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Municipal Solid Waste Incineration Residues


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The management of residues from thermal waste treatment is an integral part of waste management systems. The primary goal of managing incineration residues is to prevent any impact on our health or environment caused by unacceptable particulate, gaseous and/or solute emissions. This paper provides insight into the most important measures for putting this requirement into practice. It also offers an overview of the factors and processes affecting these mitigating measures as well as the short- and long-term behavior of residues from thermal waste treatment under different scenarios. General conditions affecting the emission rate of salts and metals are shown as well as factors relevant to mitigating measures or sources of gaseous emissions.
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Management of municipal solid waste incineration residues
T. Sabbas
, A. Polettini
*, R. Pomi
, T. Astrup
, O. Hjelmar
P. Mostbauer
, G. Cappai
, G. Magel
, S. Salhofer
, C. Speiser
S. Heuss-Assbichler
, R. Klein
, P. Lechner
(members of the pHOENIX working group on Management of MSWI Residues)
BOKU University Vienna, Department of Waste Management-Nussdorfer La
nde 29-31, A-1190, Vienna, Austria
University of Rome ‘‘La Sapienza’’, Department of Hydraulics, Transportation and Roads - Via Eudossiana 18, I-00184 Rome, Italy
Technical University of Denmark, Environment & Resources, DTU - Building 115, DK-2800 Lyngby, Denmark
DHI Water & Environment - Agern Alle
11, DK-2979 Ho
rsholm, Denmark
University of Cagliari, Department of Geoengineering and Environmental Technologies - Piazza D’Armi 1, I-09123 Cagliari, Italy
t, Institut fu
r Mineralogie, Petrologie und Geochemie - Theresienstrasse 41, D-80333 Munich, Germany
CheMin GmbH-Am Mittleren Moos 48, D-86167 Augsburg, Germany
Technical University of Munich, Department of Hydrochemistry - Marchioninistrasse 17 D-81377 Munich, Germany
Accepted 29 October 2001
The management of residues from thermal waste treatment is an integral part of waste management systems. The primary goal of
managing incineration residues is to prevent any impact on our health or environment caused by unacceptable particulate, gaseous
and/or solute emissions. This paper provides insight into the most important measures for putting this requirement into practice. It
also offers an overview of the factors and processes affecting these mitigating measures as well as the short- and long-term behavior
of residues from thermal waste treatment under different scenarios. General conditions affecting the emission rate of salts and
metals are shown as well as factors relevant to mitigating measures or sources of gaseous emissions.
# 2002 Elsevier Science Ltd. All rights reserved.
A working group named ‘‘pHOENIX’’ on the ‘‘Man-
agement of Municipal Solid Waste Incineration
(MSWI) Residues’’ was established as a result of a
workshop held in spring 2002 in Vienna, which dealt
with the practical problems, recent research findings and
solutions related to this topic. As we agreed, there are
numerous highly specific scientific articles as well as
some comprehensive studies and books with either in-
depth research or with a description of integrated waste
management in general terms or with specific MSWI
residues. However, what was missing was a short intro-
ductory overview of the management of residues from
thermal MSW treatment for operators, non-specialized
scientists and legislators. With this article, we hope to
fill the gap.
The pHOENIX working group is composed by the
following members: Peter Lechner (BOKU University
Vienna, Department of Waste Management, A), Tho-
mas Astrup (DTU Technical University of Denmark,
Environment & Resources, DK), Giovanna Cappai
(University of Cagliari, Department of Geoengineering
and Environmental Technologies, I), Holger Ecke
(Lulea University of Technology, SE), Soraya Heuss-
Assbichler (Ludwig-Maximilians-Universita
t, Institut
r Mineralogie, Petrologie und Geochemie, D), Ole
Hjelmar (DHI Water & Environment, DK), Anders Kihl
(Ragn-Sells Avfallsbehandling AB, SE), Ralf Klein
(Technical University of Munich, Department of
Hydrochemistry, D), Gabriele Magel (Ludwig-Max-
t, Institut fu
r Mineralogie, Petrologie
und Geochemie, D), Peter Mostbauer (BOKU Uni-
versity Vienna, Department of Waste Management, A),
0956-053X/02/$ - see front matter # 2002 Elsevier Science Ltd. All rights reserved.
Waste Management 23 (2003) 61–88
* Corresponding author. Tel./fax: +39-06-44-585-037.
E-mail address: (A. Polettini).
Franz Ottner (BOKU University Vienna, Department of
Waste Management, A), Alessandra Polettini (Uni-
versity of Rome ‘‘La Sapienza’’, Department of
Hydraulics, Transportation and Roads, I), Raffaella
Pomi (University of Rome ‘‘La Sapienza’’, Department
of Hydraulics, Transportation and Roads, I), Tamara
Rautner (BOKU University Vienna, Department of
Waste Management, A), Henrich Riegler (BOKU Uni-
versity Vienna, Department of Waste Management, A),
Thomas Sabbas (BOKU University Vienna, Department
of Waste Management, A), Stefan Salhofer (BOKU Uni-
versity Vienna, Department of Waste Management, A).
1. Introduction
The objective of integrated waste management is to
deal with society’s waste in an environmentally and
economically sustainable way. Under the framework of
integrated waste management, thermal treatment repre-
sents a valid option for reducing the amount of waste to
be landfilled, at the same time allowing for waste
hygienization. The relative importance of incineration
as opposed to other waste treatment and disposal
options, including mechanical/biological treatment and
sanitary landfilling, varies considerably from country
to country, depending on specific waste management
strategies as well as space availability for final land
The increasingly more stringent limits imposed in
recent years on atmospheric emissions from waste
incineration have produced a considerable shift from
the gaseous emissions to the solid residues of the pro-
cess. Thus, solid residues from thermal waste treatment
warrant significant environmental concern.
In our article, we specially focus on assessing the
environmental impacts resulting from residues from
thermal waste treatment, the treatment methods avail-
able to mitigate such impacts before and after either
utilization or final land disposal as well as the processes
and variables affecting the physical and chemical chan-
ges occurring for such residues at the utilization or dis-
posal site.
1.1. Integrated waste management
As depicted in Fig. 1, waste management systems
include all processes from waste generation to land-
filling, i.e.:
Waste generation: all processes which produce
waste during the production and distribution of
products (industry and commerce) or the con-
sumption of products (households);
Waste collection, including source separation
into different material streams;
Processing, including such steps as waste sorting,
dismantling of products (e.g. end-of-life electrical
and electronic equipment), and production of
Refuse Derived Fuel (RDF). All these steps serve
either to prepare waste for reuse or to suitably
modify waste characteristics with a view to final
land disposal;
Recycling: production of secondary materials
from waste, e.g. paper from waste paper, steel
from ferrous metal scraps etc.;
Waste treatment, including several technologies
such as thermal treatment, chemical treatment of
hazardous wastes, mechanical/biological treatment;
Waste utilization, covering all the utilization
options of waste after processing, e.g. use of
treated bottom ash for road construction, com-
post for agricultural applications or thermal uti-
lization of RDF; and
1.2. Treatment methods
Waste treatment methods strongly depend on the type
of waste. As far as municipal solid waste is concerned,
the different treatment options are aimed at recovering
materials and/or energy from the waste as well as at redu-
cing the overall amount and the impacts of waste to be
landfilled. In this framework, both mechanical/biological
Fig. 1. Integrated waste management system.
62 T. Sabbas et al. / Waste Management 23 (2003) 61–88
pretreatment and waste-to-energy incineration are sui-
table treatment options which should be combined in
order to meet the above mentioned targets.
In particular, with thermal waste treatment the fol-
lowing issues can be viewed as the main objectives:
to reduce the total organic matter content,
to destroy organic contaminants,
to concentrate the inorganic contaminants,
to reduce the mass and volume of the waste,
to recover the energy content of the waste and
to preserve raw materials and resources.
Hazardous wastes and sewage sludge are also often
treated using various thermal methods. Aside from
combustion, other thermal processes exist, including
pyrolysis, gasification, sintering, vitrification and melting.
The most common thermal treatment process for
MSW is incineration by mass-burn technology. Flui-
dized bed incineration and refuse derived fuel systems are
less common in municipal solid waste treatment. Fluidized
bed systems and multi-hearth furnaces are also widely used
for sewage sludge incineration, while major furnace types
for hazardous wastes incineration are grateless systems
such as a rotary kiln furnace, fluidized bed systems, com-
bustion chamber and multi-hearth furnace.
Non-thermal waste treatment methods consist of bio-
logical, chemical as well as physical treatment.
1.3. Input to MSWI
The quality and quantity of the MSWI input and
output are influenced by several factors:
Waste generators are households and, in addi-
tion, industrial or commercial sites.
Waste generation both in households and
industry is (theoretically) influenced by waste
prevention. In reality we can observe increasing
waste quantities. Onida (2000) reports that the
total annual production of industrial waste in five
major sectors (agriculture, mining, manufactur-
ing, municipal and energy production) increased
by 9.5% from 1990 to 1995 in the EU.
Separate collection exerts a strong influence on
the quantities and quality of waste for incine-
ration. For example, the separate collection of
small electrical appliances could reduce the Cu
content in MSWI bottom ash by up to 80%.
Through source separation of recyclables and
biogenic waste, the quantity of waste for treat-
ment is significantly reduced.
Residues from waste processing technologies
(e.g. sorting of plastics after separate collection)
and other materials can also be part of the input
to MSWI.
1.3.1. MSWI residues
As a result of the incineration process, different solid
and liquid residual materials as well as gaseous effluents
are generated. Approximately one-fourth of the waste
mass on a wet basis remains as solids. The volume of
residues corresponds to one-tenth of the initial waste
volume. Typical residues of MSWI by grate combustion
Bottom ash, which consists primarily of coarse
non-combustible materials and unburned organic
matter collected at the outlet of the combustion
chamber in a quenching/cooling tank.
Grate siftings, including relatively fine materials
passing through the grate and collected at the
bottom of the combustion chamber. Grate sift-
ings are usually combined with bottom ash, so
that in most cases it is not possible to separate the
two waste streams. Together bottom ash and
grate siftings typically represent 20–30% by mass
of the original waste on a wet basis.
Boiler and economizer ash, which represent the
coarse fraction of the particulate carried over by
the flue gases from the combustion chamber and
collected at the heat recovery section. This stream
may constitute up to 10% by mass of the original
waste on a wet basis.
Fly ash, the fine particulate matter still in the flue
gases downstream of the heat recovery units, is
removed before any further treatment of the
gaseous effluents. The amount of fly ash pro-
duced by an MSW incinerator is in the order of
1–3% of the waste input mass on a wet basis.
Air pollution control (APC) residues, including
the particulate material captured after reagent
injection in the acid gas treatment units prior to
effluent gas discharge into the atmosphere. This
residue may be in a solid, liquid or sludge form,
depending on whether dry, semi-dry or wet pro-
cesses are adopted for air pollution control. APC
residues are usually in the range of 2% to 5% of
the original waste on a wet basis.
Due to the volatilization and subsequent conden-
sation as well as concentration phenomena acting
during combustion, fly ash and APC residues bear
high concentrations of heavy metals, salts as well as
organic micro-pollutants. Iron scrap and other metals
are usually recovered from bottom ash and reused in
industry. In some European countries great efforts
are devoted to utilization of such residues. If utiliza-
tion is not possible due to regulatory constraints or
other reasons (such as a sufficient source of natural
raw materials), these residues have to be disposed in
an environmentally acceptable and economically sus-
tainable way.
T. Sabbas et al. / Waste Management 23 (2003) 61–88 63
1.4. Goal of landfilling
One objective of landfilling of waste, including MSWI
residues, is to remove from general circulation materials
and products that are no longer useful in any respect. It
is preferable to do this in a manner that ultimately
returns the basic constituents of the waste to the eco-
logical cycle, possibly after they have undergone chemi-
cal and/or physical reactions and transformations.
A second and equally important objective of waste
disposal is to ensure that the waste does not cause any
unacceptable short- or long-term impact on the
environment or on human health. Disposal methods
must ensure that this is accomplished in a sustainable
manner, i.e. without excessive and/or prolonged main-
tenance or operation requirements and without a pro-
longed need for aftercare.
The fulfillment of these objectives for MSWI landfills
will require a profound understanding and exploitation
of the short- and long-term behavior of the landfilled
MSWI residues. Based on this understanding, the design
and operation and to some extent also the location of
the landfill must be adapted to the inherent properties
of the largely inorganic MSWI residues to ensure that
long-term emissions of contaminants become or remain
environmentally acceptable. Landfills should be
designed to minimize the required lifetime of active
environmental protection systems, i.e. systems requiring
maintenance and/or operation. This means, for
instance, that a disposal strategy based on encapsula-
tion of the waste is not desirable since it merely post-
pones the impact and preserves the contamination
potential. In principle, this is true both for inorganic
contaminants (salts and metals) due to their intrinsically
conservative nature and for toxic organic micro-pollu-
tants, which are commonly regarded as persistent spe-
cies. Yet, the major environmental concerns in relation
to the short- and long-term impact of landfilling of
MSWI residues are connected with the risk of leaching
and subsequent release of potentially harmful sub-
stances, particularly inorganic salts and metals/trace
elements, into the environment. Gas production and
release may also be of some importance, even for MSWI
residues. Leaching of toxic organic compounds (espe-
cially PCDDs and PCDFs) is generally believed to be of
minor relevance due to their hydrophobic nature and
their low concentrations in residues from properly
operated waste combustion plants.
Regarding the time scales relevant to landfilling, dif-
ferent definitions can be used. The timeframes of inter-
est for landfilling can be classified on the basis of a
defined time scale, a connected activity or a dominant
In the following chapters we will refer to the defini-
tion based on the landfill activity. The basic idea is that
each human generation should take care of its own
wastes, without leaving future generations environ-
mental issues still to be resolved. This approach is
also used in the EU Landfill Directive (CEC, 1999),
which is described in Section 1.4.2. Thus, in this context
‘‘short term’’ relates to the timeframe within which
landfill operation and active aftercare (operations that
require maintenance, inspection and input of energy,
e.g. leachate and gas collection as well as leachate
treatment) are required to meet adequate environmental
protection levels. On the other hand, ‘‘long term’’
represents the timeframe within which the environ-
mental safety of the landfill no longer relies on active
protection systems, but is based on the controlled
release of contaminants at an environmentally accep-
table rate. The long-term period starts just after the
completion of active aftercare measures. The corre-
sponding time scales are shown in Fig. 2.
1.4.1. Disposal scenarios
As incineration residues are produced by high-tem-
perature processes, they are thermodynamically
unstable under ambient conditions. This renders incin-
eration residues highly reactive, especially under wet
conditions. This means that they change their mineral-
ogical and physico-chemical characteristics as well as
their leaching behavior as long as thermodynamic equi-
librium conditions with the surrounding environment
are attained. The specific environmental conditions
influence and change the leaching behavior and con-
taminant release from such materials during utilization
or final land disposal. To assess the discharge behavior
of a specific waste, it is necessary to take the specific
conditions (scenarios) into account. To arrive at a con-
clusion, the following methodology should be applied
(ENV 12920):
formulate the task and the sought-after solution,
specify the scenario,
evaluate the waste characteristics,
determine the influence of the scenario conditions
on the variation of waste characteristics over
time, as well as on their environmental behavior
model the environmental behavior of the waste
validate the model by calibration with the results
from laboratory tests and field experiments and
by comparing it to natural analogues.
Such a methodology will also help identify the most
appropriate mitigating measures to be undertaken
before, during or after utilization or final land disposal.
1.4.2. The EU Landfill Directive
Future landfilling of waste in Europe will be governed
to a large extent by the EU Landfill Directive (CEC,
1999), which was officially adopted on 16 July 1999. The
64 T. Sabbas et al. / Waste Management 23 (2003) 61–88
criteria for acceptance of waste at the different classes of
landfills are laid down in a Council Decision (expected
to be finalized in December 2002), which addresses the
requirements of Annex II to the Directive. The EU
Landfill Directive (LFD) distinguishes technically
between three main classes of landfills (landfills for inert
waste, landfills for non-hazardous waste and landfills
for hazardous waste), but only in terms of the con-
tamination potential of the waste and the environmental
protection measures required at each class of landfill.
The LFD does not include any landfill strategy or
guideline on the design and operation of landfills aiming
at the minimization of the period during which active
aftercare will be necessary. The LFD, however, to a cer-
tain extent does allow for the implementation of national
strategies and guidelines within the individual EU mem-
ber states. At a national level it will be possible to define
different sub-classes of non-hazardous waste to prevent
co-disposal of waste types with different properties (e.g.
organic biodegradable waste and inorganic mineral
waste) and different short- and long-term behavior.
2. Processes and factors
At the utilization/disposal site (hereinafter referred to
as the application site), MSWI residues will undergo a
number of processes, which will cause a set of modifi-
cations in the waste matrix at the micro-structural level.
On a macroscopic scale, the combination of the differ-
ent processes will result mainly in the following effects:
leachate production
gas production and
temperature development.
Understanding the mechanisms governing the pro-
cesses under concern and the influence of the main fac-
tors on the processes themselves will allow for the
estimation of the potential environmental impacts aris-
ing from the utilization/disposal of MSWI residues as
well as of the measures to be undertaken in order to
mitigate the extent of such impacts.
The main processes and factors of concern affecting
the utilization or disposal of MSWI residues are
strongly interrelated, so that in many cases a separate
description of each process and factor, neglecting such
interrelations, will not be exhaustive. For this reason,
the following section will discuss the main processes and
factors on the basis of the above mentioned macro-
scopic effects.
2.1. Leachate production
Leaching can be defined as the dissolution of a soluble
constituent from a solid phase into a solvent. Leaching
occurs as a consequence of the chemical reactions tak-
ing place at the scale of the individual waste particles as
well as of the contaminant transport processes via the
fluid moving through the solid particles. As far as
MSWI residues disposal is concerned (see Fig. 3), the
transport medium of pollutants is mainly represented by
water, so that the overall water balance will determine
the actual amount of water reaching the application site.
Climatic conditions and vegetation (e.g. precipitation,
solar radiation, temperature, interception, evaporation,
evapotranspiration, wind, etc.) as well as the type and
morphology of the surface soil are among the main
variables to be accounted for in the water balance. The
application site itself then modifies the water infiltration
pattern as a result of the physical and hydrological
Fig. 2. Classification of time scales relevant to landfilling.
T. Sabbas et al. / Waste Management 23 (2003) 61–88 65
characteristics of the material. Thus, the discharge pattern
also depends on the pore type, pore distribution, homo-
geneity, permeability and field capacity of the material as
well as on the presence of preferential flow paths.
Findings by Johnson et al. (1999) show that for a
four-year-old MSWI landfill in the presence of pre-
ferential flow paths, the discharge is characterized by
long periods of low and nearly constant flows inter-
spersed with increases in discharge quantity in response
to rain events. Isotope studies and tracer methods in
combination with a simple dilution model show that
fractions of rainwater, which pass through the landfill
with little interaction with the material due to pre-
ferential flow paths, make up 20–80% of the discharge
volume during summer rain events and around 10% in
winter months.
Other studies (Brechtel, 1984; Stegmann and Ehrig,
1989; Ehrig, 1990; Kru
mpelbeck, 2000) show that the
observed leachate amounts for uncovered or sparely
vegetated MSW landfills in middle Europe lie between
15 and 60% of the annual precipitation. Approximately
the same fraction of precipitation is observed for MSWI
residues as annual leachate volume (Table 1, references
herein). Based on experimental data on the influence of
the climate and vegetation on the landfill water balance
(Baumgartner and Liebscher, 1996; Lerner, 1997;
B.A.L., 2001) and given the hypothesis that climate-
specific vegetation will develop at each landfill site, a
prediction of long-term leachate production for selected
Austrian sites was carried out (Table 1, 7th and 8th
line). Water balance models, including e.g. HELP and
BOWAHALD, can also be used to analyze the effect of
different vegetation/covering scenarios on leachate gen-
eration (e.g. Berger and Dunger, 2001).
Together vegetation and physical barriers (top cover,
liners) reduce the amount of leachate from the landfill
but cannot completely prevent leachate formation over
a long time scale (refer to the following sections for
more details). Once the overall water balance of the
application site is calculated, the leachate quality needs
to be estimated.
Mobilization of constituents from inorganic wastes
into the leaching medium is the result of the interaction
between chemical and physical factors. Chemical factors
include waste composition and mineralogy, tempera-
ture, pH, redox potential and the presence of ligands,
while physical factors are represented by specific surface
area, particle size, L/S ratio, porosity, hydraulic gra-
dient and hydraulic conductivity. Some physical factors
also affect the percolation pattern (advection, diffusion)
and hence the modes of contact between leachate and
waste, which can be caused by leachate flowing around
the waste, leachate flowing through the waste or by a
combination of the two.
The processes and factors relevant to leaching can
also vary depending on the contaminant under concern;
Fig. 3. Schematic layout of water balance and geochemical processes and factors affecting the discharge and pollutant flux from a landfill containing
residues from thermal waste treatment (modified after Sabbas et al., 2001a).
66 T. Sabbas et al. / Waste Management 23 (2003) 61–88
in particular, for MSWI residues different groups of
contaminants can be identified, including metal ions,
amphoteric metals, oxyanionic species as well as salts,
which display typical leaching patterns. The total con-
tent of such contaminants can even be considerably dif-
ferent for the various residues from waste incineration,
as shown in Table 2. However, the extent of con-
taminant release from waste materials is rather a func-
tion of the so-called availability for leaching, which
represents a fraction of the total content of con-
taminants in the waste itself. When a fluid flows through
a loosely packed granular waste material, the amount of
contaminants released is dictated by solubility con-
straints, so that the leaching process is referred to as
being solubility-controlled. In the case of very soluble
mineral phases, which can completely dissolve as a
result of contact with the fluid, leaching is generally
defined as availability-controlled.
Conversely, as far as compacted granular materials or
treated (e.g., solidified) wastes are concerned, leachate
percolation occurs at the surface of the solid material,
causing molecular diffusion to be the dominant process
determining contaminant release from the waste; in this
case, the leaching process is said to be diffusion-controlled
Table 1
Observed leachate amount for lysimeters/landfills of inorganic waste and estimated long-term scenario
Description of landfill or lysimeter (or cover), location, reference Annual precipitation
(PPT, mm/a)
(% PPT or mm/a)
Pilot plant landfill containing MSWI residues and other inorganic
waste South of Sweden (Marques and Hogland, 1999)
About 600 53% Observed
Slag/ash landfill Fladsa, Denmark (Nolting et al., 1995) 468–635 53–61% Observed
Lysimeters, bottom ash or mixtures of different incineration residues
Vienna, Austria (Lechner et al., 1997)
550 22–36% (vegetated)
48–60% (uncovered)
Lysimeter, MSWI bottom ash, uncovered Innsbruck, Austria
(Lechner et al., 1997)
1000 to 1300 62–67% Observed
Lysimeter, steel smelter slag, uncovered Southern alpine region, Austria
(Lechner et al., 1997)
950 65% Observed
Prediction for scots pine forest vegetation Vienna, Austria 520 60 mm/a (10%) Predicted
Prediction for spruce forest vegetation Alpine regions, Austria 800–900 200–250 ( 25%) Predicted
Table 2
Ranges of total content of elements in MSWI residues (from IAWG, 1997)
Concentration (mg/kg)
Element Bottom ash Fly ash Dry/semi-dry APC residues Wet APC residues
Al 22,000–73,000 49,000–90,000 12,000–83,000 21,000–39,000
As 0.1–190 37–320 18–530 41–210
Ba 400–3000 330–3100 51–14,000 55–1600
Ca 370–123,000 74,000–130,000 110,000–350,000 87,000–200,000
Cd 0.3–70 50–450 140–300 150–1400
Cl 800–4200 29,000–210,000 62,000–380,000 17,000–51,000
Cr 23–3,200 140–1100 73–570 80–560
Cu 190–8200 600–3200 16–1700 440–2400
Fe 4,100–150,000 12,000–44,000 2600–71,000 20,000–97,000
Hg 0.02–8 0.7–30 0.1–51 2.2–2300
K 750–16,000 22,000–62,000 5900–40,000 810–8600
Mg 400–26,000 11,000–19,000 5100–14,000 19,000–170,000
Mn 80–2400 800–1900 200–900 5000–12,000
Mo 2–280 15–150 9–29 2–44
Na 2800–42,000 15,000–57,000 7600–29,000 720–3400
Ni 7–4200 60–260 19–710 20–310
Pb 100–13,700 5300–26,000 2500–10,000 3300–22,000
S 1000–5,000 11,000–45,000 1400–25,000 2700–6000
Sb 10–430 260–1100 300–1,100 80–200
Si 91,000–308,000 95,000–210,000 36,000–120,000 78,000
V 20–120 29–150 8–62 25–86
Zn 610–7800 9000–70,000 7000–20,000 8100–53,000
T. Sabbas et al. / Waste Management 23 (2003) 61–88 67
leaching. Due to the reduced surface area of the material
in contact with the leachate and to the extremely slow
release of contaminants from the waste, chemical equili-
brium between the solid and the liquid phase is generally
not attained in the case of diffusion-controlled release.
Again, the mechanisms controlling the leaching pro-
cess are therefore dependent on physical factors per-
taining both to the properties of the waste material
(including particle size distribution, porosity, degree of
compaction and permeability) and to the related fluid
flow characteristics (including percolation rate, percola-
tion pattern and the amount of leachate contacting the
waste). They are also dependent on chemical factors
related to the solubility of contaminants in the waste
Leachate composition is the result of reaction between
the various mineral phases in the waste and the leaching
fluid. The leachability of strongly soluble species (e.g.,
alkali salts) is almost pH-independent, whereas for a
number of contaminants a clear pH-dependence can be
observed. The influence of pH on the leaching of con-
taminants is strongly related to the nature of the parti-
cular contaminant under concern as well as the mineral
phase(s) in which this is bound. Three main typical
leaching behaviors for solubility-controlled leaching
have been identified:
cation-forming species and non-amphoteric
metal ions (e.g. Cd), see Fig. 4a),
amphoteric metals (including Al, Pb, Zn), see
Fig. 4b), and
oxyanion-forming elements (e.g. As, Cr, Mo, V,
B, Sb), see Fig. 4c).
The concentration of cation-forming species and non-
amphoteric metal ions displays fairly constant high
values at pH< 4, and decreases strongly up to pH 8 to 9,
remaining approximately constant or slightly increasing
for higher pH values. Amphoteric metals exhibit
increased solubility under both strongly acidic and
strongly alkaline conditions, resulting in a V-shaped
solubility curve. For oxyanion-forming elements usually
solubility decreases in alkaline ranges (pH> 10).
It should be emphasized that the shape of the actual
solubility curve is the result of complex competing che-
mical equilibria where common ion effects can sig-
nificantly alter the theoretical concentration calculated
for pure aqueous solutions. For example, the equili-
brium concentration of Ba in pure water (20
C, pH=7,
no CO
dissolved) saturated with barite (BaSO
) is 1.3
mg/l. In the presence of gypsum (CaSO
O), more
sulfate will dissolve, and Ba concentration will decrease
to 0.01 mg/l as a consequence of the law of mass action,
as indicated by the arrows in Fig. 5.
As shown in Figs. 4 and 5, depending on the specific
leaching behavior, critical pH regions can be identified
where minimum or maximum solubility for the indivi-
dual contaminants is attained. In light of this, a matter
of major concern is to predict the pH conditions which
are likely to occur at the application site. These depend
on the characteristics of the leaching fluid as well as on
the properties of the waste. Probably the most relevant
waste property affecting the pH of the leachate is
represented by the acid or base neutralization capacity
(ANC/BNC). ANC and BNC are measures of the abil-
ity of a system to neutralize the influence of acids or
bases. In the case of MSWI residues, which are most
often basic in their nature, alkalinity of the material is
the relevant parameter, so that ANC is the appropriate
measure of neutralization capacity. ANC assesses the
sensitivity of the material itself to external influences
and/or internal stresses (e.g. mineralization, organic
matter degradation). As a consequence, the buffering
capacity of the material affects the evolution of the pH
of the leachate over time, thus allowing the expected pH
range for the application site to be estimated. Fig. 6
depicts the ANC of a number of bottom ash and fly ash
samples from Italian municipal solid waste incinerators
(Polettini et al., 2001).
In the case of alkaline MSWI residue, the reduction in
the buffering capacity of the material over time is rela-
ted to the depletion of alkalinity, which occurs as a
consequence of progressive leaching. At the time of dis-
posal, MSWI residues will display their maximum alka-
linity level. The level will decrease as the material comes
into contact with the leachate and dissolved alkalinity is
removed from the system by the leachate. As a con-
sequence, the residual alkalinity at any time will depend
on the initial alkalinity of the material, the dissolution
of alkalinity at various pH values in the leaching sce-
narios and the infiltration through the application site.
On the other hand, dissolved alkalinity depends on the
solubility of a number of minerals (Ca(OH)
, CaCO
etc.) and thus on the leaching system pH. The pH in
turn is dependent on the system’s ability to buffer the
infiltrating leachate, i.e. on the amount of residual
alkalinity in the system itself (Astrup et al., 2001).
Other than pH, the amount of leachate that comes in
contact with a given amount of waste, usually expressed
through the so-called liquid-to-solid (L/S) ratio, also
affects the leaching behavior, especially in the case of
solubility-controlled leaching. The L/S ratio is the result
of climatic conditions, hydrology and hydrogeology of
the application site, as well as the physical charac-
teristics of the waste material.
Solubility-controlled leaching is characterized by an
approximately linear dependence of cumulative release
on the L/S ratio. In some cases the linear trend of
cumulative release of a given element as a function of
L/S can be altered by the presence of other species.
Delayed release is observed when a sparingly soluble
phase controlling solubility is present and is depleted
68 T. Sabbas et al. / Waste Management 23 (2003) 61–88
Fig. 4. Cd (a), Al (b) and B (c) concentration in eluates and leachate samples of fresh and aged ash (~=solidified MSWI residues; * MSWI
bottom ash; & MSWI bottom ash + other ashes; MSWI residues (mixed)) (Sabbas et al., 2001b).
T. Sabbas et al. / Waste Management 23 (2003) 61–88 69
after a relatively short period. In this case, the slope of
the curve of cumulative release versus L/S is lower for
low L/S ratios, and it increases at higher L/S ratios
where the sparingly soluble species are depleted. Con-
versely, enhanced initial release can be observed in the
presence of complexing agents, which increase the solu-
bility of the element under concern. In this case, a tran-
sition from a higher to a lower curve slope occurs with
increasing L/S.
On the other hand, for availability-controlled leaching
the amount of contaminants released into the solution is
at its maximum level due to the high contaminants
solubility and is not dependent on solution pH. At a
given L/S ratio, the transition from solubility-controlled
to availability-controlled leaching is evidenced by a
constant concentration in solution with decreasing pH.
Availability-controlled leaching results in rapid washout
of the soluble constituents at low L/S ratios, so that the
available amount often is attained at L/S values of 1 to 2;
for higher L/S ratios, the cumulative release remains at
this maximum. Typical examples of such leaching beha-
vior are Na, Cl and K (see Fig. 7, which depicts the results
from upflow percolation tests on weathered materials).
Leaching from compacted granular residues or
monolithic forms is neither solubility- nor availability-
controlled but could rather be ascribed to molecular
Fig. 5. Calculated Ba equilibrium concentration compared to eluates from aged or neutralized MSWI residues (A=BaSO
in pure water at 20
B=equilibrium after addition of gypsum; C=equilibrium after addition of calcite, CO
(0.04%) and NaCl (0.05 M); D=equilibrium after addition
of calcite, CO
(0.04%) and NaCl (0.05 M) and gypsum) (calculation: PHREEQC-2).
Fig. 6. Acid neutralization capacity of MSWI bottom ash (BA) and fly ash (FA).
70 T. Sabbas et al. / Waste Management 23 (2003) 61–88
diffusion and surface dissolution mechanisms. In this
case, leaching is kinetically controlled by the rate of
contaminant release via diffusion, which is measured
through the effective diffusion coefficient. Release
mechanisms and physical and chemical retardation
factors affect the diffusion process. Among the physical
retardation factors, porosity, pore structure, degree of
compaction and tortuosity can significantly slow the
rate of contaminant release. Pore solution pH and
solubility of elements/species as a function of pH can
influence the extent of sorption or co-precipitation
reactions on solid surfaces, thus acting as chemical
retardation factors.
Irrespective of the mechanism controlling leaching,
additional factors including the presence of sorbing/
complexing agents, redox reactions and the occurrence
of processes causing mineralogical changes over time
(e.g. due to aging/weathering) can also affect the
extent of contaminant release, as qualitatively illu-
strated in Fig. 8. Among the processes capable of
altering the leaching behavior of the material, sorption
includes different mechanisms of adsorption, ion
exchange, surface complexation and electrostatic
attraction of ions at the surface. During weathering of
less stable phases, new minerals with high surface
areas are formed. For instance, oxidation of iron in
Fig. 7. Leaching of Na, K and Cl from weathered MSWI bottom ash as a function of the L/S ratio.
Fig. 8. Influence of different processes on contaminant solubility as a function of pH (modified after van der Sloot et al., 1999).
T. Sabbas et al. / Waste Management 23 (2003) 61–88 71
MSWI bottom ash leads to the formation of iron oxi-
des, goethite (FeOOH) and hydrous ferric hydroxide
, often termed HFO). The resulting finely
grained phases are able to sorb heavy metals, including
Pb, Cd, Zn, Ni, Cr(III) and Cu, as well as Mo. Similar
sorptive properties are also displayed by other mineral
phases, including aluminum (hydr)oxides and amor-
phous aluminosilicates.
In the case of HFO, the general surface complexation
reaction describing sorption of divalent cations can be
simplified as:
þ Me
() Fe-OMe
þ H
where the symbol indicates bonds at the surface and
represents [Fe(OH)
Fig. 9 demonstrates the demobilizing effect and limits
of sorption. A number of sorption experiments were
carried out on pre-washed (L/S=10 achieved through
column percolation) and artificially weathered (by
means of carbonation) MSWI bottom ash ( < 2 mm).
The aged MSWI bottom ash was added with a nickel
sulfate solution at Ni concentrations varying between
0.06 and 16.4 mg/g, thereby continuously aerated to
achieve equilibrium with atmospheric CO
. The experi-
mental sorption isotherms (see Fig. 9) revealed strong
sorption phenomena, as long as the total amount of Ni
in the system was low. Sorption proceeded as long as
active sorption sites were available within the solid
material. In this case, the Ni concentration in the solu-
tion was lower than that predicted based on Ni(OH)
solubility. For high total contents of Ni, the amount of
Ni exceeding the sorption capacity of the material was
such that saturation or slight over saturation of the
solution with respect to Ni(OH)
was attained.
The presence of complexing agents can also sig-
nificantly alter the extent of contaminant leaching from
MSWI residues. Complexing agents can be either
organic or inorganic in their nature; dissolved organic
carbon (DOC) and chloride are the main complexing
agents of concern for such materials. DOC has been
extensively shown to be responsible for increasing cop-
per release from predominantly inorganic waste forms
(IAWG, 1997; Van der Sloot et al, 1999; Van der Sloot
et al., 2001).
Oxidation/reduction reactions also play a role in
determining the release of contaminants from MSWI
residues. The main oxidizing or reducing agents of rele-
vance for MSWI residue monofills are reported in
Table 3.
Leaching of contaminants from waste incineration
residues can be affected by the redox conditions accord-
ing to two main mechanisms. One mechanism relies on
the different solubility and toxicity of the contaminants
under concern for MSWI residues depending on their
oxidation state. These issues influence both the strength
of the leachate and the related potential environmental
Table 3
Relevant reducing/oxidizing agents (inorganic landfills)
Reducing agents H
; metals (Al, Fe, Zn); Fe-II metals, Fe-II
Oxidizing agents O
Fig. 9. Ni sorption for weathered MSWI bottom ash with Ni added at different concentrations.
72 T. Sabbas et al. / Waste Management 23 (2003) 61–88
impact. For example, it is well known that in an alkaline
environment Cr(III) may be rapidly oxidized by atmo-
spheric oxygen to Cr(VI), which is much more toxic and
mobile than Cr(III). However, for Cr(III) to be oxidized
to Cr(VI), high values of the redox potential are
The second mechanism through which the redox con-
ditions affect leaching is related to the fact that the sta-
bility of the mineral phases capable of immobilizing
metal ions through precipitation and/or sorption phe-
nomena is dependent on the oxidation/reduction poten-
tial. Thus, Fe(III) and Mn(IV) (hydr)oxides can be
transformed into more soluble forms of Fe(II) and
Mn(II) under moderately reducing conditions. Under
severely reducing conditions, S(VI) is reduced to ele-
mental sulfur and sulfide, resulting in the precipitation
of metal sulfides, which are among the less soluble metal
The above mentioned mechanisms can lead to either
synergistic or antagonistic interactions, so that the
influence of redox processes on leaching may result in
either mobilization or demobilization of contaminants
(see Table 4 and Fig. 8). It should be emphasized that
the observations presented in Table 4 may not reflect a
real landfill scenario, but merely indicate a significant
influence of redox processes on leaching.
Under disposal conditions, redox reactions can occur
as a result of either microbiologically mediated pro-
cesses due to the presence of organic material or abiotic
transformations leading to the formation of reducing
gases (H
). For bottom ash monofills, the presence of
unburned organic material and H
generally leads to
reducing conditions; in such cases, the leaching behavior
of contaminants is the result of, on the one hand, com-
plexation by DOC and, on the other hand, precipitation
of less soluble species, including for example insoluble
Weathering is a process, which naturally occurs in
incineration residues as a consequence of several factors
such as pH, redox potential, temperature and humidity
conditions as well as the concentration of certain com-
ponents (e.g. CO
) in the application site. Weathering
results in the occurrence of slow mineralogical changes
over time, which may alter the leaching of trace metals
from the material either in the medium or in the long
term. Due to weathering and the related neoformation
of minerals, key factors such as pH are subjected to
changes over time.
Weathering of bottom ash is a process, which deserves
particular concern. Incinerator bottom ash is composed
of high-temperature solids formed as a consequence of
rapid quenching of the material exiting the combustion
chamber, many of which are metastable under natural
conditions. Typically, such solids will thereby undergo a
number of chemical reactions while in the landfill lead-
ing to more stable mineral phases or phase assemblages
(Meima and Comans, 1997; Zevenbergen and Comans,
Weathering is the result of a complex series of several
interrelated processes, including hydrolysis, hydration,
dissolution/precipitation, carbonation, complexation
with organic and inorganic ligands, surface complexa-
tion, surface (co)precipitation, sorption, and formation
of solid solutions as well as oxidation/reduction (Belevi
et al., 1992; Bode
nan et al., 2000; Meima and Comans,
1997; Meima and Comans, 1999; Zevenbergen and
Comans, 1994). All of the mineralogical and chemical
changes caused by such processes are also accompanied
by physical changes such as pore cementation, changes
in grain size and pore size distribution, which in turn
alter the hydrological characteristics of the material.
Hydrolysis starts immediately after bottom ash
quenching and can be prolonged over the time span of
temporary storage or landfilling of the material (Belevi
et al., 1992; Johnson et al., 1995), as long as it is in
contact with water. Hydrolysis involves the transfor-
mation of oxides of Ca, Na and K and non-noble metals
like Al and Fe into the corresponding hydroxide species
(e.g. CaO!Ca(OH)
)(Belevi et al.,
1992; Speiser et al., 2000; Speiser, 2001).
As a consequence of quenching, calcium- and alumi-
num-containing phases can also dissolve and other
minerals can be formed as a result of dissolution/pre-
cipitation phenomena (Belevi et al., 1992; Meima and
Table 4
Change in leaching behavior after treatment of different residues from thermal treatments with H
(after Fa
llmann, 1997)
Oxidation increases leachability (95% significance) Oxidation decreases leachability (95% significance)
Oxidizing agent: H
Oxidizing agent: H
Blast furnace
Steel smelter
MSWI bottom
Blast furnace
Steel smelter
MSWI bottom
Cu Al, Ba As As Fe Fe Al, Fe Ba, Ca
K Cd, Cr Cr Co Mg Mg, Mn K, Mg
Na Cu, Na Cu Cr Mn Na, Si Na, S
Ni, S V Mn Zn Si, Zn
Si, V V
T. Sabbas et al. / Waste Management 23 (2003) 61–88 73
Comans, 1997); for instance, ettringite can be formed
according to the reaction (Meima and Comans, 1997):
þ 2Al
þ 3SO
þ 38 H
O !
! 12H
þ Ca
The formation of C-S-H phases was also detected
(Speiser et al., 2000) as well as the neo-formation of
clay-like minerals from the corrosion of glasses
(Comans et al., 1994; Zevenbergen and Comans, 1994;
Zevenbergen et al., 1996).
Carbonation is caused by the uptake of atmospheric
by the initially alkaline material, which leads to a
decrease in pH and to the precipitation of calcite
(Meima and Comans, 1997; Meima and Comans, 1999;
Zevenbergen and Comans, 1994). CO
results in final pH values in the range of 8 to 8.5 (Bod-
nan et al., 2000; Meima and Comans, 1999). In this
state the equilibrium between calcite and CO
with water) under the influence of gypsum dom-
inates the system as a buffer. At pH 8 the solubility
minimums are reached for most of the solid phases
controlling the leaching of such heavy metals as Cd, Pb,
Zn, Cu and Mo (Meima and Comans, 1999). Calcite
can also provide a number of sorption sites for certain
elements, e.g. Cd and Zn, that have been shown to dis-
play a high affinity for this phase (Meima and Comans,
1999). However, leaching of sulfate from weathered
bottom ash has been found to increase if compared to
fresh bottom ash (Bode
nan et al., 2000), probably as a
result of ettringite carbonation, which leads to the pre-
cipitation of gypsum.
Sorption onto the neoformed minerals, including both
adsorption and co-precipitation processes, also seems to
play a role in reducing contaminant leaching from
weathered bottom ash. Fe and Al (hydr)oxides as well
as amorphous aluminosilicates formed as a result of
weathering have been found to be reactive sorptive
minerals for e.g. Cd, Zn, Cu, Pb and Mo (Meima and
Comans, 1998; Meima and Comans, 1999).
Similar processes are observed during the weathering
of APC residues. In addition to the above mentioned
phases, aging of APC residues can lead to the formation
of Ca-Al-S-Cl-hydrate phases (e.g. hydrocalumite,
ettringite, and members of the hydrotalcite group) con-
taining varying amounts of heavy metals like Zn (Spei-
ser et al., 2001; Heuss-Assbichler et al., 2002; Speiser et
al., 2002).
Fig. 10 provides a schematic representation of the
weathering reactions and the related modifications in
leaching behavior.
2.2. Gas production
Gas generation at landfill sites with MSWI residues
can be either of a biotic or abiotic nature. The low bio-
degradable organic carbon content of MSWI residues
generally leads to the production of biogas amounts sig-
nificantly lower if compared to MSW landfill gas. Con-
versely, the evidence of significant abiotic gas generation
has been reported in a number of studies (Musselmann et
Fig. 10. Layout of mineralogical reactions as a consequence of leaching and weathering processes.
74 T. Sabbas et al. / Waste Management 23 (2003) 61–88
al., 2002; Magel et al., 2001, Lechner et al., 1997).
Abiotic gas is produced by chemical oxidation, in the
presence of water, of elemental metals including Al, Fe
and Cu. Thus, as previously observed for leachate pro-
duction, the different parameters of concern for the
water balance and the chemical properties of the waste
material must be considered for gas production as well.
Aluminum is a major constituent of bottom ash and is
also significantly concentrated in fly ash and APC resi-
dues. Due to its high solubility at pH > 9.5 and to its
lower redox potential if compared to other elements,
aluminum is regarded as the main element responsible
for abiotic gas production. In addition, a significant
fraction of Al in MSWI residues is in its elemental form,
which can undergo the following redox reactions, lead-
ing to hydrogen gas generation:
þ 3H
O ! Al
þ H
þ 2H
O ! AlOOH þ 1:5H
þ 3H
O ! Al OHðÞ
However, the chemistry of aluminum corrosion is not
completely understood at present, due to the variability
of local conditions throughout the landfill mass.
A number of studies (Fo
rster and Hirschmann, 1997;
Mizutani et al., 2000) evidenced that abiotic gas pro-
duction is almost complete after several months, so
hydrogen generation can be considered a short-term
process. However, even though no specific studies have
been carried out, some experimental data suggest that
hydrogen gas production can also evolve over a longer
term, as shown in Fig. 11 (Magel et al., 2001).
It has also been found that isolated aluminum parti-
cles in MSWI residues are generally surrounded by a
reaction rim of Al(OH)
and additionally by hydro-
calumite (Ca
O) and ettringite ([Ca
O). These coatings or by products
may result in the retardation of hydrogen production.
However, such rims can dissolve, leading to a perma-
nent release of hydrogen. Furthermore, a significant
number of ash particles are enclosed in glassy phases
formed during incineration, which act as a barrier
against the reaction between water and aluminum.
However, due to their alkaline nature, such glassy pha-
ses can be altered, so aluminum particles will come into
contact with the hydration water.
2.3. Temperature development
Recently several studies have shown that many exo-
thermic reactions may cause a temperature increase of
up to 90
C in MSWI residue landfills (e.g. Klein et al.,
2001; Heyer and Stegmann, 1997). MSWI residue sto-
rage is affected by heat generation as a result of different
exothermic reactions, such as hydration of alkaline and
alkaline earth oxides, corrosion of metals and carbona-
tion of portlandite (Huber, 1998). The main effects of
this temperature enhancement are (see Fig. 12):
acceleration of weathering/hydration reactions as
long as the material in the landfill is wet or
over a longer term, formation of salt rims as a
consequence of drying, and
modification of precipitation/dissolution and
complexation equilibria.
Moreover, temperature increase may lead to eva-
poration of some dissolved gaseous compounds (CO
Fig. 11. Hydrogen content of gas from a German MSWI residues monofill (LEL: lower explosive limit of H
T. Sabbas et al. / Waste Management 23 (2003) 61–88 75
) from the leachate and as a consequence may exert
an effect on the redox potential and the concentration of
complexing agents.
Temperature development was studied by Klein et al.
(2001) during an experimental campaign carried out at a
German bottom ash monofill. Fig. 13 shows tempera-
ture development over time in three monitored sensor
fields. In Sensor Field 1 (SF1) bottom ash was deposited
at irregular time intervals, after one to three weeks of
storage at the landfill site. SF2 was built up over three
weeks to its final height of ten meters. In SF3 bottom ash
was placed in 1 meter-thick layers every two months up
to a final height of 6 meters. Previous storage of bottom
ash in this sensor field was disregarded. Bottom ash in all
the sensor fields was not compacted and no temporary
liners were used to cover the landfill between deposits.
In every layer of the surveyed landfill the temperature
development started with an increase immediately after
deposition. Over the next three to four months the bot-
tom ash temperatures increased to a maximum which
varied depending on the layer depth. The average rate at
which the temperatures rose was between 0.16 and
C per day. In all the observed landfill layers, the
maximum temperature occurred at a time of about four to
five months after deposition. The initial temperature rises
and maximum temperatures occurred in those cases where
the ash was not stored temporarily before landfilling.
The experimental program also revealed that rain-
water percolating through the landfill body exerted no
appreciable effect on temperature development.
3. Potential environmental impacts
The main potential environmental impacts related to
the handling, utilization and disposal of MSWI residues
can be summarized as follows:
Fig. 12. Relationship between temperature, water content and rate of
weathering/hydration reactions.
Fig. 13. Temperature development in a German bottom ash monofill.
76 T. Sabbas et al. / Waste Management 23 (2003) 61–88
dust emissions,
leachate generation,
gas emissions, and
temperature increase.
Moreover, as far as landfilling is concerned, the
potential impacts arising from the construction and
operation of the landfill should be taken into account.
However, such impacts as additional traffic load, noise
pollution and site modifications in terms of topo-
graphic, hydrological and hydrogeological conditions
will not be discussed in this paper. The extent of the
most relevant impacts may differ for the handling, utili-
zation and final disposal phases and vary significantly
over time.
3.1. Dust emissions
MSWI residues contain a fine fraction of particulate
matter passing the 74 mm mesh sieve that accounts for
1–10% of bottom ash, whilst APC residues have a par-
ticle size distribution varying between 0.001 and 1 mm.
Friability of bottom ash may result in an increased per-
centage of finer particles after processing operations.
The fine bottom ash particles typically contain chloride
and sulfate salts as well as heavy metals like Pb, Cu and
Zn (IAWG, 1997).
The easily airborne nature of fine particles leads to the
dispersion of pollutants, which in turn can give rise to
health risks for exposed, unprotected workers and the
public, as well as soil contamination. To prevent or
minimize dust emissions, bottom ash and fly ash are
normally kept wet (5–15% humidity) and transported
by covered and watertight trucks. The upper limit for
humidity is regarded as the minimum value required to
prevent fugitive dust problems in open storage piles.
The ability to maintain the optimal water content in
order to minimize dust emissions is obviously related to
climatic conditions (temperature, humidity as well as
regime, intensity and frequency of the dominant local
winds), which influence the desiccation rate of the
material, the critical area of downwind dust deposition
as well as the downwind distance at which dust can be
transported. Matsuto et al. (2001) investigated the wind
dispersion of incinerated residues and found that they
are dispersed up to 50 meters from the landfill. The
atmospheric dispersion of particles may not be sig-
nificant during the after-closure period of a landfill due
to the presence of a top cover.
3.2. Leachate generation
The potential environmental impact of leaching
includes contamination of soil, groundwater and surface
water bodies. As leaching of contaminants from MSWI
residues may occur during the temporary storage,
treatment or reuse as well as during the final disposal of
the material, the following aspects should be investi-
gated: the leaching behavior of contaminants, the
environmental conditions that may occur in any of
the above mentioned scenarios, as well as their var-
iation over time. Thus, according to the discussion in
the preceding sections, the following items should be
residue characteristics, in terms of physical and
mechanical properties, particle size distribution,
acid neutralization capacity, concentration of
contaminants, availability of contaminants for
leaching, leaching mechanisms, controlling fac-
tors, and their variation over time due to
weathering reactions;
characteristics of the application site in terms of
(1) dimensions and (2) material properties (por-
osity, bulk density and permeability);
hydrological conditions of the application site in
terms of (1) net rate of infiltration, (2) properties
of the unsaturated zone and the aquifer (thick-
ness, permeability, porosity, longitudinal and
horizontal dispersivity, bulk density, flow velo-
city, etc.); moreover, as far as disposal is con-
cerned, the presence of a top cover should be
taken into account; and
mitigating effects due to leachate/soil interactions
(e.g., ionic exchange, sorption) and to dilution.
The extent of the impact depends on the rate at which
leaching occurs and on the type and concentration of
the dissolved species. The following elements must be
considered as hazardous contaminants potentially
leachable from MSWI residues: As, Al, B, Ba, Cd, Cr,
Cu, Hg, Mn, Mo, Ni, Pb, Sb, Se, Zn, Br
. Thus, their concentration
in the leachate should be compared to specific quality
criteria. Since no international guidelines for ground-
water have been proposed so far, it appears reasonable
to apply the international drinking water quality criteria
(EU Drinking Water Directive and WHO criteria on
drinking water) until a specific groundwater directive is
The distinction between the short- and the long-term
leaching behavior appears as a key factor. Whilst infor-
mation is available concerning the short-term behavior
of most MSWI residues (results of leaching tests and
field measurements), long-term behavior can be pre-
dicted only on the basis of a synthesis of information on
leaching principles, leaching tests results, field measure-
ments, simulation of mineral changes and speciation.
As far as the environmental impact assessment related
to leaching of contaminants out of the MSWI residues
is concerned, availability as opposed to the total con-
centration of contaminants in the solid matrix provides
T. Sabbas et al. / Waste Management 23 (2003) 61–88 77
an estimation of the maximum amount of contaminant
that in theory could be leached over a 1000- to 10000-
year timeframe (with the exception of highly soluble
salts, for which the maximum leachable amount can be
attained within shorter periods, typically a couple of
Table 5 shows typical ranges of the concentration of
contaminants in MSWI residue leachate.
However, availability does not account for the acid
neutralization capacity exerted by the matrix. This is an
important parameter, which determines the potential
environmental impact of MSWI residues, in that neu-
tralization processes can affect leaching reactions and
control the release of contaminants from the material.
Acid neutralization capacity allows for the evaluation of
the environmental behavior of MSWI residues, in that
ANC data can be transformed in order to estimate the
time required for the pH to drop from the ‘‘inherent’’
pH of the material to critical values for contaminants
release (Astrup et al., 2001). Such time-related informa-
tion can be gathered on the basis of the size of the
application site, hydrological and hydrogeological con-
ditions, leachate composition, and leachate flow
towards soils and groundwater.
A more detailed simulation of the potential environ-
mental behavior of MSWI residues should also consider
the attenuation phenomena caused by the interactions
between leachate and soil, such as sorption and ion
exchange (Hjelmar et al., 2001; Hjelmar et al, 1999a,b).
However, in order to predict correctly the leaching
behavior over time, additional information (pH, redox
conditions, ionic strength, complexing agents, and
mineralogy) is required. It has been observed (Hjelmar,
1996) that the first leachate produced by bottom ash has
a relatively high content of inorganic salts (chloride,
sulfate, sodium, potassium and calcium) and low con-
centrations of trace elements due to the fact that at this
stage reducing conditions are occurring and pH is
slightly or strongly alkaline, depending on the degree of
carbonation. Among the trace elements, Cu can behave
differently, as observed before, its leachability being
increased by DOC. With the exception of sulfate, the
concentration of salts in the leachate tends to decrease
over time (i.e. at increasing L/S ratios).
APC residues behave differently from one another
depending on their origin and air pollution control
devices installed into the plant. High concentrations of
readily soluble salts, such as chlorides and hydroxides of
calcium, sodium and potassium generally characterize
the first leachate from APC residues. Trace elements,
such as Pb and Mo, which are mobile under reducing
and slightly alkaline conditions, can be highly leachable
at low L/S ratios (corresponding to the first fractions of
the leachate). Thus APC residues are typically hazar-
dous materials, and their disposal requires considerable
care in order to prevent adverse environmental impacts.
As far as co-disposal of bottom ash and APC residues is
concerned, it was observed that soluble salt concentra-
tions are higher in the combined ash leachate than in
bottom ash and fly ash leachate (Hjelmar, 1996). This
behavior was also observed for Cd and is likely to be
ascribed to high complexing chloride content due to the
presence of organic acids produced by biodegradation
of residual unburned carbonaceous material in bottom
Very little data are available from the literature
regarding the long-term composition of leachate from
MSWI residues. Sabbas et al. (2001) investigated four
different mixtures including 1) MSWI bottom ash
(RAU-S2), a mixture of MSWI bottom ash and fly ash,
2) rotary drum kiln slag from hazardous waste incin-
eration and fluidized bed incineration ash (RAU-S), 3) a
mixture of rotary drum kiln slag from hazardous waste
incineration and fluidized bed incineration ash (RAU-
A) as well as 4) a cement solidified product (RAU-SB, a
mixture of MSWI bottom ash and fly ash, rotary drum
kiln slag from hazardous waste incineration and flui-
dized bed incineration ash, cement and gravel). The
materials were approximately 10–15 years old, naturally
weathered, fully carbonated and showed leachate pH
values between 7.3 and 8.9. The results shown in Fig. 14
reveal that all concentrations, except those of arsenic
Table 5
Maximum concentrations of contaminants in leachates from various MSWI residues (after Hjelmar, 1996)
Typical maximum levels of
concentration in leachate
MSWI bottom
MSWI fly ash and residues from
dry and semidry APC processes
Mixture of MSWI fly ash and sludge
from wet scrubbing process
> 100 g/l Cl
10–100 g/l Na, K, Pb Cl
, Na, K
1–10 g/l SO
, Na, K, Ca Zn SO
100–1,000 mg/l NVOC, NH
10–100 mg/l
1–10 mg/l Cu, Mo, Pb Cu, Cd, Cr, Mo NVOC, Mo
100–1,000 mg/l Mn, Zn As
10–100 mg/l As, Cd, Ni, Se As, Cr, Zn
1–10 mg/l Cr, Hg, Sn Pb
< 1 mg/l Hg Cd, Cu, Hg
78 T. Sabbas et al. / Waste Management 23 (2003) 61–88
and mercury, for the solidified product fall below the
proposed EU acceptance criteria for waste (TAC Mod-
eling Group, 2002).
3.3. Gas emissions
As described in the previous sections, gas production
occurs from both MSWI bottom ash and APC residues
as a result of metallic aluminum hydration.
As pointed out by Fo
rstner and Hirschmann (1997),
hydrogen production from unquenched bottom ash is
higher if compared to quenched bottom ash. However,
up to now no accident has been reported, which is
obviously the result of generally good aeration. Unlike
the situation of bottom ash, a number of authors
(Takatsuki, 1994; Yasuda, 1997) have reported explo-
sions resulting from hydrogen generation from APC
residues in which people were injured or died. Such
accidents occurred when ash blocks were crushed or
water was sprinkled on the ash.
It is documented that hydrogen generation can pro-
ceed even over the long term. A case of deflagration
occurred after sealing off a monofill that contained up
to 20-year-old residues (Magel et al., 2001).
3.4. Temperature development
Bottom ash storage sites are affected by heat generation
as a result of different exothermic reactions, such as
hydration of alkaline and alkaline earth oxides, corrosion
of metals and carbonation of portlandite (Huber, 1998).
As far as bottom ash landfilling and the design of
monofills is concerned, the potential impact arising
from the heat-related damage on the landfill liner and
on the leachate collection system should be taken into
account. Temperature development can last over long
periods (decades or longer) due to the low rate of heat
transfer through the residue bulk. At temperatures over
C, clay liners are liable to desiccation and therefore
cracks may form; furthermore, HDPE membrane liners
are subject to ruptures. Both effects may lead to loss of
sealing efficiency of the landfill liner system.
A secondary effect of high temperatures has been
observed in a few landfills, where, due to the desiccation
Fig. 14. Composition of leachate and discharge from four different naturally weathered residues from thermal treatment [~ RAU-SB; * RAU-S2;
& RAU-S; RAU-A; ---proposed EU acceptance criteria for non-hazardous predominantly inorganic wastes at L/S=2 (TAC Modeling Group,
T. Sabbas et al. / Waste Management 23 (2003) 61–88 79
of the residues themselves, steam generation and escape
from the surface was observed (Heyer and Stegmann,
4. Mitigating measures
This chapter will focus on the various mitigating
measures aimed at reducing the potential environmental
impacts from incineration residue reuse or disposal. It
must be stressed that when dealing with the available
treatment methods, both the environmental behavior of
the residue under consideration and its intended desti-
nation should be kept in mind. In other words, any
decision concerning reuse, treatment or disposal of
incineration residues can be taken only after their
environmental quality has been assessed. Nevertheless,
the relevant conditions for the intended application
must be evaluated carefully, which also includes the
identification of the elements to be controlled. In light
of this, as far as final disposal is concerned, the land-
filling strategies for waste combustion residues should
reflect the fundamental differences existing between
such residues and the original waste. Thus, it can be
anticipated that the common practices used for raw
municipal solid waste landfilling may not be adequate
when managing incineration residue disposal.
Various options are available for the treatment of
waste incineration residues in view of their reuse or final
disposal. Such measures can be applied at different stages
of their life cycle (from generation to reuse and/or final
disposal) and can be based on different treatment
approaches. Accordingly, one may distinguish between:
measures undertaken prior to reuse or final dis-
measures undertaken during landfilling and
active landfill operation; and
measures carried out during the passive phase of
A number of strategies affecting the environmental
properties of incineration residues may also be applied
prior to or during the combustion process. Examples of
such strategies may include either waste sorting and
selection prior to incineration in order to beneficially
modify the waste input composition (homogeneity,
chlorine content, metal content, etc.) or control of
operating parameters (temperature, oxygen concen-
tration, etc.) during the combustion process. Based on
the treatment principles, the pretreatment options can
be further grouped into three broad categories including
(IAWG, 1997; Van der Sloot et al., 2001):
physical or chemical separation processes,
solidification and/or stabilization processes, and
thermal treatment,
while the measures related to the landfilling phase can
be divided into:
landfill design options and
landfill operation strategies.
In general, as depicted in Fig. 15, it may be stated that
the basic principles of the measures to mitigate the
environmental impact of incineration residues are based
on variations in either (a) the total content, (b) the
availability for leaching or (c) the release rate of con-
taminants into the environment; combinations of one or
more of the three mechanisms shown in Fig. 15 are also
possible. For instance, washing pretreatments aimed at
removing e.g. readily soluble salts act according to
mechanism (a), so that the availability for leaching is
reduced as a consequence of the reduction in total con-
tent. Stabilization pretreatments modify the release rate
of contaminants [mechanism (c)] and may also reduce
the availability for leaching when chemical immobiliza-
tion mechanisms are involved [mechanism (b)].
Whether the mechanism acting as a consequence of
treatment pertains to type a, b or c depends both on the
specific contaminant under consideration and on the
Fig. 15. Principles of mitigating measures in respect to total content, availability and release rate.
80 T. Sabbas et al. / Waste Management 23 (2003) 61–88
nature of the process applied. Examples of treatments
for each type will be provided in the following section.
It should be mentioned that in a few cases, for
instance with a view to reducing the long-term impact,
the applied treatment may act to either enhance or
accelerate the release in the short term, instead of
decreasing or slowing it.
It is important to stress that when selecting the proper
treatment method(s) for a given incineration residue,
both its short- and long-term environmental behavior
under the expected conditions must be considered care-
fully. It should also be emphasized that most of the
available treatments generate a number of waste
streams and are also responsible for raw material and
energy consumption. Thus, the technical, environmental
and economic applicability of any treatment option
should be judged on the basis of an appropriate eval-
uation of the overall mass and energy balances under a
life-cycle assessment framework.
4.1. Treatment options
4.1.1. Measures undertaken prior to reuse or final
Table 6 reports the various treatment options applic-
able to waste incineration residues prior to their reuse or
final disposal according to the classification discussed in
the previous section. The most common treatment pro-
cesses will be dealt with in this section. Physical and chemical separation. Physical
separation methods have only a limited effect on the
quality of residues, as their principle is to separate from
the bulk of the material the individual constituents that
are already present in such residues in the same physical
and chemical form (IAWG, 1997). However, physical
separation is able to remove specific materials that ren-
der some residue streams (e.g. bottom ash) unsuitable
for a number of applications.
Among the physical separation options, particle size-
based separation is generally carried out for two main
purposes. One aim is accomplished through isolating
the fraction(s) of the material (usually the finer frac-
tion), which is more concentrated in contaminants, thus
reducing the environmental impact of the residual
stream. In this case, selecting the appropriate cut size
for separation is of crucial importance in order to
reduce effectively the leachability of the contaminants in
question. One major drawback to this kind of separa-
tion is that for some residues, e.g. bottom ash, the finer
fraction usually constitutes a considerable portion of
the total mass of the material [the percentage passing at
the 2 mm sieve is in the order of 30% by weight (IAWG,
The second aim of particle size separation is to pro-
duce a material where the engineering properties, such
as particle size gradation and hydraulic conductivity,
are more suitable for subsequent utilization. Yet fine
material can potentially create problems in bottom ash
reuse, in that it is highly sorptive for water, asphaltic
cement and Portland cement (IAWG, 1997). For
instance, when MSWI bottom ash is to be reused as a
coarse aggregate in asphalt or concrete mixtures, a
screening pretreatment to remove the 1- or 2-mm
undersize fraction and the 40-mm oversize fraction
(Wiles, 1996) is commonly applied. This has the addi-
tional benefit of reducing Cd, Cr, Cu and possibly sul-
phate leachability from the material.
Magnetic and eddy-current separation are electro-
mechanical separation processes mostly practiced on
bottom ash to reduce its ferrous and non-ferrous metal
content, respectively. According to the IAWG (1997)
and Wiles (1996), the ferrous metal content of MSWI
bottom ash ranges from 7 to 15% by weight, while non-
ferrous metals account for approximately 1–2% by
weight although such figures strongly depend on waste
sorting and selection strategies prior to the combustion
Metal separation from bottom ash may be performed
with a view to either metal scrap recovery or to
improvement of bottom ash properties for its utiliza-
tion. In this case, metal removal from bottom ash may
beneficially prevent corrosion phenomena arising from
oxidation of metals like Al, Fe and Zn, which, as dis-
cussed in the previous sections, may cause swelling and
expansion of the material under the utilization condi-
tions (IAWG, 1997; Lamers and Born, 1994).
Among the chemical separation treatments, washing
with water is one of the simpler processes for removing
highly water-soluble constituents from waste incineration
Table 6
Main options for treatment of incineration residues prior to reuse or
final disposal (after IAWG, 1997; Kosson and van der Sloot, 1997;
Lamers and Born, 1994)
Principle of treatment Process
Physical and chemical Size separation
separation Magnetic separation
Eddy-current separation
Chemical extraction/mobilisation
Chemical precipitation
Ion exchange
Solidification and/or
with hydraulic binders
Chemical stabilisation
Thermal treatment Sintering
T. Sabbas et al. / Waste Management 23 (2003) 61–88 81
residues. The soluble constituents, typically removed as
a consequence of washing, are mainly represented by
chloride and alkali ions. Conversely, due to the highly
alkaline nature of incineration residues, the effectiveness
of washing on trace metals removal has proved to be
relatively low (IAWG, 1997; Schneider et al., 1994), in
that they mostly form sparingly soluble compounds
under alkaline conditions.
Bottom ash is commonly quenched when exiting the
combustion chamber; however, in most cases the rela-
tively low L/S ratios and residence time in the quench-
ing tank prevent thermodynamic equilibrium of the
dissolution process from being attained. Thus, bottom
ash after quenching still is typically characterized by a
residual content of soluble components, which can be
further extracted through a washing treatment. Due to
the above-mentioned reasons, washing of bottom ash is
a simple measure that could easily be combined with the
quenching stage at the combustion plant. However,
washing alone may not be adequate for bottom ash to
reach a suitable level of quality for subsequent utiliza-
tion according to established regulatory limits. Thus,
the washing treatment may be carried out beneficially in
combination with other processes, e.g. chemical mobili-
zation or aging (see below) (IAWG, 1997; Lahl, 1992),
although a number of studies have indicated that
mobilization of soluble salts mostly occurs during the
initial washing stage (Schneider et al., 1994).
As far as APC residue is concerned, washing with
water can also be applied, commonly as a pretreatment
stage prior to further chemical stabilization processes
(Derie, 1996; Lundtorp et al., 1999; Mangialardi et al.,
1999; Nzihou and Sharrock, 2002) in order to remove
soluble salts. Such salts may account for up to approxi-
mately 20% of the material and are responsible for
much of the negative properties of such residues (e.g.
high leachability, high water absorption and corrosive-
ness). It has been reported (Derie, 1996; Laethem et al.,
1994; Nzihou and Sharrock, 2002), particularly for dry
and semi-dry APC residues, that the high pH of the
material coupled with the large concentrations of
highly-soluble heavy metal chlorides are also respon-
sible for the partial extraction of such metals as lead,
zinc and cadmium as a consequence of the washing
process. However, the extent of chemical mobilization is
not suitable for adequate APC residue detoxification
levels to be attained (Lamers and Born, 1994), and
typically the material still needs additional treatment
prior to final disposal. Such treatment will most often
include either chemical stabilization or solidification
with hydraulic binders (see the following section).
Derie, 1996; Nzihou and Sharrock, 2002 have
demonstrated that an L/S ratio of 10 allows for extrac-
tion of most (90%) of the highly soluble salts (mainly
chlorides). For other anionic species, e.g. sulfates, which
are solubility-controlled, the efficiency of dissolution
relies on the L/S ratio adopted during the washing pro-
cess. The washing solution can subsequently be evapo-
rated, yielding a crystalline mass which is mainly
composed of halite, sylvite and gypsum (Nzihou and
Sharrock, 2002). To prevent contamination of the
washing solution by heavy metals, the use of additives
which are able to form insoluble heavy metal com-
pounds has also been suggested (Nzihou and Sharrock,
2002). The benefits of washing have also been demon-
strated with a view to APC residue solidification and/or
stabilization treatments (Derie, 1996; Mangialardi et
al., 1999; Nzihou and Sharrock, 2002). In this case,
removal of chloride and water-soluble sulfate as well as
alkali ions (which are known to negatively interfere
with binder hydration) may allow for significant
improvements in the physical and mechanical proper-
ties of the final products. However, when evaluating
the overall benefits of a washing treatment, the shift of
the pollution problems from solid to liquid waste
streams must be carefully considered by designers and
decision makers.
Chemical extraction/chemical mobilization processes
may be applied both to bottom ash and to APC resi-
dues. As far as bottom ash is concerned, treatments
consisting of sodium carbonate or sodium bicarbonate
addition are beneficial. This form of treatment has the
effect of mobilizing sulphate through the formation of
soluble Na
and precipitation of CaCO
Several chemical extraction processes have been
proposed for APC residues (Hong et al., 2000b; IAWG,
1997; Katsuura et al., 1996; Laethem et al., 1994). The
aim of both processes is to recover heavy metals and
to detoxify the material. A number of treatment
methods have been proposed using inorganic acids,
including hydrochloric, nitric or sulfuric acid and aqua
regia (Hong et al., 2000a; Hong et al., 2000b; Katsuura
et al., 1996; Laethem et al., 1994), as well as chelating
agents including nitrilotriacetic acid (NTA), ethylen-
diamine-tetraacetate (EDTA), diethylen-triamine-pen-
taacetate (DTPA) and saponins (Hong et al., 2000a,b;
Laethem et al., 1994). The efficiency of heavy metal
extraction has been found to be strongly dependent on
the pH and L/S ratio as well as on both the nature of
the extracting agent used and the particular metal of
concern. Solidification and stabilization. Solidification/
stabilization treatments are among the most widespread
processes used for waste incineration residues, mainly
APC ash (e.g. Conner, 1990; Gilliam and Wiles, 1996).
The main purpose of solidification/stabilization is to
produce a material whose physical (specific surface area,
porosity, tortuosity, etc.), mechanical (durability,
mechanical strength, etc.) and chemical properties are
more favorable with respect to reducing the leachability
of contaminants out of the waste matrix.
82 T. Sabbas et al. / Waste Management 23 (2003) 61–88
The most common processes make use of hydraulic
binders including cement, lime and/or pozzolanic mate-
rials. In general, it can be stated that improvement in
the leaching behavior of solidified/stabilized incinera-
tion residues is attained through either physical or che-
mical immobilization mechanisms, depending on the
specific contaminant of concern as well as on the type of
binder used. However, weak stabilization efficiencies
typically have been recorded for soluble salts. Further-
more, due to their strong amphoteric behavior, treat-
ment of zinc and lead with cement- and lime-based
processes may be problematic, unless incorporation in
the crystal lattice of the hydration products occurs or
appropriate additives are used.
Chemical stabilization processes have been proposed,
which basically involve chemical precipitation of heavy
metal-incorporating insoluble compounds and/or heavy
metal substitution/adsorption into various mineral spe-
cies. The principal forms of chemical agents used
include sulfides (IAWG, 1997; Katsuura et al., 1996),
soluble phosphates (Derie, 1996; Eighmy et al., 1997;
Hjelmar et al., 1999a,b; Nzihou and Sharrock, 2002),
ferrous iron sulfate (Lundtorp et al., 1999) and carbo-
nates (Hjelmar et al., 1999a,b).
Treatments with hydraulic or chemical binders gen-
erally yield good leaching properties at relatively low
costs. However, solidification/stabilization with
hydraulic binders results in increased amounts to be
landfilled and the physical encapsulation from the bin-
der cannot be considered to last in the long term.
Aging and weathering processes are applied to pro-
mote mineralogical changes as a consequence of altering
mineralogical phases in MSWI residues over time. Such
changes may lead to significant reductions in trace ele-
ments (including heavy metals as Cd, Cu, Pb, Zn and
Mo) and leaching (Meima and Comans, 1999; Zeven-
bergen and Comans, 1994; Zevenbergen et al., 1996)as
a result of hydration, carbonation or oxidation/reduc-
tion. These can give rise to pH decrease, contaminant
sorption processes as well as formation of more stable
mineral species (Meima and Comans, 1997; Meima and
Comans, 1999). Aging and weathering can be benefi-
cially applied particularly to reactive materials such as
bottom ash, as this is composed of high-temperature
solids which are metastable under natural conditions
and are therefore likely to undergo a number of miner-
alogical changes. From this perspective, bottom ash,
prior to utilization, is commonly aged through storage
in stockpiles open to the atmosphere for periods ranging
from several weeks to a couple of months.
Aging and weathering can also be artificially
enhanced to accelerate the chemical reactions respon-
sible for the fixation of contaminants within the waste
matrix. With a view to this, accelerated carbonation has
been proposed as an efficient treatment for reducing
leaching of soluble salts, Pb and Zn, although it has
been shown to have the potential of mobilizing sulfate
nan et al., 2000). Thermal treatment. The thermal treatment of
incineration residues is used extensively in some coun-
tries to obtain reduced leaching from the residues and
reduced volume as well as a treated material that is sui-
table for reuse. Thermal treatment can be grouped into
three categories: vitrification, melting and sintering
(IAWG, 1997).
Vitrification is a process whereby residues are mixed
with glass precursor materials and then combined at
high temperatures into a single-phase amorphous,
glassy product. Typical vitrification temperatures are at
C. The retention mechanisms are chemical
bonding of inorganic species in the residues with glass-
forming materials, such as silica, and encapsulation of
residue constituents by a layer of glassy material.
Melting is similar to vitrifying, but this process does
not include the addition of glass materials and results in
a multiple-phased product. Often several molten metal
phases are produced. It is possible to separate specific
metal phases from the melted product and recycle these
metals, perhaps after refinement. Temperatures are
similar to those used in vitrifying.
Sintering involves heating the residues to a level at
which bonding of particles occurs and chemical phases
in the residues reconfigure. This leads to a denser pro-
duct with less porosity and a higher strength. Typical
temperatures are around 900
C. When MSW is incin-
erated, some level of sintering will typically take place in
the incineration furnace. This is especially the case if a
rotary kiln is used as part of the incineration process.
Regardless of the process, thermal treatment of
incineration residues in most cases results in a more
homogeneous, denser product with improved leaching
properties. Vitrifying also adds the benefits of physical
encapsulation of contaminants in the glass matrix. A
major drawback to these methods, however, is that they
require substantial amounts of energy and especially
vitrifying and melting result in the mobilization of
volatile elements such as Hg, Pb and Zn during the
thermal treatment process. Combined methods. On the basis of the results
from batch-scale extraction treatments, a number of
multistage processes have also been developed either at
a pilot- or full-scale level. For the implementation of
such processes, which typically involve different combi-
nations of the above-mentioned treatments, efforts have
been devoted to integrating the proposed treatment(s)
with the combustion and flue gas treatment process.
Thus, the residues from one stage of the process com-
monly are used in one or more subsequent stages to
keep the net discharge from the overall treatment to a
minimum. Among the several processes developed so
T. Sabbas et al. / Waste Management 23 (2003) 61–88 83
far, probably the most used are the 3-R (Vehlow et al.,
1990), the MR (Stubenvoll, 1989), the AES (Katsuura
et al., 1996) and the VKI process (Hjelmar et al.,
1999a). As a detailed description of such processes is
beyond the scope of this paper, the reader can refer to
the literature cited for additional information.
4.1.2. Measures undertaken during landfilling and active
landfill operation
The landfilling phase includes the period of time in
which the landfill receives waste. The time span of this
phase will often amount to about 5–30 years depending
on the capacity of the landfill. When the last section of
the landfill is closed and the site is no longer used for
active landfilling, a period of active post closure care
begins. This period can include active measures to
reduce the environmental impact of the landfilled waste
as well as monitoring of emissions, and it will be pro-
longed until the emissions from the landfill are con-
sidered acceptable with respect to the surrounding
environment. The extent of this period can be very dif-
ficult to estimate and depends on the characteristics of
the specific landfill and landfill site in question.
The design and operation of the landfill have the
potential to diminish or enhance the possible environ-
mental impact from landfilling the treated incineration
residues. Any landfilling of residues should include an
assessment of the environmental impact, both in the
short- and long-term perspective, and the landfilling
should be part of a proper disposal strategy which takes
the pretreatment of the residues into account. However,
it is evident that the landfill design should assist in mini-
mizing the total lifetime of required active environmental
protection systems. This should include consideration of
the properties of the waste, the potential risks related to
handling and landfilling the residues (as discussed in the
previous chapter) and thereby also the long-term effects
of keeping the residues at the disposal site as well as any
derived consequence of operating the landfill (for exam-
ple, effects of leachate management). At the termination
of the active care period, the emissions from the residues
remaining at the landfill site must be at an environmen-
tally acceptable level—even given the long-term perspec-
tive—without requiring any active operation.
The most important issues regarding the design and
operation of a landfill with incineration residues will be
discussed in the following section. First are the issues
concerning the design phase:
siting of the landfill,
size of landfill sections,
height of landfill, and
liner systems and disposal strategy.
Next come issues regarding the operation of the
control of waste types,
compaction of waste,
covering of waste,
leachate collection and treatment,
infiltration (leachate recirculation and/or irri-
gation), and
gas collection or venting. Landfill design. The geological, hydrological and
geotechnical characteristics of the location of a planned
landfill are by far the most important issues in relation to
the potential impact of landfill emissions on the surround-
ing environment. Issues such as climate, quality and vul-
nerability of ground and surface water, the ability of in-
situ geological formations to attenuate migration of
released contaminants, soil bearing capacity, seismic activ-
ity, etc. should be considered when choosing the site. It is
obviously desired to place the landfill at a location that is
capable of coping with the expected emissions from the
The sectioning of the landfill provides for the possibi-
lity of disposing of specific waste types separately and of
closing certain parts of the landfill before other parts.
The sectioning will most often depend on the number of
different waste types as well as the amount of these
wastes that are expected to be deposited at the landfill
during its lifetime. However, the construction of new
sections within the landfill area typically is carried out
sequentially according to the capacity needs to landfill a
specific waste type, and it is thereby possible for the
operators to adapt the sectioning plan according to
future needs. From an environmental as well as opera-
tional point of view, it would be reasonable to keep to a
minimum the number of open sections and the time they
are open, thereby minimizing the potential impact on
the surroundings.
The height of the landfill in many cases is determined
by legal regulations, economic considerations or limits
imposed by the physical properties of the site. With a
given net infiltration to the landfill, the height, on the
other hand, will determine the overall time to reach a
specific L/S ratio and therefore will be a factor in con-
trolling the relationship between leaching and time. In
most cases, landfill designers seek to keep the height to a
maximum, thus prolonging the leaching process as
much as possible into the future. This, however, may
have the effect of extending the time required for the
active operation of the landfill site and may not always
be the best solution. In any case it is desirable to
minimize the period in which leaching of contaminants
is at unacceptable levels and thereby decrease the need
for the active operation and maintenance of landfill
Once the location, size and shape of the landfill are
established, the choice of liner system is usually the next
most important decision in terms of controlling the
84 T. Sabbas et al. / Waste Management 23 (2003) 61–88
emissions from the landfilled residues. The choice of
liner system is typically the outcome of the overall dis-
posal strategy of the residues, which can be categorized
as follows: total containment (i.e. ‘‘dry’’ storage), lea-
chate collection by use of bottom liner systems, con-
trolled release of contaminants by controlling leachate
production, and finally uncontrolled release. Concern-
ing incineration residues, it is not suitable to follow a
strategy of total encapsulation of the waste in the land-
fill by restricting any contact with water, as this will
keep the pollution potential of the waste unchanged and
consequently involve active care and maintenance of the
landfill indefinitely. In most cases the appropriate solu-
tion is to have some type of bottom liners—either
naturally occurring or artificial—and leachate collection
during the active period.
At the same time some type of top cover will usually
be applied to reduce infiltration to the waste. When the
leachate quality is considered environmentally accep-
table, i.e. the waste has reached the so-called ‘‘final sto-
rage quality’’ (Hjelmar, 1996a,b), the collection of
leachate can be terminated and leaching will be allowed;
however, leachate production will still be controlled by
the top cover. At some point the integrity of the top
cover may no longer be guaranteed and the landfill
will then be subject to unrestricted leaching. At this
point, the leachate quality allows any active environ-
mental protection system to be abandoned safely. In
this sense, the appropriate use of top and bottom liner
systems will correspond to the various phases of landfill
operation. Landfill operation. A number of issues involving
the daily operation of the landfill can have an effect on
the emissions from the landfill and should be considered
according to the overall disposal strategy applied to the
residues. Notably, controlling the incoming waste types
and redirecting these to appropriate landfills or landfill
sections is important because mixing of different types
of wastes can have undesirable effects on leaching.
Waste types should be landfilled separately if they exhi-
bit different environmental behavior, as the disposal
strategies for such waste are in general mutually incom-
patible. Thus, for example, co-disposal of MSWI resi-
dues and organic wastes (such as MSW) generally
would not represent a sustainable option, as leaching of
contaminants such as Cu can be enhanced in the pre-
sence of organic matter. Similarly, co-disposal of differ-
ent MSWI residue streams, e.g. bottom ash and APC
residues, generally is not advisable, as a high content of
soluble contaminants in certain types of residues
requires specific landfill operation strategies that are not
necessary for other residue types.
The daily procedures of compaction and covering
incineration residues may have an effect on leachate
production and dust problems during landfill operation.
It is generally desired to keep covering areas to a mini-
mum and to be aware of the fact that using a certain
type of daily cover (such as topsoil) can introduce com-
ponents into the landfill that can potentially enhance—
or possibly even reduce—leaching from the residues.
Leachate collection and treatment is an option that
should be used under the explicit condition that it could
be terminated within a reasonable timeframe. As such,
collection and treatment is a measure that can be used
to control leaching while waiting for a certain leachate
quality to be attained. In some cases it can be beneficial
to accelerate the leaching by irrigating the residues with
new water or recirculated leachate. Following this pro-
cedure, it is possible to reach an acceptable level of
contaminant release from the landfill within shorter
timeframes and thereby decrease the time required for
active care of the landfill.
For the evaluation of proper leachate collection and
treatment strategies, consideration should be given to
the final fate of contaminants. In this regard, the con-
servative nature of most of the contaminants released by
MSWI residues must be kept in mind. Thus, if MSWI
landfill leachate treatment is accomplished at a waste-
water treatment plant through biological processes, the
control of pollutants will mostly rely on dilution. On the
other hand, if specific processes are applied for heavy
metal removal, additional residues will be generated
which require proper final disposal. In both cases,
recirculation of contaminants to the environment will
occur. For these reasons, it appears more advisable to
follow landfill operation strategies involving the con-
trolled return of contaminants to the ecological cycle at
environmentally acceptable levels (Hjelmar, 1996),
especially if coupled with appropriate waste pretreat-
ment methods.
At most landfills containing MSW, it is common
practice to consider the need for systems to control the
migration of landfill gas. This, however, is rarely the
case at landfills for incineration residues. As discussed in
previous chapters, production of hydrogen from incin-
eration residues can occur and potentially can create
operational problems at the landfill. It is therefore sug-
gested that the potential production of hydrogen be
kept in mind and gas production and composition be
carefully monitored.
4.1.3. Measures acting during the passive phase of
By definition, the passive phase of landfilling requires
no active maintenance or operation of facilities at the
landfill site in order to achieve acceptable emissions.
Any active environmental protection system at the
landfill, such as leachate collection systems or artificial
liner systems, cannot be trusted to work effectively
without proper maintenance. It is therefore essential—
and necessary—that emissions from the landfill are
T. Sabbas et al. / Waste Management 23 (2003) 61–88 85
reduced to an acceptable level at the time the passive
phase is entered.
Some measures introduced in the active phase, can
have an effect within reasonable limits on the natural
processes that will occur in the passive phase. Notably,
a number of naturally occurring weathering processes
(as discussed in the previous sections), especially in the
case of APC residues, will slowly change the miner-
alogical compositions of the residues and thereby
change the conditions for leaching of chemical species
(for example trace metals) in the long term. Also the
ability of in-situ geological formations to reduce migra-
tion of contaminants released from the landfill site
potentially can have an effect in the long term.
5. Conclusions and recommendations
Over the last decades important progress has been
made in integrated waste management systems. Treat-
ment of waste, landfilling and utilization of residues
from waste treatment, and mitigating measures are
integral elements of these systems. In particular these
measures focus on minimizing the threats to our health
and environment. Most environmental and health
threats as well as economic risks associated with resi-
dues from thermally-treated wastes are caused by metals
or salts due to liquid discharge, by particulate airborne
transport and by production of potentially dangerous
gases. Now these threats can be mitigated by various
measures including options prior, during and/or after
In some cases there still remain some residual emis-
sions at environmentally unacceptable levels, making
active aftercare systems indispensable. To keep this
aftercare period as short