managementofmunicipalsolidwasteincinerationresidues · 2014-04-17 ·...

Management of municipal solid waste incineration residues

T. Sabbasa, A. Polettinib,*, R. Pomib, T. Astrupc, O. Hjelmard,P. Mostbauera, G. Cappaie, G. Magelf, S. Salhofera, C. Speiserg,

S. Heuss-Assbichlerf, R. Kleinh, P. Lechnera

(members of the pHOENIX working group on Management of MSWI Residues)aBOKU University Vienna, Department of Waste Management-Nussdorfer Lande 29-31, A-1190, Vienna, Austria

bUniversity of Rome ‘‘La Sapienza’’, Department of Hydraulics, Transportation and Roads - Via Eudossiana 18, I-00184 Rome, ItalycTechnical University of Denmark, Environment & Resources, DTU - Building 115, DK-2800 Lyngby, Denmark

dDHI Water & Environment - Agern Alle 11, DK-2979 Horsholm, DenmarkeUniversity of Cagliari, Department of Geoengineering and Environmental Technologies - Piazza D’Armi 1, I-09123 Cagliari, Italy

fLudwig-Maximilians-Universitat, Institut fur Mineralogie, Petrologie und Geochemie - Theresienstrasse 41, D-80333 Munich, GermanygCheMin GmbH-Am Mittleren Moos 48, D-86167 Augsburg, Germany

h Technical University of Munich, Department of Hydrochemistry - Marchioninistrasse 17 D-81377 Munich, Germany

Accepted 29 October 2001

Abstract

The management of residues from thermal waste treatment is an integral part of waste management systems. The primary goal ofmanaging incineration residues is to prevent any impact on our health or environment caused by unacceptable particulate, gaseousand/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 behaviorof residues from thermal waste treatment under different scenarios. General conditions affecting the emission rate of salts andmetals are shown as well as factors relevant to mitigating measures or sources of gaseous emissions.

# 2002 Elsevier Science Ltd. All rights reserved.

Preface

A working group named ‘‘pHOENIX’’ on the ‘‘Man-agement of Municipal Solid Waste Incineration(MSWI) Residues’’ was established as a result of aworkshop held in spring 2002 in Vienna, which dealtwith the practical problems, recent research findings andsolutions related to this topic. As we agreed, there arenumerous highly specific scientific articles as well assome comprehensive studies and books with either in-depth research or with a description of integrated wastemanagement in general terms or with specific MSWIresidues. However, what was missing was a short intro-ductory overview of the management of residues fromthermal MSW treatment for operators, non-specialized

scientists and legislators. With this article, we hope tofill the gap.The pHOENIX working group is composed by the

following members: Peter Lechner (BOKU UniversityVienna, Department of Waste Management, A), Tho-mas Astrup (DTU Technical University of Denmark,Environment & Resources, DK), Giovanna Cappai(University of Cagliari, Department of Geoengineeringand Environmental Technologies, I), Holger Ecke(Lulea University of Technology, SE), Soraya Heuss-Assbichler (Ludwig-Maximilians-Universitat, Institutfur Mineralogie, Petrologie und Geochemie, D), OleHjelmar (DHI Water & Environment, DK), Anders Kihl(Ragn-Sells Avfallsbehandling AB, SE), Ralf Klein(Technical University of Munich, Department ofHydrochemistry, D), Gabriele Magel (Ludwig-Max-imilians-Universitat, Institut fur Mineralogie, Petrologieund 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.

doi:10.1016/S0956-053X(02)00161-7

Waste Management 23 (2003) 61–88

www.elsevier.com/locate/wasman

* Corresponding author. Tel./fax: +39-06-44-585-037.

E-mail address: [email protected] (A. Polettini).

http://www.sciencedirect.com



http://www.elsevier.com/locate/wasman/a4.3d

mailto:[email protected]

Franz Ottner (BOKU University Vienna, Department ofWaste Management, A), Alessandra Polettini (Uni-versity of Rome ‘‘La Sapienza’’, Department ofHydraulics, Transportation and Roads, I), RaffaellaPomi (University of Rome ‘‘La Sapienza’’, Departmentof Hydraulics, Transportation and Roads, I), TamaraRautner (BOKU University Vienna, Department ofWaste Management, A), Henrich Riegler (BOKU Uni-versity Vienna, Department of Waste Management, A),Thomas Sabbas (BOKU University Vienna, Departmentof Waste Management, A), Stefan Salhofer (BOKU Uni-versity Vienna, Department of Waste Management, A).

1. Introduction

The objective of integrated waste management is todeal with society’s waste in an environmentally andeconomically sustainable way. Under the framework ofintegrated waste management, thermal treatment repre-sents a valid option for reducing the amount of waste tobe landfilled, at the same time allowing for wastehygienization. The relative importance of incinerationas opposed to other waste treatment and disposaloptions, including mechanical/biological treatment andsanitary landfilling, varies considerably from countryto country, depending on specific waste managementstrategies as well as space availability for final landdisposal.The increasingly more stringent limits imposed in

recent years on atmospheric emissions from wasteincineration have produced a considerable shift fromthe gaseous emissions to the solid residues of the pro-cess. Thus, solid residues from thermal waste treatmentwarrant significant environmental concern.In our article, we specially focus on assessing the

environmental impacts resulting from residues fromthermal waste treatment, the treatment methods avail-able to mitigate such impacts before and after eitherutilization or final land disposal as well as the processesand 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 systemsinclude all processes from waste generation to land-filling, i.e.:

� Waste generation: all processes which producewaste during the production and distribution ofproducts (industry and commerce) or the con-sumption of products (households);

� Waste collection, including source separationinto different material streams;

� Processing, including such steps as waste sorting,dismantling of products (e.g. end-of-life electricaland electronic equipment), and production ofRefuse Derived Fuel (RDF). All these steps serveeither to prepare waste for reuse or to suitablymodify waste characteristics with a view to finalland disposal;

� Recycling: production of secondary materialsfrom waste, e.g. paper from waste paper, steelfrom ferrous metal scraps etc.;

� Waste treatment, including several technologiessuch as thermal treatment, chemical treatment ofhazardous wastes, mechanical/biological treatment;

� Waste utilization, covering all the utilizationoptions of waste after processing, e.g. use oftreated bottom ash for road construction, com-post for agricultural applications or thermal uti-lization of RDF; and

� Landfilling.

1.2. Treatment methods

Waste treatment methods strongly depend on the typeof waste. As far as municipal solid waste is concerned,the different treatment options are aimed at recoveringmaterials and/or energy from the waste as well as at redu-cing the overall amount and the impacts of waste to belandfilled. 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 inorder 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 oftentreated using various thermal methods. Aside fromcombustion, other thermal processes exist, includingpyrolysis, gasification, sintering, vitrification andmelting.The most common thermal treatment process for

MSW is incineration by mass-burn technology. Flui-dized bed incineration and refuse derived fuel systems areless common in municipal solid waste treatment. Fluidizedbed systems andmulti-hearth furnaces are also widely usedfor sewage sludge incineration, while major furnace typesfor hazardous wastes incineration are grateless systemssuch 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 andoutput are influenced by several factors:

� Waste generators are households and, in addi-tion, industrial or commercial sites.

� Waste generation both in households andindustry is (theoretically) influenced by wasteprevention. In reality we can observe increasingwaste quantities. Onida (2000) reports that thetotal annual production of industrial waste in fivemajor sectors (agriculture, mining, manufactur-ing, municipal and energy production) increasedby 9.5% from 1990 to 1995 in the EU.

� Separate collection exerts a strong influence onthe quantities and quality of waste for incine-ration. For example, the separate collection ofsmall electrical appliances could reduce the Cucontent in MSWI bottom ash by up to 80%.Through source separation of recyclables andbiogenic 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 inputto MSWI.

1.3.1. MSWI residuesAs a result of the incineration process, different solid

and liquid residual materials as well as gaseous effluentsare generated. Approximately one-fourth of the wastemass on a wet basis remains as solids. The volume ofresidues corresponds to one-tenth of the initial wastevolume. Typical residues of MSWI by grate combustionare:

� Bottom ash, which consists primarily of coarsenon-combustible materials and unburned organicmatter collected at the outlet of the combustionchamber in a quenching/cooling tank.

� Grate siftings, including relatively fine materialspassing through the grate and collected at thebottom of the combustion chamber. Grate sift-ings are usually combined with bottom ash, sothat in most cases it is not possible to separate thetwo waste streams. Together bottom ash andgrate siftings typically represent 20–30% by massof the original waste on a wet basis.

� Boiler and economizer ash, which represent thecoarse fraction of the particulate carried over bythe flue gases from the combustion chamber andcollected at the heat recovery section. This streammay constitute up to 10% by mass of the originalwaste on a wet basis.

� Fly ash, the fine particulate matter still in the fluegases downstream of the heat recovery units, isremoved before any further treatment of thegaseous effluents. The amount of fly ash pro-duced by an MSW incinerator is in the order of1–3% of the waste input mass on a wet basis.

� Air pollution control (APC) residues, includingthe particulate material captured after reagentinjection in the acid gas treatment units prior toeffluent gas discharge into the atmosphere. Thisresidue 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. APCresidues are usually in the range of 2% to 5% ofthe original waste on a wet basis.

Due to the volatilization and subsequent conden-sation as well as concentration phenomena actingduring combustion, fly ash and APC residues bearhigh concentrations of heavy metals, salts as well asorganic micro-pollutants. Iron scrap and other metalsare usually recovered from bottom ash and reused inindustry. In some European countries great effortsare devoted to utilization of such residues. If utiliza-tion is not possible due to regulatory constraints orother reasons (such as a sufficient source of naturalraw materials), these residues have to be disposed inan 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 MSWIresidues, is to remove from general circulation materialsand products that are no longer useful in any respect. Itis preferable to do this in a manner that ultimatelyreturns 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 anyunacceptable short- or long-term impact on theenvironment or on human health. Disposal methodsmust ensure that this is accomplished in a sustainablemanner, 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 exploitationof the short- and long-term behavior of the landfilledMSWI residues. Based on this understanding, the designand operation and to some extent also the location ofthe landfill must be adapted to the inherent propertiesof the largely inorganic MSWI residues to ensure thatlong-term emissions of contaminants become or remainenvironmentally acceptable. Landfills should bedesigned to minimize the required lifetime of activeenvironmental protection systems, i.e. systems requiringmaintenance and/or operation. This means, forinstance, that a disposal strategy based on encapsula-tion of the waste is not desirable since it merely post-pones the impact and preserves the contaminationpotential. In principle, this is true both for inorganiccontaminants (salts and metals) due to their intrinsicallyconservative nature and for toxic organic micro-pollu-tants, which are commonly regarded as persistent spe-cies. Yet, the major environmental concerns in relationto the short- and long-term impact of landfilling ofMSWI residues are connected with the risk of leachingand subsequent release of potentially harmful sub-stances, particularly inorganic salts and metals/traceelements, into the environment. Gas production andrelease may also be of some importance, even for MSWIresidues. Leaching of toxic organic compounds (espe-cially PCDDs and PCDFs) is generally believed to be ofminor relevance due to their hydrophobic nature andtheir low concentrations in residues from properlyoperated 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 adefined time scale, a connected activity or a dominantprocess.In the following chapters we will refer to the defini-

tion based on the landfill activity. The basic idea is thateach human generation should take care of its own

wastes, without leaving future generations environ-mental issues still to be resolved. This approach isalso 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 whichlandfill operation and active aftercare (operations thatrequire maintenance, inspection and input of energy,e.g. leachate and gas collection as well as leachatetreatment) are required to meet adequate environmentalprotection levels. On the other hand, ‘‘long term’’represents the timeframe within which the environ-mental safety of the landfill no longer relies on activeprotection systems, but is based on the controlledrelease of contaminants at an environmentally accep-table rate. The long-term period starts just after thecompletion of active aftercare measures. The corre-sponding time scales are shown in Fig. 2.

1.4.1. Disposal scenariosAs incineration residues are produced by high-tem-

perature processes, they are thermodynamicallyunstable under ambient conditions. This renders incin-eration residues highly reactive, especially under wetconditions. This means that they change their mineral-ogical and physico-chemical characteristics as well astheir leaching behavior as long as thermodynamic equi-librium conditions with the surrounding environmentare attained. The specific environmental conditionsinfluence and change the leaching behavior and con-taminant release from such materials during utilizationor final land disposal. To assess the discharge behaviorof a specific waste, it is necessary to take the specificconditions (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 conditionson the variation of waste characteristics overtime, as well as on their environmental behavior

� model the environmental behavior of the wasteand

� validate the model by calibration with the resultsfrom laboratory tests and field experiments andby comparing it to natural analogues.

Such a methodology will also help identify the mostappropriate mitigating measures to be undertakenbefore, during or after utilization or final land disposal.

1.4.2. The EU Landfill DirectiveFuture 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


criteria for acceptance of waste at the different classes oflandfills are laid down in a Council Decision (expectedto be finalized in December 2002), which addresses therequirements of Annex II to the Directive. The EULandfill Directive (LFD) distinguishes technicallybetween three main classes of landfills (landfills for inertwaste, landfills for non-hazardous waste and landfillsfor hazardous waste), but only in terms of the con-tamination potential of the waste and the environmentalprotection measures required at each class of landfill.The LFD does not include any landfill strategy orguideline on the design and operation of landfills aimingat the minimization of the period during which activeaftercare will be necessary. The LFD, however, to a cer-tain extent does allow for the implementation of nationalstrategies and guidelines within the individual EU mem-ber states. At a national level it will be possible to definedifferent sub-classes of non-hazardous waste to preventco-disposal of waste types with different properties (e.g.organic biodegradable waste and inorganic mineralwaste) and different short- and long-term behavior.

2. Processes and factors

At the utilization/disposal site (hereinafter referred toas the application site), MSWI residues will undergo anumber 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 theestimation of the potential environmental impacts aris-ing from the utilization/disposal of MSWI residues aswell as of the measures to be undertaken in order tomitigate the extent of such impacts.The main processes and factors of concern affecting

the utilization or disposal of MSWI residues arestrongly interrelated, so that in many cases a separatedescription of each process and factor, neglecting suchinterrelations, will not be exhaustive. For this reason,the following section will discuss the main processes andfactors on the basis of the above mentioned macro-scopic effects.

2.1. Leachate production

Leaching can be defined as the dissolution of a solubleconstituent from a solid phase into a solvent. Leachingoccurs as a consequence of the chemical reactions tak-ing place at the scale of the individual waste particles aswell as of the contaminant transport processes via thefluid moving through the solid particles. As far asMSWI residues disposal is concerned (see Fig. 3), thetransport medium of pollutants is mainly represented bywater, so that the overall water balance will determinethe 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 andmorphology of the surface soil are among the mainvariables to be accounted for in the water balance. Theapplication site itself then modifies the water infiltrationpattern as a result of the physical and hydrological

Fig. 2. Classification of time scales relevant to landfilling.


characteristics of the material. Thus, the discharge patternalso depends on the pore type, pore distribution, homo-geneity, permeability and field capacity of the material aswell 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 bylong periods of low and nearly constant flows inter-spersed with increases in discharge quantity in responseto rain events. Isotope studies and tracer methods incombination with a simple dilution model show thatfractions of rainwater, which pass through the landfillwith little interaction with the material due to pre-ferential flow paths, make up 20–80% of the dischargevolume during summer rain events and around 10% inwinter months.Other studies (Brechtel, 1984; Stegmann and Ehrig,

1989; Ehrig, 1990; Krumpelbeck, 2000) show that theobserved leachate amounts for uncovered or sparelyvegetated MSW landfills in middle Europe lie between15 and 60% of the annual precipitation. Approximatelythe same fraction of precipitation is observed for MSWIresidues as annual leachate volume (Table 1, referencesherein). Based on experimental data on the influence ofthe 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 selectedAustrian sites was carried out (Table 1, 7th and 8thline). Water balance models, including e.g. HELP andBOWAHALD, can also be used to analyze the effect ofdifferent 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 landfillbut cannot completely prevent leachate formation overa long time scale (refer to the following sections formore details). Once the overall water balance of theapplication site is calculated, the leachate quality needsto be estimated.Mobilization of constituents from inorganic wastes

into the leaching medium is the result of the interactionbetween chemical and physical factors. Chemical factorsinclude waste composition and mineralogy, tempera-ture, pH, redox potential and the presence of ligands,while physical factors are represented by specific surfacearea, particle size, L/S ratio, porosity, hydraulic gra-dient and hydraulic conductivity. Some physical factorsalso affect the percolation pattern (advection, diffusion)and hence the modes of contact between leachate andwaste, which can be caused by leachate flowing aroundthe waste, leachate flowing through the waste or by acombination 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).


in particular, for MSWI residues different groups ofcontaminants 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, whichrepresents a fraction of the total content of con-taminants in the waste itself. When a fluid flows througha loosely packed granular waste material, the amount of

contaminants released is dictated by solubility con-straints, so that the leaching process is referred to asbeing solubility-controlled. In the case of very solublemineral phases, which can completely dissolve as aresult of contact with the fluid, leaching is generallydefined as availability-controlled.Conversely, as far as compacted granular materials or

treated (e.g., solidified) wastes are concerned, leachatepercolation occurs at the surface of the solid material,causing molecular diffusion to be the dominant processdetermining contaminant release from the waste; in thiscase, 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)

Leachate

(% PPT or mm/a)

Remark

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)

Observed

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


leaching. Due to the reduced surface area of the materialin contact with the leachate and to the extremely slowrelease of contaminants from the waste, chemical equili-brium between the solid and the liquid phase is generallynot 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 ofcompaction and permeability) and to the related fluidflow characteristics (including percolation rate, percola-tion pattern and the amount of leachate contacting thewaste). They are also dependent on chemical factorsrelated to the solubility of contaminants in the wastematerial.Leachate composition is the result of reaction between

the various mineral phases in the waste and the leachingfluid. The leachability of strongly soluble species (e.g.,alkali salts) is almost pH-independent, whereas for anumber of contaminants a clear pH-dependence can beobserved. 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 mineralphase(s) in which this is bound. Three main typicalleaching behaviors for solubility-controlled leachinghave been identified:

� cation-forming species and non-amphotericmetal ions (e.g. Cd), see Fig. 4a),

� amphoteric metals (including Al, Pb, Zn), seeFig. 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 highvalues at pH<4, and decreases strongly up to pH 8 to 9,remaining approximately constant or slightly increasingfor higher pH values. Amphoteric metals exhibitincreased solubility under both strongly acidic andstrongly alkaline conditions, resulting in a V-shapedsolubility curve. For oxyanion-forming elements usuallysolubility 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 calculatedfor pure aqueous solutions. For example, the equili-brium concentration of Ba in pure water (20 �C, pH=7,no CO2 dissolved) saturated with barite (BaSO4) is 1.3mg/l. In the presence of gypsum (CaSO4.2H2O), moresulfate will dissolve, and Ba concentration will decreaseto 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 matterof major concern is to predict the pH conditions whichare likely to occur at the application site. These dependon the characteristics of the leaching fluid as well as onthe properties of the waste. Probably the most relevantwaste property affecting the pH of the leachate isrepresented 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 orbases. In the case of MSWI residues, which are mostoften basic in their nature, alkalinity of the material isthe relevant parameter, so that ANC is the appropriatemeasure of neutralization capacity. ANC assesses thesensitivity of the material itself to external influencesand/or internal stresses (e.g. mineralization, organicmatter degradation). As a consequence, the bufferingcapacity of the material affects the evolution of the pHof the leachate over time, thus allowing the expected pHrange for the application site to be estimated. Fig. 6depicts the ANC of a number of bottom ash and fly ashsamples 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 aconsequence 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 comesinto contact with the leachate and dissolved alkalinity isremoved from the system by the leachate. As a con-sequence, the residual alkalinity at any time will dependon the initial alkalinity of the material, the dissolutionof 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 thesolubility of a number of minerals (Ca(OH)2, CaCO3,etc.) and thus on the leaching system pH. The pH inturn is dependent on the system’s ability to buffer theinfiltrating leachate, i.e. on the amount of residualalkalinity 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 expressedthrough the so-called liquid-to-solid (L/S) ratio, alsoaffects the leaching behavior, especially in the case ofsolubility-controlled leaching. The L/S ratio is the resultof climatic conditions, hydrology and hydrogeology ofthe 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 releaseon the L/S ratio. In some cases the linear trend ofcumulative release of a given element as a function ofL/S can be altered by the presence of other species.Delayed release is observed when a sparingly solublephase controlling solubility is present and is depleted


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).


after a relatively short period. In this case, the slope ofthe curve of cumulative release versus L/S is lower forlow L/S ratios, and it increases at higher L/S ratioswhere the sparingly soluble species are depleted. Con-versely, enhanced initial release can be observed in thepresence 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 withincreasing L/S.On the other hand, for availability-controlled leaching

the amount of contaminants released into the solution isat its maximum level due to the high contaminantssolubility and is not dependent on solution pH. At a

given L/S ratio, the transition from solubility-controlledto availability-controlled leaching is evidenced by aconstant concentration in solution with decreasing pH.Availability-controlled leaching results in rapid washoutof the soluble constituents at low L/S ratios, so that theavailable amount often is attained at L/S values of 1 to 2;for higher L/S ratios, the cumulative release remains atthis maximum. Typical examples of such leaching beha-vior are Na, Cl and K (see Fig. 7, which depicts the resultsfrom 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=BaSO4 in pure water at 20�C;

B=equilibrium after addition of gypsum; C=equilibrium after addition of calcite, CO2 (0.04%) and NaCl (0.05 M); D=equilibrium after addition

of calcite, CO2 (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).


diffusion and surface dissolution mechanisms. In thiscase, leaching is kinetically controlled by the rate ofcontaminant release via diffusion, which is measuredthrough the effective diffusion coefficient. Releasemechanisms and physical and chemical retardationfactors affect the diffusion process. Among the physicalretardation factors, porosity, pore structure, degree ofcompaction and tortuosity can significantly slow therate of contaminant release. Pore solution pH andsolubility of elements/species as a function of pH caninfluence the extent of sorption or co-precipitationreactions on solid surfaces, thus acting as chemicalretardation factors.

Irrespective of the mechanism controlling leaching,additional factors including the presence of sorbing/complexing agents, redox reactions and the occurrenceof processes causing mineralogical changes over time(e.g. due to aging/weathering) can also affect theextent of contaminant release, as qualitatively illu-strated in Fig. 8. Among the processes capable ofaltering the leaching behavior of the material, sorptionincludes different mechanisms of adsorption, ionexchange, surface complexation and electrostaticattraction of ions at the surface. During weathering ofless stable phases, new minerals with high surfaceareas 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).


MSWI bottom ash leads to the formation of iron oxi-des, goethite (FeOOH) and hydrous ferric hydroxide([Fe(OH)3]n, often termed HFO). The resulting finelygrained phases are able to sorb heavy metals, includingPb, Cd, Zn, Ni, Cr(III) and Cu, as well as Mo. Similarsorptive properties are also displayed by other mineralphases, including aluminum (hydr)oxides and amor-phous aluminosilicates.In the case of HFO, the general surface complexation

reaction describing sorption of divalent cations can besimplified as:

� Fe-OH0 þMe2þ () � Fe-OMeþ þHþ

where the symbol � indicates bonds at the surface and�FeOH0 represents [Fe(OH)3]n.Fig. 9 demonstrates the demobilizing effect and limits

of sorption. A number of sorption experiments werecarried out on pre-washed (L/S=10 achieved throughcolumn percolation) and artificially weathered (bymeans of carbonation) MSWI bottom ash (�<2 mm).The aged MSWI bottom ash was added with a nickelsulfate solution at Ni concentrations varying between0.06 and 16.4 mg/g, thereby continuously aerated toachieve equilibrium with atmospheric CO2. The experi-mental sorption isotherms (see Fig. 9) revealed strongsorption phenomena, as long as the total amount of Niin the system was low. Sorption proceeded as long asactive sorption sites were available within the solidmaterial. In this case, the Ni concentration in the solu-tion was lower than that predicted based on Ni(OH)2solubility. For high total contents of Ni, the amount ofNi exceeding the sorption capacity of the material was

such that saturation or slight over saturation of thesolution with respect to Ni(OH)2 was attained.The presence of complexing agents can also sig-

nificantly alter the extent of contaminant leaching fromMSWI residues. Complexing agents can be eitherorganic or inorganic in their nature; dissolved organiccarbon (DOC) and chloride are the main complexingagents of concern for such materials. DOC has beenextensively shown to be responsible for increasing cop-per release from predominantly inorganic waste forms(IAWG, 1997; Van der Sloot et al, 1999; Van der Slootet al., 2001).Oxidation/reduction reactions also play a role in

determining the release of contaminants from MSWIresidues. The main oxidizing or reducing agents of rele-vance for MSWI residue monofills are reported inTable 3.Leaching of contaminants from waste incineration

residues can be affected by the redox conditions accord-ing to two main mechanisms. One mechanism relies onthe different solubility and toxicity of the contaminantsunder concern for MSWI residues depending on theiroxidation state. These issues influence both the strengthof the leachate and the related potential environmental

Table 3

Relevant reducing/oxidizing agents (inorganic landfills)

Short

timescale

Medium/long

timescale

Reducing agents H2; metals (Al, Fe, Zn); Fe-II metals, Fe-II

Oxidizing agents O2; H2O O2

Fig. 9. Ni sorption for weathered MSWI bottom ash with Ni added at different concentrations.


impact. For example, it is well known that in an alkalineenvironment Cr(III) may be rapidly oxidized by atmo-spheric oxygen to Cr(VI), which is much more toxic andmobile than Cr(III). However, for Cr(III) to be oxidizedto Cr(VI), high values of the redox potential arerequired.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 immobilizingmetal 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 betransformed into more soluble forms of Fe(II) andMn(II) under moderately reducing conditions. Underseverely reducing conditions, S(VI) is reduced to ele-mental sulfur and sulfide, resulting in the precipitationof metal sulfides, which are among the less soluble metalforms.The above mentioned mechanisms can lead to either

synergistic or antagonistic interactions, so that theinfluence of redox processes on leaching may result ineither mobilization or demobilization of contaminants(see Table 4 and Fig. 8). It should be emphasized thatthe observations presented in Table 4 may not reflect areal landfill scenario, but merely indicate a significantinfluence 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 abiotictransformations leading to the formation of reducinggases (H2). For bottom ash monofills, the presence ofunburned organic material and H2 generally leads toreducing conditions; in such cases, the leaching behaviorof contaminants is the result of, on the one hand, com-plexation by DOC and, on the other hand, precipitationof less soluble species, including for example insolublesulfides.Weathering is a process, which naturally occurs in

incineration residues as a consequence of several factorssuch as pH, redox potential, temperature and humidityconditions as well as the concentration of certain com-ponents (e.g. CO2) in the application site. Weathering

results in the occurrence of slow mineralogical changesover time, which may alter the leaching of trace metalsfrom the material either in the medium or in the longterm. Due to weathering and the related neoformationof minerals, key factors such as pH are subjected tochanges over time.Weathering of bottom ash is a process, which deserves

particular concern. Incinerator bottom ash is composedof high-temperature solids formed as a consequence ofrapid quenching of the material exiting the combustionchamber, many of which are metastable under naturalconditions. Typically, such solids will thereby undergo anumber of chemical reactions while in the landfill lead-ing to more stable mineral phases or phase assemblages(Meima and Comans, 1997; Zevenbergen and Comans,1994).Weathering is the result of a complex series of several

interrelated processes, including hydrolysis, hydration,dissolution/precipitation, carbonation, complexationwith organic and inorganic ligands, surface complexa-tion, surface (co)precipitation, sorption, and formationof solid solutions as well as oxidation/reduction (Beleviet al., 1992; Bodenan et al., 2000; Meima and Comans,1997; Meima and Comans, 1999; Zevenbergen andComans, 1994). All of the mineralogical and chemicalchanges caused by such processes are also accompaniedby physical changes such as pore cementation, changesin grain size and pore size distribution, which in turnalter the hydrological characteristics of the material.Hydrolysis starts immediately after bottom ash

quenching and can be prolonged over the time span oftemporary storage or landfilling of the material (Beleviet al., 1992; Johnson et al., 1995), as long as it is incontact with water. Hydrolysis involves the transfor-mation of oxides of Ca, Na and K and non-noble metalslike Al and Fe into the corresponding hydroxide species(e.g. CaO!Ca(OH)2, Al2O3!Al(OH)3) (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 otherminerals 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 H2O2 (after Fallmann, 1997)

Oxidation increases leachability (95% significance) Oxidation decreases leachability (95% significance)

Oxidizing agent: H2O2 Oxidizing agent: H2O2

Blast furnace

slag

Steel smelter

slag

MSWI bottom

ash

Wood

ash

Blast furnace

slag

Steel smelter

slag

MSWI bottom

ash

Wood

ash

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


Comans, 1997); for instance, ettringite can be formedaccording to the reaction (Meima and Comans, 1997):

6Ca2þ þ 2Al3þ þ 3SO24 þ 38 H2O !

! 12Hþ þ Ca6Al2ðSO4Þ3ðOHÞ12 �26H2OðsÞ:

The formation of C-S-H phases was also detected(Speiser et al., 2000) as well as the neo-formation ofclay-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

CO2 by the initially alkaline material, which leads to adecrease in pH and to the precipitation of calcite(Meima and Comans, 1997; Meima and Comans, 1999;Zevenbergen and Comans, 1994). CO2 absorptionresults in final pH values in the range of 8 to 8.5 (Bod-enan et al., 2000; Meima and Comans, 1999). In thisstate the equilibrium between calcite and CO2 (formingHCO3

- with water) under the influence of gypsum dom-inates the system as a buffer. At pH ffi 8 the solubilityminimums are reached for most of the solid phasescontrolling the leaching of such heavy metals as Cd, Pb,Zn, Cu and Mo (Meima and Comans, 1999). Calcitecan also provide a number of sorption sites for certainelements, 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 weatheredbottom ash has been found to increase if compared tofresh bottom ash (Bodenan 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 toplay a role in reducing contaminant leaching fromweathered bottom ash. Fe and Al (hydr)oxides as wellas amorphous aluminosilicates formed as a result ofweathering have been found to be reactive sorptiveminerals for e.g. Cd, Zn, Cu, Pb and Mo (Meima andComans, 1998; Meima and Comans, 1999).Similar processes are observed during the weathering

of APC residues. In addition to the above mentionedphases, aging of APC residues can lead to the formationof 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 etal., 2002).Fig. 10 provides a schematic representation of the

weathering reactions and the related modifications inleaching behavior.

2.2. Gas production

Gas generation at landfill sites with MSWI residuescan be either of a biotic or abiotic nature. The low bio-degradable organic carbon content of MSWI residuesgenerally leads to the production of biogas amounts sig-nificantly lower if compared to MSW landfill gas. Con-versely, the evidence of significant abiotic gas generationhas been reported in a number of studies (Musselmann et

Fig. 10. Layout of mineralogical reactions as a consequence of leaching and weathering processes.


al., 2002; Magel et al., 2001, Lechner et al., 1997).Abiotic gas is produced by chemical oxidation, in thepresence of water, of elemental metals including Al, Feand Cu. Thus, as previously observed for leachate pro-duction, the different parameters of concern for thewater balance and the chemical properties of the wastematerial must be considered for gas production as well.Aluminum is a major constituent of bottom ash and isalso significantly concentrated in fly ash and APC resi-dues. Due to its high solubility at pH > 9.5 and to itslower redox potential if compared to other elements,aluminum is regarded as the main element responsiblefor abiotic gas production. In addition, a significantfraction of Al in MSWI residues is in its elemental form,which can undergo the following redox reactions, lead-ing to hydrogen gas generation:

2Al0 þ 3H2O ! Al2O3 þH2

Al0 þ 2H2O ! AlOOH þ 1:5H2

Al0 þ 3H2O ! Al OHð Þ3þ1:5H2

However, the chemistry of aluminum corrosion is notcompletely understood at present, due to the variabilityof local conditions throughout the landfill mass.A number of studies (Forster and Hirschmann, 1997;

Mizutani et al., 2000) evidenced that abiotic gas pro-duction is almost complete after several months, sohydrogen generation can be considered a short-termprocess. However, even though no specific studies havebeen carried out, some experimental data suggest thathydrogen gas production can also evolve over a longerterm, 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)3 and additionally by hydro-calumite (Ca2Al(OH)6Cl.2H2O) and ettringite ([Ca3Al(OH)6]2(SO4)3.26H2O). These coatings or by productsmay result in the retardation of hydrogen production.However, such rims can dissolve, leading to a perma-nent release of hydrogen. Furthermore, a significantnumber of ash particles are enclosed in glassy phasesformed during incineration, which act as a barrieragainst the reaction between water and aluminum.However, due to their alkaline nature, such glassy pha-ses can be altered, so aluminum particles will come intocontact with the hydration water.

2.3. Temperature development

Recently several studies have shown that many exo-thermic reactions may cause a temperature increase ofup 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 differentexothermic reactions, such as hydration of alkaline andalkaline earth oxides, corrosion of metals and carbona-tion of portlandite (Huber, 1998). The main effects ofthis temperature enhancement are (see Fig. 12):

� acceleration of weathering/hydration reactions aslong as the material in the landfill is wet orhumid,

� over a longer term, formation of salt rims as aconsequence of drying, and

� modification of precipitation/dissolution andcomplexation equilibria.

Moreover, temperature increase may lead to eva-poration of some dissolved gaseous compounds (CO2,

Fig. 11. Hydrogen content of gas from a German MSWI residues monofill (LEL: lower explosive limit of H2).


O2) from the leachate and as a consequence may exertan effect on the redox potential and the concentration ofcomplexing agents.Temperature development was studied by Klein et al.

(2001) during an experimental campaign carried out at aGerman bottom ash monofill. Fig. 13 shows tempera-ture development over time in three monitored sensorfields. In Sensor Field 1 (SF1) bottom ash was deposited

at irregular time intervals, after one to three weeks ofstorage at the landfill site. SF2 was built up over threeweeks to its final height of ten meters. In SF3 bottom ashwas placed in 1 meter-thick layers every two months upto a final height of 6 meters. Previous storage of bottomash in this sensor field was disregarded. Bottom ash in allthe sensor fields was not compacted and no temporaryliners were used to cover the landfill between deposits.In every layer of the surveyed landfill the temperature

development started with an increase immediately afterdeposition. Over the next three to four months the bot-tom ash temperatures increased to a maximum whichvaried depending on the layer depth. The average rate atwhich the temperatures rose was between 0.16 and1.02 �C per day. In all the observed landfill layers, themaximum temperature occurred at a time of about four tofive months after deposition. The initial temperature risesandmaximum temperatures occurred in those cases wherethe ash was not stored temporarily before landfilling.The experimental program also revealed that rain-

water percolating through the landfill body exerted noappreciable effect on temperature development.

3. Potential environmental impacts

The main potential environmental impacts related tothe handling, utilization and disposal of MSWI residuescan 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.


� dust emissions,� leachate generation,� gas emissions, and� temperature increase.

Moreover, as far as landfilling is concerned, thepotential impacts arising from the construction andoperation of the landfill should be taken into account.However, such impacts as additional traffic load, noisepollution and site modifications in terms of topo-graphic, hydrological and hydrogeological conditionswill not be discussed in this paper. The extent of themost relevant impacts may differ for the handling, utili-zation and final disposal phases and vary significantlyover time.

3.1. Dust emissions

MSWI residues contain a fine fraction of particulatematter passing the 74 mm mesh sieve that accounts for1–10% 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 chlorideand sulfate salts as well as heavy metals like Pb, Cu andZn (IAWG, 1997).The easily airborne nature of fine particles leads to the

dispersion of pollutants, which in turn can give rise tohealth risks for exposed, unprotected workers and thepublic, as well as soil contamination. To prevent orminimize dust emissions, bottom ash and fly ash arenormally kept wet (5–15% humidity) and transportedby covered and watertight trucks. The upper limit forhumidity is regarded as the minimum value required toprevent fugitive dust problems in open storage piles.The ability to maintain the optimal water content inorder to minimize dust emissions is obviously related toclimatic conditions (temperature, humidity as well asregime, intensity and frequency of the dominant localwinds), which influence the desiccation rate of thematerial, the critical area of downwind dust depositionas well as the downwind distance at which dust can betransported. Matsuto et al. (2001) investigated the winddispersion of incinerated residues and found that theyare dispersed up to 50 meters from the landfill. Theatmospheric dispersion of particles may not be sig-nificant during the after-closure period of a landfill dueto the presence of a top cover.

3.2. Leachate generation

The potential environmental impact of leachingincludes contamination of soil, groundwater and surfacewater bodies. As leaching of contaminants from MSWIresidues may occur during the temporary storage,

treatment or reuse as well as during the final disposal ofthe material, the following aspects should be investi-gated: the leaching behavior of contaminants, theenvironmental conditions that may occur in any ofthe above mentioned scenarios, as well as their var-iation over time. Thus, according to the discussion inthe preceding sections, the following items should beconsidered:

� residue characteristics, in terms of physical andmechanical properties, particle size distribution,acid neutralization capacity, concentration ofcontaminants, availability of contaminants forleaching, leaching mechanisms, controlling fac-tors, and their variation over time due toweathering 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 interms of (1) net rate of infiltration, (2) propertiesof the unsaturated zone and the aquifer (thick-ness, permeability, porosity, longitudinal andhorizontal dispersivity, bulk density, flow velo-city, etc.); moreover, as far as disposal is con-cerned, the presence of a top cover should betaken 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 whichleaching occurs and on the type and concentration ofthe dissolved species. The following elements must beconsidered as hazardous contaminants potentiallyleachable from MSWI residues: As, Al, B, Ba, Cd, Cr,Cu, Hg, Mn, Mo, Ni, Pb, Sb, Se, Zn, Br, Cl, CN,F, NH4

+, NO3, NO2

, SO42. Thus, their concentration

in the leachate should be compared to specific qualitycriteria. Since no international guidelines for ground-water have been proposed so far, it appears reasonableto apply the international drinking water quality criteria(EU Drinking Water Directive and WHO criteria ondrinking water) until a specific groundwater directive isproposed.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 behaviorof most MSWI residues (results of leaching tests andfield measurements), long-term behavior can be pre-dicted only on the basis of a synthesis of information onleaching 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 residuesis concerned, availability as opposed to the total con-centration of contaminants in the solid matrix provides


an estimation of the maximum amount of contaminantthat in theory could be leached over a 1000- to 10000-year timeframe (with the exception of highly solublesalts, for which the maximum leachable amount can beattained within shorter periods, typically a couple ofyears).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 animportant parameter, which determines the potentialenvironmental impact of MSWI residues, in that neu-tralization processes can affect leaching reactions andcontrol the release of contaminants from the material.Acid neutralization capacity allows for the evaluation ofthe environmental behavior of MSWI residues, in thatANC data can be transformed in order to estimate thetime required for the pH to drop from the ‘‘inherent’’pH of the material to critical values for contaminantsrelease (Astrup et al., 2001). Such time-related informa-tion can be gathered on the basis of the size of theapplication site, hydrological and hydrogeological con-ditions, leachate composition, and leachate flowtowards soils and groundwater.A more detailed simulation of the potential environ-

mental behavior of MSWI residues should also considerthe attenuation phenomena caused by the interactionsbetween leachate and soil, such as sorption and ionexchange (Hjelmar et al., 2001; Hjelmar et al, 1999a,b).However, in order to predict correctly the leachingbehavior over time, additional information (pH, redoxconditions, ionic strength, complexing agents, andmineralogy) is required. It has been observed (Hjelmar,1996) that the first leachate produced by bottom ash hasa 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 thisstage reducing conditions are occurring and pH isslightly or strongly alkaline, depending on the degree ofcarbonation. Among the trace elements, Cu can behavedifferently, as observed before, its leachability being

increased by DOC. With the exception of sulfate, theconcentration of salts in the leachate tends to decreaseover time (i.e. at increasing L/S ratios).APC residues behave differently from one another

depending on their origin and air pollution controldevices installed into the plant. High concentrations ofreadily soluble salts, such as chlorides and hydroxides ofcalcium, sodium and potassium generally characterizethe first leachate from APC residues. Trace elements,such as Pb and Mo, which are mobile under reducingand slightly alkaline conditions, can be highly leachableat low L/S ratios (corresponding to the first fractions ofthe leachate). Thus APC residues are typically hazar-dous materials, and their disposal requires considerablecare in order to prevent adverse environmental impacts.As far as co-disposal of bottom ash and APC residues isconcerned, it was observed that soluble salt concentra-tions are higher in the combined ash leachate than inbottom ash and fly ash leachate (Hjelmar, 1996). Thisbehavior was also observed for Cd and is likely to beascribed to high complexing chloride content due to thepresence of organic acids produced by biodegradationof residual unburned carbonaceous material in bottomash.Very little data are available from the literature

regarding the long-term composition of leachate fromMSWI residues. Sabbas et al. (2001) investigated fourdifferent 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) amixture of rotary drum kiln slag from hazardous wasteincineration and fluidized bed incineration ash (RAU-A) as well as 4) a cement solidified product (RAU-SB, amixture of MSWI bottom ash and fly ash, rotary drumkiln slag from hazardous waste incineration and flui-dized bed incineration ash, cement and gravel). Thematerials were approximately 10–15 years old, naturallyweathered, fully carbonated and showed leachate pHvalues between 7.3 and 8.9. The results shown in Fig. 14reveal 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

ash

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, Ca

10–100 g/l Na, K, Pb Cl, Na, K

1–10 g/l SO42, Cl, Na, K, Ca Zn SO4

2, Ca

100–1,000 mg/l NVOC, NH4N NVOC, SO4

2

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


and mercury, for the solidified product fall below theproposed EU acceptance criteria for waste (TAC Mod-eling Group, 2002).

3.3. Gas emissions

As described in the previous sections, gas productionoccurs from both MSWI bottom ash and APC residuesas a result of metallic aluminum hydration.As pointed out by Forstner and Hirschmann (1997),

hydrogen production from unquenched bottom ash ishigher if compared to quenched bottom ash. However,up to now no accident has been reported, which isobviously the result of generally good aeration. Unlikethe situation of bottom ash, a number of authors(Takatsuki, 1994; Yasuda, 1997) have reported explo-sions resulting from hydrogen generation from APCresidues in which people were injured or died. Suchaccidents occurred when ash blocks were crushed orwater 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 upto 20-year-old residues (Magel et al., 2001).

3.4. Temperature development

Bottom ash storage sites are affected by heat generationas a result of different exothermic reactions, such ashydration of alkaline and alkaline earth oxides, corrosionof metals and carbonation of portlandite (Huber, 1998).As far as bottom ash landfilling and the design of

monofills is concerned, the potential impact arisingfrom the heat-related damage on the landfill liner andon the leachate collection system should be taken intoaccount. Temperature development can last over longperiods (decades or longer) due to the low rate of heattransfer through the residue bulk. At temperatures over50 �C, clay liners are liable to desiccation and thereforecracks may form; furthermore, HDPE membrane linersare subject to ruptures. Both effects may lead to loss ofsealing 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,

2002)].


of the residues themselves, steam generation and escapefrom the surface was observed (Heyer and Stegmann,1997).

4. Mitigating measures

This chapter will focus on the various mitigatingmeasures aimed at reducing the potential environmentalimpacts from incineration residue reuse or disposal. Itmust be stressed that when dealing with the availabletreatment methods, both the environmental behavior ofthe residue under consideration and its intended desti-nation should be kept in mind. In other words, anydecision concerning reuse, treatment or disposal ofincineration residues can be taken only after theirenvironmental quality has been assessed. Nevertheless,the relevant conditions for the intended applicationmust be evaluated carefully, which also includes theidentification of the elements to be controlled. In lightof this, as far as final disposal is concerned, the land-filling strategies for waste combustion residues shouldreflect the fundamental differences existing betweensuch residues and the original waste. Thus, it can beanticipated that the common practices used for rawmunicipal solid waste landfilling may not be adequatewhen managing incineration residue disposal.Various options are available for the treatment of

waste incineration residues in view of their reuse or finaldisposal. Such measures can be applied at different stagesof their life cycle (from generation to reuse and/or finaldisposal) and can be based on different treatmentapproaches. Accordingly, one may distinguish between:

� measures undertaken prior to reuse or final dis-posal;

� measures undertaken during landfilling andactive landfill operation; and

� measures carried out during the passive phase oflandfilling.

A number of strategies affecting the environmentalproperties of incineration residues may also be appliedprior to or during the combustion process. Examples ofsuch strategies may include either waste sorting andselection prior to incineration in order to beneficiallymodify the waste input composition (homogeneity,chlorine content, metal content, etc.) or control ofoperating parameters (temperature, oxygen concen-tration, etc.) during the combustion process. Based onthe treatment principles, the pretreatment options canbe 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 canbe divided into:

� landfill design options and� landfill operation strategies.

In general, as depicted in Fig. 15, it may be stated thatthe basic principles of the measures to mitigate theenvironmental impact of incineration residues are basedon variations in either (a) the total content, (b) theavailability for leaching or (c) the release rate of con-taminants into the environment; combinations of one ormore of the three mechanisms shown in Fig. 15 are alsopossible. For instance, washing pretreatments aimed atremoving e.g. readily soluble salts act according tomechanism (a), so that the availability for leaching isreduced as a consequence of the reduction in total con-tent. Stabilization pretreatments modify the release rateof contaminants [mechanism (c)] and may also reducethe 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 thespecific contaminant under consideration and on the

Fig. 15. Principles of mitigating measures in respect to total content, availability and release rate.


nature of the process applied. Examples of treatmentsfor 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 oraccelerate the release in the short term, instead ofdecreasing 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 behaviorunder the expected conditions must be considered care-fully. It should also be emphasized that most of theavailable treatments generate a number of wastestreams and are also responsible for raw material andenergy consumption. Thus, the technical, environmentaland economic applicability of any treatment optionshould be judged on the basis of an appropriate eval-uation of the overall mass and energy balances under alife-cycle assessment framework.

4.1. Treatment options

4.1.1. Measures undertaken prior to reuse or finaldisposalTable 6 reports the various treatment options applic-

able to waste incineration residues prior to their reuse orfinal disposal according to the classification discussed inthe previous section. The most common treatment pro-cesses will be dealt with in this section.

4.1.1.1. Physical and chemical separation. Physicalseparation methods have only a limited effect on thequality of residues, as their principle is to separate fromthe bulk of the material the individual constituents that

are already present in such residues in the same physicaland chemical form (IAWG, 1997). However, physicalseparation is able to remove specific materials that ren-der some residue streams (e.g. bottom ash) unsuitablefor a number of applications.Among the physical separation options, particle size-

based separation is generally carried out for two mainpurposes. One aim is accomplished through isolatingthe fraction(s) of the material (usually the finer frac-tion), which is more concentrated in contaminants, thusreducing the environmental impact of the residualstream. In this case, selecting the appropriate cut sizefor separation is of crucial importance in order toreduce effectively the leachability of the contaminants inquestion. One major drawback to this kind of separa-tion is that for some residues, e.g. bottom ash, the finerfraction usually constitutes a considerable portion ofthe total mass of the material [the percentage passing atthe 2 mm sieve is in the order of 30% by weight (IAWG,1997)].The second aim of particle size separation is to pro-

duce a material where the engineering properties, suchas particle size gradation and hydraulic conductivity,are more suitable for subsequent utilization. Yet finematerial can potentially create problems in bottom ashreuse, in that it is highly sorptive for water, asphalticcement and Portland cement (IAWG, 1997). Forinstance, when MSWI bottom ash is to be reused as acoarse aggregate in asphalt or concrete mixtures, ascreening pretreatment to remove the 1- or 2-mmundersize 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 onbottom ash to reduce its ferrous and non-ferrous metalcontent, respectively. According to the IAWG (1997)and Wiles (1996), the ferrous metal content of MSWIbottom ash ranges from 7 to 15% by weight, while non-ferrous metals account for approximately 1–2% byweight although such figures strongly depend on wastesorting and selection strategies prior to the combustionprocess.Metal separation from bottom ash may be performed

with a view to either metal scrap recovery or toimprovement of bottom ash properties for its utiliza-tion. In this case, metal removal from bottom ash maybeneficially prevent corrosion phenomena arising fromoxidation of metals like Al, Fe and Zn, which, as dis-cussed in the previous sections, may cause swelling andexpansion 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 removinghighly 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

Washing

Chemical extraction/mobilisation

Chemical precipitation

Ion exchange

Adsorption

Crystallisation/evaporation

Solidification and/or

stabilisation

Solidification/stabilisation

with hydraulic binders

Chemical stabilisation

Ageing/weathering

Thermal treatment Sintering

Vitrification

Melting


residues. The soluble constituents, typically removed asa consequence of washing, are mainly represented bychloride and alkali ions. Conversely, due to the highlyalkaline nature of incineration residues, the effectivenessof washing on trace metals removal has proved to berelatively low (IAWG, 1997; Schneider et al., 1994), inthat they mostly form sparingly soluble compoundsunder 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 thedissolution process from being attained. Thus, bottomash after quenching still is typically characterized by aresidual content of soluble components, which can befurther extracted through a washing treatment. Due tothe above-mentioned reasons, washing of bottom ash isa simple measure that could easily be combined with thequenching stage at the combustion plant. However,washing alone may not be adequate for bottom ash toreach a suitable level of quality for subsequent utiliza-tion according to established regulatory limits. Thus,the washing treatment may be carried out beneficially incombination with other processes, e.g. chemical mobili-zation or aging (see below) (IAWG, 1997; Lahl, 1992),although a number of studies have indicated thatmobilization of soluble salts mostly occurs during theinitial washing stage (Schneider et al., 1994).As far as APC residue is concerned, washing with

water can also be applied, commonly as a pretreatmentstage prior to further chemical stabilization processes(Derie, 1996; Lundtorp et al., 1999; Mangialardi et al.,1999; Nzihou and Sharrock, 2002) in order to removesoluble salts. Such salts may account for up to approxi-mately 20% of the material and are responsible formuch 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 dryand semi-dry APC residues, that the high pH of thematerial coupled with the large concentrations ofhighly-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 washingprocess. However, the extent of chemical mobilization isnot suitable for adequate APC residue detoxificationlevels to be attained (Lamers and Born, 1994), andtypically the material still needs additional treatmentprior to final disposal. Such treatment will most ofteninclude either chemical stabilization or solidificationwith 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 (mainlychlorides). For other anionic species, e.g. sulfates, whichare 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 mainlycomposed of halite, sylvite and gypsum (Nzihou andSharrock, 2002). To prevent contamination of thewashing solution by heavy metals, the use of additiveswhich 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/orstabilization treatments (Derie, 1996; Mangialardi etal., 1999; Nzihou and Sharrock, 2002). In this case,removal of chloride and water-soluble sulfate as well asalkali ions (which are known to negatively interferewith binder hydration) may allow for significantimprovements in the physical and mechanical proper-ties of the final products. However, when evaluatingthe overall benefits of a washing treatment, the shift ofthe pollution problems from solid to liquid wastestreams must be carefully considered by designers anddecision 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, treatmentsconsisting of sodium carbonate or sodium bicarbonateaddition are beneficial. This form of treatment has theeffect of mobilizing sulphate through the formation ofsoluble Na2SO4 and precipitation of CaCO3.Several chemical extraction processes have been

proposed for APC residues (Hong et al., 2000b; IAWG,1997; Katsuura et al., 1996; Laethem et al., 1994). Theaim of both processes is to recover heavy metals andto detoxify the material. A number of treatmentmethods have been proposed using inorganic acids,including hydrochloric, nitric or sulfuric acid and aquaregia (Hong et al., 2000a; Hong et al., 2000b; Katsuuraet al., 1996; Laethem et al., 1994), as well as chelatingagents 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 metalextraction has been found to be strongly dependent onthe pH and L/S ratio as well as on both the nature ofthe extracting agent used and the particular metal ofconcern.

4.1.1.2. Solidification and stabilization. Solidification/stabilization treatments are among the most widespreadprocesses used for waste incineration residues, mainlyAPC ash (e.g. Conner, 1990; Gilliam and Wiles, 1996).The main purpose of solidification/stabilization is toproduce a material whose physical (specific surface area,porosity, tortuosity, etc.), mechanical (durability,mechanical strength, etc.) and chemical properties aremore favorable with respect to reducing the leachabilityof contaminants out of the waste matrix.


The most common processes make use of hydraulicbinders including cement, lime and/or pozzolanic mate-rials. In general, it can be stated that improvement inthe leaching behavior of solidified/stabilized incinera-tion residues is attained through either physical or che-mical immobilization mechanisms, depending on thespecific contaminant of concern as well as on the type ofbinder used. However, weak stabilization efficienciestypically have been recorded for soluble salts. Further-more, due to their strong amphoteric behavior, treat-ment of zinc and lead with cement- and lime-basedprocesses may be problematic, unless incorporation inthe crystal lattice of the hydration products occurs orappropriate additives are used.Chemical stabilization processes have been proposed,

which basically involve chemical precipitation of heavymetal-incorporating insoluble compounds and/or heavymetal substitution/adsorption into various mineral spe-cies. The principal forms of chemical agents usedinclude 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 lowcosts. However, solidification/stabilization withhydraulic binders results in increased amounts to belandfilled 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 alteringmineralogical phases in MSWI residues over time. Suchchanges may lead to significant reductions in trace ele-ments (including heavy metals as Cd, Cu, Pb, Zn andMo) and leaching (Meima and Comans, 1999; Zeven-bergen and Comans, 1994; Zevenbergen et al., 1996) asa result of hydration, carbonation or oxidation/reduc-tion. These can give rise to pH decrease, contaminantsorption processes as well as formation of more stablemineral species (Meima and Comans, 1997; Meima andComans, 1999). Aging and weathering can be benefi-cially applied particularly to reactive materials such asbottom ash, as this is composed of high-temperaturesolids which are metastable under natural conditionsand are therefore likely to undergo a number of miner-alogical changes. From this perspective, bottom ash,prior to utilization, is commonly aged through storagein stockpiles open to the atmosphere for periods rangingfrom 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 wastematrix. With a view to this, accelerated carbonation hasbeen proposed as an efficient treatment for reducingleaching of soluble salts, Pb and Zn, although it has

been shown to have the potential of mobilizing sulfate(Bodenan et al., 2000).

4.1.1.3. Thermal treatment. The thermal treatment ofincineration residues is used extensively in some coun-tries to obtain reduced leaching from the residues andreduced volume as well as a treated material that is sui-table for reuse. Thermal treatment can be grouped intothree categories: vitrification, melting and sintering(IAWG, 1997).Vitrification is a process whereby residues are mixed

with glass precursor materials and then combined athigh temperatures into a single-phase amorphous,glassy product. Typical vitrification temperatures are at1000–1500 �C. The retention mechanisms are chemicalbonding of inorganic species in the residues with glass-forming materials, such as silica, and encapsulation ofresidue 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 ina multiple-phased product. Often several molten metalphases are produced. It is possible to separate specificmetal phases from the melted product and recycle thesemetals, perhaps after refinement. Temperatures aresimilar to those used in vitrifying.Sintering involves heating the residues to a level at

which bonding of particles occurs and chemical phasesin the residues reconfigure. This leads to a denser pro-duct with less porosity and a higher strength. Typicaltemperatures are around 900 �C. When MSW is incin-erated, some level of sintering will typically take place inthe incineration furnace. This is especially the case if arotary kiln is used as part of the incineration process.Regardless of the process, thermal treatment of

incineration residues in most cases results in a morehomogeneous, denser product with improved leachingproperties. Vitrifying also adds the benefits of physicalencapsulation of contaminants in the glass matrix. Amajor drawback to these methods, however, is that theyrequire substantial amounts of energy and especiallyvitrifying and melting result in the mobilization ofvolatile elements such as Hg, Pb and Zn during thethermal treatment process.

4.1.1.4. Combined methods. On the basis of the resultsfrom batch-scale extraction treatments, a number ofmultistage processes have also been developed either ata pilot- or full-scale level. For the implementation ofsuch processes, which typically involve different combi-nations of the above-mentioned treatments, efforts havebeen 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 tokeep the net discharge from the overall treatment to aminimum. Among the several processes developed so


far, probably the most used are the 3-R (Vehlow et al.,1990), the MR (Stubenvoll, 1989), the AES (Katsuuraet al., 1996) and the VKI process (Hjelmar et al.,1999a). As a detailed description of such processes isbeyond the scope of this paper, the reader can refer tothe literature cited for additional information.

4.1.2. Measures undertaken during landfilling and activelandfill operationThe landfilling phase includes the period of time in

which the landfill receives waste. The time span of thisphase will often amount to about 5–30 years dependingon the capacity of the landfill. When the last section ofthe landfill is closed and the site is no longer used foractive landfilling, a period of active post closure carebegins. This period can include active measures toreduce the environmental impact of the landfilled wasteas 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 surroundingenvironment. The extent of this period can be very dif-ficult to estimate and depends on the characteristics ofthe 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 incinerationresidues. Any landfilling of residues should include anassessment of the environmental impact, both in theshort- and long-term perspective, and the landfillingshould be part of a proper disposal strategy which takesthe 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 environmentalprotection systems. This should include consideration ofthe properties of the waste, the potential risks related tohandling and landfilling the residues (as discussed in theprevious chapter) and thereby also the long-term effectsof keeping the residues at the disposal site as well as anyderived consequence of operating the landfill (for exam-ple, effects of leachate management). At the terminationof the active care period, the emissions from the residuesremaining 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 bediscussed in the following section. First are the issuesconcerning 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 thelandfill:

� 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.

4.1.2.1. Landfill design. The geological, hydrological andgeotechnical characteristics of the location of a plannedlandfill are by far the most important issues in relation tothe 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 ofreleased contaminants, soil bearing capacity, seismic activ-ity, etc. should be considered when choosing the site. It isobviously desired to place the landfill at a location that iscapable of coping with the expected emissions from thelandfill.The sectioning of the landfill provides for the possibi-

lity of disposing of specific waste types separately and ofclosing certain parts of the landfill before other parts.The sectioning will most often depend on the number ofdifferent waste types as well as the amount of thesewastes that are expected to be deposited at the landfillduring its lifetime. However, the construction of newsections within the landfill area typically is carried outsequentially according to the capacity needs to landfill aspecific waste type, and it is thereby possible for theoperators to adapt the sectioning plan according tofuture needs. From an environmental as well as opera-tional point of view, it would be reasonable to keep to aminimum the number of open sections and the time theyare open, thereby minimizing the potential impact onthe surroundings.The height of the landfill in many cases is determined

by legal regulations, economic considerations or limitsimposed by the physical properties of the site. With agiven net infiltration to the landfill, the height, on theother hand, will determine the overall time to reach aspecific L/S ratio and therefore will be a factor in con-trolling the relationship between leaching and time. Inmost cases, landfill designers seek to keep the height to amaximum, thus prolonging the leaching process asmuch as possible into the future. This, however, mayhave the effect of extending the time required for theactive operation of the landfill site and may not alwaysbe the best solution. In any case it is desirable tominimize the period in which leaching of contaminantsis at unacceptable levels and thereby decrease the needfor the active operation and maintenance of landfillfacilities.Once the location, size and shape of the landfill are

established, the choice of liner system is usually the nextmost important decision in terms of controlling the


emissions from the landfilled residues. The choice ofliner system is typically the outcome of the overall dis-posal strategy of the residues, which can be categorizedas follows: total containment (i.e. ‘‘dry’’ storage), lea-chate collection by use of bottom liner systems, con-trolled release of contaminants by controlling leachateproduction, and finally uncontrolled release. Concern-ing incineration residues, it is not suitable to follow astrategy of total encapsulation of the waste in the land-fill by restricting any contact with water, as this willkeep the pollution potential of the waste unchanged andconsequently involve active care and maintenance of thelandfill indefinitely. In most cases the appropriate solu-tion is to have some type of bottom liners—eithernaturally occurring or artificial—and leachate collectionduring the active period.At the same time some type of top cover will usually

be applied to reduce infiltration to the waste. When theleachate quality is considered environmentally accep-table, i.e. the waste has reached the so-called ‘‘final sto-rage quality’’ (Hjelmar, 1996a,b), the collection ofleachate can be terminated and leaching will be allowed;however, leachate production will still be controlled bythe top cover. At some point the integrity of the topcover may no longer be guaranteed and the landfillwill then be subject to unrestricted leaching. At thispoint, the leachate quality allows any active environ-mental protection system to be abandoned safely. Inthis sense, the appropriate use of top and bottom linersystems will correspond to the various phases of landfilloperation.

4.1.2.2. Landfill operation. A number of issues involvingthe daily operation of the landfill can have an effect onthe emissions from the landfill and should be consideredaccording to the overall disposal strategy applied to theresidues. Notably, controlling the incoming waste typesand redirecting these to appropriate landfills or landfillsections is important because mixing of different typesof wastes can have undesirable effects on leaching.Waste types should be landfilled separately if they exhi-bit different environmental behavior, as the disposalstrategies for such waste are in general mutually incom-patible. Thus, for example, co-disposal of MSWI resi-dues and organic wastes (such as MSW) generallywould not represent a sustainable option, as leaching ofcontaminants 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 APCresidues, generally is not advisable, as a high content ofsoluble contaminants in certain types of residuesrequires specific landfill operation strategies that are notnecessary for other residue types.The daily procedures of compaction and covering

incineration residues may have an effect on leachateproduction 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 certaintype 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 couldbe terminated within a reasonable timeframe. As such,collection and treatment is a measure that can be usedto control leaching while waiting for a certain leachatequality to be attained. In some cases it can be beneficialto accelerate the leaching by irrigating the residues withnew water or recirculated leachate. Following this pro-cedure, it is possible to reach an acceptable level ofcontaminant release from the landfill within shortertimeframes and thereby decrease the time required foractive care of the landfill.For the evaluation of proper leachate collection and

treatment strategies, consideration should be given tothe final fate of contaminants. In this regard, the con-servative nature of most of the contaminants released byMSWI residues must be kept in mind. Thus, if MSWIlandfill leachate treatment is accomplished at a waste-water treatment plant through biological processes, thecontrol of pollutants will mostly rely on dilution. On theother hand, if specific processes are applied for heavymetal removal, additional residues will be generatedwhich require proper final disposal. In both cases,recirculation of contaminants to the environment willoccur. For these reasons, it appears more advisable tofollow landfill operation strategies involving the con-trolled return of contaminants to the ecological cycle atenvironmentally 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 themigration of landfill gas. This, however, is rarely thecase at landfills for incineration residues. As discussed inprevious chapters, production of hydrogen from incin-eration residues can occur and potentially can createoperational problems at the landfill. It is therefore sug-gested that the potential production of hydrogen bekept in mind and gas production and composition becarefully monitored.

4.1.3. Measures acting during the passive phase oflandfillingBy definition, the passive phase of landfilling requires

no active maintenance or operation of facilities at thelandfill site in order to achieve acceptable emissions.Any active environmental protection system at thelandfill, such as leachate collection systems or artificialliner systems, cannot be trusted to work effectivelywithout proper maintenance. It is therefore essential—and necessary—that emissions from the landfill are


reduced to an acceptable level at the time the passivephase is entered.Some measures introduced in the active phase, can

have an effect within reasonable limits on the naturalprocesses that will occur in the passive phase. Notably,a number of naturally occurring weathering processes(as discussed in the previous sections), especially in thecase of APC residues, will slowly change the miner-alogical compositions of the residues and therebychange the conditions for leaching of chemical species(for example trace metals) in the long term. Also theability of in-situ geological formations to reduce migra-tion of contaminants released from the landfill sitepotentially can have an effect in the long term.

5. Conclusions and recommendations

Over the last decades important progress has beenmade in integrated waste management systems. Treat-ment of waste, landfilling and utilization of residuesfrom waste treatment, and mitigating measures areintegral elements of these systems. In particular thesemeasures focus on minimizing the threats to our healthand environment. Most environmental and healththreats as well as economic risks associated with resi-dues from thermally-treated wastes are caused by metalsor salts due to liquid discharge, by particulate airbornetransport and by production of potentially dangerousgases. Now these threats can be mitigated by variousmeasures including options prior, during and/or afterlandfilling.In some cases there still remain some residual emis-

sions at environmentally unacceptable levels, makingactive aftercare systems indispensable. To keep thisaftercare period as short as possible, it is beneficial toaccelerate the processes that lead to the production ofemissions. The overall strategy should be to minimizeand facilitate the controlled release of contaminants tothe environment without any additional aftercare. Thisrequires the profound understanding of the waste beha-vior, the environmental scenario, the various factorsand the particular processes as well as the implemen-tation of this knowledge for the treatment processes,landfill design and landfill operation.Many factors and processes as well as their impact on

residues from waste treatment are well known. But thereis still a lack of understanding, particularly in control-ling all the interactions especially under natural condi-tions. Therefore future (both laboratory and field)research needs to consider final disposal as part of anintegrated waste management system.Even though some aspects related to management,

treatment and disposal of MSWI residues still necessi-tate deeper understanding, nevertheless the current levelof knowledge allows some conclusions on the disposal

criteria to be drawn. In general, it can be stated that theproper disposal strategies to be adopted are stronglydependent on the characteristics of the MSWI residuesto be landfilled. For MSWI residues containing onlylimited amounts of trace contaminants (such as bottomash) the disposal strategy may be based on a single-stageoperation with the leachate release to the environmentbeing controlled at an environmentally acceptable rate.In this case pollution prevention may be well attainedwith passive environmental protection systems only.For residues containing significant amounts of readilysoluble species in addition to trace elements (such assome kinds of APC residues) it may be advisable toadopt disposal criteria based on a two-stage operation.The first stage will rely on active environmental protec-tion systems with leachate collection and treatment; inthis case, the leaching rate may also be enhanced inorder to reduce the duration of the active care period.After depletion of the readily leachable constituents, themigration of contaminants in the environment can becontrolled by passive environmental protection systemsonly. Treating MSWI residues prior to landfilling canhave the result to reduce the length of the first operatingstage or to remove it at all.Nevertheless, in addition to the impacts from leachate

emissions, the potential for gas production from MSWIresidues monofills should be carefully considered, sothat gas collection systems may be required during theearly years of landfill operation.

Acknowledgements

The authors wish to thank the other pHOENIXmembers for stimulating fruitful discussion on the sub-jects dealt with in the present paper during the groupmeetings.

References

Astrup, T., Hansen, J.B., Hjelmar, O., Christensen, T., Long-term

development in APC ash landfills with respect to pH, Christensen,

T.H., Cossu, R., Stegmann, R. (Eds.). Proceedings Sardinia 2001,

Eighth International Waste Management and Landfill Symposium,

vol. I, 2001, pp. 453–459.

B.A.L, 9th Lysimeter conference of the Gumpenstein society of alpine

agriculture, Gebietsbilanzen bei unterschiedlicher Landnutzung.

Gumpenstein (A), 24–25 April 2001, Herausgeber: Bundesanstalt

fur Alpenlandische Landwirtschaft, Styria, 2001.

Baumgartner, A., Liebscher, H.J., 1996. Comprehensive Hydro-

geology. Allgemeine Hydrologie. Band 2: Quantitative Hydrologie.

Verlag Gebruder Borntraeger, Berlin/Stuttgart.

Belevi, H., Stampfli, D.M., Baccini, P., 1992. Chemical behaviour of

municipal solid waste incinerator bottom ash in monofills. Waste

Manage. Res. 10, 153–167.

Berger, K., Dunger, V., Vergleichende Simulationsrechnungen mittels

der Deponie- und Wasserhaushaltsmodelle HELP und BOWA-

HALD. DGZF-Seminar ‘‘Simulation zu Halden-/Deponie-

wasserhaushalt’’, 7–8 April 2001, Dresden, 2001.


Bodenan, F., Azaroual, M., Piantone, P., 2000. Forecasting the long-

term behaviour of municipal solid waste incineration bottom ash:

rapid combined tests. In: Woolley, G.R., Goumans, J.J.J.M.,

Wainwright, P. (Eds.), Waste Materials in Construction, Studies in

Environmental Sciences 71. Elsevier Sci, Amsterdam, pp. 475–482.

Brechtel, H.M., Beeinflussung des Wasserhaushaltes von Mulldepo-

nien. Mullhandbuch, Hrsg: HOSEL/SCHENKEL/SCHNURER,

Nr.4623, Erganzungslieferung 5/84, 1984.

CEC, Council Directive 1999/31/EC of 26 April 1999 on the landfill of

waste. Official Journal of the European Communities, 16/7/1999,

L182/1-19.

Conner, J.R., 1990. Chemical Fixation and Solidification of Hazar-

dous Wastes. Van Nostrand-Reinhold, New York.

Derie, R., 1996. A new way to stabilise fly ash from municipal incin-

erators. Waste Manage 16 (8), 711–716.

Ehrig, H.J., 1990. Sickerwasserentstehung in Altlasten und ihre Pro-

blematik. Handbuch der Altlastensanierung. Hrsg: FRANZIUS/

STEGMANN/WOLF/BRANDT. R.v. Decker’s Verlag, Heidel-

berg.

Eighmy, T.T., Crannel, B.S., Cartledge, F.K., Emery, E.F., Oblas, D.,

Krzanowski, J.E., Shaw, E.L., Francis, C.A., 1997. Heavy metal

stabilization in municipal solid waste combustion dry scrubber residues

using soluble phosphate. Environ. Sci. Technol. 31 (11), 3330–3338.

ENV 12920, Characterization of waste––methodology for the deter-

mination of the leaching behaviour of waste under specified condi-

tions. European Prestandard, CEN 1997.

Forstner, U., Hirschmann, G. Verbundvorhaben Deponiekorper:

Langfristiges Deponieverhalten von Mullverbrennungsschlacken,

Umweltbundesamt, Projekttrager Abfallwirtschaft und Altlastensa-

nierung des BMBF, 202, 1997.

Gilliam, T.M., Wiles, C.C. (Eds.), 1996. Stabilization and Solidifica-

tion of Hazardous, Radioactive and Mixed Wastes, Vol. III, ASTM

STP 1240. ASTM, Philadelphia, PA.

Heyer, K.U, Stegmann, R., Landfilling of bottom ashes from incin-

eration municipal solid waste: a general survey. Christensen, T.H.,

Cossu, R., Stegmann, R. (Eds.). Proc. Sardinia 97, Sixth interna-

tional Landfill Symposium, Vol. V, 1997, pp. 521–533.

Hjelmar, O., 1996. Disposal strategies for municipal solid waste

incineration residues. J. Hazard. Mater. 47, 345–368.

Hjelmar, O., Birch, H., Hansen, J.B. Development of a process for

treatment of APC residues from MSW incinerators prior to land-

filling. Proc. Sardinia 99, Seventh International Landfill Sympo-

sium, Vol. I, 543–548, 1999a.

Hjelmar, O., Holm, P.E., Christensen, T.H., Lehmann, N.K.J., An

approach to development of rational limit values for regulatory

testing of waste prior to disposal or utilisation. Christensen, T.H.,

Cossu, R., Stegmann, R. (Eds.) Proc. Sardinia 99, Seventh Interna-

tional Waste Management and Landfill Symposium, Vol. I , 1999b,

pp. 581–588.

Hjelmar, O., Van der Sloot, H.A., Guyonett, D., Rietra, R.P.J.J.,

Brun, A., Hall, D. Development of acceptance criteria for landfilling

of waste: an approach based on impact modelling and scenario cal-

culations. Christensen, T.H., Cossu, R., Stegmann, R. (Eds.) Proc.

Sardinia 2001, Eighth International Waste Management and Land-

fill Symposium, Vol. III, 2001, pp. 711–721.

Hong, K.-J., Tokunaga, S., Ishigami, Y., Kajiuchi, T., 2000. Extrac-

tion of heavy metals from MSW incinerator fly ashes using sapo-

nins. Chemosphere 41, 345–352.

Hong, K.-J., Tokunaga, S., Kajiuchi, T., 2000. Extraction of heavy

metals from MSW incinerator fly ashes by chelating agents. J.

Hazard. Mater., B75 57–73.

Huber H. Development of pH and alkalinity in bottom ash—simu-

lation of the long-term behaviour. Waste Reports no. 8, November

1998. ABF-BOKU, Vienna, 1998.

IAWG (International Ash Working Group: A.J. Chandler, T.T.

Eighmy, J. Hartlen, O. Hjelmar, D. Kosson, S.E. Sawell, H.A. van

der Sloot, J. Vehlow), In: Municipal Solid Waste Incinerator Resi-

dues, Studies in Environmental Sciences 67, Elsevier Sci., Amster-

dam, 1997.

Johnson, C.A., Brandenberger, S., Baccini, P., 1995. Acid neutralizing

capacity of municipal waste incinerator bottom ash. Environ. Sci.

Technol. 29, 142–147.

Johnson, A., Kaeppeli, M., Brandenberger, S., Ulrich, A., Baumann,

W., 1999. Hydrogeological and geochemical factors affecting lea-

chate composition in municipal solid waste incinerator bottom ash.

J. Contam. Hydrol. 40, 239. –259, .

Katsuura, H., Inoue, T., Hiraoka, M., Sakai, S., 1996. Full-scale plant

study on fly ash treatment by the acid extraction process. Waste

Manage. 16 (5-6), 491–499.

Klein, R., Baumann, T., Kahapka, E., Niessner, R., 2001. Tempera-

ture development in a modern municipal solid waste incineration

(MSWI) landfill with regard to a sustainable waste management. J.

Hazard. Mater. B83, 265–280.

Kosson, D.S., Van der Sloot, H.A., 1997. Integration of testing pro-

tocols for evaluation of contaminant release from monolithic and

granular wastes. In: Goumans, J.J.J.M., Senden, G.J., van der

Sloot, H.A. (Eds.), Waste Materials in Construction. Putting The-

ory into Practice, Studies in Environmental Sciences 71. Elsevier Sci,

Amsterdam, pp. 201–216.

Krumpelbeck, I., 2000. Untersuchungen zum langfristigen Verhalten

von Siedlungsabfalldeponien, Ph.D. Thesis. University Wuppertal,

Germany.

Laethem, B., Van Herck, P., Geuzens, P., Vandecasteele, C., 1994.

Integrated treatment of MSWI-residues. Treatment of fly ash in

view of metal recovery. In: Goumans, J.J.J.M., van der Sloot, H.A.,

Aalbers, Th.G. (Eds.), Environmental Aspects of Construction with

Waste Materials, Studies in Environmental Sciences 60. Elsevier Sci,


Lahl, U., Verwertung von MV - Schlacken - durch Optimierung kon-

ventioneller Aufbereitung, Mull und Abfall, 24, 1992, 619.

Lamers, F.J.M., Born, J.G.P., 1994. Upgrading techniques for the qual-

ity improvement of municipal waste incineration residues. In: Gou-

mans, J.J.J.M., van der Sloot, H.A., Aalbers, Th.G. (Eds.),

Environmental Aspects of Construction withWasteMaterials, Studies

in Environmental Sciences 60. Elsevier Sci, Amsterdam, pp. 645–654.

Lechner, P., Huber, H., Jaros, M., Assessment of the long-term beha-

viour of MSW incinerator slag. Proc. Sardinia 97, Sixth Interna-

tional Landfill Symposium, Vol. V, 1997, pp. 553–563.

Lerner, D.N. Groundwater recharge. In: Saether, O.M., de Caritat, P.

Geochemical Processes, Weathering and Groundwater Recharge in

Catchments. Balkema, Rotterdam, 1997.

Lundtorp, K., Jensen, D.L., Sørensen, M.A., Mogensen, E.P.B.,

Christensen, T.H., Stabilization of APC-Residues with FeSO4. Proc.

Sardinia 99, Seventh International Landfill Symposium, Vol. I,

1999, pp. 549–556.

Magel, G., Heuss-Assbichler, S., Fehr, K.T., Abiotic H2 generating

reactions in a MSWI monofill. Christensen, T.H., Cossu, R., Steg-

mann, R. (Eds.) Proc. Sardinia 2001, Eighth International Waste

Management and Landfill Symposium, Vol. I, 2001, pp. 489–498.

Mangialardi, T., Paolini, A.E., Polettini, A., Sirini, P., 1999. Optimi-

zation of the solidification/stabilization process of MSW fly ash in

cementitious matrices. J. Hazard. Mater. 70 (1-2), 53–70.

Marques, M., Hogland, W., Leachate modelling in full-scale cells

containing predominately MSW incineration residues. Proc. Sardi-

nia ’99, Seventh International Landfill Symposium, Vol. I, 1999, pp.

613–620.

Matsuto, T., Tamaru, T., Takata, M., Matsushita, M., Matsufuji, Y.,

Dispersion of incineration residue from MSW landfill. Christensen,

T.H., Cossu, R., Stegmann, R. (Eds.). Proc. Sardinia 2001, Eighth

International Waste Management and Landfill Symposium, Vol. I,

2001, pp. 469–478.

Meima, J.A., Comans, R.N.J., 1997. Geochemical modeling of

weathering reactions in municipal solid waste incinerator bottom

ash. Environ. Sci. Technol. 31, 1269–1276.


Meima, J.A., Comans, R.N.J., 1998. Application of surface com-

plexation/precipitation modeling to contaminant leaching from

weathered municipal solid waste incinerator bottom ash. Environ.

Sci. Technol. 32, 688–693.

Meima, J.A., Comans, R.N.J., 1999. The leaching of trace elements

from municipal solid waste incinerator bottom ash at different

stages of weathering. Appl. Geochem. 14, 159–171.

Mizutani, S., Sakai, S., Takatsuki, H., 2000. Investigation of hydrogen

generation from municipal solid waste incineration fly ash. J.

Material Cycles and Waste Manage. 2, 16–23.

Musselmann, C.N., Straub, W.A., Bidwell, J.N., Carpenter, J., Gas

generation at municipal waste combustor ash monofill-Franklin,

New Hampshire. In Sustainable Construction: Use of Incineration

Ash, 2000.

Nolting, B., Gossele, P., Wefer, H., Bender, M. Use of water balances

for landfill site monitoring. Proc. Sardinia 95, Fifth International

Landfill Symposium, Vol. I, 1995, pp. 263–274.

Nzihou, A., Sharrock, P., 2002. Calcium phosphate stabilisation of fly

ash with chloride extraction. Waste Manage 22, 235–239.

OnidaM., Innovation in prevention of quantitative and qualitative waste

production. In: Centro di Ingegneria per la Protezione dell Ambiente

(C.I.P.A.): Innovation in Waste Management, Proc. IV European

Waste Forum, Milano, 30 November–1 December 2000, pp. 91–105.

Polettini, A., Pomi, R., Sirini, P., Testa, F., 2001. Properties of Port-

land cement-stabilised MSWI fly ashes. J. Hazard. Mater. 1 (88),

123–138.

Sabbas T., Mostbauer P., Rautner T., Lechner P., A prediction model

for the assessment of the middle and long-term emission flux of

inorganic wastes, Proc. ISWA World Congress 2001, Stavanger,

Norway, 2001a.

Sabbas, T., Lechner, P., Developing an assessment method for the

middle to long-term pollutant flux of inorganic residues, Proc.

SWANA Conference, San Diego, CA, 2001b.

Schneider, J., Vehlow, J., Vogg, H., 1994. Improving the MSWI bot-

tom ash quality by simple in-plant measures. In: Goumans, J.J.J.M.,

van der Sloot, H.A., Aalbers, Th.G. (Eds.), Environmental Aspects

of Construction with Waste Materials, Studies in Environmental

Sciences 60. Elsevier Sci, Amsterdam, pp. 605–620.

Speiser, C., Bauman, T., Niessner, R., 2000. Morphology and chemi-

cal characterization of calcium-hydrate phases formed in alteration

processes of deposited municipal solid waste incinerator bottom

ashes, Environ. Sci. Technol. 34, 5030–5037.

Speiser, C. Exothermer Stoffumsatz in MVA-Schlackedeponien:

Mineralogische und geochemische Charakterisierung von Mullver-

brennungsschlacken, Stoff- und Warmebilanz, Dissertation Tech-

nische Universitat Munchen, 2001.

Speiser, C., Heuss-Assbichler, S., Spiegel, W., Long-term stability of

boiler ashes and gas cleaning residues of municipal solid waste.

Christensen, T. H., Cossu, R., Stegmann, R. (Eds.) Proc. Sardinia

2001, Eighth International Waste Management and Landfill Sym-

posium, Vol. 1, 2001, pp. 607–616.

Speiser, C., Heuss-Assbichler, S., Spiegel, W. The significance of

ettringite for the long-term stability of boiler ashes and gas cleaning

residues of municipal solid waste incinerator, Proc. R’02 6th World

Congress an Integrated Resources Management, 12–15 February,

2002, Geneva, Switzerland, No. 284, 2002.

Stegmann, R., Ehrig, H.J., Leachate production and quality – results

of landfill processes and operation. Proc. Sardinia 89, Second

International Landfill Symposium, 1989.

Stubenvoll, J. The MR-process treatment of MSW residues with

reclamation of heavy metals. Proc. Recycling International, Berlin,

1989, pp. 1476–1481.

TAC Modeling Group. Development of acceptance criteria for wastes

– EU landfill directive–Annex II, Draft-March 12th, 2002.

Takatsuki, H., 1994. Conclusions drawn from an explosion at a dust

bunker (in Japanese). Haikibutsu 5, 441–448 .

Van der Sloot, H.A., Cnubben, P.A.J.P., Scharff, H., Predominantly

inorganic equilibrium disposal: part of the total concept sustainable

recycling and storage of solid waste. Proc. Sardinia 99, Seventh

Waste Management and Landfill Symposium. S. Margherita di Pula

(CA), Italy, Vol. I, 1999, pp. 103-110.

Van der Sloot, H.A., Kosson, D., Hjelmar, O., 2001. Characteristics,

treatment and utilization of residues from municipal waste incin-

eration. Waste Manage 21, 753–765.

Vehlow, J., Braun, H., Horch, K., Merz, A., Schneider, J., Stieglitz, L.,

Vogg, H., 1990. Semi-technical Demonstration of the 3-R Process,

Waste Manage. Res. 8, 461–472.

Wiles, C.C., 1996. Municipal solid waste combustion ash: state-of-the-

knowledge. J. Hazard. Mater. 47, 325–344.

Yasuda, K., Cause of explosion accident at a dust chute - counter-

measure for prevention (in Japanese). Proc. 18th Toshi Seisou

Kenkyu Happyokai Kouen Ronbunshu, 148, 1997.

Zevenbergen, C., Comans, R.N.J., 1994. Geochemical factors con-

trolling the mobilization of major elements during weathering of

MSWI bottom ash. In: Goumans, J.J.J.M., van der Sloot, H.A.,

Aalbers, Th.G. (Eds.), Environmental Aspects of Construction with

Waste Materials, Studies in Environmental Sciences 60. Elsevier Sci,


Zevenbergen, C., Van Reeuwijk, L.P., Bradley, J.P., Comans,

R.N.J., 1996. Mechanism and conditions of clay formation during

natural weathering of MSWI bottom ash. Clays Clay Miner 44,

546–552.


managementofmunicipalsolidwasteincinerationresidues · 2014-04-17 ·...

Documents