Review of leaching behavior of municipal solid waste incineration (MSWI) ash

Abstract

Incineration is widely adopted in modern waste management because it provides an effective way to minimize municipal solid waste that needs to be disposed of in landfills. The ash residue is often disposed by landfilling. Alternatively, the incineration ash may be recycled and reused for various applications. The crucial issues, however, are the leaching of harmful elements during the use and the end-of-life phases. This review summarizes extensive studies on leaching behavior of municipal solid waste incineration ash. Specifically, pollutants generated through leaching, factors governing leaching, methodologies to study leaching, leaching mechanisms, and treatments to reduce leaching. Many types of pollutants are generated through leaching from municipal solid waste incineration ash, in which heavy metals and organic contaminants are the most toxic and concerned. Ash properties, pH and liquid to solid ratio are the main factors governing municipal solid waste incineration ash leaching. Leaching behavior of municipal solid waste incineration ash is complicated and existing methods to evaluate leaching may not be able to represent the field conditions. Solubility and sorption are the two major leaching mechanisms. Many treatment methods have been proposed. However, not all methods are effective and some approaches are associated with high energy and high cost, which makes them less economically feasible and attractive.

Full text

Review of leaching behavior of municipal solid waste incineration (MSWI) ash
Hongwei Luoa,b, Ying Chenga, Dongqin Hea,*, En-Hua Yangc,*
aCollege of Environment, Zhejiang University of Technology, Hangzhou, 310014,
China
bEnergy Research Institute, Nanyang Technological University, Singapore 637553
cSchool of Civil and Environmental Engineering, Nanyang Technological University,
Singapore 639798
* Corresponding author:
Dr. Dongqin He, Tel: +86 571 88871571; Email: dqhe123@zjut.edu.cn
Prof. En-Hua Yang, Tel: +65 67905291; Email: ehyang@ntu.edu.sg


1
Abstract 1
Incineration is widely adopted in modern waste management because it provides an 2
effective way to minimize municipal solid waste that needs to be disposed of in landfills. The 3
ash residue is often disposed by landfilling. Alternatively, the incineration ash may be recycled 4
and reused for various applications. The crucial issues, however, are the leaching of harmful 5
elements during the use and the end-of-life phases. This review summarizes extensive studies 6
on leaching behavior of municipal solid waste incineration ash. Specifically, pollutants 7
generated through leaching, factors governing leaching, methodologies to study leaching, 8
leaching mechanisms, and treatments to reduce leaching. Many types of pollutants are 9
generated through leaching from municipal solid waste incineration ash, in which heavy metals 10
and organic contaminants are the most toxic and concerned. Ash properties, pH and liquid to 11
solid ratio are the main factors governing municipal solid waste incineration ash leaching. 12
Leaching behavior of municipal solid waste incineration ash is complicated and existing 13
methods to evaluate leaching may not be able to represent the field conditions. Solubility and 14
sorption are the two major leaching mechanisms. Many treatment methods have been proposed. 15
However, not all methods are effective and some approaches are associated with high energy 16
and high cost, which makes them less economically feasible and attractive. 17
18
Keywords: MSWI ash; leaching behavior; treatment methods; landfill; pollutants; heavy 19
metals 20

2
Abbreviations 21
22
MSW: Municipal solid waste 23
MSWI: Municipal solid waste incineration 24
WTE: Waste-to-energy 25
DOM: Dissolved organic matter 26
COD: Chemical oxygen demand 27
BOD: Biochemical oxygen demand 28
TOC: Total organic carbon 29
VFAs: Volatile fatty acids 30
HA: Humic acid 31
FA: Fulvic acid 32
HMs: Heavy metals 33
XOCs: Xenobiotic organic compounds 34
L/S: Liquid to solid ratio 35
LCA: Life cycle assessment 36
S/S: Solidification/stabilization 37
EDTA: Ethylene diamine tetraacetic acid 38
39

3
1. Introduction 40
Growing generation and disposal of municipal solid waste (MSW) have become a great 41
burden to the society, which pose serious environmental and economic concerns (Bai and 42
Sutanto, 2002; Chen et al., 2010). Landfill is an expedient and inexpensive method for MSW 43
management. However, various problems, such as groundwater pollution, odor emission, and 44
soil contamination are concerns of this approach (Kjeldsen et al., 2002; Mukherjee et al., 2014). 45
Incineration is an alternative approach to treat MSW, which reduces waste mass and volume 46
by 70% and 90% (Hjelmar, 1996), respectively, and recovers energy in the form of heat and 47
electricity (Fig. 1) (Allegrini et al., 2015b). The ash residues after incineration are classified as 48
bottom ash and fly ash, with a mass ratio of 4:1 to 5:1 (Wiles, 1996). Major elements in 49
municipal solid waste incineration (MSWI) ash are Ca, Si, Al, Fe, Mg, Na, K and Cl, and 50
common oxides in MSWI ash are CaO, SiO2, Al2O3, Fe2O3, Na2O and K2O (Kirby and Rimstidt, 51
1993; Luo et al., 2019). As shown in Table 1, CaO and SiO2 are often the most abundant 52
compounds in MSWI ash. Heavy metals such as Cr, Ni, Cu, Zn, Cd, Hg and Pb are commonly 53
found in MSWI ash (Tang et al., 2015; Yin et al., 2018b). 54
MSWI ash is often disposed of in landfills (Lo and Liao, 2007). It has been reported that 55
MSWI ash may be recycled for the production of cement and concrete, road pavement, glasses 56
and ceramics, agricultural nutrients, stabilizing agents, adsorbents, and zeolites (Dabo et al., 57
2009; Ferreira et al., 2003; Lam et al., 2010). However, leaching of harmful elements, such as 58
soluble heavy metals and toxic organic compounds (e.g., dioxins and furans), is the main 59
concern for MSWI ash disposal and recycling (Baun et al., 2004; Gao et al., 2017; Yin et al., 60
2018b). The leaching characteristics are influenced by a number of factors, including pH, liquid 61

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to solid (L/S) ratio, ash properties, weathering and aging, leaching methods, contact time, and 62
investigation scale (Dou et al., 2017; Phoungthong et al., 2016). Static batch tests and dynamic 63
column tests have been used to evaluate the leaching of MSWI ash (Di Gianfilippo et al., 64
2016b). Column leaching tests provide more reliable information because the percolation-65
based data are obtained from a flow-through pattern similar to field conditions (Lopez Meza et 66
al., 2010). 67
Geochemical models such as Visual MINTEQ, ORCHESTRA, and PHREEQC are widely 68
used to predict metal release from MSWI ash and further describe its leaching behavior (Hyks 69
et al., 2009; van der Sloot et al., 2017; Wang et al., 2016a). Potential mechanisms governing 70
leaching of major and trace elements are solubility and sorption (Komonweeraket et al., 2015). 71
Solubility-controlled leaching is associated with the dissolution of metal oxides present in 72
MSWI ash, such as aluminum oxide, iron oxide, and zinc oxide (Kosson et al., 1996). Sorption 73
determines the release of elements that exhibit sorptive affinity to the active sites on the ash 74
surface, such as (hydr)oxides, organic matters, and clay (Cornelis et al., 2008; Zhao et al., 2017). 75
Several methods, including physical or chemical separation, solidification and stabilization, 76
and thermal treatment have been investigated to treat MSWI ash (Dou et al., 2017; Lindberg et 77
al., 2015; Quina et al., 2008). The treatment aims to immobilize, remove or eliminate harmful 78
components, so that the ash can be more usable or safer for disposal (Sabbas et al., 2003). 79
However, compositions and properties of MSWI ash from different countries and plants 80
can vary significantly due to social structure and culture difference, seasonal changes, waste 81
recycling strategies, pre-treatment of MSW before incineration, combustion technology, and 82
post-treatment of MSWI ash (Bai and Sutanto, 2002; Chen et al., 2010). As a result, leaching 83

5
characteristics of MSWI ash from different geographic locations can be very different. While 84
a few review articles on MSWI ash have been published (Dou et al., 2017; Ferreira et al., 2003; 85
Lam et al., 2010), the foci are on potential applications of the ash. It is therefore necessary to 86
have a comprehensive review to understand the state-of-the-art on the leaching behavior of 87
MSWI ash and related treatments to reduce its leachability. Specifically, this article summarizes 88
pollutants generated through leaching from MSWI ash, factors affecting leaching behavior of 89
MSWI ash, methodologies to study leaching from MSWI ash, potential mechanisms behind 90
leaching, and treatments to reduce leaching from MSWI ash. In this paper, more than 200 91
literatures over the past 30 years were reviewed. Among the articles, 51% of the studies were 92
carried out in Europe and 38% of the investigations were conducted in Asia. This review shall 93
greatly contribute to future sustainable MSWI ash management. 94
95
2. Pollutants generated through leaching 96
97
2.1. Dissolved organic matter (DOM) 98
DOM in varying degrees was reported from both the lab leaching tests and the sanitary 99
landfill leachates. DOM is quantified by means of chemical oxygen demand (COD), 100
biochemical oxygen demand (BOD), or total organic carbon (TOC). It is well known that 101
MSWI ash mainly consists of a mineral matrix together with a small fraction approximately 4% 102
w/w in concentration of unburnt organic matters and organic by-products which are hazardous 103
(Dugenest et al., 1999; Nito and Ishizaki, 1997). DOM leaching from MSWI ash often relates 104
to mineral leaching which is caused by interplay of various processes (Guimaraes et al., 2006). 105

6
The leached DOM consists of a number of compounds such as volatile fatty acids (VFAs), 106
alcohols, aldehydes, and more refractory compounds like humic substances (Harmsen, 1983). 107
Among these compounds, VFAs (e.g., acetic acid, propionic acid, isobutyric acid, and 108
isovaleric acid) account for about 95% of organic matter and are readily biodegradable (Kang 109
et al., 2002; Zhang et al., 2004). Humic substances are primarily composed of humic acid (HA) 110
and fulvic acid (FA), which have been verified as an important fraction of DOM contributing 111
to the complexation with contaminants (Meima et al., 1999; Olsson et al., 2007). According to 112
previous studies, HA and FA were responsible for 0.3-0.6% and 14.3-25.6% of dissolved 113
organic carbon (DOC) in the leachate from MSWI bottom ash, respectively (Van Zomeren and 114
Comans, 2004). 115
Meanwhile, other aliphatic acids (e.g., lactic, formic, oxalic, and maleic acids) and 116
aromatic acids (e.g., terephthalic, phthalic, and benzoic acids) which account for 5-13% of 117
DOC were also identified. Many factors may affect the chemical compositions of DOM in the 118
leachates. For example, the concentrations of DOM (COD, BOD, and TOC) decrease with 119
increasing landfill age (Mandal et al., 2017). A low BOD/COD ratio suggests a low VFAs 120
concentration and relatively high HA and FA amounts in the leachate (Kjeldsen et al., 2002). 121
Therefore, DOM in leachate is a bulk parameter that covers a variety of organic reaction 122
products ranging from small volatile acids to refractory humic and fulvic-like compounds 123
(Chian and DeWalle, 1977). The presence of DOM can affect other leachate compositions 124
remarkably due to its complexing properties by the high-molecular-weight component in DOM 125
(Kjeldsen et al., 2002). 126
127

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2.2. Heavy metals (HMs) 128
Cu, Zn, Cr, Ni, Cd, Hg, and Pb are common HMs in MSWI ash. Among the HMs, Cu, Zn, 129
and Pb are usually the most abundant (Table 1). Without proper treatment, these metals may 130
leach out which can lead to serious environmental problems (Pan et al., 2013). The 131
concentration of HMs ranges from less than 0.01 mg kg-1 (e.g., Hg) to more than 1.0×104 mg 132
kg-1 (e.g., Cu and Zn). Fly ash often contains more HMs than bottom ash (Table 2) owing to 133
vaporization of metals during combustion, followed by adsorption of vaporized metals on the 134
surface of fly ash particles (Lam et al., 2010). Overall, HMs are at trace amounts and they 135
account for only less than 0.5% by weight of MSWI ash (as compared to 1-4 wt.% of DOM in 136
MSWI ash). However, HMs raise more concerns than organic pollutants due to their higher 137
leaching and contamination potentials (Dou et al., 2017). It has been reported that DOM has a 138
high affinity toward metals and enhances HMs’ mobility in water (Baun and Christensen, 2004; 139
Li et al., 2009). Previous studies suggest DOM (with 15.2, 87.2, and 214.1 mg C/L in leachates) 140
facilitate leaching of Cu from MSWI bottom ash (Olsson et al., 2007; Van Zomeren and 141
Comans, 2004). In MSWI fly ash, leaching of Cr and Cd is dissolution controlled, whereas 142
leaching of Cu, Zn, and Pb is governed by precipitation and sorption (Jiao et al., 2016). 143
Extensive investigations have been carried out to study leaching of HMs from materials, 144
products, and systems incorporating MSWI ash, such as cement and concrete production 145
(Keulen et al., 2016; Shi and Kan, 2009), road pavement (Cetin et al., 2012; De Windt et al., 146
2011), glasses and ceramics (Silva et al., 2017), stabilizing agents (Qian et al., 2006), 147
adsorbents and zeolite production (Jing et al., 2007a; Luo et al., 2017), and land reclamation 148
(Chan et al., 2018; Yin et al., 2018b). After treatment (e.g., solidification, stabilization, and 149

8
thermal methods), the concentrations of most HMs in leachate can meet the strict environment 150
regulation requirements (Gong et al., 2017; Jing et al., 2013). 151
152
2.3. Inorganic components 153
Ammonium (NH4+), nitrate (NO3–), nitrite (NO2–), bicarbonates (HCO3–), carbonates 154
(CO32–), sulfate (SO42–), chloride (Cl–), sodium (Na+), potassium (K+), calcium (Ca2+), 155
magnesium (Mg2+), iron (Fe3+), and manganese (Mn2+) are the common inorganic components 156
in leachates, which are released from the waste mainly by the biological and chemical processes 157
(Mandal et al., 2017; Mukherjee et al., 2014). 158
159
2.4. Xenobiotic organic compounds (XOCs) 160
XOCs are organic chemicals originated from anthropogenic sources such as household 161
waste and non-hazardous industrial waste. These compounds include a variety of aromatic 162
hydrocarbons, chlorinated aliphatics, phenols, pesticides, and pollutants like dioxins and furans 163
(Kjeldsen et al., 2002; Liu et al., 2008). XOCs have relatively low concentrations (usually less 164
than 1 mg L-1 of individual compound) in leachates. However, XOCs constitute a potential risk 165
to the receiving water (e.g., surface water and groundwater) if leachates are released into the 166
environment (Baun et al., 2004). 167
168
3. Factors affecting leaching behavior 169
170
3.1. Ash properties 171

9
Ash properties greatly depend on the operational condition (e.g., incineration temperature), 172
combustion technology, and composition of the waste input (Rendek et al., 2007). Leaching 173
behavior of MSWI ash can be influenced by ash properties such as the bulk content, particle 174
size distribution, and chemical speciation (Saqib and Backstrom, 2015). For instance, the 175
presence of DOM, sulfate (SO42–), and chloride (Cl–) imposes an effect on leaching of HMs 176
(Olsson et al., 2009; Zhao et al., 2017). DOM originated from MSWI bottom ash enhances Cu 177
leaching and has a large impact on the speciation of Cu in leachate by forming Cu-DOM 178
complexes (Meima et al., 1999; Olsson et al., 2007; Van Zomeren and Comans, 2004). It was 179
found that the Cu release would reduce if the TOC (e.g., fulvic acid) in ash decreased (Arickx 180
et al., 2010). The existence of SO42– stabilizes most of the HMs into the form of sulfate phases, 181
and therefore contributes to the stabilization of HMs (Verhulst et al., 1996). During incineration, 182
a higher content of Cl- in MSW enhances the volatility of HMs, leading to the lower 183
concentration of HMs in the bottom ash (Li et al., 2010). Meanwhile, the leaching behavior of 184
HMs is highly associated with Cl- or NaCl. This is because Cl- mobilizes HMs and facilitates 185
leaching of HMs (Olsson et al., 2009; van der Sloot et al., 2001; Weibel et al., 2018). 186
Particle size distribution also affects the leaching of pollutants. On one hand, some toxic 187
contaminants are preferentially concentrated in smaller particles (Table 1) as reported in 188
previous studies (Xia et al., 2017; Yao et al., 2014). For instance, MSWI bottom ash with size 189
fraction below 4 mm is the most contaminated one that contains potentially toxic elements (e.g., 190
Cu, Cr, Hg, and Pb), sulfates and chlorides (Tang and Steenari, 2016; Wang et al., 2002). On 191
the other hand, smaller particles possess larger specific surface areas, resulting in faster 192
leaching kinetics (Luo et al., 2017). The fine fractions of MSWI ash are therefore responsible 193

10
for the enhanced release of HMs and other salts (Abbas et al., 2003; Astrup, 2007). Separation 194
of fine fractions by sieving and washing is effective, which significantly reduces the 195
concentration and leaching of contaminants to be lower than the legal limit (Alam et al., 2017). 196
Table 3 presents the regulatory limits of various leaching elements from MSWI bottom ash in 197
different countries. As can be seen, different leaching criteria have been adopted to regulate the 198
leachate from bottom ash and the values are very different due to different scopes and 199
requirements from governments. 200
201
3.2. Liquid to solid ratio (L/S) 202
The effects of L/S ratio on leaching behavior have been intensively studied through 203
mathematical models and sequential and column/batch tests (Grathwohl and Susset, 2009; 204
Guyonnet et al., 2008; Quina et al., 2011). L/S ratio is a critical factor affecting leaching of 205
inorganic constituents (e.g., HMs) from MSWI ash as leaching of HMs is primarily controlled 206
by solubility. A higher L/S ratio promotes the dissolution of minerals and accelerates the release 207
of HMs (Mizutani et al., 1996) as shown in Fig. 2. Mathematical models were developed to 208
predict concentrations from column and batch experiments as a function of L/S, and results 209
indicated the mobility of metals was low under a low L/S ratio (Aberg et al., 2006). The effect 210
of L/S ratio on leaching is also element-specific. For example, leaching of some HMs such as 211
Ba, Pb, Sb, Zn, and V is mainly affected by L/S due to solubility-controlled release of these 212
HMs, i.e. high L/S leads to increased dissolution of mineral phase which governs the release 213
of target elements, resulting in a higher cumulative release (Allegrini et al., 2015a). In contrast, 214
As, Cd, Cu, Ni, and Se are mainly affected by the leaching procedure, e.g., the choice of column 215

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or batch test. 216
217
3.3. Impact of pH 218
It is well recognized that pH is one of the most important factors that controls the leaching 219
of both inorganic and organic constituents from solid phase into solution (Cornelis et al., 2012; 220
Komonweeraket et al., 2015). The pH values of fresh MSWI ash are often high (~10-12), which 221
can be attributed to the formation of Ca(OH)2 as incurred by its hydration reaction (Dou et al., 222
2017). Moreover, leaching of trace elements is highly pH-dependent, in particular, the 223
solubility of HMs is sensitive to pH (Table 4). The degree of sensitivity; however, differs among 224
different metal species (Quina et al., 2009; Yakubu et al., 2018). Most metal species (e.g., Ca, 225
Cu, Mn, Zn, Cd, and Pb) follow a cationic leaching pattern and the concentrations of leached 226
elements decrease with increasing solution pH (Fig. 3). In comparison, the maximum leaching 227
of Al and As is observed at extremely acidic (pH~2) and alkaline (pH~12) conditions, assigned 228
to an amphoteric leaching pattern. Se and Ba show an oxyanionic leaching pattern and the 229
highest concentrations are reached at a very alkaline condition (pH >11) (Zhang et al., 2016b). 230
Leaching of Na, K, SO42-, and Cl- is less pH dependent and the released amounts are mainly 231
availability-controlled (Quina et al., 2009). According to previous studies, the pH-dependent 232
leaching of elements, except As, Se, Ag, and Cr, is controlled by the dissolution of common 233
minerals from MSWI ash (Dijkstra et al., 2006a) or by the precipitation of their (hydr)oxides, 234
carbonate, and sulfate solids (Cetin et al., 2012). Sorption-controlled leaching mechanisms are 235
proposed for As, Se, and Ag, while Cr leaching is likely controlled by BaCrO4(s) minerals 236
(Komonweeraket et al., 2015). Aging may decrease pH of bottom ash from 10-12 to 8-8.5, and 237

12
thus leaching is inhibited (Meima and Comans, 1997). Leaching of HMs in weathered bottom 238
ash is presumably controlled by metal sorption to the newly formed Fe and Al (hydr)oxides 239
(Dijkstra et al., 2002; Meima and Comans, 1999). However, aging does not seem to 240
significantly affect leaching from fly ash (Zhang et al., 2016b). 241
242
3.4. Weathering and aging 243
Due to high temperature incineration process, MSWI ash is not thermodynamically stable 244
under atmospheric conditions. Weathering and aging treatments can remarkably change its 245
chemical and mineralogical characteristics by hydrolysis of Na, K, Al, and Ca, 246
dissolution/precipitation of hydroxides and salts, carbonation, neutralization of pH, 247
oxidation/reduction reactions, and formation of clay-like minerals (Meima and Comans, 1997; 248
Meima and Comans, 1998; Yao et al., 2010). These geochemical changes caused by weathering 249
alter the macroscopic properties, including pH and acid-neutralizing capacity, redox potential, 250
as well as sorption and ion-exchange capacity, of MSWI ash (Polettini and Pomi, 2004). 251
Meanwhile, the neo-formation of reactive and sorptive minerals in turn affects the leaching, 252
solubility and complexation of HMs in MSWI ash (Chimenos et al., 2005b). For instance, 253
weathering is able to increase the aluminum (hydr)oxides in bottom ash, thereby forming 254
reactive surface to keep HMs such as Cu and Zn from leaching (Dijkstra et al., 2006b). 255
During weathering process, carbonation is identified as the primary reaction and causes 256
the pH to decrease (Arickx et al., 2006; Santos et al., 2013). As a result, formation of calcite 257
and ettringite may stabilize HMs like Cu, Zn, Cr, Cd, and Pb, which are tightly encapsulated 258
into Ca/Si matrix by replacing Ca (Pandey et al., 2012; Zhen et al., 2013). Bottom ash is widely 259

13
co-disposed with refuse in the form of landfill cover or mixed landfill, while accelerated 260
weathering provides an effective pre-treatment to reduce its leaching (Su et al., 2013). In 261
addition, accelerated carbonation is also effective in washing out Cl- ions, and thus may permit 262
the utilization of treated bottom ash in construction applications where low chloride 263
concentration is required (Santos et al., 2013). A previous study showed that organic pollutants 264
(e.g., PAHs) in weathered MSWI bottom ash exceeded the local generic guidelines for sensitive 265
land use, but carcinogenic PAHs were strongly bound to the ash and did not release to the 266
environment during deposition on a small scale open air landfill (Johansson and van Bavel, 267
2003). Other studies showed that TOC in bottom ash and in the leachate decreased to 70% and 268
40%, respectively, after natural weathering, and further reduced to 55% and 25%, respectively, 269
after accelerated carbonation (Arickx et al., 2010). 270
271
3.5. Use of chemical reagents 272
Many different chemical reagents have been used in the leaching process of MSWI ash, 273
aiming to extract and recover HMs from ash, and therefore minimizing the metal load to 274
landfills. Among the reagents, mineral acids (e.g., HCl, HNO
3, and H2SO4) and alkaline 275
solutions (e.g., NaOH and aqueous NH3) are commonly used (Okada et al., 2007). Extensive 276
investigations show that the use of hydrochloric acid (HCl) is feasible to remove Cr, Cu, Zn, 277
and Pb from MSWI fly ash (Tang and Steenari, 2016; Weibel et al., 2018). HCl and HNO3 can 278
extract almost all metallic elements, but HCl is considered as a more practical and economically 279
feasible choice for ash leaching than HNO3. H2SO4 is able to dissolve many of the metals 280
except Ca and Pb, because their sulfates can precipitate as a secondary compound (Tang and 281

14
Steenari, 2016; Wu and Ting, 2006; Zhang and Itoh, 2006). Table 4 shows the leached amounts 282
of major and trace elements from MSWI bottom ash and fly ash using different acids. As can 283
be seen, the leaching behavior is significantly influenced by the types of acid and pH values. 284
Compared with mineral acids, organic acids such as formic acid, acetic acid, lactic acid, and 285
oxalic acid are not effective as leaching agents for metals, even though these acids can form 286
soluble metal complexes (Fedje et al., 2010). However, citric acid is an exception and it was 287
found to be quite effective, especially taking account of its environmental friendliness (Huang 288
et al., 2011; Wang et al., 2018). 289
Alkaline leaching was reported to selectively extract amphoteric metals (e.g., Zn and Pb) 290
from fly ash, leaving all other impurities (e.g., Fe, Mg, and Al) in the solid residue (Mizutani 291
et al., 1996; Nagib and Inoue, 2000). In addition, a combined use of HCl and NaCl exhibited 292
excellent leaching performance toward Cu/Zn (70-80%) and Pb/Cd (≥ 90%) since the 293
formation of metal-chloride-complexes (e.g., PbCl42-) increased the solubility of HMs (Weibel 294
et al., 2018). Similar results were observed when using seawater as the leachant. This may be 295
attributed to the fact that an increased ionic strength (NaCl) could decrease the surface 296
negativity of ash and release the metal ions from the solid surface into aqueous solution by the 297
electrostatic interaction (Cetin et al., 2012; Lin et al., 2017; Yin et al., 2018c). It should be 298
noted that the use of these hazardous chemicals has adverse effects, and therefore an eco-299
friendly approach for leaching of MSWI ash still needs to be developed in the future. 300
Besides, chelating agents for selective leaching of specific metals were studied. For 301
instance, EDTA could be an effective agent for extracting Ca, Cu, Zn, and Pb (Garrabrants and 302
Kosson, 2000; Youcai et al., 2002). NH4NO3 was an interesting alternative for selective 303

15
leaching of Cu. Other complexing agents like NTA, DTPA, DOM, and saponins were also used 304
(Hong et al., 2000a; Hong et al., 2000b; Olsson et al., 2009). In MSWI fly ash, leaching of 305
HMs by using HCl was dependent on pH, whereas the use of chelating agents for HMs leaching 306
was independent of pH. The maximum extraction of Cr, Cu, Zn, and Pb from fly ash by 307
chelating agents was achieved at a concentration of 0.3-1.0% (Hong et al., 2000b). 308
309
3.6. Use of bacteria and fungi 310
Microbial leaching (bioleaching) is based on the natural ability of microorganisms, i.e., 311
bacteria and fungi to transform solid compounds to a soluble and extractable form through the 312
production of organic or inorganic acids (Krebs et al., 1997). Bioleaching process is widely 313
used for extraction of metals in the mining industry and likely involves enzymatic 314
oxidation/reduction toward solid compounds or interactions with metabolic products (Wu and 315
Ting, 2006). Via this approach, it might be feasible to obtain metals from natural ores and 316
industrial residues that cannot be processed economically by conventional methods (Bosecker, 317
1997; Schinner and Burgstaller, 1989). While conventional techniques (e.g., thermal treatment, 318
chloride evaporation, and chemical leaching) may be applied for efficient removal or recovery 319
of metals from MSWI ash, bioleaching is considered as a cleaner production technology to 320
recover resources with lower costs and energy (Bosshard et al., 1996). 321
Researchers have used three main groups of microorganisms for bioleaching process, 322
which are autotrophic bacteria (e.g., Thiobacilli spp.), heterotrophic bacteria (e.g., 323
Pseudomonas spp. and Bacillus spp.) and heterotrophic fungi (e.g., Aspergillus spp. and 324
Penicillium spp.) (Krebs et al., 1997; Schinner and Burgstaller, 1989). These microorganisms 325

16
are isolated from natural environments and have high relative species abundance therein. Use 326
of these kinds of species for bioleaching process of MSWI ash should have minimum effects 327
on the environment or human health. Bioleaching of MSWI ash is challenging owing to the 328
alkaline nature and toxic heavy metal content of ash, both of which are detrimental to microbial 329
growth and bioleaching activity (Ramanathan and Ting, 2016). Therefore, fungal leaching is 330
more suitable than bacterial leaching in treatment of alkaline MSWI ash because fungi can 331
survive under a higher pH environment. However, operating costs of fungal leaching are 332
relatively higher due to the demand of organic carbon sources for their growth and organic acid 333
excretion (Burgstaller and Schinner, 1993). 334
The most commonly used fungus for bioleaching of MSWI ash is Aspergillus niger, which 335
was reported to produce organic acids (e.g., citric acid, oxalic acid, and gluconic acid) and 336
involve acidolysis as the leaching mechanism (Bosshard et al., 1996; Wu and Ting, 2006; Xu 337
and Ting, 2009; Yang et al., 2008). Bacteria commonly used in bioleaching of MSWI ash are 338
mainly acidophiles which require sulfur or Fe(II) for energy production and simultaneous 339
acidification of the alkaline ash to pH of 1-2 (Funari et al., 2017; Ishigaki et al., 2005). Fig. 4 340
shows that chemical leaching and bioleaching of MSWI fly ash resulted in comparable yields 341
toward Mg and Zn (>90%), Al and Mn (>85%), Cr (~65%), Ga (~60%), and Ce (~50%). 342
Surprisingly, chemical leaching exhibited the best extraction performance for Cu (95%), Fe 343
(91%), and Ni (93%), whereas bioleaching was more effective for Nd (76%), Pb (59%), and 344
Co (55%) (Funari et al., 2017; Wu and Ting, 2006; Yang et al., 2008). 345
Moreover, the organic matter in bottom ash can be oxidized by microbes during 346
respiration and its oxidation provides energy for microbial growth and releases carbon dioxide 347

17
(Zhang et al., 2004). The released CO2 then reacts with bottom ash, further supplementing the 348
carbonation treatment and affecting metals leaching (Rendek et al., 2006a; Rendek et al., 349
2006b). Importantly, the residual organic matter in ash has been shown to support the 350
development of microbial biofilms, which potentially may trap and immobilize the toxic metals 351
(Aouad et al., 2008). In essence, the microbial respiration plays a crucial role in stabilizing the 352
organic matter. Without it, DOM leaching can occur and consequently metal leaching is 353
enhanced through its complexation (Ilyas et al., 2015; van Zomeren and Comans, 2009). Thus, 354
bioleaching of metals from MSWI ash is largely associated with the microbial uptake of organic 355
matters in the ash. 356
357
4. Methodologies to study leaching 358
359
4.1. Leaching methods 360
Applications of MSWI ash depend mainly on its impact to the environment, which is 361
experimentally verified by performing a series of appropriate leaching tests. It is necessary to 362
develop application-oriented leaching tests with corresponding leaching criteria to preclude 363
discriminations between different applications, e.g., terrestrial applications vs. land reclamation 364
(Yin et al., 2018c). Leaching is defined as the dissolution process whereby a solid gets into 365
contact with a liquid (leachant), after which mass transport occurs between these two phases 366
(Lin et al., 2017). Leachate can be formed in this process by dissolution of the soluble 367
constituents from the solid phase into the leachant. Investigation of leaching is important, 368
which enables the evaluation of potential environment risks of MSWI ash during its practical 369

18
applications (Di Gianfilippo et al., 2018; Naveen et al., 2017; Shi and Kan, 2009). Many lab-370
based leaching methods have been developed to explore the leaching potential of MSWI ash 371
when it is applied in the field. These leaching methods are classified into several categories, 372
e.g., batch leaching tests, column leaching tests, and pH-static leaching tests (Yin et al., 2018c). 373
Based on different applications, the most commonly used leaching test procedures include the 374
batch leaching tests (EN 12457-1:2002 at L/S 2 and EN 12457-2:2002 at L/S 10) and the 375
column leaching tests (DD CEN/TS 14405:2004 and DD CEN/TS 14405:2017). 376
Batch leaching tests are static leaching method which generates chemical data at 377
equilibrium for mechanistic applications. The test conditions such as contact time (generally 378
24-48 h), L/S, and pH are fixed in batch tests. As a simple and fast test method, batch leaching 379
tests cannot provide information on release kinetics under dynamic conditions (Grathwohl and 380
Susset, 2009; Kalbe et al., 2008). On the contrary, column leaching tests are considered as 381
simulating the flow of percolating groundwater through a porous bed of the ash. Column tests 382
can be used to assess the environmental impact of ash, since the results are percolation-based 383
and are collected from a flow-through pattern similar to the field conditions (Lopez Meza et 384
al., 2010; Yin et al., 2018b). Column leaching tests account for contaminants washed out at 385
lower L/S ratios and changes of solubility-controlled phases in the ash (Di Gianfilippo et al., 386
2016b). Fig. 2 shows the effects of L/S ratio and leaching methods on the released amounts of 387
Cr, Cu, Zn, and Pb from two different MSWI fly ash samples. As can be seen, the cumulative 388
release in column tests is significantly lower than that in batch tests. Another type of leaching 389
tests to quantify the leachability of MSWI ash is the pH-static leaching tests, which intend to 390
establish the complicated leaching profile as a function of pH values as compared to batch 391

19
leaching tests under natural pHs (Dijkstra et al., 2006a; Yin et al., 2018c). 392
A good correlation between batch and column leaching tests has been found in previous 393
studies (Delay et al., 2007; Meza et al., 2008). Using these methods, the leaching performance 394
of MSWI ash from perspectives of instant leaching strength and the long-term impact can be 395
predicted (Astrup et al., 2006b). Column tests may provide more reliable information, but are 396
more time-consuming and require more equipment, thereby resulting in higher costs for testing 397
(Pecorini et al., 2017). As such, the selection of leaching methods should consider the balance 398
among the type of information needed, the scientific basis behind the procedure, and the time 399
frame of the test (Lopez Meza et al., 2010). For instance, when it comes to the handling of large 400
test portions, which is required to assess a representative sample volume, especially in the case 401
of coarse grained materials such as MSWI ash, column tests present an advantage compared to 402
batch tests (Kalbe et al., 2008). 403
404
4.2. Life cycle assessment 405
Life cycle assessment (LCA) is a methodology to assess potential environmental impacts 406
of a product or a system by accounting for the environmental exchanges (emissions, 407
consumption of reagents, and energy) over the entire life cycle of the product or system (Butera 408
et al., 2015). LCA has always been applied to quantify and compare the potential environmental 409
impacts related to recovery, utilization, and final disposal of waste materials (Finnveden et al., 410
1995; Kosson et al., 2002). As a comprehensive assessment methodology, LCA is capable of 411
identifying environmental benefits, problem shifting and breakeven points, and criticality for 412
management of MSWI ash (Allegrini et al., 2015b; Margallo et al., 2014). Using LCA, several 413

20
studies have assessed the environmental impacts of an ash management system and identified 414
the critical aspects such as toxic impact and time horizon (Birgisdóttir et al., 2007; Boesch et 415
al., 2014). The release of toxic pollutants from MSWI ash into the environment upon its 416
utilization or disposal has been addressed through LCA (Toller et al., 2009). 417
During leaching assessment, LCA studies mainly apply a generic approach to convert the 418
liquid to solid ratio (L/S) of a leaching test to a projected time frame, by employing basic 419
assumptions about the application scenarios (Allegrini et al., 2015a; Kosson et al., 1996). 420
Accordingly, a number of application scenarios for MSWI ash (e.g., landfill disposal and road 421
construction) were established and leaching data were included in LCA (Allegrini et al., 2015b; 422
Birgisdóttir et al., 2007; Wang et al., 2017). For example, researchers developed an LCA model 423
dedicated to road pavement, with and without the inclusion of MSWI bottom ash as aggregates. 424
Emission data in the LCA accounted for the cumulative release extrapolated at the relevant L/S 425
based on two-stage batch leaching tests (Birgisdottir et al., 2006). Others applied the result of 426
simple batch leaching tests at L/S of 2 or 10 to consider potential life cycle environmental 427
impacts of leaching. Data for leaching as a function of pH were also included in the LCA studies 428
(Mroueh et al., 2001). Hence, leaching behavior is critical for the outcomes of LCA studies on 429
utilization and management of MSWI ash. 430
431
4.3. Geochemical modeling 432
Geochemical models such as Visual MINTEQ, ORCHESTRA, and PHREEQC have been 433
widely used to predict metal release and further describe metal leaching from MSWI ash 434
(Astrup et al., 2006a; De Windt et al., 2011; Hyks et al., 2009; van der Sloot et al., 2017; Wang 435

21
et al., 2016a; Yin et al., 2018a; Zhang et al., 2008b). More specifically, Visual MINTEQ is a 436
geochemical thermodynamic equilibrium model capable of calculating equilibrium aqueous 437
speciation, adsorption, solid-phase saturation states, and precipitation/dissolution of metals in 438
leachate (Dijkstra et al., 2008; Zhang et al., 2008a). The ORCHESTRA modelling framework 439
embedded in LeachXS is commonly applied to evaluate the extent of equilibrium with regard 440
to mineral dissolution/precipitation and adsorption processes for a wide range of major and 441
trace elements (Dijkstra et al., 2006a). A two-step geochemical modeling in PHREEQC has 442
been employed to identify solubility-controlled minerals and investigate their interactions in a 443
water-percolated column leaching system (Hyks et al., 2009). Leaching of MSWI ash under 444
native and adjusted pH values was evaluated using the PHREEQC geochemical speciation-445
modeling program to identify possible solubility controls (Cornelis et al., 2012; Wang et al., 446
2016a). In the leachate, adsorption of metal ions onto organic matters (e.g., humic substances) 447
was calculated with the NICA-Donnan model (Kinniburgh et al., 1999). Overall, researchers 448
perform geochemical modeling to calculate the compositions of leachate in equilibrium with 449
solubility/sorption-controlled minerals and to calculate the leaching potential of elements under 450
different circumstances. These models thus provide a rapid estimation on metal release 451
potential of MSWI ash upon its application scenarios, such as landfill disposal and road 452
construction. In most cases, the modeling accuracy can be further improved by adjusting 453
various parameters as inputs in the models (Yin et al., 2018a), which may serve as a platform 454
to testify the effectiveness of different assumed variables within the boundary via comparison 455
with experimental leaching data. 456
457

22
4.4. Environmental impacts 458
So far, it still remains unclear that which toxicity test should be implemented in regulations 459
for MSWI ash. A simple and fast approach is to analyze its leaching components and compare 460
that to the existing toxicity data or standards in literatures. However, this method tends to either 461
overestimate the risk caused by individual element or underestimate the additive, synergistic 462
and antagonistic effect caused by coexisting compositions (Stiernström et al., 2011). Besides, 463
in some cases, the concentrations of toxic substances are below the detection limits. Hence, the 464
ecotoxicological tests using bioassays attract increasing concerns (Stiernstrom et al., 2014). 465
The different leachants used can affect the biological toxicity, since the final pH values of the 466
leachates vary from acidic to alkaline and therefore it is challenging for ecotoxicity study. 467
Researchers have observed higher toxicity for small particle size of MSWI ash (< 0.125 mm) 468
as compared to particle sizes less than 1 and 4 mm, which was attributed to the higher release 469
amount of HMs during leaching tests of smaller particle size of MSWI ash (Stiernstrom et al., 470
2014). Others evaluated the toxicity effect of solid MSWI ash on the germination of white 471
clover T. repens and results showed that dosing over 50% of ash to garden soil completely 472
inhibited the germination process (Ribé et al., 2014). In the same study, a considerably toxic 473
effect at 18-29% of leachate from MSWI ash was observed for water flea D. magna. The 474
leachate from an aged MSWI ash exhibited no toxicity toward algae P. subcapitata, indicating 475
that aging may decrease its toxicity. Further analysis of the chronic effect of the leachate 476
showed that the larvae development for copepod N. spinipes was inhibited and the hatching 477
time for zebra fish D. rerio was delayed (Stiernström et al., 2011). For ecotoxicity tests of 478
higher-order species and human health risk assessment, the bioaccumulation in individual 479

23
organism and the trophic transfer of trace elements in interspecies make it more difficult to 480
evaluate the toxic effect solely caused by MSWI ash and its leachate. Nevertheless, the existing 481
ecotoxicological tests demonstrate that MSWI ash may pose an environmental risk during 482
utilization. Further studies are necessary to investigate the long term leaching behavior, 483
ecotoxicology, health risk assessment, and bioaccumulation and to develop corresponding 484
evaluation methods. 485
486
5. Potential mechanisms behind leaching 487
488
5.1. Solubility control 489
Generally, solubility control is associated with the dissolution of metal oxides present in 490
the solid, such as aluminum oxide, iron oxide, and zinc oxide. As such, solubility control occurs 491
when the solution in contact with the solid is saturated with regard to the constituent species of 492
interest (Komonweeraket et al., 2015; Kosson et al., 1996). Studies showed that the leaching 493
of Cr and Cd in fly ash was controlled by a dissolution mechanism (Jiao et al., 2016), while 494
others demonstrated that the pH-dependent leaching of elements, except As, Se, Ag, and Cr, 495
was controlled by the dissolution of common minerals from MSWI ash (Dijkstra et al., 2006a) 496
or by the precipitation of their (hydr)oxides, carbonate, and sulfate solids (Cetin et al., 2012). 497
In comparison, leaching of Na, K, SO42-, and Cl- is less pH dependent and the released amounts 498
are mainly availability-controlled (Quina et al., 2009). Particularly, the surface layer of MSWI 499
ash particles contains a large number of readily leachable elements and can be more susceptible 500
to leaching in aqueous environment. While elements enriched in the cores are not directly 501

24
exposed to the aqueous environment, leaching of metals from the cores is diffusion-controlled 502
and dependent on the dissolution rates of surface layers (Gunasekara et al., 2015). 503
504
5.2. Sorption control 505
Sorption process controls the release of elements that exhibit sorptive affinity to the active 506
sites on the solid surface. Usually, the leaching of trace metals from MSWI ash is controlled 507
by sorption processes onto site surfaces, such as (hydr)oxides, organic matter, and clay (Zhao 508
et al., 2017). For example, sorption-controlled leaching mechanisms were previously proposed 509
for As, Se, and Ag in ash (Komonweeraket et al., 2015). The fate of Cr, Cu, Zn, and Pb in 510
leachate was controlled by the sorption or precipitation. To be specific, Cr leaching was likely 511
controlled by BaCrO4 minerals and Pb mainly existed as sulfate and phosphate in leachate. 512
Based on the results of leaching tests, Cu and Zn in fly ash have been proven to associate with 513
Ca-bearing compounds through precipitation and sorption (Jiao et al., 2016). Both sorption and 514
co-precipitation are the mechanisms for HMs immobilization, leading to the decreased 515
concentrations in the leachate (Cornelis et al., 2008). Understanding these mechanisms is a key 516
step for predicting the release of trace elements and quantifying the potential risks during 517
MSWI ash utilization and landfilling (Saqib and Bäckström, 2016). 518
519
6. Treatments to reduce leaching 520
Table 5 presents a summary of treatment principles and methods for MSWI ash. As can 521
be seen, treatment methods are classified into three groups: chemical and physical separation, 522
stabilization and solidification, and thermal treatment. 523

25
524
6.1. Separation techniques 525
In order to improve the quality of MSWI ash and to enhance its utilization, separation 526
processes are employed to minimize the leaching of constituents. Separation methods include 527
washing and electrochemical process (Lam et al., 2010). 528
529
6.1.1. Washing 530
Washing treatment aims to reduce the contents of soluble chlorides, salts, alkali, and HMs 531
in MSWI ash by using water or acid as the leachants (Kirby and Rimstidt, 1994). Previous 532
studies showed the removal of 72.8% of Na, K, Ca, and Cl at a L/S ratio of 10, and about 12.3% 533
of Cr was removed based on a water-extraction process (Jiang et al., 2009). Through a counter-534
current process with two washing steps and one rinse step, the release of HMs and chloride 535
after water-washing at a L/S ratio of 3 and reaction time of 1 h was found to significantly 536
decrease (Chimenos et al., 2005a). As a result, the leachability of the washed ash was below 537
the limits. Other studies demonstrated that the water-washing process as a pre-treatment can 538
remove most of the chlorides, leachable salts and amphoteric HMs from MSWI fly ash, 539
resulting in a lowered chlorine content in the washed ash (Chiang and Hu, 2010; Zhu et al., 540
2010). Acid-extraction process is able to extract the HMs from MSWI ash and further recover 541
them from the leachant solutions. This process depends mainly on the type of extraction solvent, 542
the pH and the L/S ratio (Mizutani et al., 1996). 543
544
6.1.2. Electrochemical process 545

26
Electrochemical process is a novel separation technique for MSWI ash, which also aims 546
to extract critical elements and reduce their leaching. This process applies an electric potential 547
to trigger the reduction/oxidation reactions on the surface of two electrodes, i.e., cathode and 548
anode. These electrodes are connected with a voltage source and placed in the sample area to 549
generate an electric field which helps the migration of contaminants (Acar and Alshawabkeh, 550
1993). Metals are deposited on the surface of cathode during this process, but the efficiencies 551
are usually low and a long remediation period is needed (Ferreira et al., 2005). The remediation 552
mechanisms include four steps, i.e., acidification, desorption, migration, and precipitation. 553
After electrodialytic remediation, significant leaching reduction of target elements (e.g., Cr, 554
Mn, Cu, Zn, Cd, Pb, SO42-, and Cl-) has been observed (Jensen et al., 2015; Kirkelund et al., 555
2015; Pedersen et al., 2005). Meanwhile, researchers proposed a treatment method combined 556
with washing and electrodialytic remediation, and the results showed that a great reduction of 557
HMs in fly ash was achieved (Ferreira et al., 2008). Using electrokinetic remediation, the 558
concentration of pollutants in the leachate was reduced by 31-83% (Traina et al., 2007). Acid 559
pre-treatment and prolonged reaction duration could enhance the removal of HMs via 560
electrokinetic remediation (Li et al., 2018). 561
562
6.2. Solidification and stabilization 563
Various additive, stabilizer and binder have been used to immobilize the hazardous 564
constituents present in MSWI ash during the solidification/stabilization (S/S) process. The 565
main purpose of this process is to minimize the solubility, leachability, and toxicity of 566
contaminants. The S/S treatment commonly includes accelerated carbonation, hydrothermal 567

27
solidification, and chemical stabilization (Dou et al., 2017; Jing et al., 2007b; Zhang et al., 568
2016a). 569
570
6.2.1. Accelerated carbonation 571
Accelerated carbonation is a thermodynamically favored process primarily derived from 572
natural weathering and aging, in which mineralogical characteristics change accordingly since 573
carbonates are formed via the reaction between CO2 and alkali like Ca(OH)2 (Rendek et al., 574
2006a). Two important effects of carbonation have been identified in previous works. The most 575
obvious effect was the pH neutralization of the alkaline MSWI ash, while the other was the 576
immobilization of major and trace metals by sorption to neo-formed clay minerals, e.g., Al-577
(hydr)oxides, calcite, and ettringite (Meima et al., 2002; Nilsson et al., 2016). After accelerated 578
carbonation, the release of soluble salts such as SO42- and Cl- was also reduced (Li et al., 2007), 579
and meanwhile TOC remarkably decreased both in the solid phase and in the leachate (Arickx 580
et al., 2010). Consequently, accelerated carbonation is an effective treatment to reduce the 581
leaching of MSWI ash (Ni et al., 2017; Van Gerven et al., 2005). 582
583
6.2.2. Hydrothermal solidification 584
Hydrothermal treatment is an attractive method for MSWI ash solidification because it is 585
capable of not only solidifying MSWI ash at a relatively low temperature (150-200°C) but also 586
treating MSWI ash on a large scale (Jing et al., 2010; Shi et al., 2017). After hydrothermal 587
processing, the leaching tests showed that the concentration of dissolved HMs from MSWI ash 588
was drastically reduced below the regulatory levels for the environmental quality standards 589

28
(Jing et al., 2007b; Pena et al., 2006). Studies reported that the concentration of metals in the 590
leachate was two orders of magnitude lower after hydrothermal treatment (Etoh et al., 2009). 591
Hence, hydrothermal solidification provides an effective way to stabilize HMs in MSWI ash 592
and to reduce their leachability (Hu et al., 2015). Also, hydrothermal treatment provides a 593
sustainable means to convert wastes into valuable products (Luo et al., 2018). For example, 594
using MSWI bottom ash as a starting material, new mineral phases like katoite, zeolite, and 595
tobermorite are formed upon hydrothermal treatment under an appropriate Ca/Al/Si ratio (Luo 596
et al., 2019; Penilla et al., 2003). These mineral products can be further used as sorbent 597
materials to remove organic pollutants and metal ions from water with no significant toxicity 598
risk (Luo et al., 2017; Wang et al., 2016b). As such, the hydrothermal treatment may have a 599
high potential for recycling MSWI ash on a large scale. 600
601
6.2.3. Chemical stabilization 602
Chemical stabilization is one of the main treatment technologies to reduce leaching, which 603
has shown satisfactory results for immobilization of toxic metals in MSWI ash. During this 604
process, inorganic and organic additives have been employed to convert metallic minerals that 605
are highly soluble into less soluble forms (Wang et al., 2015). The commonly used inorganic 606
additives are phosphates, silicates, sulfates, sulfides, and iron oxides (Li et al., 2014; Sun et al., 607
2010), while organic additives include chelating agents (e.g., EDTA, NTA, and DOM), 608
pyrrolidines, imines, carbamates, and thiols (Youcai et al., 2002). Recently, organic additives 609
have attracted increasing attention due to low cost and high tolerance for different 610
environments. Related research showed that MSWI fly ash after chemical stabilization 611

29
treatment could meet the landfill standards as evaluated by leaching tests (Quina et al., 2010). 612
In this aspect, the batch leaching tests (EN 12457-1:2002 at L/S 2 and EN 12457-2:2002 at L/S 613
10) and the column leaching tests (DD CEN/TS 14405:2004 and DD CEN/TS 14405:2017) are 614
widely adopted. Chemical stabilization of fly ash using sodium sulfide and thiourea converted 615
soluble and leachable toxic metals into non-leachable and insoluble forms (Youcai et al., 2002). 616
Colloidal aluminate oxide was another effective stabilizer to immobilize lead in fly ash, and a 617
high reduction ratio of 94.8% for the leachability was obtained (Huang and Lo, 2004). The 618
introduction of chemical reagents into cement mortar mixed with ash is also a feasible solution 619
to minimize leaching (Tasneem et al., 2017). Overall, chemical stabilization/solidification 620
followed by disposing of in landfills is the most adopted approach for the management of 621
MSWI ash (Vavva et al., 2017). 622
623
6.3. Thermal treatment 624
Thermal treatment aims to reduce the leachability of harmful residue constituents, remove 625
toxic compounds, minimize residue volume and produce useful materials. For MSWI ash, the 626
major thermal treatment methods are vitrification, melting/fusion, and sintering (Lindberg et 627
al., 2015). Vaporization is another type of thermal treatment method capable of vaporizing 628
harmful trace elements from MSWI ash. These thermal treatment processes are mainly 629
differentiated based on the characteristics of the process product, rather than the process itself. 630
For example, a glassy phase is produced in vitrification, whereas a crystalline or heterogeneous 631
product is formed in fusion process by melting MSWI ash. During sintering, the ash residues 632
are heated to achieve a reconfiguration of solid materials (Sabbas et al., 2003). 633

30
634
6.3.1. Vitrification 635
Vitrification is a process whereby the residue is melted with additives or other waste solids 636
to form a homogenous liquid phase, which is subsequently cooled to form an amorphous and 637
homogenous single phase glass. Herein, typical melting temperatures are between 1100°C and 638
1500°C (Quina et al., 2008). Vitrification is meant to separate specific elements from melted 639
ash through volatilization or to incorporate them in the glass phase, making harmful 640
constituents less likely to leach out. The main reactions related to ash treatment during 641
vitrification are incorporating the ash compounds into the glass matrix by chemical bonding or 642
encapsulation (Colombo et al., 2003). In MSWI fly ash, the main glass forming component 643
(SiO2) is generally lower than 35% by weight, which may require supplement of glass cullet 644
or pure additives to increase the silica content in glass. Studies showed that vitrification was 645
effective for reducing the toxicity of fly ash and simultaneously lowered the leaching of Cr, Cd, 646
and Pb (Park and Heo, 2002). It is worth noting that, although vitrification of MSWI ash has 647
been shown to be technically feasible, heterogeneous nature of MSWI ash with varied ch emical 648
composition is the fundamental difficulty in recycling the vitrified products. 649
650
6.3.2. Melting/fusion 651
Different from vitrification process, no additives are used in melting/fusion process and 652
the final product is a heterogeneous slag mixture, which consists of glassy material and 653
crystalline phases. Here, the melting temperatures are similar to those in vitrification process, 654
typically between 1200°C and 1600°C (Sakai and Hiraoka, 2000). Products generated by 655

31
melting of MSWI ash can be applied as construction materials for road or pavement 656
construction. The leaching behavior of the slag from ash melting systems is similar to that of 657
the vitrified slag produced in vitrification process, both are significantly lower than that from 658
the original MSWI ash (Lindberg et al., 2015). 659
660
6.3.3. Sintering 661
Sintering process is heat-induced coalescence and densification of porous solid particles, 662
which is operated below the melting points of their major components. Temperatures involved 663
in this process are typically in the range of 700-1200°C (Cheeseman et al., 2003; Wey et al., 664
2006). Sintering treatment can form a product with lower porosity and higher strength and 665
density compared to the original ash. For MSWI ash, the above changes are mainly caused by 666
solid phase transformation and recrystallization, solid-state reactions, and sometimes even 667
reactions with minor amounts of a liquid phase. After sintering, the leachability of harmful 668
components from the end-product decreases due to reduced porosity of the sintered ash (Sabbas 669
et al., 2003). Furthermore, toxic metals such as Cd, Pb, and Hg vaporize during the heat 670
treatment and are not detected in the sintered products (Nowak et al., 2010). Combined with 671
washing treatment, the thermal process can serve as a dechlorination procedure and enhance 672
the final Cl removal rate up to 90% (Yang et al., 2014). Among thermal treatment, sintering is 673
more sustainable because it not only minimizes the leaching behavior of MSWI ash, but also 674
produces various types of products such as concrete aggregates, ceramic tiles, and other solid-675
monolithic ceramics (Mangialardi, 2001; Zhang et al., 2007). 676
677

32
7. Conclusion remarks 678
Incineration is widely adopted in modern waste management because it provides an 679
effective way to minimize the MSW that needs to be disposed of in landfills (Hu et al., 2018; 680
Ouda et al., 2016). The ash residues are often disposed by landfilling or reused for various 681
applications. The widely concerning issues are the leaching of harmful elements. In particular, 682
the leaching behavior occurs continually in a very long time scale (tens and hundreds of years), 683
making it difficult to accurately predict the concentrations of pollutants and evaluate their 684
environmental impacts. This review summarizes extensive study on MSWI ash leaching and 685
treatment, specifically, pollutants generated through leaching, factors governing leaching, 686
methodologies to study leaching, leaching mechanisms, and treatments to reduce leaching. It 687
was concluded that hundreds of pollutants are generated through leaching from MSWI ash, in 688
which heavy metals and organic contaminants are the most toxic and concerned. Ash properties, 689
pH and L/S ratio are the main factors governing MSWI ash leaching. Leaching behavior of 690
MSWI ash is complicated and existing methods to evaluate leaching may not be able to 691
represent the field conditions. Solubility and sorption are the two major leaching mechanisms. 692
Many treatment methods have been proposed. Not all methods are effective and some 693
approaches are associated with high energy and high cost, which ma kes them less economi cally 694
feasible and attractive. In-depth investigations on the long term leaching behavior should 695
arouse enough attention. Furthermore, bacteria and fungi induced leaching is not well-696
documented and needs further study. 697
698
Acknowledgements 699

33
The authors gratefully acknowledge financial supports from the Environment Technology 700
Research Program (ETRP), National Environment Agency, Singapore (ETRP 1301 104), 701
National Science Foundation of China (21806140), China Postdoctoral Science Foundation 702
(Z741290007), Postdoctoral Science Foundation of Zhejiang Province (Z87129002) and Start-703
Up Grant from Zhejiang University of Technology (2017129004429). Dr. Bin Cao is gratefully 704
acknowledged for helpful discussion. 705
706
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48
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1245
1246

58
Ta bl e 1 Effects of particle size distribution on chemical compositions of
MSWI bottom ash (Samples collected from a local Waste-to-Energy
incineration plant in Singapore).
Component
Content (wt.%) as a function of particle size
< 0.30 mm 0.30~1.18 mm 1.18~6.30 mm > 6.30 mm
CaO 54.06 29.74 15.7 11.65
SiO2 8.24 23.13 49.43 61.22
SO3 7.89 2.27 0.61 0.27
Al2O3 6.85 9.22 6.81 6.03
P2O5 4.83 7.28 3.63 0.92
Fe2O3 4.59 15.77 10.83 5.35
Cl 3.69 2.81 0.69 0.34
Na2O 2.55 3.64 7.88 10.61
MgO 2.38 1.47 1.33 1.23
TiO2 1.75 1.23 0.44 0.24
K2O 1.37 1.43 0.99 1.09
ZnO 0.79 0.76 0.36 0.16
CuO 0.25 0.46 0.38 0.11
BaO 0.24 0.15 0.22 0.09
MnO 0.16 0.16 0.18 0.07
PbO 0.10 0.12 0.09 0.02
Cr2O3 0.08 0.17 0.22 0.41
SrO 0.07 0.05 0.07 0.03
SnO2 0.05 0.05 0.03 0.01
NiO 0.02 0.05 0.06 0.11
ZrO2 0.02 0.02 0.02 0.03
Total 99.98 99.98 99.97 99.99

59
Ta bl e 2 Concentration range of major and trace elements in MSWI
bottom ash and fly ash. Data obtained from reference (Lindberg et al.,
2015).
Element Bottom ash Fly ash
Minimum Maximum Minimum Maximum
P 1.4 (g/kg) 6.4 1.7 9.6
S 1 5 1.4 32
Cl 0.8 4.19 45 380
Si 91 308 36 190
Na 2.87 42 6.2 84
K 0.75 16 17 109
Ca 0.37 123 46 361
Mg 0.4 26 1.1 19
Al 21.9 72.8 6.4 93
Fe 4.12 150 0.76 71
Mn 0.083 2.4 0.2 1.7
Ti 2.6 9.5 0.7 12
Cr 23 (mg/kg) 3170 72 570
Co 6 350 1.9 300
Ni 7 4280 19 710
Cu 190 8240 16 2220
Zn 613 7770 4308 41000
Se 0.05 10 0.7 31
Mo 2.5 276 9.3 49
Sr 85 1000 80 500
V 20 122 4 150
Pb 98 13700 254 27000
Cd 0.3 70.5 16 1660
Hg 0.02 7.75 0.1 51
As 0.12 189 18 960
Ag 0.28 36.9 0.9 192
Sn 2 380 367 5900
Ba 400 3000 34 14000

60
Ta bl e 3 Regulatory limits of various leaching elements from MSWI bottom ash in different countries. Data
obtained from references (Chiang et al., 2014; Dou et al., 2017).
Country SO42-
mg/L Cl- Cr Cu Zn Pb As Cd Hg Ni Se Ba
China - - 300 2000 2000 100 100 20 2 100 20 2000
USA - - 50 - - 50 50 10 2 - 10 1000
Netherlands 3300 560 0.35 0.32 2.3 0.97 0.83 0.022 0.017 0.7 0.031 6
Denmark 8000 6000 1 4 3 0.2 0.1 0.08 0.002 0.14 - 8
Germany 6000 2500 2 3 3 0.5 - 0.05 0.001 0.4 - -
Italy 1250 1000 0.25 0.25 15 0.25 0.25 0.025 0.005 0.05 0.05 5
Spain 1000 800 0.5 2 4 0.5 0.5 0.04 0.01 0.4 0.1 20
France 10000 - 1.5 - - 10 2 1 0.2 - - -
Korea - - 12 200 - 400 20 12 - - - -
Belgium - - 0.1 2 0.9 0.2 0.1 0.1 - 0.2 - -

61
Ta bl e 4 The leached amounts of major and trace elements from MSWI
b
ottom ash and fly ash using different acids.
Results are given in % of total amount of each element in dry ash. Data obtained from reference (Tang and Steenari,
2016).
Leached amounts from fly ash (wt.%) Leached amounts from bottom ash (%)
Elements
HCl HNO3 H
2SO4 HNO3
pH=2 pH=3 pH=4 pH=2 pH=3 pH=4 pH=2 pH=2 pH=4
Na 90.0 88.6 84.0 100 100 92.4 81.3 23.5 13.2
K 97.5 87.6 85.5 82.3 74 71.2 82.7 23.5 9.5
Ca 51.8 47.1 33.0 40.5 37 26.2 5.7 58.0 37.3
Fe 17.4 4.5 - 9.5 9.0 3.4 24.5 30.9 8.5
Al 62.0 36.9 23.8 50.5 41 8.1 64.2 36.1 2.0
Si 51.5 20.3 7.3 51.5 39.5 9.5 50.5 18.8 3.2
Mn 52.2 29.1 16.4 37.5 30.5 14.9 60.5 49.5 19.4
Cu 67.1 28.1 9.5 47.5 30.0 5.0 50.9 30.2 5.3
Zn 74.2 69.0 56.1 65.0 63.4 56.5 80.0 39.0 23.7
Cd 93.3 82.7 78.3 77.0 - - 88.9 - -
Pb 34.9 8.9 8.3 9.2 9.0 7.1 2.2 2.4 -

62
Ta bl e 5 Summary of treatment principles and methods for MSWI ash (Lindberg et al., 2015;
Sabbas et al., 2003; van der Sloot et al., 2001).
Treatment principles Processes/methods
Chemical and physical separation Washing
Electrokinetic remediation
Eddy-current separation
Electrolysis
Chemical extraction/mobilization
Chemical precipitation
Crystallization/evaporation
Adsorption
Distillation
Ion exchange
Magnetic separation
Density and particle size based separation
Stabilization and solidification Chemical stabilization
Hydrothermal solidification
Accelerated carbonation
Aging/weathering
Using binders and additives
Thermal treatment Vitrification
Melting/fusion
Sintering
Vaporization/condensation

63
Figure Captions
Figure 1. The solid waste management system in Singapore (Data obtained from
National Environment Agency, Singapore).
Figure 2. Effects of L/S ratio and leaching method on the released amounts of Cr, Cu,
Zn and Pb from two different MSWI fly ash samples. Adapted from reference (Quina
et al., 2011).
Figure 3. Typical pH-dependent leaching behavior of Cr, Cu, Zn and Pb from different
sources of MSWI bottom ash. Adapted from references (Di Gianfilippo et al., 2016a;
Dijkstra et al., 2006a; Dou et al., 2017; Lin et al., 2015; Luo et al., 2017; Polettini and
Pomi, 2004; Stiernstrom et al., 2014; Xia et al., 2015).
Figure 4. Comparison of leaching yields (%) between chemical leaching and
bioleaching for major and trace elements from MSWI fly ash. Data obtained from
reference (Funari et al., 2017).

64
Figure 1. The solid waste management system in Singapore (Data obtained from
National Environment Agency, Singapore).

65
Figure 2. Effects of L/S ratio and leaching method on the released amounts of Cr, Cu,
Zn and Pb from two different MSWI fly ash samples. Adapted from reference (Quina
et al., 2011).
0.1 1 10 100 1000
0.1
1
10
100
0.1 1 10 100 1000
0.1
1
10
100
1000
0.1 1 10 100 1000
1
10
100
1000
10000
0.1 1 10 100 1000
1
10
100
1000
10000
Batch tests
Column tests
Total content in ash
Leached amount (mg/kg)
Cr
Zn
Batch tests
Column tests
Cu
Batch tests
Column tests
Leached amount (mg/kg)
L/S ratio
Batch tests
L/S ratio
Pb
Column tests

66
Figure 3. Typical pH-dependent leaching behavior of Cr, Cu, Zn and Pb from different
sources of MSWI bottom ash. Adapted from references (Di Gianfilippo et al., 2016a;
Dijkstra et al., 2006a; Dou et al., 2017; Lin et al., 2015; Luo et al., 2017; Polettini and
Pomi, 2004; Stiernstrom et al., 2014; Xia et al., 2015).
2468101214 2 4 6 8 10 12 14
2468101214 2468101214
Leached amount (mg/kg)
1.0E+02
1.0E+01
1.0E+00
1.0E-01
1.0E-02
1.0E-03
1.0E-04
Cr
1.0E+04
1.0E+03
1.0E+02
1.0E+01
1.0E+00
1.0E-01
1.0E-02
1.0E-03
1.0E-04
Cu
1.0E+04
Leached amount (mg/kg)
pH value
Zn
1.0E+02
1.0E+01
1.0E+00
1.0E-01
1.0E-02
1.0E-03
1.0E+03
pH value
1.0E+02
1.0E+01
1.0E+00
1.0E-01
1.0E-02
1.0E-03
1.0E+03
Pb

67
Figure 4. Comparison of leaching yields (%) between chemical leaching and
bioleaching for major and trace elements from MSWI fly ash. Data obtained from
reference (Funari et al., 2017).
0
20
40
60
80
100
Chemical leaching Bioleaching
Leaching yields (%)
Zn
SbNdNb
GaCu
Cr
Co
Ce
P
Mg
Al
0
20
40
60
80
100
V
SrPbNi
MoBa
AsTi
Si
Mn
Fe
Leaching yields (%)
Ca