge treatment methods is heavily dependent on the presence (or absence) of fuel/energy substitution. Thus, the GWP scores assigned for anaerobic digestion, pyrolysis, and drying ranged from the negative to the positive, as some studies considered scenarios with fuel/energy substitution while others did not. (Not enough studies examined composting, incineration, or SCWO; hence, a range of scores could not be given.) Large ranges of scores were also assigned for TPs for anaerobic digestion, composting, lime stabilisation, and drying. This was reflective of the high variability in TPs across studies for these sludge treatment methods, as can be seen from Fig. 5. The variability was due to a few factors, including the large impact of end uses and differences in HTP, TETP, and AETP for the same sludge treatment method and end-use. For example, Peters and Rowley (2009) showed that using sludge (treated by anaerobic digestion and drying) as cement kiln fuel resulted in negative HTP and TETP, whereas applying the same treated sludge to land for agriculture resulted in positive HTP and TETP. Xu et al. (2011) showed that applying anaerobically digested sludge to agricultural land would produce a significantly different HTP, TETP, and AETP profile, than landfilling the same treated sludge. The scores in Table 2 for

37

effectiveness in sludge volume/weight reduction and in pollutant reduction were given based on a direct comparison of percent reductions. Overall, the best-performing treatment methods are anaerobic digestion, pyrolysis, and SCWO, as these are comparatively more effective in reducing sludge volume/weight and pollutants, and they have lower GWP than most of the other methods studied. However, deficiencies in their ability to reduce some pollutants as well as in their environmental impact potential persist. Anaerobic digestion has been shown to substantially reduce sludge dry weight, VS, TSS, and the levels of some pollutants such as pharmaceuticals and PCBs (Table 5). The use of sludgederived biogas as an energy source substantially reduces the net GWP of the anaerobic digestion process. On the other hand, this treatment method cannot reduce, remove, or stabilise heavy metals. Thus, agricultural use of sludge treated by anaerobic digestion results in significant TPs owing to the leaching of heavy metals from treated sludge. Pyrolysis has been shown to reduce sludge dry weight and the levels of some pollutants such as PCDD/Fs, while the use of pyrolysis products (e.g., bio-oil and biochar) as fuel/material substitutes reduces GWP (Table 5). However, there is some evidence that pollutants accumulated in bio-oil may be released when these products are consumed as fuel. SCWO (combined with SWG) has been shown to be even more effective than anaerobic digestion or pyrolysis in reducing sludge volume/weight (Table 5), as well as in destroying COD and a wide range of organics in contaminated soil. However, more studies would need to be pursued in order to elucidate the effectiveness of SCWO in reducing pollutant levels in sludge as well as its overall TPs. While anaerobic digestion of sludge is widely used in the EU at the least (European Biogas Association, 2018), it should be noted that use of pyrolysis and SCWO at full commercial scale to treat sewage sludge is limited. The pyrolysis process is less developed than incineration (Viana et al., 2016) while only three

38

demonstration or commercial scale SCWO facilities have been reported in literature (Qian et al., 2016). In-depth analysis of the contributing factors is beyond the scope of this review. This study developed a methodology to determine the comparative feasibility of sewage sludge treatment methods, on the basis of a broad survey of the literature from the last two decades, in terms of their effectiveness and environmental impacts. However, the methodology is limited to a semi-quantitative one owing to the difficulty in making fully quantitative comparisons across studies. As pointed out earlier, the differences across studies in terms of their choices of LCA methods, system boundaries, inventory analysis, and other parameters influence the apparent significance of the factors affecting the impact potentials, even for the same sludge treatment method (Table SI-5). For example, it is not clear how significantly the overall GWPs vary simply because of the differences in the assumption of the magnitudes of the greenhouse gas emission factors for the land application of treated sludge (Table SI-6), or how sensitive the impact potentials may be to the differences in sludge transport options and distances assumed in modelling alternate scenarios (e.g. Hospido et al., 2005). This makes it challenging to compare environmental impacts across studies without bias. Future studies could employ the systematic design of alternate LCA scenarios in order to eliminate such biases as far as possible, or include sensitivity analyses to explore how the impact potentials may vary if the parameters are assumed differently (e.g. see Brown et al., 2010). Further LCA studies could also involve the incorporation of utility functions that account for uncertainty as well as stakeholder preferences. This would enable Likert-type scaling to be a more reliable predictor for overall feasibility. Future studies should also aim to bridge other gaps in existing knowledge. Carballa et al. (2011) noted that, to the best of their knowledge, no studies at the time had analysed the environmental consequences of pre-treatments before the anaerobic digestion of sludge, such as ozonation. The

39

influence of possible pre-treatments on sludge volume/weight reduction and on environmental impacts, not covered by the scope of this review, can be further explored. In addition, as has been shown in this review, studies in recent decades have uncovered a body of evidence on pollutant levels (especially pertaining to heavy metals) that remain in the products and byproducts of sludge treatment. More recent studies have also explored the effectiveness of sludgebased adsorbents or soil amendments (e.g. biochar produced from sludge pyrolysis) in binding and stabilising various pollutants (e.g. Bian et al., 2018; Frišták et al., 2018). Future studies can further probe the toxicity impacts of a wider array of pollutants (particularly emerging ones) arising from different sludge treatment methods from lifecycle perspectives, in order to facilitate fair comparison of the relative magnitudes across multiple factors. This will enable decisionmakers and practitioners to select appropriate treatment methods according to their geographical and contextual needs.

5. •

Conclusions The best-performing treatment methods were anaerobic digestion, pyrolysis, and SCWO, as they were comparatively more effective in reducing sludge volume/weight and pollutants and had lower GWP than most other methods studied.

•

Agricultural application of sludge treated by anaerobic digestion resulted in significant TPs owing to the leaching of heavy metals from treated sludge, even though anaerobic digestion substantially reduced organic pollutants.

40

•

Pollutants could accumulate in the products of sludge pyrolysis, which could be released if these products were consumed as fuel.

•

Fuel/energy substitution, for example the use of biogas from anaerobic digestion or bio-oil and biochar from pyrolysis as fuel, had a significant contribution in determining the GWP of sludge treatment methods.

•

Challenges remain in comparing environmental impacts across studies without bias. The differences across studies in choices of LCA methods, system boundaries, inventory analysis, and other parameters influence the apparent significance of the factors affecting impact potentials, even for the same sludge treatment method. Future studies could employ systematic design of alternate scenarios in order to eliminate such biases as much as possible, or include sensitivity analyses to explore how the impact potentials may vary if the parameters are assumed differently.

•

The toxicity impacts of a wider array of pollutants (particularly emerging ones) arising from different sludge treatment methods could be probed further from a lifecycle perspective, in order to facilitate fair comparison of the relative magnitudes across multiple factors. This would build on the work of past studies on pollutant levels that remain in the products and by-products of sludge treatment, as well as on the effectiveness of sludge-based adsorbents or soil amendments in binding and stabilising various pollutants.

•

The influence of possible pre-treatments on sludge volume/weight reduction and on environmental impacts could be further explored.

41

Acknowledgements The authors thank Dr Omar Swei for reviewing and providing valuable suggestions on life-cycle assessment in the final stages of the preparation of the manuscript, as well as Drs Ken Halls and John R. Grace, Professor of Emeritus of UBC for proof-reading the manuscript in the early stages. Our thanks also extend to Elsevier Webshop Support editors for language editing.

Funding: This research did not receive any specific grant from funding agencies in the public, commercial, or not-for-profit sectors.

Declaration of interest: None

42

Table 1. Assessment Matrix for Effectiveness and Environmental Impact Effectiveness

Environmental Impact (LCA Impact Category)

Sludge volume/weight reduction

Pollutant reduction, removal, or stabilisation

Global Warming Potential (GWP)

+2

High volume/weight reduction relative to other sludge treatment methods

High reduction, removal, or stabilisation for a wide range of organic pollutants and heavy metals relative to other sludge treatment methods

Very low or negative GWP or TP relative to other sludge treatment methods

+1

Moderate volume/weight reduction relative to other sludge treatment methods

Moderate reduction, removal, or stabilisation for some pollutants relative to other sludge treatment methods

Moderately low or negative GWP or TP relative to other sludge treatment methods

0

Little or no change in volume/weight relative to sludge that has not been treated with the method in question

Little or no change or improvement in pollutant concentrations, quantities, leachability, or destabilisation, or effects uncertain or variable, relative to sludge that has not been treated with the method in question

Average GWP or TP relative to other sludge treatment methods

-1

Moderate increase in volume/weight relative to other sludge treatment methods

Moderate increase in pollutant concentrations, quantities, leachability, or destabilisation for some pollutants relative to other sludge treatment methods

Moderately high GWP or TP relative to other sludge treatment methods

-2

Large increase in volume/weight relative to other sludge treatment methods

Large increase in pollutant concentrations, quantities, leachability, or destabilisation for many pollutants relative to other sludge treatment methods

Very high GWP or TP relative to other sludge treatment methods

Score

Toxicity Potentials (TP)

1

Table 2. Summary of Effectiveness in Volume/Weight and Pollutant Reduction Sludge Treatment Method

Volume/Weight Reduction Effectiveness

Score

Pollutant Reduction Effectiveness

Score

VS destruction = 40%–50% (European Commission, 2001b) TSS removal yield = 66%–86.2% (Salsabil et al., 2010) Sludge dry wt. reduction after 6 days = 20% (Kurahashi et al., 2017)

+1

Pharmaceuticals = 30% reduction (Malmborg and Magnér, 2015) PCBs after 21 days = 12%–32% of original concentration (Rosínska and Dąbrowska, 2014) Effects uncertain for other organic pollutants (Mailler et al., 2014b; Mailler et al., 2017)

+1

Biological treatment Anaerobic digestion

No biodegradation of heavy metals Composting/aerobic digestion

TSS removal yield = 57%–76% (Salsabil et al., 2010)

+1

12 organic pollutants experienced mass reductions ranging from 13%–89% (Poulsen and Bester, 2010) No biodegradation of heavy metals

+1

Constructed wetlands

No data

0

Ibuprofen and caffeine reduction = >80%; partial or poor removal of other organic pollutants (Zhu and Chen, 2014)

0

No biodegradation of heavy metals Chemical treatment Lime stabilisation

Addition of 20%–40% or 50%–90% CaO or equivalent Ca(OH)2 per unit dry solids (European Lime Association), with corresponding volume/weight increase

-2

Reduction of some heavy metals by 6%–23%, but no reduction of others. Dosages may be too high for land application (Wong and Selvam, 2006; Wong and Fang, 2000)

0

None reported for organic pollutants Thermal or thermo-chemical treatment Incineration

Reduction of sludge cake by up to 96% (Vesilind and Ramsey, 1996) Weight reduction = 62% (Hwang et al., 2007)

+2

Net formation of PCDD/Fs (Van Caneghem et al., 2010; Dai et al., 2014)

0

Net destruction of dioxin-like PCBs, PCBs, and PAHs for sludge co-incinerated with other waste;

2

(Van Caneghem et al., 2010) Heavy metals contained in solid residue (ash) Pyrolysis

Reduction of dry sludge (5.2 wt.% moisture) by 35%–50% (Inguanzo et al., 2002)

+2

Weight reduction = 63% (Hwang et al., 2007)

Reduction of PCDD/Fs to <5 wt.% of original (Dai et al., 2014)

+1

Reduced leaching of heavy metals from biochar, from 0.43%–88.87% to 0.09%–13.24% (Lu et al., 2016) Accumulation of 5%–20% of heavy metal content in bio-oil, with risk of exchange and leaching (Leng et al., 2015)

Gasification

Expected to be similar to pyrolysis

+2

Low rate of retention of metals (as low as 13.1% for Cd) on biochar when tested for leaching with 50% nitric acid (Marrero et al., 2004)

0

Hydrothermal carbonisation

60% of input solid mass in hydrochar (He et al., 2013)

+2

+1

Hydrothermal liquefaction

Solid residues around 12.8–22.6 wt.% of total product weight (Qian et al., 2017)

+2

Leaching rates of Cu, Cd, Ni and Zn reduced from 2.04%–7.31% to 0.14%–2.30% after treatment (Huang et al., 2011). Zn leaching still above TCLP limit by 1.6–4.1 times (Yuan et al., 2015)

Sub/supercritical water gasification (SWG)

Varies: Solid residues of around 68%–69% of dry weight of sludge (Li et al., 2012); <30% of original sludge dry weight (Zhang et al., 2010)

+2

Supercritical water oxidation (SCWO)

Primarily inorganic residues

+1

Combined SWG and SCWO reduced weight to 3.5% of initial (Qian et al., 2015) No other numerical evidence uncovered

Heavy metal distribution into bio-oil (5%–20%) poses risks for some metals (Leng et al., 2015; Yuan et al., 2015; Li et al., 2012) Formation of small amounts of PAHs during gasification (Xu et al., 2013) 95% destruction of COD and >95% destruction of wide range of organics in diesel fuel and waste landfill leachate contaminated soil (Williams and Onwudili, 2006; Zou et al., 2013) (Similar process application)

+1

+1

+1

Reduced leachability of Cr, Cu, Zn, and Fe by as much as 99% but increased leachability of Ni by 13 times (Zou et al., 2013) No relevant studies uncovered for sewage sludge Wet oxidation

Primarily inorganic residues

+1

Maximum COD destruction of 89% achieved for pulp and paper mill effluent treated by catalytic wet

+1

3

No numerical evidence uncovered for volume/weight reduction

air oxidation (Garg et al., 2007) (similar process application) No relevant studies uncovered for wet oxidation of sewage sludge

Melting

Higher combustion temperature than incineration; glass-like slag with higher density than ash

+1

No numerical evidence uncovered for volume/weight reduction Drying

Reduces moisture by as much as 85–95 wt.% dry solids (i.e. 5%–15% moisture)

Reduced leachability of wide range of heavy metals to very low amounts (≤2.89 mg L-1) (Idris and Saed, 2002)

+1

Dioxin production reported to be reduced (Hong et al., 2009) (no numerical evidence uncovered) +1

No studies uncovered

0

4

Table 3. Summary of Net Global Warming Potential Sludge Treatment Method

End-Use/Disposal

Energy/Fuel Consumption/Substitution

References

Score

AD

Landfill

Partial use of biogas from AD for heat and/or power generation (Poulsen and Hansen, 2003; Brown et al., 2010); None (Peters and Rowley, 2009)

Poulsen and Hansen, 2003 Peters and Rowley, 2009 Brown et al., 2010

-1

AD

Agricultural application

None

Peters and Lundie, 2001 Hospido et al., 2005

0

AD

Agricultural application

Partial use of biogas from AD for heat and/or power generation

Poulsen and Hansen, 2003 Murray et al., 2008 Brown et al., 2010

+1

AD

Agricultural application

Use of biogas from AD for power generation

Peters and Lundie, 2001

+2

AD + composting

Agricultural application

Partial use of biogas from AD for heat and/or power generation

Poulsen and Hansen, 2003

0

Composting

Agricultural application

None

Liu et al., 2013

0

Lime stabilisation

Landfill

None

Houillon and Jolliet, 2005

-2

Lime stabilisation

Agricultural application

None

Peters and Lundie, 2001 Houillon and Jolliet, 2005 Murray et al., 2008 Peters and Rowley, 2009

-1

Biological treatment

Chemical treatment

Thermal or thermo-chemical treatment Drying

Agricultural application

None

Peters and Rowley, 2009

-2

Drying

Fuel for cement kiln firing

Partially replace coal

Peters and Rowley, 2009

+2

5

Incineration

Ash landfilled

None

Hospido et al., 2005

0

Pyrolysis (w/ pre-drying)

Fuel/raw material

Use of syngas only

Hospido et al., 2005

-1

Pyrolysis (w/ pre-drying)

Fuel/raw material

Use of syngas, char, tar

Hospido et al., 2005

0

Pyrolysis

Fuel/substitute for fertiliser

Use of bio-oil for heat and power generation Use of bio-char to replace fertiliser

Cao and Pawłowski, 2013

+1

SCWO

Land

Excess heat used for district heating

Johansson et al., 2008

+1

Combinations of treatment methods AD + co-incineration

Fuel/additive for cement kiln firing/production

Partial use of biogas from AD for heat and power generation Partially replace cement kiln primary fuel mix

Poulsen and Hansen, 2003

+2

AD + co-incineration

Fuel in MSW incinerator Ash landfilled

Partial use of biogas from AD for heat and power generation Heat and power generation from MSW coincineration

Poulsen and Hansen, 2003

+1

AD + drying

Agricultural application

None

Peters and Lundie, 2001 Peters and Rowley, 2009

-2

AD + drying

Agricultural application

Use of biogas from AD to power dryers

Peters and Lundie, 2001

+1

AD + drying

Fuel for cement kiln firing

Partially replace coal

Peters and Rowley, 2009

+2

AD + drying + coincineration

Fuel in MSW incinerator Ash landfilled

Partial use of biogas from AD for heat and power generation Heat and power generation from MSW coincineration

Poulsen and Hansen, 2003

+2

AD + pyrolysis

Fuel/substitute for fertiliser

Use of biogas for heat and power generation and to replace diesel Use of bio-oil for heat and power generation Use of bio-char to replace fertiliser

Cao and Pawłowski, 2013

+2

6

Drying + composting

Agricultural application

None

Murray et al., 2008

-1

Abbreviations: AD = anaerobic digestion; MSW = municipal solid waste; SCWO = super-critical water oxidation

7

Table 4. Summary of Net Toxicity Potentials Sludge Treatment Method

EndUse/Disposal

HTP

Score

TETP

Score

AETP

Score

Heavy metal release in agricultural application (Peters and Rowley, 2009; Hospido et al., 2010; Harder et al., 2017)

-1

Heavy metal release in agricultural application (Peters and Rowley, 2009; Hospido et al., 2010; Xu et al., 2014)

-2

(Xu et al., 2014)

0

(Xu et al., 2014)

+2

Assumption of a small but wellmanaged landfill with leachate trapping (Peters and Rowley, 2009), leading to average performance compared to other sludge treatment methods studied

0

Assumption of a small but wellmanaged landfill with leachate trapping (Peters and Rowley, 2009), leading to average performance compared to other sludge treatment methods studied

0

(Xu et al., 2014)

0

Biological treatment AD

AD

Agricultural application

Landfill

-2

Mercury, lead release into water (Xu et al., 2014), leading to poor performance compared to other sludge treatment methods studied

(Xu et al., 2014)

AD + composting

Land application

Heavy metal release in land application (Peters and Rowley, 2009)

-1

Heavy metal release in land application (Peters and Rowley, 2009)

-2

--

--

Composting

Agricultural application

(Usapein and Chavalparit, 2017)

0

Heavy metal release in agricultural application (Usapein and Chavalparit, 2017)

-2

Heavy metal release in agricultural application (Usapein and Chavalparit, 2017)

-2

Heavy metal release in

-1

Heavy metal release in

-2

--

--

Chemical treatment Lime

Agricultural

8

stabilisation

application

agricultural application (Peters and Rowley, 2009)

agricultural application (Peters and Rowley, 2009)

Thermal or thermo-chemical treatment Drying

Agricultural application

Fuel use without fuel substitution, heavy metal release in agricultural application (Peters and Rowley, 2009)

-2

Fuel use without fuel substitution, heavy metal release in agricultural application (Peters and Rowley, 2009)

-2

--

--

Drying

Cement kiln

Coal substitution (Peters and Rowley, 2009)

+2

Coal substitution (Peters and Rowley, 2009)

+2

--

--

Coincineration with coal

Cement kiln

Coal substitution (Usapein and Chavalparit, 2017)

0

Coal substitution (Usapein and Chavalparit, 2017)

0

Coal substitution (Usapein and Chavalparit, 2017)

0

Fuel use without fuel substitution, heavy metal release in agricultural application (Peters and Rowley, 2009)

-1

Fuel use without fuel substitution, heavy metal release in agricultural application (Peters and Rowley, 2009)

-2

(Xu et al., 2014)

0

(Xu et al., 2014)

+2

Heavy metal release in agricultural application (Xu et al., 2014)

Coal substitution (Peters and Rowley, 2009)

+1

Coal substitution (Peters and Rowley, 2009)

+1

--

--

Combinations of treatment methods AD + drying

Agricultural application

AD + drying

Cement kiln

AD + drying

Landfill

Disposal in landfill (Buonocore et al., 2018)

-1

--

--

--

--

AD + drying + gasification

Fuel

Fuel substitution (Buonocore et al., 2018)

+1

--

--

--

--

AD + incineration

Landfill

Energy recovery (Xu et al., 2014)

+2

(Xu et al., 2014)

0

(Xu et al, 2014)

+2

Abbreviations: AD = anaerobic digestion; AETP = aquatic ecotoxicity potential; HTP = human toxicity potential; TETP = terrestrial ecotoxicity potential

9

10

Table 5. Overall Assessment of Effectiveness and Environmental Impact Sludge Treatment Method

Effectiveness/Environmental Impact

Score

Biological treatment Anaerobic Digestion

+1

Volume/Weight Reduction

VS destruction = 40%–50% (European Commission, 2001b) TSS removal yield = 66%–86.2% (Salsabil et al., 2010) Sludge dry wt. reduction after 6 days = 20% (Kurahashi et al., 2017)

+1

Pollutant Reduction

Pharmaceuticals = 30% reduction (Malmborg and Magnér, 2015) PCBs after 21 days = 12%–32% of original concentration (Rosínska and Dąbrowska, 2014) Effects uncertain for other organic pollutants (Mailler et al., 2014b; Mailler et al., 2017) No biodegradation of heavy metals

+1

GWP

Not using biogas from AD to substitute fuel/energy increases GWP (Peters and Lundie, 2001; Hospido et al., 2005); Landfilling of treated sludge (as opposed to agricultural application) results in higher GWP (Poulsen and Hansen, 2003; Peters and Rowley, 2009; Brown et al., 2010) Using biogas from AD to substitute fuel/energy reduces GWP (Peters and Lundie, 2001; Poulsen and Hansen, 2003; Murray et al., 2008; Brown et al., 2010)

-1 to +2

TPs

Heavy metal release in agricultural application (Peters and Rowley, 2009; Hospido et al., 2010; Harder et al., 2017) Mercury, lead release into water in the case of landfilling (Xu et al., 2014)

-2 to +2

Composting

0

Volume/Weight Reduction

TSS removal yield = 57%–76% (Salsabil et al., 2010)

+1

Pollutant Reduction

12 organic pollutants experienced mass reductions ranging from 13%–89% (Poulsen and Bester, 2010) No biodegradation of heavy metals

+1

GWP

CH4 and N2O emissions during composting increases GWP, but this is partially offset by the use of treated sludge to replace fertilisers (Liu et al., 2013)

0

TPs

Heavy metal release in agricultural application (Usapein and Chavalparit, 2017)

0 to -2

11

Chemical treatment Lime stabilisation

-1

Volume/Weight Reduction

Addition of 20%–40% or 50%–90% CaO or equivalent Ca(OH)2 per unit dry solids (European Lime Association) with corresponding volume/weight increase

-2

Pollutant Reduction

Reduction of some heavy metals by 6%–23%, but no reduction of others. Dosages may be too high for land application (Wong and Selvam, 2006; Wong and Fang, 2000) None reported for organic pollutants

0

GWP

Lime production increases GWP significantly (Peters and Lundie, 2001; Houillon and Jolliet, 2005; Murray et al., 2008; Peters and Rowley, 2009)

-1 to -2

TPs

Heavy metal release in agricultural application (Peters and Rowley, 2009)

-1 to -2

Thermal or thermo-chemical treatment Drying Volume/Weight Reduction Pollutant Reduction

0 Reduces moisture by as much as 85–95 wt.% dry solids (i.e. 5%–15% moisture)

+1

--

GWP

Use of treated sludge to substitute fuel produces beneficial effects; otherwise, fuel use is detrimental to GWP (Peters and Rowley, 2009)

-2 to +2

TPs

Use of treated sludge to substitute fuel produces beneficial effects; otherwise, fuel use is detrimental to TP (Peters and Rowley, 2009) Heavy metal release in agricultural application (Peters and Rowley, 2009)

-2 to +2

Incineration

0

Volume/Weight Reduction

Reduction of sludge cake by up to 96% (Vesilind and Ramsey, 1996) Weight reduction = 62% (Hwang et al., 2007)

+2

Pollutant Reduction

Net formation of PCDD/Fs (Van Caneghem et al., 2010; Dai et al., 2014)

0

Net destruction of dioxin-like PCBs, PCBs, and PAHs for sludge co-incinerated with other waste; (Van

12

Caneghem et al., 2010) Heavy metals contained in solid residue (ash) GWP TPs

(Hospido et al., 2005)

0 --

Pyrolysis

+1

Volume/Weight Reduction

Reduction of dry sludge (5.2 wt.% moisture) by 35%–50% (Inguanzo et al., 2002) Weight reduction = 63% (Hwang et al., 2007)

+2

Pollutant Reduction

Reduction of PCDD/Fs to <5 wt.% of original (Dai et al., 2014) Reduced leaching of heavy metals from biochar, from 0.43%–88.87% to 0.09%–13.24% (Lu et al., 2016) Accumulation of 5%–20% of heavy metal content in bio-oil, with risk of exchange and leaching (Leng et al., 2015)

+1

GWP

Increasing the use of pyrolysis products to substitute fuel/material produces a beneficial effect (Hospido et al., 2005; Cao and Pawłowski, 2013)

-1 to +1

TPs

--

SCWO

+1

Volume/Weight Reduction

Primarily inorganic residues Combined SWG and SCWO reduced weight to 3.5% of initial (Qian et al., 2015) No other numerical evidence uncovered

+1

Pollutant Reduction

95% destruction of COD and >95% destruction of wide range of organics in diesel fuel and waste landfill leachate contaminated soil (Williams and Onwudili, 2006; Zou et al., 2013) (similar process application) Reduced leachability of Cr, Cu, Zn, and Fe by as much as 99% but increased leachability of Ni by 13 times (Zou et al., 2013) No relevant studies uncovered for sewage sludge

+1

GWP

Use of excess heat for district heating (Johansson et al., 2008)

+1

TPs

--

Abbreviations: AD = anaerobic digestion; AETP = aquatic ecotoxicity potential; GWP = global warming potential; HTP = human toxicity potential; TETP = terrestrial ecotoxicity potential; SWG = super-critical water gasification; SCWO = super-critical water oxidation

13

Sewage sludge from secondary wastewater treatment

Reduction of sludge volume or weight Reduction, removal, or stabilisation of pollutants in sludge

(i) Effectiveness (a) Sludge volume or weight reduction (b) Reduction, removal, or stabilisation of pollutants (ii) Environmental impact (LCA) (a) Global warming potential (b) Toxicity potentials

Aims of Sludge Treatment Methods Feasibility Assessment Factors for Sludge Treatment Methods

Fig. 1 Feasibility Assessment Methodology

Waste-activated sludge

A. Pre-Treatment 1. Thickening 2. Dewatering

B1. Biological Treatment 1. Anaerobic digestion 2. Co-digestion 3. Composting/aerobic digestion 4. Constructed wetlands

B2. Chemical Treatment 1. Lime stabilisation

B3. Thermal/Thermo-chemical Treatment 1. 2. 3. 4. 5. 6. 7. 9. 8.

Drying Incineration Co-incineration Melting Pyrolysis Gasification Supercritical water oxidation Wet oxidation Hydrothermal processes

C. End-use or Disposal 1. 2. 3. 4.

Landfill Land/agricultural application Fuel Material

Fig. 2 Processes for Recovery of Material and Energy from Sludge

4000

3000 [CELLRANGE]

GWP (kg-CO2e/t-DS)

[CELLRANGE]

2000

1000

0 Dry. [CELLRANGE] + [CELLRANGE] [CELLRANGE] [CELLRANGE] [CELLRANGE] [CELLRANGE] comp. [CELLRANGE] [CELLRANGE] [CELLRANGE] [CELLRANGE]

-1000

[CELLRANGE]

[CELLRANGE]

-2000 Houillon and Jolliet,

Johansson et al., 2008

Murray et al., 2008

Brown et al., 2010

Carballa et al., 2011 Liu et al., 2013

Righi et al., 2013

Usapein and Chavalparit, 2017

GWP due to carbon sequestration GWP due to fertiliser avoidance GWP due to release of methane and nitrous oxide following land application GWP due to other factors

Note: Findings from studies are plotted in chronological order of study. Abbreviations: AD (fertiliser; no fertiliser; lime; no lime) = anaerobic digestion (with replacement of fertiliser by treated sludge; without replacement of fertiliser by treated sludge; with lime addition; without lime addition); Co-AD = co-digestion (anaerobic) with organic fraction of municipal solid waste; Comp. (low; high) = composting (low estimate for methane and nitrous oxide release; high estimate for methane and nitrous oxide release); Dry. = drying; Lime (agri.; lf.) = lime stabilisation (with agricultural application as the end-use; with landfill disposal); SCWO (land) = supercritical water oxidation with land application as the end-use

Fig. 3 Contribution to GWP by Factors due to Land Application of Treated Sludge

GWP (kg-CO2e/t-DS) or Net Energy Consumption (kWh/t-DS)

5000 4000 3000 [CELLRANGE]

2000

[CELLRANGE]

1000

[CELLRANGE] [CELLRANGE]

[CELLRANGE] [CELLRANGE] [CELLRANGE] [CELLRANGE] [CELLRANGE] [CELLRANGE] [CELLRANGE]

[CELLRANGE] [CELLRANGE]

[CELLRANGE] [CELLRANGE]

0 [CELLRANGE][CELLRANGE] [CELLRANGE] [CELLRANGE] [CELLRANGE]

[CELLRANGE] [CELLRANGE] -1000[CELLRANGE]

-2000 -3000 -4000 -5000 Peters and Lundie, 2001

Poulsen and Hansen, 2003

Hospido et al., 2005

Peters and Rowley, 2009

Cao and Pawłowski, 2013

Net Energy Consumption (kWh/t-DS) GWP (kg CO2e/t-DS)

Note: Findings from studies are plotted in chronological order of study. Abbreviations: AD (agri.; alternate; Bondi; Malabar) = anaerobic digestion (with agricultural application end-use; alternate scenario; process at Bondi facility; process at Malabar facility); Comp. = composting; Co-inc. (cement; MSW) = co-incineration (with cement kiln primary fuel mix; with municipal solid waste); Comp. = composting; Dry. (agri.; cement) = drying (with agricultural application end-use; with end-use as cement kiln fuel); Inc. = incineration; Lime (North Head) = lime stabilisation (process at North Head facility); Pyr. (partial reuse; full reuse) = pyrolysis (with partial reuse of products (syngas); with full reuse of products (syngas, char, tar))

Fig. 4 GWP and Net Energy Consumption

400

HTP, TETP, AETP (kg 1,4-DCBe/t-DS)

300 200 100 [CELLRANGE] [CELLRANGE] [CELLRANGE] [CELLRANGE]

10 [CELLRANGE] [CELLRANGE] 5[CELLRANGE] [CELLRANGE] [CELLRANGE] [CELLRANGE] [CELLRANGE] 0 [CELLRANGE] -5

[CELLRANGE]

-10 -100 -200 -300 -400 Peters and Rowley, 2009

Xu et al., 2011

Human Toxicity Potential (HTP)

500,000

Terrestrial Ecotoxicity Potential (TETP)

400,000

TETP, AETP (kg TEG/t-DS)

Carballa et al., 2011

Aquatic Ecotoxicity Potential (AETP)

300,000 200,000

Note: Findings from studies are plotted in chronological order of study.

100,000 20,000 10,000 0

[CELLRAN GE]

-10,000

Abbreviations: AD (agri.; dewater.; lf.) = anaerobic digestion (with dewatering; with agricultural application end-use; with landfill disposal); Carb. = carbonisation; Co-inc. = coincineration; Comp. = composting; Dry. (agri.; cement) = drying (with agricultural application end-use; with end-use as cement kiln fuel); Gas. (biomass cogen.) = gasification (with biomass used in co-generation); Lime = lime stabilisation

-20,000 -30,000

[CELLRAN GE] Wang et al., 2013

[CELLRAN GE] Usapein and Chavalparit, 2017

Fig. 5 Toxicity Potentials of Sludge Treatment Methods: (a) HTP, TETP, and AETP expressed in kg 1,4-DCBe/t-DS; (b) TETP and AETP expressed in kg TEG/t-DS

Highlights •

Sewage sludge treatment methods are assessed from a lifecycle perspective.

•

Key aspects assessed are sludge volume, pollutants, global warming, and toxicity.

•

Semi-quantitative scoring matrices aid holistic decision-making.

•

Best methods are anaerobic digestion, pyrolysis, and supercritical water oxidation.

•

More work is needed to reduce pollutant release from sludge treatment processes.