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

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Journal of Cleaner Production 247 (2020) 119495
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Review
Feasibility of alternative sewage sludge treatment methods from a
lifecycle assessment (LCA) perspective
Soon Kay Teoh a, Loretta Y. Li b, *
a 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
a r t i c l e i n f o
Article history:
Received 26 June 2019
Received in revised form
8 October 2019
Accepted 27 November 2019
Available online 29 November 2019
Handling editor. Prof. Jiri Jaromir Kleme�s
Keywords:
Sewage sludge treatment
Lifecycle assessment (LCA)
Environmental pollutants
Toxicity potentials
Decision-making score
* Corresponding author.
E-mail address: lli@civil.ubc.ca (L.Y. Li).
https://doi.org/10.1016/j.jclepro.2019.119495
0959-6526/© 2019 The Authors. Published by Elsevier
a b s t r a c t
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. Variousbiological, chemical, thermal, and thermo-chemical sludge treatmentmethods
described in the literature are reviewedand evaluatedholistically byadopting the developedmethodology 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 treat-
ment 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.
© 2019 The Authors. Published by Elsevier Ltd. This is an open access article under the CC BY-NC-ND
license (http://creativecommons.org/licenses/by-nc-nd/4.0/).
Contents
1. Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2
2. Approach and methodology . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3
2.1. Feasibility assessment approach . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3
2.2. Methodology for assessing effectiveness and environmental impact . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3
2.3. Five-point scale scoring system . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4
3. Results and discussion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5
3.1. Effectiveness in reducing sludge volume/weight . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5
3.2. Effectiveness in reducing, removing, or stabilising pollutants . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6
3.3. Global warming potential (GWP) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9
3.3.1. Effects of end-use and final disposal . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9
3.3.2. Net energy/fuel consumption/substitution . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12
3.4. Toxicity potentials (TPs) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13
4. Overall assessment and limitations . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 15
5. Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18
Funding . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18
Declaration of competing interest . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18
Acknowledgements . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18
Supplementary data . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18
References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18
Ltd. This is an open access article under the CC BY-NC-ND license (http://creativecommons.org/licenses/by-nc-nd/4.0/).
http://creativecommons.org/licenses/by-nc-nd/4.0/
mailto:lli@civil.ubc.ca
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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
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 hydrocarbonPBDE 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
S.K. Teoh, L.Y. Li / Journal of Cleaner Production 247 (2020) 1194952
1. Introduction
Population growth, urbanisation, industrialisation, and inade-
quate 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 twomajor 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 (H€ohne and Püttmann,
2008; Gao and Wen, 2016). In the EU, the extent of use of land-
filling as a sludge disposal method declined from 15% in 2005 to 7%
in 2015 and that of agricultural application decreased from 43% in
2005 to 28% in 2015, while that of sludge incineration increased
from 21% to 38% over the same period (Gutjahr andMüller-Schaper,
2018).
Second, conventional secondary wastewater treatment pro-
cesses 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,
2015), hydrophilic or refractory organic compounds remain in the
treated wastewater at ng/Lemg/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 hy-
drophobicity, and limited biodegradability persist. For example,
over 60%e90% 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 SIe1 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 so-
lutions has been recognised (Moe and Gangarosa, 2009), alongwith
other factors such as cost, especially in less developed economies.
In their assessment of progress towards achieving the United Na-
tions (UN) Millennium Development Goal (MDG) of doubling ac-
cess to sanitation facilities, UNICEF and World Health Organization
(WHO), 2015 reported that despite some progress, large disparities
in 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 fac-
tors, 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
S.K. Teoh, L.Y. Li / Journal of Cleaner Production 247 (2020) 119495 3
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 imple-
mentation in a particular case.
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 “life-
cycle 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 SIe2. 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
Sewage sludge 
from secondary 
wastewater 
treatment
Reduc on of sludge volume o
Aims of Sludge Treatment Me
Reduc on, removal, or stabili
pollutants in sludge
Fig. 1. Feasibility assessm
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 nitro-
gen 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 (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 re-
sults 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 facili-
tating an examination of trade-offs by considering lifecycle im-
pacts. 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 measure-
ment and is thus useful for measuring the overall impact. As the
numerical magnitudes of LCA impact potentials (or category in-
dicators) 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.
Guine�e 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 consid-
erable attention in 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
(i) Effec veness
r weight
thods 
Feasibility Assessment Factors for 
Sludge Treatment Methods
(ii) Environmental impact (LCA)
(a) Sludge volume or 
weight reduc on
(b) Reduc on, removal, or 
stabilisa on of pollutants
(a) Global warming poten al
(b) Toxicity poten als
sa on of 
ent methodology.
http://apps.webofknowledge.com
http://apps.webofknowledge.com
A. Pre-Treatment
1. Thickening
Waste-ac vated sludge
2. Dewatering
B2. Chemical Treatment
2. Incinera on
4. Mel ng
5. Pyrolysis
2 Incinera
1. Drying
B1. Biological Treatment
1. Anaerobic diges on
2. Co-diges on
C. End-use or Disposal
2. Land/agricultural applica on
3. Fuel
1. Landfill
4. Material
1 D i
B3. Thermal/Thermo-chemical Treatment
1. Lime stabilisa on
7. Supercri cal water oxida on
8. Wet oxida on
3. Compos ng/aerobic diges on
4. Constructed wetlands
3. Co-incinera on
6. Gasifica on
9. Hydrothermal processes
Fig. 2. Processes for recovery of material and energy from sludge.
S.K. Teoh, L.Y. Li / Journal of Cleaner Production 247 (2020) 1194954
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 com-
parisons 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-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 performwell 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) toþ2 (very positive). As themagnitudes 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 treat-
ment 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 fac-
tors. Semi-quantitative numerical scores for the two effectiveness
S.K. Teoh, L.Y. Li / Journal of Cleaner Production 247 (2020) 119495 5
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 feasibilityappraisal. The maximum potential
overall score for any treatment method is þ2; conversely, the
minimum is �2. Decision-makers can use the same scoring system
and tailor it to their needs by assigning their ownweights 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 thor-
ough 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 com-
binations 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 SIe5 summarises the func-
tional units, lifecycle stages, and system boundaries considered in
each study. It should be noted that the sludge treatmentmethods 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 typical range of values for VS destruction in mesophilic
anaerobic digestion was reported to be 40%e50% (European
Commission, 2001b). In a study to investigate the effect of adding
Table 1
Assessment matrix for effectiveness and environmental impact.
Score Effectiveness
Sludge volume/weight reduction Pollutant reduction, removal, or s
þ2 High volume/weight reduction relative
to other sludge treatment methods
High reduction, removal, or stabi
range of organic pollutants and h
relative to other sludge treatmen
þ1 Moderate volume/weight reduction
relative to other sludge treatment
methods
Moderate reduction, removal, or
some pollutants relative to other
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 improveme
concentrations, quantities, leacha
destabilisation, or effects uncerta
relative to sludge that has not bee
method in question
�1 Moderate increase in volume/weight
relative to other sludge treatment
methods
Moderate increase in pollutant co
quantities, leachability, or destab
pollutants relative to other sludg
methods
�2 Large increase in volume/weight
relative to other sludge treatment
methods
Large increase in pollutant conce
quantities, leachability, or destab
pollutants relative to other sludg
methods
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 anaer-
obic 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%e76% while that for anaerobic
digestion was 66%e86%. 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 destruc-
tion. 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%).
The main functions of lime stabilisation are the reduction of the
microbial content of sludge and reduction of heavy metal avail-
ability (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 (dis-
countingmoisture 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%e90% CaO per unit dry
solids for 75 min to treat sludge at >55 �C and pH > 12, or the
addition of 20%e40% 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
Environmental Impact (LCA Impact Category)
tabilisation Global Warming Potential
(GWP)
Toxicity Potentials (TP)
lisation for a wide
eavy metals
t methods
Very low or negative GWP or TP relative to other
sludge treatment methods
stabilisation for
sludge treatment
Moderately low or negative GWP or TP relative to
other sludge treatment methods
nt in pollutant
bility, or
in or variable,
n treated with the
Average GWP or TP relative to other sludge
treatment methods
ncentrations,
ilisation for some
e treatment
Moderately high GWP or TP relative to other
sludge treatment methods
ntrations,
ilisation for many
e treatment
Very high GWP or TP relative to other sludge
treatment methods
S.K. Teoh, L.Y. Li / Journal of Cleaner Production 247 (2020) 1194956
disposal can produce an adverse environmental impact, necessi-
tating additional control steps to bring the pH within environ-
mental regulation limits (e.g. B.C. Reg. 63/88 O.C. 268/88).
In general, thermo-chemical treatment methods are among the
most effectivemethods 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 300e900 �C (or even higher), reduced the
volume of sewage sludge at 5.2 wt% moisture by around 40%e50%
to carbonaceous residues (Inguanzo et al., 2002). Hwang et al.
(2007) found that the reductions of sewage sludge weight by py-
rolysis 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, oxida-
tion, and reduction. Thus, while reliable estimates of volume
reduction by gasification have not been uncovered in the literature
review, they are expected to be similar to those achieved by
pyrolysis.
Hydrothermal carbonisation (~180e250 �C)and hydrothermal
liquefaction (~250e400 �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 carbon-
isation 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 12.8e22.6 wt% of the total
product weight (Qian et al., 2017). Sub/supercritical water gasifi-
cation (�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%e69% 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 (Svanstr€om et al., 2004, 2005; Houillon and
Jolliet, 2005). The combined processes of supercritical water gasi-
fication and 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 thermo-chemical methods have not been un-
covered 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 1200e1500 �C; at such temperatures,
organic matter is burnt and the remaining inorganic matter be-
comes a liquid, which solidifies into a glass-like slag upon cooling
(Smith, 1992). As combustion temperatures are higher than incin-
eration 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 sludgemoisture 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 et 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 ex-
pected 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 gasifi-
cation, 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 (þ2).
A score of þ2 is given for the combined process of SCWO and wet
oxidation owing to the significant volume/weight reduction re-
ported (Qian et al., 2015). These methods are followed closely by
hydrothermal carbonisation, liquefaction, and sub/supercritical
water gasification, which are given the same score of þ2. 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 (þ1), 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 re-uptake inhibitors (SSRIs) and
oestrogens/endocrine disruptors), with an average reduction of
around 30% (Malmborg and Magn�er, 2015). Rosi�nska 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 151.0e375.3 mg kg�1 dry matter. After 21 days,
the PCB congener concentrations decreased to 45.3e48.5 mg 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 concen-
trations remained the same (150.2e374.1 mg kg�1 dry matter),
while only partial degradation was attained after 14 days (to
67.9e231.8 mg kg�1 dry matter). This may pose problems if suffi-
cient 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).
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), phe-
nolics, 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
Table 2
Summary of effectiveness in volume/weight and pollutant reduction.
Sludge Treatment Method Volume/Weight Reduction Effectiveness Score Pollutant Reduction Effectiveness Score
Biological treatment
Anaerobic digestion VS destruction ¼ 40%e50% (European Commission,
2001b)
TSS removal yield ¼ 66%e86.2% (Salsabil et al.,
2010)
Sludge dry wt. reduction after 6 days ¼ 20%
(Kurahashi et al., 2017)
þ1 Pharmaceuticals ¼ 30% reduction (Malmborg and Magn�er,
2015)
PCBs after 21 days ¼ 12%e32% of original concentration
(Rosi�nska and Dąbrowska, 2014)
Effects uncertain for other organic pollutants (Mailler et al.,
2014b, 2017)
No biodegradation of heavy metals
þ1
Composting/aerobic digestion TSS removal yield ¼ 57%e76% (Salsabil et al., 2010) þ1 12 organic pollutants experienced mass reductions ranging
from 13% to 89% (Poulsen and Bester, 2010)
No biodegradation of heavy metals
þ1
Constructed wetlands No data 0 Ibuprofen and caffeine reduction ¼ >80%; partial or poor
removal of other organic pollutants (Zhu and Chen, 2014)
No biodegradation of heavy metals
0
Chemical treatment
Lime stabilisation Addition of 20%e40% or 50%e90% 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%e23%, 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
Thermal or thermo-chemical
treatment
Incineration Reduction of sludge cake by up to 96% (Vesilind and
Ramsey, 1996)
Weight reduction ¼ 62% (Hwang et al., 2007)
þ2 Net formation of PCDD/Fs (Van Caneghem et al., 2010; Dai
et al., 2014)
Net destruction of dioxin-like PCBs, PCBs, and PAHs for
sludge co-incinerated with other waste; (Van Caneghem
et al., 2010)
Heavy metals contained in solid residue (ash)
0
Pyrolysis Reduction of dry sludge (5.2 wt% moisture) by 35%
e50% (Inguanzo et al., 2002)
Weight reduction ¼ 63% (Hwang et al., 2007)
þ2 Reduction of PCDD/Fs to <5 wt% of original (Dai et al., 2014)
Reduced leaching of heavy metals from biochar, from 0.43%
e88.87% to 0.09%e13.24% (Lu et al., 2016)
Accumulation of 5%e20% of heavy metal content in bio-oil,
with risk of exchange and leaching (Leng et al., 2015)
þ1
Gasification Expected to be similar to pyrolysis þ2 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)
þ2 Leaching rates of Cu, Cd, Ni and Zn reduced from 2.04%
e7.31% to 0.14%e2.30% after treatment (Huang et al., 2011).
Zn leaching still above TCLP limit by 1.6e4.1 times (Yuan
et al., 2015)
Heavy metal distribution into bio-oil (5%e20%) poses risks
for somemetals (Leng et al., 2015; Yuan et al., 2015; Li et al.,
2012)
Formation of small amounts of PAHs during gasification (Xu
et al., 2013)
þ1
Hydrothermal liquefaction Solid residues around 12.8e22.6 wt% of total
product weight (Qian et al., 2017)
þ2 þ1
Sub/supercritical water
gasification (SWG)
Varies: Solid residues of around 68%e69% of dry
weight of sludge (Li et al., 2012); <30% of original
sludge dry weight (Zhang et al., 2010)
þ2 þ1
Supercritical water
oxidation (SCWO)
Primarily inorganic residues
Combined SWG and SCWO reduced weight to 3.5%
of initial (Qian et al., 2015)
No other numerical evidence uncovered
þ1 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
þ1
Wet oxidation Primarily inorganic residues
No numerical evidence uncovered for volume/
weight reduction
þ1 Maximum COD destruction of 89% achieved for pulp and
paper mill effluent treated by catalytic wet air oxidation
(Garg et al., 2007) (similar process application)
No relevant studies uncovered for wet oxidation of sewage
sludge
þ1
Melting Higher combustion temperature than incineration;
glass-like slag with higher density than ash
No numerical evidence uncovered for volume/
weight reduction
þ1 Reduced leachability of wide range of heavy metals to very
low amounts (�2.89 mg/L) (Idris and Saed, 2002)
Dioxin production reported to be reduced (Hong et al.,
2009) (no numerical evidence uncovered)
þ1
Drying Reduces moisture by as much as 85e95 wt% dry
solids (i.e. 5%e15% moisture)
þ1 No studies uncovered 0
S.K. Teoh, L.Y. Li / Journal of Cleaner Production 247 (2020) 119495 7
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%e95% of
nonylphenols, nonylphenol monoethoxylate, nonylphenol dieth-
oxylate, octylphenol, DEHP, and BDE-209 were removed, leading to
lower concentrations in the final solid product; however, the au-
thors 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 alkylben-
zene sulphonate, alkylphenols, phthalates, PAHs, PCBs, and other
pollutants after anaerobic digestion. They found that the
S.K. Teoh, L.Y. Li / Journal of Cleaner Production 247 (2020) 1194958
concentrations of some pharmaceuticals (e.g. azithromycin, dom-
peridone, lidocaine, sulphamethoxazole, tramadol) decreased from
40 to 130 mg kg�1 dry matter to as low as undetectable levels after
digestion, while those of perfluorinated acids (PFOA, PFOS)
decreased as well (from 316 mg kg�1 dry matter to 49 mg 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 mg kg�1 dry matter in raw sludge to 58,100 mg kg�1 dry
matter in digested sludge, while those of nonylphenols, non-
ylphenol monoethoxylate, and diethoxylate increased from 940 to
1720 mg kg�1 dry matter to 1300e4520 mg kg�1 dry matter. The
authors attributed this increase in concentration to the greater
removal of both drymatter andmoisture (i.e. a decrease in themass
of the substrate). Considering the findings of the 2014 and 2017
studies together, the effect of anaerobic digestion on the concen-
trations 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 whichwere statistically significant), withmass reductions
of 13%e89%. 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 pol-
lutants. 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 prod-
ucts, such as ibuprofen and caffeine (Zhu and Chen, 2014), with
reported removal efficiencies of >80%, whereas other pharmaceu-
ticals were partially removed (e.g. DEET by 32.3%e78.4%, sulpha-
methoxazole by 33.6%e41.6%) or not removed significantly (e.g.
carbamazepine and diclofenac sodium salt by <30%) (Zhu and
Chen, 2014).
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 0.55mg/L, 2.42mg/L, and
1.09mg/L to 0.13mg/L,1.54mg/L, and 0.01mg/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%e7%. 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%e23%.
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 mobilityof 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 thermo-
chemical conversion or destruction and be released primarily in
stack gases. Several studies have shown that, in general, incinera-
tion 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 (700e950 �C) produced 2
to 13 times as much toxic polychlorinated dibenzo-p-dioxins and
dibenzofurans (PCDD/Fs) in gaseous emissions as the amount
originally present in the untreated sludge. When sewage sludge
was co-incinerated with various other waste types, such as auto-
motive 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, 2014), with the effect largely in-
dependent 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 5e14,
1200e3,900, and 70e110 for dioxin-like PCBs, PCBs, and PAHs,
respectively, and 0.03e0.1 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 150e380, 4900e6,900, and
1000e8,200, respectively) was also reported, while the net change
in the mass of PCDD/Fs was small (input/output mass ratio
0.95e3.35) (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%e97.5% 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 thermo-
chemically 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 to
be due to the volatilisation of most oestrogens at such tempera-
tures, 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 py-
rolysis at temperatures of 400e600 �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
<0.01e0.1, 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%e88.87% to 0.06%e13.24% 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%e20% of metals including Cu, Zn, Pb,
Cd, Cr, Ni, V, Mn, Ba, Co, Ti, Sn, As, and Hg could be distributed in
bio-oil, while Yuan et al. (2015) found that metals such as Zn, Ni,
S.K. Teoh, L.Y. Li / Journal of Cleaner Production 247 (2020) 119495 9
and Cd were at risk of exchange and leaching from bio-oil, sug-
gesting that bio-oils produced by pyrolysis from metal-rich
biomass such as sewage sludge should be pre-treated or upgra-
ded 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 carbona-
ceous 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.
The leaching rates of Cu, Cd, Ni, and Zn from liquefaction resi-
dues were suppressed after hydrothermal liquefaction compared
with untreated sewage sludge, from 2.04% to 7.31% in the untreated
sludge to 0.14%e2.30% in the treated sludge (Huang et al., 2011).
However, leachable Zn concentrations based on the toxicity char-
acteristic leaching procedure (TCLP) were still above the US EPA
threshold limit by 1.6e4.1 times. Leng et al. (2015) reported that
heavy metals were distributed mainly into biochar, with around
5%e20% 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 4-ring 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 re-
fractory organic pollutants in industrial wastewater effluents. Wet
oxidation is the aqueous oxidation of thickened sludgewith oxygen
at elevated temperature and pressure (e.g. 235 �C and 40 bar)
(Houillon and Jolliet, 2005). It transforms organic matter primarily
into carbon dioxide and water vapour, destroying organic pollut-
ants 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 in-
termediates, carbon dioxide, water, and innocuous end products at
elevated temperature (125e320 �C) and pressure (0.5e20 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 0.85MPa (Garg et al., 2007).
Williams and Onwudili (2006) assessed the effect of SCWO on or-
ganics in dieselfuel and waste landfill leachate. Organic species in
diesel fuel spiked into the sandmatrix at concentrations of 4e20wt
% were decomposed at 96.6%e99.8%. 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 supercrit-
ical water. Concentrations of Cr, for example, were found to lie in
the range of 9.33e11.21 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 to 0.12 mg/L) was observed owing to SCWO at
400 �C at an oxidant ratio of 3:1. Similar trends were observed for
Cu, Zn, and Fe. However, Ni experienced an increase in leachability,
from 0.28 mg/L in raw sludge to 3.51 mg/L 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 incin-
eration 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
compared to the standard limits of 1.0e100.0 mg/L. Hong et al.
(2009) reported sludge melting to be advantageous over incinera-
tion, as dioxin production is reduced owing to crystallisation at
high temperature.
Table 2 shows the scores assigned for the comparative effec-
tiveness of the treatment methods in reducing, removing, or sta-
bilising 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 þ1 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 leach-
ability. Methods that are moderately effective in reducing pollutant
levels, such as constructed wetlands, incineration, and gasification,
were assigned a score of 0.
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 con-
sumption/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 biologi-
cally 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 SIe6).
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
Table 3
Summary of net global warming potential.
Sludge Treatment Method End-Use/Disposal Energy/Fuel Consumption/Substitution References Score
Biological treatment
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)
þ1
AD Agricultural application Use of biogas from AD for power generation Peters and Lundie (2001) þ2
AD þ 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
Chemical treatment
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
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) þ2
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) þ1
SCWO Land Excess heat used for district heating Johansson et al. (2008) þ1
Combinations of treatment methods
AD þ 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) þ2
AD þ 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 co-
incineration
Poulsen and Hansen (2003) þ1
AD þ drying Agricultural application None Peters and Lundie (2001)
Peters and Rowley (2009)
�2
AD þ drying Agricultural application Use of biogas from AD to power dryers Peters and Lundie (2001) þ1
AD þ drying Fuel for cement kiln firing Partially replace coal Peters and Rowley (2009) þ2
AD þ drying þ 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 co-
incineration
Poulsen and Hansen (2003) þ2
AD þ 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) þ2
Drying þ composting Agricultural application None Murray et al. (2008) �1
Abbreviations: AD ¼ anaerobic digestion; MSW ¼ municipal solid waste; SCWO ¼ super-critical water oxidation.
S.K. Teoh, L.Y. Li / Journal of Cleaner Production 247 (2020) 11949510
amendment, treated sludge may also reduce the GWP by seques-
tering soil carbon. (3) Land application of treated sludge may in-
crease 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 theGWP.
Table SIe6 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 dis-
counted by some studies. Emission factors vary widely across
studies, with ranges of 0.02e6.3 kg CH4/t-DS and 0.00011e1.80 kg
N2O/t-DS for methane and nitrous oxide emissions owing to the
degradation of treated sludge in soil, and 50e328 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 SIe6, 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 SIe6). In the first of these studies, Brown et al. (2010)
AD 
(lime)
Lime 
(agri.)
Lime (lf.)
AD + 
SCWO 
(land)
Comp. 
(low)
Comp. 
(high) AD 
(no 
lime)
Lime
AD 
(fer lizer)
AD (no 
fer lizer)
AD
Comp.
Co-AD + 
comp.
Comp.
-2000
-1000
0
1000
2000
3000
4000
GW
P 
(k
g-
CO
2e
/t
-D
S)
 
Dry. +
comp.
Houillon and 
Jolliet, 2005
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 sequestra on 
GWP due to fer liser avoidance 
GWP due to release of methane and nitrous oxide following land applica on 
GWP due to other factors 
Fig. 3. Contribution to GWP by Factors due to Land Application of Treated Sludge
Note: Findings from studies are plotted in chronological order of study.
Abbreviations: AD (fertiliser; no fertiliser; lime; no lime) ¼ anaerobic digestion (with replacement of fertiliser by treated sludge; without replacement of fertiliser by treated sludge;
with lime addition; without lime addition); Co-AD ¼ co-digestion (anaerobic) with organic fraction of municipal solid waste; Comp. (low; high) ¼ composting (low estimate for
methane and nitrous oxide release; high estimate for methane and nitrousoxide release); Dry. ¼ drying; Lime (agri.; lf.) ¼ lime stabilisation (with agricultural application as the
end-use; with landfill disposal); SCWO (land) ¼ supercritical water oxidation with land application as the end-use.
S.K. Teoh, L.Y. Li / Journal of Cleaner Production 247 (2020) 119495 11
applied a greenhouse gas calculator tool (Biosolids Emissions
Assessment Model, BEAM), developed for the Canadian Council of
Ministers of the Environment (CCME) using data from peer-
reviewed literature and municipalities, to nine scenarios in Can-
ada. 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 fertil-
iser) 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 seques-
tration was considered, based on findings by other authors
(including Peters and Rowley (2009) and Brown et al. (2010)).
Some studies listed in Table SIe3 (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 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 SIe6 (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 consid-
ered 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 applica-
tion, 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 themethodologies adopted by the above-
S.K. Teoh, L.Y. Li / Journal of Cleaner Production 247 (2020) 11949512
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 sig-
nificant 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
biosolidsproduce 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 bio-
logically treated sludge. Of the studies reviewed, only drying/pas-
teurisation, SCWO, and wet oxidation processes led to land/
AD 
(Malabar)
AD 
(alternate)
AD + dry. 
(alternate)
Lime 
(North 
Head)
AD 
(Bondi)
AD + co-
inc. 
(cement)
AD + 
co-inc. 
(MSW)
AD + 
inc.
AD 
(agri.)
AD + 
comp. AD
-5000
-4000
-3000
-2000
-1000
0
1000
2000
3000
4000
5000
GW
P 
(k
g-
CO
2e
/t
-
t/h
Wk(
noitp
musnoC
ygrenEte
Nro)SD
-D
S)
Peters and Lundie, 2001 Poulsen and Hansen, 2003 Hosp
GWP (kg CO2e/t-DS) 
Net Energy Consump
Fig. 4. GWP and Net Energy Consumption.
Note: Findings from studies are plotted in chronological order of study.
Abbreviations: AD (agri.; alternate; Bondi; Malabar) ¼ anaerobic digestion (with agricultura
facility); Comp. ¼ composting; Co-inc. (cement; MSW) ¼ co-incineration (with cement k
cement) ¼ drying (with agricultural application end-use; with end-use as cement kiln fue
facility); Pyr. (partial reuse; full reuse) ¼ pyrolysis (with partial reuse of products (syngas)
agricultural application. This is not unusual, as drying/pasteurisa-
tion 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 (Svanstr€om et al., 2004; Johansson et al., 2008). Residues
from sludge incineration or co-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. Simi-
larly, 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
Inc.
Pyr. 
(full 
reuse)
Pyr. 
(par al 
reuse)
Dry. 
(cement)
AD + dry. 
(cement)
AD
AD + 
comp.
Lime AD + 
dry. 
(agri.)
Dry. 
(agri.)
AD + 
pyr. Pyr.
ido et al., 2005 Peters and Rowley, 2009
Cao and Pawłowski, 
2013
on (kWh/t-DS)
l application end-use; alternate scenario; process at Bondi facility; process at Malabar
iln primary fuel mix;with municipal solid waste); Comp. ¼ composting; Dry. (agri.;
l); Inc. ¼ incineration; Lime (North Head) ¼ lime stabilisation (process at North Head
; with full reuse of products (syngas, char, tar)).
S.K. Teoh, L.Y. Li / Journal of Cleaner Production 247 (2020) 119495 13
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 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 py-
rolysis 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 pro-
cesses, 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 energy-
consumptive 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 diges-
tion, 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 py-
rolysis. 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 and Lundie (2001), transport of limed sludge by truck from
North Head for a distance of 250 kmwas 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,