Waste Management

Waste Management 23 (2003) 61–88 www.elsevier.com/locate/wasman

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)
BOKU University Vienna, Department of Waste Management-Nussdorfer La¨nde 29-31, A-1190, Vienna, Austria University of Rome ‘‘La Sapienza’’, Department of Hydraulics, Transportation and Roads - Via Eudossiana 18, I-00184 Rome, Italy c Technical University of Denmark, Environment & Resources, DTU - Building 115, DK-2800 Lyngby, Denmark d ´ DHI Water & Environment - Agern Alle 11, DK-2979 Ho ¨rsholm, Denmark e University of Cagliari, Department of Geoengineering and Environmental Technologies - Piazza D’Armi 1, I-09123 Cagliari, Italy f Ludwig-Maximilians-Universita Institut fu¨r Mineralogie, Petrologie und Geochemie - Theresienstrasse 41, D-80333 Munich, Germany ¨t, g CheMin GmbH-Am Mittleren Moos 48, D-86167 Augsburg, Germany h Technical University of Munich, Department of Hydrochemistry - Marchioninistrasse 17 D-81377 Munich, Germany b a

Accepted 29 October 2001

Abstract The management of residues from thermal waste treatment is an integral part of waste management systems. The primary goal of managing incineration residues is to prevent any impact on our health or environment caused by unacceptable particulate, gaseous and/or solute emissions. This paper provides insight into the most important measures for putting this requirement into practice. It also offers an overview of the factors and processes affecting these mitigating measures as well as the short- and long-term behavior of residues from thermal waste treatment under different scenarios. General conditions affecting the emission rate of salts and metals are shown as well as factors relevant to mitigating measures or sources of gaseous emissions. # 2002 Elsevier Science Ltd. All rights reserved.

Preface A working group named ‘‘pHOENIX’’ on the ‘‘Management of Municipal Solid Waste Incineration (MSWI) Residues’’ was established as a result of a workshop held in spring 2002 in Vienna, which dealt with the practical problems, recent research findings and solutions related to this topic. As we agreed, there are numerous highly specific scientific articles as well as some comprehensive studies and books with either indepth research or with a description of integrated waste management in general terms or with specific MSWI residues. However, what was missing was a short introductory overview of the management of residues from thermal MSW treatment for operators, non-specialized

* Corresponding author. Tel./fax: [phone]. E-mail address: [e-mail] (A. Polettini).

scientists and legislators. With this article, we hope to fill the gap. The pHOENIX working group is composed by the following members: Peter Lechner (BOKU University Vienna, Department of Waste Management, A), Thomas Astrup (DTU Technical University of Denmark, Environment & Resources, DK), Giovanna Cappai (University of Cagliari, Department of Geoengineering and Environmental Technologies, I), Holger Ecke (Lulea University of Technology, SE), Soraya HeussAssbichler (Ludwig-Maximilians-Universitat, Institut ¨ fur Mineralogie, Petrologie und Geochemie, D), Ole ¨ Hjelmar (DHI Water & Environment, DK), Anders Kihl (Ragn-Sells Avfallsbehandling AB, SE), Ralf Klein (Technical University of Munich, Department of Hydrochemistry, D), Gabriele Magel (Ludwig-Maximilians-Universitat, Institut fur Mineralogie, Petrologie ¨ ¨ und Geochemie, D), Peter Mostbauer (BOKU University 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

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Franz Ottner (BOKU University Vienna, Department of Waste Management, A), Alessandra Polettini (University of Rome ‘‘La Sapienza’’, Department of Hydraulics, Transportation and Roads, I), Raffaella Pomi (University of Rome ‘‘La Sapienza’’, Department of Hydraulics, Transportation and Roads, I), Tamara Rautner (BOKU University Vienna, Department of Waste Management, A), Henrich Riegler (BOKU University Vienna, Department of Waste Management, A), Thomas Sabbas (BOKU University Vienna, Department of Waste Management, A), Stefan Salhofer (BOKU University Vienna, Department of Waste Management, A).

1. Introduction The objective of integrated waste management is to deal with society’s waste in an environmentally and economically sustainable way. Under the framework of integrated waste management, thermal treatment represents a valid option for reducing the amount of waste to be landfilled, at the same time allowing for waste hygienization. The relative importance of incineration as opposed to other waste treatment and disposal options, including mechanical/biological treatment and sanitary landfilling, varies considerably from country to country, depending on specific waste management strategies as well as space availability for final land disposal. The increasingly more stringent limits imposed in recent years on atmospheric emissions from waste incineration have produced a considerable shift from the gaseous emissions to the solid residues of the process. Thus, solid residues from thermal waste treatment warrant significant environmental concern. In our article, we specially focus on assessing the environmental impacts resulting from residues from thermal waste treatment, the treatment methods available to mitigate such impacts before and after either utilization or final land disposal as well as the processes and variables affecting the physical and chemical changes occurring for such residues at the utilization or disposal site. 1.1. Integrated waste management As depicted in Fig. 1, waste management systems include all processes from waste generation to landfilling, i.e.:  Waste generation: all processes which produce waste during the production and distribution of products (industry and commerce) or the consumption of products (households);  Waste collection, including source separation into different material streams;

Fig. 1. Integrated waste management system.

 Processing, including such steps as waste sorting, dismantling of products (e.g. end-of-life electrical and electronic equipment), and production of Refuse Derived Fuel (RDF). All these steps serve either to prepare waste for reuse or to suitably modify waste characteristics with a view to final land disposal;  Recycling: production of secondary materials from waste, e.g. paper from waste paper, steel from ferrous metal scraps etc.;  Waste treatment, including several technologies such as thermal treatment, chemical treatment of hazardous wastes, mechanical/biological treatment;  Waste utilization, covering all the utilization options of waste after processing, e.g. use of treated bottom ash for road construction, compost for agricultural applications or thermal utilization of RDF; and  Landfilling. 1.2. Treatment methods Waste treatment methods strongly depend on the type of waste. As far as municipal solid waste is concerned, the different treatment options are aimed at recovering materials and/or energy from the waste as well as at reducing the overall amount and the impacts of waste to be landfilled. In this framework, both mechanical/biological

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pretreatment and waste-to-energy incineration are suitable treatment options which should be combined in order to meet the above mentioned targets. In particular, with thermal waste treatment the following issues can be viewed as the main objectives:       to to to to to to reduce the total organic matter content, destroy organic contaminants, concentrate the inorganic contaminants, reduce the mass and volume of the waste, recover the energy content of the waste and preserve raw materials and resources.

1.3.1. MSWI residues As a result of the incineration process, different solid and liquid residual materials as well as gaseous effluents are generated. Approximately one-fourth of the waste mass on a wet basis remains as solids. The volume of residues corresponds to one-tenth of the initial waste volume. Typical residues of MSWI by grate combustion are:  Bottom ash, which consists primarily of coarse non-combustible materials and unburned organic matter collected at the outlet of the combustion chamber in a quenching/cooling tank.  Grate siftings, including relatively fine materials passing through the grate and collected at the bottom of the combustion chamber. Grate siftings are usually combined with bottom ash, so that in most cases it is not possible to separate the two waste streams. Together bottom ash and grate siftings typically represent 20–30% by mass of the original waste on a wet basis.  Boiler and economizer ash, which represent the coarse fraction of the particulate carried over by the flue gases from the combustion chamber and collected at the heat recovery section. This stream may constitute up to 10% by mass of the original waste on a wet basis.  Fly ash, the fine particulate matter still in the flue gases downstream of the heat recovery units, is removed before any further treatment of the gaseous effluents. The amount of fly ash produced by an MSW incinerator is in the order of 1–3% of the waste input mass on a wet basis.  Air pollution control (APC) residues, including the particulate material captured after reagent injection in the acid gas treatment units prior to effluent gas discharge into the atmosphere. This residue may be in a solid, liquid or sludge form, depending on whether dry, semi-dry or wet processes are adopted for air pollution control. APC residues are usually in the range of 2% to 5% of the original waste on a wet basis. Due to the volatilization and subsequent condensation as well as concentration phenomena acting during combustion, fly ash and APC residues bear high concentrations of heavy metals, salts as well as organic micro-pollutants. Iron scrap and other metals are usually recovered from bottom ash and reused in industry. In some European countries great efforts are devoted to utilization of such residues. If utilization is not possible due to regulatory constraints or other reasons (such as a sufficient source of natural raw materials), these residues have to be disposed in an environmentally acceptable and economically sustainable way.

Hazardous wastes and sewage sludge are also often treated using various thermal methods. Aside from combustion, other thermal processes exist, including pyrolysis, gasification, sintering, vitrification and melting. The most common thermal treatment process for MSW is incineration by mass-burn technology. Fluidized bed incineration and refuse derived fuel systems are less common in municipal solid waste treatment. Fluidized bed systems and multi-hearth furnaces are also widely used for sewage sludge incineration, while major furnace types for hazardous wastes incineration are grateless systems such as a rotary kiln furnace, fluidized bed systems, combustion chamber and multi-hearth furnace. Non-thermal waste treatment methods consist of biological, chemical as well as physical treatment. 1.3. Input to MSWI The quality and quantity of the MSWI input and output are influenced by several factors:  Waste generators are households and, in addition, industrial or commercial sites.  Waste generation both in households and industry is (theoretically) influenced by waste prevention. In reality we can observe increasing waste quantities. Onida (2000) reports that the total annual production of industrial waste in five major sectors (agriculture, mining, manufacturing, municipal and energy production) increased by 9.5% from 1990 to 1995 in the EU.  Separate collection exerts a strong influence on the quantities and quality of waste for incineration. For example, the separate collection of small electrical appliances could reduce the Cu content in MSWI bottom ash by up to 80%. Through source separation of recyclables and biogenic waste, the quantity of waste for treatment is significantly reduced.  Residues from waste processing technologies (e.g. sorting of plastics after separate collection) and other materials can also be part of the input to MSWI.

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1.4. Goal of landfilling One objective of landfilling of waste, including MSWI residues, is to remove from general circulation materials and products that are no longer useful in any respect. It is preferable to do this in a manner that ultimately returns the basic constituents of the waste to the ecological cycle, possibly after they have undergone chemical and/or physical reactions and transformations. A second and equally important objective of waste disposal is to ensure that the waste does not cause any unacceptable short- or long-term impact on the environment or on human health. Disposal methods must ensure that this is accomplished in a sustainable manner, i.e. without excessive and/or prolonged maintenance or operation requirements and without a prolonged need for aftercare. The fulfillment of these objectives for MSWI landfills will require a profound understanding and exploitation of the short- and long-term behavior of the landfilled MSWI residues. Based on this understanding, the design and operation and to some extent also the location of the landfill must be adapted to the inherent properties of the largely inorganic MSWI residues to ensure that long-term emissions of contaminants become or remain environmentally acceptable. Landfills should be designed to minimize the required lifetime of active environmental protection systems, i.e. systems requiring maintenance and/or operation. This means, for instance, that a disposal strategy based on encapsulation of the waste is not desirable since it merely postpones the impact and preserves the contamination potential. In principle, this is true both for inorganic contaminants (salts and metals) due to their intrinsically conservative nature and for toxic organic micro-pollutants, which are commonly regarded as persistent species. Yet, the major environmental concerns in relation to the short- and long-term impact of landfilling of MSWI residues are connected with the risk of leaching and subsequent release of potentially harmful substances, particularly inorganic salts and metals/trace elements, into the environment. Gas production and release may also be of some importance, even for MSWI residues. Leaching of toxic organic compounds (especially PCDDs and PCDFs) is generally believed to be of minor relevance due to their hydrophobic nature and their low concentrations in residues from properly operated waste combustion plants. Regarding the time scales relevant to landfilling, different definitions can be used. The timeframes of interest for landfilling can be classified on the basis of a defined time scale, a connected activity or a dominant process. In the following chapters we will refer to the definition based on the landfill activity. The basic idea is that each human generation should take care of its own

wastes, without leaving future generations environmental issues still to be resolved. This approach is also used in the EU Landfill Directive (CEC, 1999), which is described in Section 1.4.2. Thus, in this context ‘‘short term’’ relates to the timeframe within which landfill operation and active aftercare (operations that require maintenance, inspection and input of energy, e.g. leachate and gas collection as well as leachate treatment) are required to meet adequate environmental protection levels. On the other hand, ‘‘long term’’ represents the timeframe within which the environmental safety of the landfill no longer relies on active protection systems, but is based on the controlled release of contaminants at an environmentally acceptable rate. The long-term period starts just after the completion of active aftercare measures. The corresponding time scales are shown in Fig. 2. 1.4.1. Disposal scenarios As incineration residues are produced by high-temperature processes, they are thermodynamically unstable under ambient conditions. This renders incineration residues highly reactive, especially under wet conditions. This means that they change their mineralogical and physico-chemical characteristics as well as their leaching behavior as long as thermodynamic equilibrium conditions with the surrounding environment are attained. The specific environmental conditions influence and change the leaching behavior and contaminant release from such materials during utilization or final land disposal. To assess the discharge behavior of a specific waste, it is necessary to take the specific conditions (scenarios) into account. To arrive at a conclusion, the following methodology should be applied (ENV 12920): formulate the task and the sought-after solution, specify the scenario, evaluate the waste characteristics, determine the influence of the scenario conditions on the variation of waste characteristics over time, as well as on their environmental behavior  model the environmental behavior of the waste and  validate the model by calibration with the results from laboratory tests and field experiments and by comparing it to natural analogues.     Such a methodology will also help identify the most appropriate mitigating measures to be undertaken before, during or after utilization or final land disposal. 1.4.2. The EU Landfill Directive Future landfilling of waste in Europe will be governed to a large extent by the EU Landfill Directive (CEC, 1999), which was officially adopted on 16 July 1999. The

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

criteria for acceptance of waste at the different classes of landfills are laid down in a Council Decision (expected to be finalized in December 2002), which addresses the requirements of Annex II to the Directive. The EU Landfill Directive (LFD) distinguishes technically between three main classes of landfills (landfills for inert waste, landfills for non-hazardous waste and landfills for hazardous waste), but only in terms of the contamination potential of the waste and the environmental protection measures required at each class of landfill. The LFD does not include any landfill strategy or guideline on the design and operation of landfills aiming at the minimization of the period during which active aftercare will be necessary. The LFD, however, to a certain extent does allow for the implementation of national strategies and guidelines within the individual EU member states. At a national level it will be possible to define different sub-classes of non-hazardous waste to prevent co-disposal of waste types with different properties (e.g. organic biodegradable waste and inorganic mineral waste) and different short- and long-term behavior.

Understanding the mechanisms governing the processes under concern and the influence of the main factors on the processes themselves will allow for the estimation of the potential environmental impacts arising from the utilization/disposal of MSWI residues as well as of the measures to be undertaken in order to mitigate the extent of such impacts. The main processes and factors of concern affecting the utilization or disposal of MSWI residues are strongly interrelated, so that in many cases a separate description of each process and factor, neglecting such interrelations, will not be exhaustive. For this reason, the following section will discuss the main processes and factors on the basis of the above mentioned macroscopic effects. 2.1. Leachate production Leaching can be defined as the dissolution of a soluble constituent from a solid phase into a solvent. Leaching occurs as a consequence of the chemical reactions taking place at the scale of the individual waste particles as well as of the contaminant transport processes via the fluid moving through the solid particles. As far as MSWI residues disposal is concerned (see Fig. 3), the transport medium of pollutants is mainly represented by water, so that the overall water balance will determine the actual amount of water reaching the application site. Climatic conditions and vegetation (e.g. precipitation, solar radiation, temperature, interception, evaporation, evapotranspiration, wind, etc.) as well as the type and morphology of the surface soil are among the main variables to be accounted for in the water balance. The application site itself then modifies the water infiltration pattern as a result of the physical and hydrological

2. Processes and factors At the utilization/disposal site (hereinafter referred to as the application site), MSWI residues will undergo a number of processes, which will cause a set of modifications in the waste matrix at the micro-structural level. On a macroscopic scale, the combination of the different processes will result mainly in the following effects:  leachate production  gas production and  temperature development.

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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).

characteristics of the material. Thus, the discharge pattern also depends on the pore type, pore distribution, homogeneity, permeability and field capacity of the material as well as on the presence of preferential flow paths. Findings by Johnson et al. (1999) show that for a four-year-old MSWI landfill in the presence of preferential flow paths, the discharge is characterized by long periods of low and nearly constant flows interspersed with increases in discharge quantity in response to rain events. Isotope studies and tracer methods in combination with a simple dilution model show that fractions of rainwater, which pass through the landfill with little interaction with the material due to preferential flow paths, make up 20–80% of the discharge volume during summer rain events and around 10% in winter months. Other studies (Brechtel, 1984; Stegmann and Ehrig, 1989; Ehrig, 1990; Krumpelbeck, 2000) show that the ¨ observed leachate amounts for uncovered or sparely vegetated MSW landfills in middle Europe lie between 15 and 60% of the annual precipitation. Approximately the same fraction of precipitation is observed for MSWI residues as annual leachate volume (Table 1, references herein). Based on experimental data on the influence of the climate and vegetation on the landfill water balance (Baumgartner and Liebscher, 1996; Lerner, 1997; B.A.L., 2001) and given the hypothesis that climatespecific vegetation will develop at each landfill site, a

prediction of long-term leachate production for selected Austrian sites was carried out (Table 1, 7th and 8th line). Water balance models, including e.g. HELP and BOWAHALD, can also be used to analyze the effect of different vegetation/covering scenarios on leachate generation (e.g. Berger and Dunger, 2001). Together vegetation and physical barriers (top cover, liners) reduce the amount of leachate from the landfill but cannot completely prevent leachate formation over a long time scale (refer to the following sections for more details). Once the overall water balance of the application site is calculated, the leachate quality needs to be estimated. Mobilization of constituents from inorganic wastes into the leaching medium is the result of the interaction between chemical and physical factors. Chemical factors include waste composition and mineralogy, temperature, pH, redox potential and the presence of ligands, while physical factors are represented by specific surface area, particle size, L/S ratio, porosity, hydraulic gradient and hydraulic conductivity. Some physical factors also affect the percolation pattern (advection, diffusion) and hence the modes of contact between leachate and waste, which can be caused by leachate flowing around the waste, leachate flowing through the waste or by a combination of the two. The processes and factors relevant to leaching can also vary depending on the contaminant under concern;

T. Sabbas et al. / Waste Management 23 (2003) 61–88 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 Pilot plant landfill containing MSWI residues and other inorganic waste South of Sweden (Marques and Hogland, 1999) Slag/ash landfill Fladsa, Denmark (Nolting et al., 1995) Lysimeters, bottom ash or mixtures of different incineration residues Vienna, Austria (Lechner et al., 1997) Lysimeter, MSWI bottom ash, uncovered Innsbruck, Austria (Lechner et al., 1997) Lysimeter, steel smelter slag, uncovered Southern alpine region, Austria (Lechner et al., 1997) Prediction for scots pine forest vegetation Vienna, Austria Prediction for spruce forest vegetation Alpine regions, Austria Annual precipitation (PPT, mm/a) About 600 468–635 550 1000 to 1300 950 520 800–900 Leachate (% PPT or mm/a) 53% 53–61% 22–36% (vegetated) 48–60% (uncovered) 62–67% 65% 60 mm/a ( $10%) 200–250 ( $25%)

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Remark Observed Observed Observed Observed Observed Predicted Predicted

in particular, for MSWI residues different groups of contaminants can be identified, including metal ions, amphoteric metals, oxyanionic species as well as salts, which display typical leaching patterns. The total content of such contaminants can even be considerably different for the various residues from waste incineration, as shown in Table 2. However, the extent of contaminant release from waste materials is rather a function of the so-called availability for leaching, which represents a fraction of the total content of contaminants in the waste itself. When a fluid flows through a loosely packed granular waste material, the amount of
Table 2 Ranges of total content of elements in MSWI residues (from IAWG, 1997) Concentration (mg/kg) Element Al As Ba Ca Cd Cl Cr Cu Fe Hg K Mg Mn Mo Na Ni Pb S Sb Si V Zn Bottom ash 22,000–73,000 0.1–190 400–3000 370–123,000 0.3–70 800–4200 23–3,200 190–8200 4,100–150,000 0.02–8 750–16,000 400–26,000 80–2400 2–280 2800–42,000 7–4200 100–13,700 1000–5,000 10–430 91,000–308,000 20–120 610–7800 Fly ash 49,000–90,000 37–320 330–3100 74,000–130,000 50–450 29,000–210,000 140–1100 600–3200 12,000–44,000 0.7–30 22,000–62,000 11,000–19,000 800–1900 15–150 15,000–57,000 60–260 5300–26,000 11,000–45,000 260–1100 95,000–210,000 29–150 9000–70,000

contaminants released is dictated by solubility constraints, so that the leaching process is referred to as being solubility-controlled. In the case of very soluble mineral phases, which can completely dissolve as a result of contact with the fluid, leaching is generally defined as availability-controlled. Conversely, as far as compacted granular materials or treated (e.g., solidified) wastes are concerned, leachate percolation occurs at the surface of the solid material, causing molecular diffusion to be the dominant process determining contaminant release from the waste; in this case, the leaching process is said to be diffusion-controlled

Dry/semi-dry APC residues 12,000–83,000 18–530 51–14,000 110,000–350,000 140–300 62,000–380,000 73–570 16–1700 2600–71,000 0.1–51 5900–40,000 5100–14,000 200–900 9–29 7600–29,000 19–710 2500–10,000 1400–25,000 300–1,100 36,000–120,000 8–62 7000–20,000

Wet APC residues 21,000–39,000 41–210 55–1600 87,000–200,000 150–1400 17,000–51,000 80–560 440–2400 20,000–97,000 2.2–2300 810–8600 19,000–170,000 5000–12,000 2–44 720–3400 20–310 3300–22,000 2700–6000 80–200 78,000 25–86 8100–53,000

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leaching. Due to the reduced surface area of the material in contact with the leachate and to the extremely slow release of contaminants from the waste, chemical equilibrium between the solid and the liquid phase is generally not attained in the case of diffusion-controlled release. Again, the mechanisms controlling the leaching process are therefore dependent on physical factors pertaining both to the properties of the waste material (including particle size distribution, porosity, degree of compaction and permeability) and to the related fluid flow characteristics (including percolation rate, percolation pattern and the amount of leachate contacting the waste). They are also dependent on chemical factors related to the solubility of contaminants in the waste material. Leachate composition is the result of reaction between the various mineral phases in the waste and the leaching fluid. The leachability of strongly soluble species (e.g., alkali salts) is almost pH-independent, whereas for a number of contaminants a clear pH-dependence can be observed. The influence of pH on the leaching of contaminants is strongly related to the nature of the particular contaminant under concern as well as the mineral phase(s) in which this is bound. Three main typical leaching behaviors for solubility-controlled leaching have been identified:  cation-forming species and non-amphoteric metal ions (e.g. Cd), see Fig. 4a),  amphoteric metals (including Al, Pb, Zn), see Fig. 4b), and  oxyanion-forming elements (e.g. As, Cr, Mo, V, B, Sb), see Fig. 4c). The concentration of cation-forming species and nonamphoteric metal ions displays fairly constant high values at pH < 4, and decreases strongly up to pH 8 to 9, remaining approximately constant or slightly increasing for higher pH values. Amphoteric metals exhibit increased solubility under both strongly acidic and strongly alkaline conditions, resulting in a V-shaped solubility curve. For oxyanion-forming elements usually solubility decreases in alkaline ranges (pH > 10). It should be emphasized that the shape of the actual solubility curve is the result of complex competing chemical equilibria where common ion effects can significantly alter the theoretical concentration calculated for pure aqueous solutions. For example, the equilibrium concentration of Ba in pure water (20  C, pH=7, no CO2 dissolved) saturated with barite (BaSO4) is 1.3 mg/l. In the presence of gypsum (CaSO4.2H2O), more sulfate will dissolve, and Ba concentration will decrease to 0.01 mg/l as a consequence of the law of mass action, as indicated by the arrows in Fig. 5. As shown in Figs. 4 and 5, depending on the specific leaching behavior, critical pH regions can be identified

where minimum or maximum solubility for the individual contaminants is attained. In light of this, a matter of major concern is to predict the pH conditions which are likely to occur at the application site. These depend on the characteristics of the leaching fluid as well as on the properties of the waste. Probably the most relevant waste property affecting the pH of the leachate is represented by the acid or base neutralization capacity (ANC/BNC). ANC and BNC are measures of the ability of a system to neutralize the influence of acids or bases. In the case of MSWI residues, which are most often basic in their nature, alkalinity of the material is the relevant parameter, so that ANC is the appropriate measure of neutralization capacity. ANC assesses the sensitivity of the material itself to external influences and/or internal stresses (e.g. mineralization, organic matter degradation). As a consequence, the buffering capacity of the material affects the evolution of the pH of the leachate over time, thus allowing the expected pH range for the application site to be estimated. Fig. 6 depicts the ANC of a number of bottom ash and fly ash samples from Italian municipal solid waste incinerators (Polettini et al., 2001). In the case of alkaline MSWI residue, the reduction in the buffering capacity of the material over time is related to the depletion of alkalinity, which occurs as a consequence of progressive leaching. At the time of disposal, MSWI residues will display their maximum alkalinity level. The level will decrease as the material comes into contact with the leachate and dissolved alkalinity is removed from the system by the leachate. As a consequence, the residual alkalinity at any time will depend on the initial alkalinity of the material, the dissolution of alkalinity at various pH values in the leaching scenarios and the infiltration through the application site. On the other hand, dissolved alkalinity depends on the solubility of a number of minerals (Ca(OH)2, CaCO3, etc.) and thus on the leaching system pH. The pH in turn is dependent on the system’s ability to buffer the infiltrating leachate, i.e. on the amount of residual alkalinity in the system itself (Astrup et al., 2001). Other than pH, the amount of leachate that comes in contact with a given amount of waste, usually expressed through the so-called liquid-to-solid (L/S) ratio, also affects the leaching behavior, especially in the case of solubility-controlled leaching. The L/S ratio is the result of climatic conditions, hydrology and hydrogeology of the application site, as well as the physical characteristics of the waste material. Solubility-controlled leaching is characterized by an approximately linear dependence of cumulative release on the L/S ratio. In some cases the linear trend of cumulative release of a given element as a function of L/S can be altered by the presence of other species. Delayed release is observed when a sparingly soluble phase controlling solubility is present and is depleted

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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; Â MSWI residues (mixed)) (Sabbas et al., 2001b).

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after a relatively short period. In this case, the slope of the curve of cumulative release versus L/S is lower for low L/S ratios, and it increases at higher L/S ratios where the sparingly soluble species are depleted. Conversely, enhanced initial release can be observed in the presence of complexing agents, which increase the solubility of the element under concern. In this case, a transition from a higher to a lower curve slope occurs with increasing L/S. On the other hand, for availability-controlled leaching the amount of contaminants released into the solution is at its maximum level due to the high contaminants solubility and is not dependent on solution pH. At a

given L/S ratio, the transition from solubility-controlled to availability-controlled leaching is evidenced by a constant concentration in solution with decreasing pH. Availability-controlled leaching results in rapid washout of the soluble constituents at low L/S ratios, so that the available amount often is attained at L/S values of 1 to 2; for higher L/S ratios, the cumulative release remains at this maximum. Typical examples of such leaching behavior are Na, Cl and K (see Fig. 7, which depicts the results from upflow percolation tests on weathered materials). Leaching from compacted granular residues or monolithic forms is neither solubility- nor availabilitycontrolled 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).

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Fig. 7. Leaching of Na, K and Cl from weathered MSWI bottom ash as a function of the L/S ratio.

diffusion and surface dissolution mechanisms. In this case, leaching is kinetically controlled by the rate of contaminant release via diffusion, which is measured through the effective diffusion coefficient. Release mechanisms and physical and chemical retardation factors affect the diffusion process. Among the physical retardation factors, porosity, pore structure, degree of compaction and tortuosity can significantly slow the rate of contaminant release. Pore solution pH and solubility of elements/species as a function of pH can influence the extent of sorption or co-precipitation reactions on solid surfaces, thus acting as chemical retardation factors.

Irrespective of the mechanism controlling leaching, additional factors including the presence of sorbing/ complexing agents, redox reactions and the occurrence of processes causing mineralogical changes over time (e.g. due to aging/weathering) can also affect the extent of contaminant release, as qualitatively illustrated in Fig. 8. Among the processes capable of altering the leaching behavior of the material, sorption includes different mechanisms of adsorption, ion exchange, surface complexation and electrostatic attraction of ions at the surface. During weathering of less stable phases, new minerals with high surface areas are formed. For instance, oxidation of iron in

Fig. 8. Influence of different processes on contaminant solubility as a function of pH (modified after van der Sloot et al., 1999).

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MSWI bottom ash leads to the formation of iron oxides, goethite (FeOOH) and hydrous ferric hydroxide ([Fe(OH)3]n, often termed HFO). The resulting finely grained phases are able to sorb heavy metals, including Pb, Cd, Zn, Ni, Cr(III) and Cu, as well as Mo. Similar sorptive properties are also displayed by other mineral phases, including aluminum (hydr)oxides and amorphous aluminosilicates. In the case of HFO, the general surface complexation reaction describing sorption of divalent cations can be simplified as:  Fe-OH0 þ Me2þ (  Fe-OMeþ þ Hþ ) where the symbol  indicates bonds at the surface and FeÀOH0 represents [Fe(OH)3]n. Fig. 9 demonstrates the demobilizing effect and limits of sorption. A number of sorption experiments were carried out on pre-washed (L/S=10 achieved through column percolation) and artificially weathered (by means of carbonation) MSWI bottom ash (È < 2 mm). The aged MSWI bottom ash was added with a nickel sulfate solution at Ni concentrations varying between 0.06 and 16.4 mg/g, thereby continuously aerated to achieve equilibrium with atmospheric CO2. The experimental sorption isotherms (see Fig. 9) revealed strong sorption phenomena, as long as the total amount of Ni in the system was low. Sorption proceeded as long as active sorption sites were available within the solid material. In this case, the Ni concentration in the solution was lower than that predicted based on Ni(OH)2 solubility. For high total contents of Ni, the amount of Ni exceeding the sorption capacity of the material was

such that saturation or slight over saturation of the solution with respect to Ni(OH)2 was attained. The presence of complexing agents can also significantly alter the extent of contaminant leaching from MSWI residues. Complexing agents can be either organic or inorganic in their nature; dissolved organic carbon (DOC) and chloride are the main complexing agents of concern for such materials. DOC has been extensively shown to be responsible for increasing copper release from predominantly inorganic waste forms (IAWG, 1997; Van der Sloot et al, 1999; Van der Sloot et al., 2001). Oxidation/reduction reactions also play a role in determining the release of contaminants from MSWI residues. The main oxidizing or reducing agents of relevance for MSWI residue monofills are reported in Table 3. Leaching of contaminants from waste incineration residues can be affected by the redox conditions according to two main mechanisms. One mechanism relies on the different solubility and toxicity of the contaminants under concern for MSWI residues depending on their oxidation state. These issues influence both the strength of the leachate and the related potential environmental
Table 3 Relevant reducing/oxidizing agents (inorganic landfills) Short timescale Reducing agents Oxidizing agents H2; metals (Al, Fe, Zn); Fe-II O2; H2O Medium/long timescale metals, Fe-II O2

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

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impact. For example, it is well known that in an alkaline environment Cr(III) may be rapidly oxidized by atmospheric oxygen to Cr(VI), which is much more toxic and mobile than Cr(III). However, for Cr(III) to be oxidized to Cr(VI), high values of the redox potential are required. The second mechanism through which the redox conditions affect leaching is related to the fact that the stability of the mineral phases capable of immobilizing metal ions through precipitation and/or sorption phenomena is dependent on the oxidation/reduction potential. Thus, Fe(III) and Mn(IV) (hydr)oxides can be transformed into more soluble forms of Fe(II) and Mn(II) under moderately reducing conditions. Under severely reducing conditions, S(VI) is reduced to elemental sulfur and sulfide, resulting in the precipitation of metal sulfides, which are among the less soluble metal forms. The above mentioned mechanisms can lead to either synergistic or antagonistic interactions, so that the influence of redox processes on leaching may result in either mobilization or demobilization of contaminants (see Table 4 and Fig. 8). It should be emphasized that the observations presented in Table 4 may not reflect a real landfill scenario, but merely indicate a significant influence of redox processes on leaching. Under disposal conditions, redox reactions can occur as a result of either microbiologically mediated processes due to the presence of organic material or abiotic transformations leading to the formation of reducing gases (H2). For bottom ash monofills, the presence of unburned organic material and H2 generally leads to reducing conditions; in such cases, the leaching behavior of contaminants is the result of, on the one hand, complexation by DOC and, on the other hand, precipitation of less soluble species, including for example insoluble sulfides. Weathering is a process, which naturally occurs in incineration residues as a consequence of several factors such as pH, redox potential, temperature and humidity conditions as well as the concentration of certain components (e.g. CO2) in the application site. Weathering

results in the occurrence of slow mineralogical changes over time, which may alter the leaching of trace metals from the material either in the medium or in the long term. Due to weathering and the related neoformation of minerals, key factors such as pH are subjected to changes over time. Weathering of bottom ash is a process, which deserves particular concern. Incinerator bottom ash is composed of high-temperature solids formed as a consequence of rapid quenching of the material exiting the combustion chamber, many of which are metastable under natural conditions. Typically, such solids will thereby undergo a number of chemical reactions while in the landfill leading 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, complexation with organic and inorganic ligands, surface complexation, surface (co)precipitation, sorption, and formation of solid solutions as well as oxidation/reduction (Belevi ´ et al., 1992; Bodenan et al., 2000; Meima and Comans, 1997; Meima and Comans, 1999; Zevenbergen and Comans, 1994). All of the mineralogical and chemical changes caused by such processes are also accompanied by physical changes such as pore cementation, changes in grain size and pore size distribution, which in turn alter the hydrological characteristics of the material. Hydrolysis starts immediately after bottom ash quenching and can be prolonged over the time span of temporary storage or landfilling of the material (Belevi et al., 1992; Johnson et al., 1995), as long as it is in contact with water. Hydrolysis involves the transformation of oxides of Ca, Na and K and non-noble metals like Al and Fe into the corresponding hydroxide species (e.g. CaO!Ca(OH)2, Al2O3!Al(OH)3) (Belevi et al., 1992; Speiser et al., 2000; Speiser, 2001). As a consequence of quenching, calcium- and aluminum-containing phases can also dissolve and other minerals can be formed as a result of dissolution/precipitation 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) Oxidizing agent: H2O2 Blast furnace slag Cu K Na Steel smelter slag Al, Ba Cd, Cr Cu, Na Ni, S Si, V MSWI bottom ash As Cr Cu V Wood ash As Co Cr Mn V Oxidation decreases leachability (95% significance) Oxidizing agent: H2O2 Blast furnace slag Fe Steel smelter slag Fe Mg Mn MSWI bottom ash Al, Fe Mg, Mn Na, Si Zn Wood ash Ba, Ca K, Mg Na, S Si, Zn

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Comans, 1997); for instance, ettringite can be formed according to the reaction (Meima and Comans, 1997): 6Ca2þ þ 2Al3þ þ 3SO2À þ 38 H2 O 4 !

! 12Hþ þ Ca6 Al2 ðSO4 Þ3 ðOHÞ12 Á26H2 OðsÞ : The formation of C-S-H phases was also detected (Speiser et al., 2000) as well as the neo-formation of clay-like minerals from the corrosion of glasses (Comans et al., 1994; Zevenbergen and Comans, 1994; Zevenbergen et al., 1996). Carbonation is caused by the uptake of atmospheric CO2 by the initially alkaline material, which leads to a decrease in pH and to the precipitation of calcite (Meima and Comans, 1997; Meima and Comans, 1999; Zevenbergen and Comans, 1994). CO2 absorption results in final pH values in the range of 8 to 8.5 (Bod´ enan et al., 2000; Meima and Comans, 1999). In this state the equilibrium between calcite and CO2 (forming HCO- with water) under the influence of gypsum dom3 inates the system as a buffer. At pH ffi 8 the solubility minimums are reached for most of the solid phases controlling the leaching of such heavy metals as Cd, Pb, Zn, Cu and Mo (Meima and Comans, 1999). Calcite can also provide a number of sorption sites for certain elements, e.g. Cd and Zn, that have been shown to display a high affinity for this phase (Meima and Comans, 1999). However, leaching of sulfate from weathered bottom ash has been found to increase if compared to ´ fresh bottom ash (Bodenan et al., 2000), probably as a

result of ettringite carbonation, which leads to the precipitation of gypsum. Sorption onto the neoformed minerals, including both adsorption and co-precipitation processes, also seems to play a role in reducing contaminant leaching from weathered bottom ash. Fe and Al (hydr)oxides as well as amorphous aluminosilicates formed as a result of weathering have been found to be reactive sorptive minerals for e.g. Cd, Zn, Cu, Pb and Mo (Meima and Comans, 1998; Meima and Comans, 1999). Similar processes are observed during the weathering of APC residues. In addition to the above mentioned phases, aging of APC residues can lead to the formation of Ca-Al-S-Cl-hydrate phases (e.g. hydrocalumite, ettringite, and members of the hydrotalcite group) containing varying amounts of heavy metals like Zn (Speiser et al., 2001; Heuss-Assbichler et al., 2002; Speiser et al., 2002). Fig. 10 provides a schematic representation of the weathering reactions and the related modifications in leaching behavior. 2.2. Gas production Gas generation at landfill sites with MSWI residues can be either of a biotic or abiotic nature. The low biodegradable organic carbon content of MSWI residues generally leads to the production of biogas amounts significantly lower if compared to MSW landfill gas. Conversely, the evidence of significant abiotic gas generation has been reported in a number of studies (Musselmann et

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

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al., 2002; Magel et al., 2001, Lechner et al., 1997). Abiotic gas is produced by chemical oxidation, in the presence of water, of elemental metals including Al, Fe and Cu. Thus, as previously observed for leachate production, the different parameters of concern for the water balance and the chemical properties of the waste material must be considered for gas production as well. Aluminum is a major constituent of bottom ash and is also significantly concentrated in fly ash and APC residues. Due to its high solubility at pH > 9.5 and to its lower redox potential if compared to other elements, aluminum is regarded as the main element responsible for abiotic gas production. In addition, a significant fraction of Al in MSWI residues is in its elemental form, which can undergo the following redox reactions, leading to hydrogen gas generation: 2Al0 þ 3H2 O À ! Al2 O3 þ H2 À Al0 þ 2H2 O À ! AlOOH þ 1:5H2 À Al0 þ 3H2 O À ! AlðOHÞ3 þ1:5H2 À However, the chemistry of aluminum corrosion is not completely understood at present, due to the variability of local conditions throughout the landfill mass. A number of studies (Forster and Hirschmann, 1997; ¨ Mizutani et al., 2000) evidenced that abiotic gas production is almost complete after several months, so hydrogen generation can be considered a short-term process. However, even though no specific studies have been carried out, some experimental data suggest that hydrogen gas production can also evolve over a longer term, as shown in Fig. 11 (Magel et al., 2001). It has also been found that isolated aluminum particles in MSWI residues are generally surrounded by a

reaction rim of Al(OH)3 and additionally by hydrocalumite (Ca2Al(OH)6Cl.2H2O) and ettringite ([Ca3Al (OH)6]2(SO4)3.26H2O). These coatings or by products may result in the retardation of hydrogen production. However, such rims can dissolve, leading to a permanent release of hydrogen. Furthermore, a significant number of ash particles are enclosed in glassy phases formed during incineration, which act as a barrier against the reaction between water and aluminum. However, due to their alkaline nature, such glassy phases can be altered, so aluminum particles will come into contact with the hydration water. 2.3. Temperature development Recently several studies have shown that many exothermic reactions may cause a temperature increase of up to 90  C in MSWI residue landfills (e.g. Klein et al., 2001; Heyer and Stegmann, 1997). MSWI residue storage is affected by heat generation as a result of different exothermic reactions, such as hydration of alkaline and alkaline earth oxides, corrosion of metals and carbonation of portlandite (Huber, 1998). The main effects of this temperature enhancement are (see Fig. 12):  acceleration of weathering/hydration reactions as long as the material in the landfill is wet or humid,  over a longer term, formation of salt rims as a consequence of drying, and  modification of precipitation/dissolution and complexation equilibria. Moreover, temperature increase may lead to evaporation of some dissolved gaseous compounds (CO2,

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

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Fig. 12. Relationship between temperature, water content and rate of weathering/hydration reactions.

O2) from the leachate and as a consequence may exert an effect on the redox potential and the concentration of complexing agents. Temperature development was studied by Klein et al. (2001) during an experimental campaign carried out at a German bottom ash monofill. Fig. 13 shows temperature development over time in three monitored sensor fields. In Sensor Field 1 (SF1) bottom ash was deposited

at irregular time intervals, after one to three weeks of storage at the landfill site. SF2 was built up over three weeks to its final height of ten meters. In SF3 bottom ash was placed in 1 meter-thick layers every two months up to a final height of 6 meters. Previous storage of bottom ash in this sensor field was disregarded. Bottom ash in all the sensor fields was not compacted and no temporary liners were used to cover the landfill between deposits. In every layer of the surveyed landfill the temperature development started with an increase immediately after deposition. Over the next three to four months the bottom ash temperatures increased to a maximum which varied depending on the layer depth. The average rate at which the temperatures rose was between 0.16 and 1.02  C per day. In all the observed landfill layers, the maximum temperature occurred at a time of about four to five months after deposition. The initial temperature rises and maximum temperatures occurred in those cases where the ash was not stored temporarily before landfilling. The experimental program also revealed that rainwater percolating through the landfill body exerted no appreciable effect on temperature development.

3. Potential environmental impacts The main potential environmental impacts related to the handling, utilization and disposal of MSWI residues can be summarized as follows:

Fig. 13. Temperature development in a German bottom ash monofill.

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

Moreover, as far as landfilling is concerned, the potential impacts arising from the construction and operation of the landfill should be taken into account. However, such impacts as additional traffic load, noise pollution and site modifications in terms of topographic, hydrological and hydrogeological conditions will not be discussed in this paper. The extent of the most relevant impacts may differ for the handling, utilization and final disposal phases and vary significantly over time. 3.1. Dust emissions MSWI residues contain a fine fraction of particulate matter passing the 74 mm mesh sieve that accounts for 1–10% of bottom ash, whilst APC residues have a particle size distribution varying between 0.001 and 1 mm. Friability of bottom ash may result in an increased percentage of finer particles after processing operations. The fine bottom ash particles typically contain chloride and sulfate salts as well as heavy metals like Pb, Cu and Zn (IAWG, 1997). The easily airborne nature of fine particles leads to the dispersion of pollutants, which in turn can give rise to health risks for exposed, unprotected workers and the public, as well as soil contamination. To prevent or minimize dust emissions, bottom ash and fly ash are normally kept wet (5–15% humidity) and transported by covered and watertight trucks. The upper limit for humidity is regarded as the minimum value required to prevent fugitive dust problems in open storage piles. The ability to maintain the optimal water content in order to minimize dust emissions is obviously related to climatic conditions (temperature, humidity as well as regime, intensity and frequency of the dominant local winds), which influence the desiccation rate of the material, the critical area of downwind dust deposition as well as the downwind distance at which dust can be transported. Matsuto et al. (2001) investigated the wind dispersion of incinerated residues and found that they are dispersed up to 50 meters from the landfill. The atmospheric dispersion of particles may not be significant during the after-closure period of a landfill due to the presence of a top cover. 3.2. Leachate generation The potential environmental impact of leaching includes contamination of soil, groundwater and surface water bodies. As leaching of contaminants from MSWI residues may occur during the temporary storage,

treatment or reuse as well as during the final disposal of the material, the following aspects should be investigated: the leaching behavior of contaminants, the environmental conditions that may occur in any of the above mentioned scenarios, as well as their variation over time. Thus, according to the discussion in the preceding sections, the following items should be considered:  residue characteristics, in terms of physical and mechanical properties, particle size distribution, acid neutralization capacity, concentration of contaminants, availability of contaminants for leaching, leaching mechanisms, controlling factors, and their variation over time due to weathering reactions;  characteristics of the application site in terms of (1) dimensions and (2) material properties (porosity, bulk density and permeability);  hydrological conditions of the application site in terms of (1) net rate of infiltration, (2) properties of the unsaturated zone and the aquifer (thickness, permeability, porosity, longitudinal and horizontal dispersivity, bulk density, flow velocity, etc.); moreover, as far as disposal is concerned, the presence of a top cover should be taken into account; and  mitigating effects due to leachate/soil interactions (e.g., ionic exchange, sorption) and to dilution. The extent of the impact depends on the rate at which leaching occurs and on the type and concentration of the dissolved species. The following elements must be considered as hazardous contaminants potentially leachable from MSWI residues: As, Al, B, Ba, Cd, Cr, Cu, Hg, Mn, Mo, Ni, Pb, Sb, Se, Zn, BrÀ, ClÀ, CNÀ, FÀ, NH+, NOÀ, NOÀ, SO2À. Thus, their concentration 4 3 2 4 in the leachate should be compared to specific quality criteria. Since no international guidelines for groundwater have been proposed so far, it appears reasonable to apply the international drinking water quality criteria (EU Drinking Water Directive and WHO criteria on drinking water) until a specific groundwater directive is proposed. The distinction between the short- and the long-term leaching behavior appears as a key factor. Whilst information is available concerning the short-term behavior of most MSWI residues (results of leaching tests and field measurements), long-term behavior can be predicted only on the basis of a synthesis of information on leaching principles, leaching tests results, field measurements, simulation of mineral changes and speciation. As far as the environmental impact assessment related to leaching of contaminants out of the MSWI residues is concerned, availability as opposed to the total concentration of contaminants in the solid matrix provides

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an estimation of the maximum amount of contaminant that in theory could be leached over a 1000- to 10000year timeframe (with the exception of highly soluble salts, for which the maximum leachable amount can be attained within shorter periods, typically a couple of years). Table 5 shows typical ranges of the concentration of contaminants in MSWI residue leachate. However, availability does not account for the acid neutralization capacity exerted by the matrix. This is an important parameter, which determines the potential environmental impact of MSWI residues, in that neutralization processes can affect leaching reactions and control the release of contaminants from the material. Acid neutralization capacity allows for the evaluation of the environmental behavior of MSWI residues, in that ANC data can be transformed in order to estimate the time required for the pH to drop from the ‘‘inherent’’ pH of the material to critical values for contaminants release (Astrup et al., 2001). Such time-related information can be gathered on the basis of the size of the application site, hydrological and hydrogeological conditions, leachate composition, and leachate flow towards soils and groundwater. A more detailed simulation of the potential environmental behavior of MSWI residues should also consider the attenuation phenomena caused by the interactions between leachate and soil, such as sorption and ion exchange (Hjelmar et al., 2001; Hjelmar et al, 1999a,b). However, in order to predict correctly the leaching behavior over time, additional information (pH, redox conditions, ionic strength, complexing agents, and mineralogy) is required. It has been observed (Hjelmar, 1996) that the first leachate produced by bottom ash has a relatively high content of inorganic salts (chloride, sulfate, sodium, potassium and calcium) and low concentrations of trace elements due to the fact that at this stage reducing conditions are occurring and pH is slightly or strongly alkaline, depending on the degree of carbonation. Among the trace elements, Cu can behave differently, as observed before, its leachability being

increased by DOC. With the exception of sulfate, the concentration of salts in the leachate tends to decrease over time (i.e. at increasing L/S ratios). APC residues behave differently from one another depending on their origin and air pollution control devices installed into the plant. High concentrations of readily soluble salts, such as chlorides and hydroxides of calcium, sodium and potassium generally characterize the first leachate from APC residues. Trace elements, such as Pb and Mo, which are mobile under reducing and slightly alkaline conditions, can be highly leachable at low L/S ratios (corresponding to the first fractions of the leachate). Thus APC residues are typically hazardous materials, and their disposal requires considerable care in order to prevent adverse environmental impacts. As far as co-disposal of bottom ash and APC residues is concerned, it was observed that soluble salt concentrations are higher in the combined ash leachate than in bottom ash and fly ash leachate (Hjelmar, 1996). This behavior was also observed for Cd and is likely to be ascribed to high complexing chloride content due to the presence of organic acids produced by biodegradation of residual unburned carbonaceous material in bottom ash. Very little data are available from the literature regarding the long-term composition of leachate from MSWI residues. Sabbas et al. (2001) investigated four different mixtures including 1) MSWI bottom ash (RAU-S2), a mixture of MSWI bottom ash and fly ash, 2) rotary drum kiln slag from hazardous waste incineration and fluidized bed incineration ash (RAU-S), 3) a mixture of rotary drum kiln slag from hazardous waste incineration and fluidized bed incineration ash (RAUA) as well as 4) a cement solidified product (RAU-SB, a mixture of MSWI bottom ash and fly ash, rotary drum kiln slag from hazardous waste incineration and fluidized bed incineration ash, cement and gravel). The materials were approximately 10–15 years old, naturally weathered, fully carbonated and showed leachate pH values between 7.3 and 8.9. The results shown in Fig. 14 reveal that all concentrations, except those of arsenic

Table 5 Maximum concentrations of contaminants in leachates from various MSWI residues (after Hjelmar, 1996) Typical maximum levels of concentration in leachate >100 g/l 10–100 g/l 1–10 g/l 100–1,000 mg/l 10–100 mg/l 1–10 mg/l 100–1,000 mg/l 10–100 mg/l 1–10 mg/l