Effects of enhanced metal recovery on the recycling potential of MSWI bottom ash fractions in various legal frameworks

Abstract

In recent years, complex new bottom ash treatment processes for enhanced metal recovery have been implemented in Switzerland, producing residual bottom ash fractions with various qualities. This study focusses on three different treatment processes by characterizing all arising fractions in detail. Thereby the factors influencing the composition of these fractions are identified and their recycling potential in Switzerland is investigated. However, high legislative requirements on total contents of heavy metals represent a high barrier for bottom ash recycling in Switzerland. Therefore, the recycling potential is further evaluated based on the waste legislation applied in the Netherlands, where recycling of bottom ash has a long tradition. There, threshold values for bottom ash recycling are based on leachate concentrations and not on total contents as in Switzerland. However, Swiss Waste Legislation also knows threshold values based on leachate concentrations for certain waste materials. The leaching tests applied in these two countries, however, are different. The comparison of both leaching tests reveals that the setup and conditions, especially the considered pH range, significantly influence the leaching of heavy metals. With emphasis on problematic pollutants, the possibilities for new applications of these fractions are evaluated based on Swiss and Dutch legal threshold values. The comparison within the legal frameworks of these two countries allows recognizing opportunities and risks related to bottom ash recycling.

Keywords

Municipal solid waste incineration bottom ash metal recovery process waste legislation recycling potential leaching tests

Introduction

Recycling of municipal solid waste incineration (MSWI) bottom ash offers an opportunity towards circular economy and is established to different extent in European countries (Blasenbauer et al., 2020). In Switzerland separate collection and recycling of municipal solid waste (MSW) has a long tradition and approx. 50% of total MSW is recycled (FOEN, 2019). Strict regulations demand the incineration of the remaining MSW, presently accounting for 4.2 million tonnes, in Switzerland’s 30 MSWI plants (FOEN, 2019; Swiss Confederation, 2016). The benefit of thermal treatment are mass and volume reduction, destruction of organic compounds and energy recovery. However, about 20 wt.% (800,000 t y⁻¹) remain as bottom ash, with high contents of ferrous (approx. 9 wt.%) and non-ferrous metals (approx. 3.5 wt.%) (Bunge, 2014).

With this composition bottom ash is predestined for metal recovery, which is demanded by the Swiss Waste Ordinance (Swiss Confederation, 2016). Recovering metals from bottom ash is both economically and ecologically beneficial (Allegrini et al., 2014, 2015; Bunge, 2018; Haupt et al., 2018; Mehr et al., 2020). However, after metal recovery the residual bottom ash is removed from the material cycle and disposed of on specific landfills (Type D). There, only threshold values for elemental non-ferrous metal content in the grain size spectrum 2−16 mm (<1 wt.%) and total organic carbon (TOC) content (<2 wt.%) have to be met. In contrast, on landfills of a lower category, additional threshold values for total contents or leachate concentrations based on batch leaching tests according to Swiss Waste Legislation would apply (FOEN, 2017). The same applies for the recycling of suitable waste fractions as secondary raw material for the production of cement clinker. Due to the elevated heavy metal contents of bottom ash, these options are very limited. An opportunity to do so however, lies in the variety of new enhanced metal recovery processes, which have been implemented in Swiss MSWI plants and on landfills in recent years. The influence of these modern processes on the residual bottom ash has not been studied in detail although they potentially improve the quality of sub-fractions. However, the pressure to recycle residual bottom ash is low in Switzerland as primary mineral raw materials are highly abundant. Switzerland is rich in gravel deposits, and additionally vast amounts of unpolluted excavation materials arise, for example, from tunnel construction.

In contrast, other European countries pursue a strategy of bottom ash treatment with partial or full recycling in the construction sector (Blasenbauer et al., 2020; Verbinnen et al., 2017). Based on national regulations, applications for suitable bottom ash fractions exist, for example, as foundation layers in road construction (Danish EPA, 2016; LAGA, 1994, 2004; MATTM, 2006; MEDDTL, 2011). Further, applications as secondary raw material in the cement production are possible in some countries (MATTM, 2006; Republic of Austria, 2017). Especially the Netherlands has a long tradition and a leading role in bottom ash recycling, due to lack of primary mineral raw materials. Therefore, bottom ash is an important source of alternative raw materials, and based on the Green Deal, by 2020, all bottom ash has to be recycled as secondary raw material (Government of the Netherlands, 2012). Suitable fractions are used as unbound, granular construction material, the remaining shares in shaped materials like concrete. For both applications, bottom ash has to comply with threshold values based on leachate concentrations in order to assess the release potential of pollutants into the environment (Soil Quality Decree, 2013). For unbound granular construction materials, the leaching of pollutants is determined with threshold values based on column tests according to NEN 7383 (2004). This Dutch standard is the base for the standardized European column test (CEN/TS 14405, 2017), and the two tests are comparable. Threshold values for shaped materials are determined on the products by diffusion tests according to NEN 7375 (2004).

The aim of this study is to investigate how new enhanced metal recovery processes applied in Switzerland influence the composition of residual bottom ash fractions. For this, two Swiss MSWI plants operating with innovative processes have been selected. All fractions produced in these plants after metal recovery have been characterized, and the factors influencing their composition have been identified. Focusing on constituents relevant for Swiss Waste Legislation allows recognizing fractions suitable for applications as secondary raw material. To assess the suitability of these fractions in more detail, regulations according to the Swiss and Dutch Waste Legislations have been considered and compared. The sum of all these investigations creates a detailed overview of the possibilities and limitations emerging for Swiss bottom ash fractions remaining after enhanced metal recovery. The gained insight on two different legal frameworks shows the respective recycling options for the same material and helps to find potential improvements to the existing national regulations.

Materials and methods

Bottom ash treatment processes

In Switzerland MSWI plants traditionally operate with wet discharge of bottom ash and subsequent conventional metal recovery (Figure 1a). In recent years, individual plants developed and implemented enhanced metal recovery processes such as the wet-based recovery of metals in fine fractions or electrodynamic fragmentation for the liberation of metals (Figure 1b and c).

The conventional metal recovery process generally requires ageing of the wet discharged raw bottom ash (weeks to months), which is then crushed and sieved into sub-fractions (Figure 1a). Using magnetic and eddy current separation, as well as optical sensor sorting, metals are recovered from these fractions, which are subsequently reunited at the end of the process to one fraction, conventional residual bottom ash (cBA).

In contrast, plants operating with enhanced metal recovery usually do not require ageing but treat the raw bottom ash immediately after discharge, producing at least two separate, residual bottom ash fractions (Figure 1b and c): (1) finer fractions, which are removed at the beginning of the process by washing and/or sieving, where no metals are recovered (rBA) and (2) coarser fractions, undergoing enhanced treatment, in order to optimize the amount and quality of metals (eBA). The metals in these fractions are then recovered with the same methods as applied in the conventional process.

Apart from wet-based bottom ash discharge and treatment, another modern technology operating with dry discharge and subsequent enhanced dry metal recovery exists in Switzerland. Recent publications focus on this technology (Blatter et al., 2014; Mehr et al., 2020).

Specification of the bottom ash fractions

The MSWI Plant 1 performs conventional metal recovery (Process A) from bottom ash up to a grain size of 80 mm (Fraction A). Metal pieces >80 mm are manually separated, and the uncombusted coarse material >80 mm is returned to the waste bunker (Figure 1a). From 2016 to 2017 enhanced metal recovery based on Process B was tested in Plant 1, where the bottom ash fraction <0.033 mm (B-rBA) is initially removed by wet sieving, while wet-based metal recovery is applied to the fraction 0.033–2 mm with a separate module (Figure 1b). The remaining bottom ash fraction 2–80 mm undergoes conventional metal recovery along Process A and is subsequently united with the fraction 0.033–2 mm to the enhanced treated residual bottom ash fraction 0.033–80 mm (B-eBA). Plant 1 was able to switch between metal recovery Process A and B at any time, producing a total of approx. 21,000 t y⁻¹ of residual bottom ash.

In Plant 2, the enhanced metal recovery Process C is implemented, where the bottom ash fraction <2 mm is removed initially in two steps by dry (C-rBA1) and wet sieving (C-rBA2), without further metal recovery (Figure 1c). The bottom ash fraction >40 mm (C-cBA) is removed initially as well and metals are separated conventionally. This fraction is not considered in this study. The enhanced treatment is therefore applied to the fraction 2–40 mm, where metals are liberated with electrodynamic fragmentation and subsequently recovered in two fractions (C-eBA1 and C-eBA2). Plant 2 produces a total of approx. 16,000 t y⁻¹ of residual bottom ash.

Sampling and sample preparation

In 2017, a total of approx. 1 tonne of bottom ash has been sampled during three campaigns in Plants 1 and 2, where for 2 days 5–6 increments of ±20 kg of every residual bottom ash fraction have been sampled after metal recovery. Fractions >5 mm were subsequently crushed and sieved to a grain size <5 mm, while metals and unburnt material >5 mm were manually separated. No additional grain size reduction was conducted for the finer fractions. Finally, all bottom ash fractions were split into representative samples for further analysis. Treated like this, the material was used for leaching experiments. For chemical analysis the material was dried at 105°C, crushed and milled to a grain size <0.25 mm, while metals and unburnt material were manually separated. The results on the chemical composition thus correspond to metal depleted bottom ash.

Chemical analysis

Chemical composition

Chemical composition was determined by wavelength dispersive X-ray fluorescence analysis (WD-XRF) using a PANalytical Omnian spectrometer with matrix adjusted calibration. All XRF measurements were performed on pressed powder pellets (40 mm diameter) using 6.40g material <0.25 mm and 1.44 g Hoechstwax as binder. For quality control, two pellets of each fraction have been measured.

TOC was determined using a CHNS-analyser, while the loss on ignition (LOI) was determined at 1050°C for 1.5 hours to account for additional water of crystallization and CO₂ from carbonates. As TOC is accounted for separately, it is subtracted from LOI, which is thus shown in the results as ‘LOI (without TOC)’.

Leaching behaviour

Leaching experiments have been performed on bottom ash fractions <5 mm or in their original condition if finer. Batch leaching tests have been performed according to Swiss Waste Legislation (FOEN, 2017) during 24 hours with a liquid to solid (L/S) ratio of 10 l kg⁻¹. Further, Swiss Waste Legislation demands CO₂-saturated water for the elution of cationic heavy metals, simulating a stage of advanced weathering due to acid rain conditions. For this reason, batch tests have to be performed in two parts: Test 1 − Mixing of bottom ash with CO₂-saturated water (injection of approx. 50 ml CO₂ min⁻¹) for the determination of heavy metals and Test 2 − Mixing of bottom ash with deionized water for the determination of all other constituents. Column tests have been performed according to the European standard CEN/TS 14405 (CEN/TS 14405, 2017), where bottom ash is packed 30 ± 5 cm high into a glass column and the leachate is sampled at specified L/S ratios (0.1, 0.2, 0.5, 1, 2, 5 and 10 l kg⁻¹). As eluent, deionized water under up-flow-percolation with a flowrate of 12 ml hour⁻¹ is used.

The leachates were analysed for Cu, Pb and Zn by inductively coupled plasma optical emission spectroscopy (iCAP 7000 SERIES), calibrated with certified standard solutions (ICP Multi-element Standard Merck IV, Sigma-Aldrich 6 and RotiStar). Ion chromatography (IC) was used to analyse Cl⁻, using a Metrohm 850 Professional IC system. Additionally, DOC was determined by non-dispersive infrared spectroscopy (NDIR) using a multi N/C 2100 S analyser.

Enrichment factor

Enhanced metal recovery leads to fractionation trends among the resulting fractions of a process. To study these fractionation trends with emphasis on trace elements, all fractions of a process are normalized to the overall bottom ash composition. The resulting enrichment factor ∆_f is based on equation (1):

Δ_{f} = \frac{c_{f r a c t i o n}}{c_{o v e r a l l}}

(1)

where ∆_f = enrichment factor

c_fraction = concentration of an element in fractions from an enhanced bottom ash treatment process (mg kg⁻¹ or mg l⁻¹)

c_overall = concentration of an element in the overall bottom ash (sum of all fractions) from an enhanced bottom ash treatment process (mg kg⁻¹ or mg l⁻¹)

Thus, ∆_f > 1 indicates enrichment and ∆_f < 1 depletion of a trace elements in a bottom ash fraction. Enrichment factors ∆_f are applied to total contents as well as leachate concentrations.

Results and discussion

Characterization of residual bottom ash

The major oxide composition of the overall bottom ash from Processes A−C show significant differences (Figure 2a). In Process A and B especially, the content of SiO₂, CaO, SO₃, Fe₂O₃ and Na₂O differ. As the bottom ash of these two processes derives from Plant 1 the differences are induced by the various treatment steps involved in these processes. Extensive wet-based treatment in Process B, leads to partial dissolution of carbonates and sulphates, while Fe₂O₃ is partially removed by the enhanced metal separation. With respect to Process A, CaO, SO₃ and Fe₂O₃ are thus depleted, while SiO₂ and Na₂O are enriched. Further the ageing treatment of Process A leads to carbonation of the bottom ash, resulting in an increased CaO content. Based on a different waste input, the overall bottom ash from Process C in Plant 2, stands out with high contents of CaO and LOI. This indicates a waste input more abundant in carbonates, possibly due to high quantities of waste from construction materials. Regardless of these differences, in all processes approx. 90 wt.% of the overall bottom ash consist of the major oxides SiO₂, CaO, Fe₂O₃, Al₂O₃, SO₃ and Na₂O. The remaining shares consist to approx. ⅔ of the major oxides MgO, TiO₂, K₂O, P₂O₅ and MnO and to approx. ⅓ of trace elements including TOC.

Concerning trace elements, the influence of enhanced metal recovery (Process B and C) is demonstrated by the concentrations of Cu, Zn, Cr and Pb in the overall bottom ash, which are 20–50% lower than in conventionally treated residual bottom ash (Process A) (Figure 2b). The same effect is observed for TOC, due to the removal of unburnt components by floating in the course of treatment with process water. The remaining shares are composed of Cl (20–30 wt.%), alkaline earth metals (10–15 wt.%) and remaining heavy metals (5–8 wt.%). Of the latter As, Cd, Co, Ni, Sb, Sn are relevant for Swiss Waste Legislation.

Fractionation trends

When individual bottom ash fractions after metal recovery are considered the distribution pattern of major oxides changes. The conventional metal recovery Process A produces all residual bottom ash in a single Fraction A, where the two major components SiO₂ and CaO account for approx. 30 wt.% each (Figure 2a). In contrast, treating the same input material with the enhanced metal recovery Process B leads to a fractionation of these constituents. The removed fine fraction (B-rBA) is enriched in CaO and depleted in SiO₂, while for the enhanced treated fraction (B-eBA) the inverse pattern applies (Figure 2c). In fact, for many major oxides such fractionation trends are observed, showing similarities between Process B and C (Figure 2c–f): removed fine fractions (rBA) are enriched in CaO and SO₃, enhanced treated fractions (eBA) in SiO₂, Fe₂O₃, Al₂O₃ and Na₂O. Strong negative correlations (R² ⩾ 0.8) emphasize these observations, indicating that grain size is the main factor for fractionation. The chemical composition of removed fine fractions is strongly related to fly ash, with a high abundance of carbonates and sulphates (Wan et al., 2006; Weibel et al., 2017). In contrast, the chemical composition of enhanced treated bottom ash fractions is related to transit minerals like quartz, feldspar and pyroxene, abundant in coarse particles.

While major oxides influence the general physical and chemical composition of bottom ash fractions, trace elements are relevant concerning environmental properties. Based on the enrichment factor (∆_f) according to equation (1) all fractions of a process are normalized to the overall bottom ash composition. By this, the relevant treatment steps of enhanced metal recovery processes that influence the enrichment or depletion of trace elements in bottom ash fractions are highlighted (Figure 3a and b). The following factors are crucial: (1) grain size, (2) vapour pressure of heavy metal bearing compounds and (3) washing during bottom ash treatment. For the distribution of TOC, grain size is the only controlling factor, resulting in ∆_f > 1 for fine fractions (rBA) and ∆_f < 1 for coarse fractions (eBA). For heavy metals, the two factors grain size and vapour pressure control the fractionation. Heavy metals or heavy metal bearing compounds with high vapour pressures (e.g. chlorides) are generally enriched in the fine fractions (rBA) due to their volatile character (Chandler et al., 1997). These chemical compounds tend to be vaporized during combustion and subsequently accumulate on fly ash particles in the flue gas (Morf et al., 2000). How these compounds precipitate on fine fractions of the bottom ash on the grate remains unclear. In any case, Pb, Sb, Cd and Zn clearly follow a tendency with strong enrichment of ∆_f > 2.5 in the fine fraction of Process B (B-rBA). Also in Process C, these elements are enriched in the fine fractions (C-rBA1 & 2). Additionally Pb, Sb and Zn correlate with SO₃ (R² ⩾ 0.95), possibly due to incorporation of these elements in sulphates such as ettringites abundant in fine fractions (Gougar et al., 1996; Chrysochoou and Dermatas, 2006; Cornelis et al., 2006). The volatile heavy metals Sn and As, however, are only partially enriched in the fine fractions.

In contrast, the lithophilic heavy metals V, Ni, Cr and Co with low vapour pressures tend to remain in the bottom ash and are therefore enriched in the coarse, enhanced treated fractions (eBA). Concerning the lithophilic Cu, no distinct fractionation is observed, resulting in a similar abundance of Cu in all fractions with ∆_f approx. 1. Finally, the distribution of Cl is generally controlled by grain size, showing increasing ∆_f with decreasing grain size. However, washing of a fraction during bottom ash treatment strongly influences ∆_f of Cl, as it is readily soluble: the washed fine fraction C-rBA2 shows 50% lower Cl concentrations (8000 mg kg⁻¹) than the dry removed fraction C-rBA1 (17,000 mg kg⁻¹) (Figure 3b).

Leaching behaviour

Leaching tests of different setup and operating conditions provide information on the availability of species to the environment. How the leaching tests applied in Swiss and Dutch Waste Legislations influence leaching and mobility of species is studied on Cl, DOC, Cu, Zn and Pb. These species have been selected because of their importance in waste legislation and because they represent groups with a characteristic leaching behaviour: availability-controlled main components (Cl), organic constituents (DOC) and heavy metals with differing leaching behaviour (Cu, Zn and Pb).

Influence of leaching tests on the leaching behaviour

The pH of column tests and batch tests with deionized H₂O both lie in the range of 10.8–12.5 (Figure 4a–f), while for CO₂-saturated batch tests pH is 5.6–6.6 (Figure 4c–e), indicating the importance of the applied conditions. For Cl and DOC batch and column tests are performed with deionized H₂O resulting in the same alkaline pH range and strongly correlating results (R² ⩾ 0.95). The average concentrations are approx. 300 mg l⁻¹ Cl and approx. 40 mg l⁻¹ DOC, with deviations of approx. 10% between the two tests (Figure 4a and b). In contrast, heavy metal leaching is determined under CO₂-saturated conditions (Test 1) in Switzerland and thus not comparable with results from column tests relevant in the Netherlands. To study the influence of the test setup on heavy metal leaching nevertheless, concentrations from the batch test with deionized H₂O (Test 2) are considered. For these similar conditions, the average concentrations of Cu and Zn in the batch test (0.45 mg l⁻¹ Cu and 0.20 mg l⁻¹ Zn) are ±30% with respect to the column test (Figure 4c and d). Thus, especially for Cu, the results of the two tests strongly correlate (R² = 0.95), while for Zn correlation is not pronounced (R² = 0.65). For Pb compliance of the results is less distinct, with average concentrations in the batch test with deionized H₂O (0.55 mg l⁻¹ Pb) that are 2.5 times higher than in the column test (Figure 4e). Consequently, correlation between the results of the two tests is low (R² = 0.55). Thus, while compliance between column and batch tests for Cl, DOC and Cu is given, the results of Zn and Pb deviate. Possibly this is due to slight differences in pH as observed in other studies (Lopez Meza et al., 2008; Di Gianfilippo et al., 2016).

To study the effect of CO₂-saturated conditions on heavy metal leaching, the concentrations from both parts of the Swiss batch test (Tests 1 and 2) are compared. Especially Zn is strongly influenced with an average concentration of 3 mg l⁻¹, which is 15 times higher than in the batch test with deionized H₂O (Figure 4d). This highlights the pH sensitivity of Zn, which is sorbed to reactive surfaces of hydrous ferric oxides (HFO) above pH 7.5 (Dzombak and Morel, 1990). At the present CO₂-saturated conditions with pH 6, Zn is mobilized as Zn²⁺-ions, leading to the observed elevated concentrations. In fact, this observation is characteristic for most heavy metals. While soluble under strongly alkaline (hydroxo complexes) and acidic conditions (cations), heavy metal solubility is lowest around neutral pH (Chandler et al., 1997). However, often there are other factors involved. The average concentrations of Cu under CO₂-saturated conditions are 25% higher than under the alkaline conditions, present in batch tests with deionized H₂O (Figure 4c). This is because at pH 6 Cu is still mostly sorbed to HFO, while under alkaline conditions Cu leaching is enhanced by complexation of Cu with organic ligands such as DOC (Johnson et al., 1996; van Zomeren and Comans, 2004; Dijkstra et al., 2006; Glauser et al., 2020). In fact, Cu concentrations in fractions with high DOC concentrations are even higher under alkaline than under CO₂-saturated conditions. This is highlighted by a strong correlation of Cu and DOC (R² = 0.9). Similarly, average concentrations of Pb are higher under alkaline conditions, where Pb is mobilized as Pb(OH)₃⁻ (Holleman and Wiberg, 2007) (Figure 4e). Additionally, complexation with inorganic ligands such as Cl, may contribute to further mobilization of Pb under alkaline conditions (Byrne and Miller, 1984; Powell et al., 2007; Powell et al., 2009). This is supported by the correlation of Pb and Cl (R² = 0.75). Under acidic conditions mobilization of Pb²⁺ only starts below pH 5, which is not reached under CO₂-saturated conditions (Holleman and Wiberg, 2007). However, under these conditions leaching of Pb is likely to be limited by the formation of PbCO₃ (Simon et al., 1995). In sum, these findings demonstrate the importance of pH for the mobility of the studied heavy metals. However, additional factors influence the behaviour of each heavy metal and are responsible for the differences in mobility between the two batch tests.

Influence of bottom ash treatment on the leaching behaviour

Apart from the applied leaching test, the bottom ash treatment influences the leaching behaviour of each fraction, which is highlighted by the enrichment factor ∆_f according to equation (1). As observed for total contents, grain size proves to be an important factor once more. Especially under alkaline conditions, coarse fractions (eBA) clearly show decreased leaching in Cl, DOC and heavy metals with ∆_f predominantly <0.5 (Figure 5a and b). In terms of absolute concentration, the fractions C-eBA1 and C-eBA2 from Process C in Plant 2 show the lowest concentrations (Figure 4). There, another factor contributes to the decreased leaching: the intense washing with clean process water (conductivity approx. 4 mS cm⁻¹) for the electrodynamic fragmentation leading to an additional solution of salts. Especially for availability-controlled Cl this results in strongly decreased leaching. The importance of both, grain size and washing becomes apparent, when the enhanced treated fraction B-eBA from Process B in Plant 1 is considered: with a wide grain size range (0.033−80 mm) and process water circulated in a closed loop, concentrations tend to be higher with respect to fractions C-eBA1 and C-eBA2 (Figure 4).

In contrast, removed fine fractions (rBA) show increased leaching in these studied constituents with ∆_f predominantly >2 (Figure 5a and b). The factors grain size and process water define the leaching behaviour also within these fine fractions: Fraction B-rBA with the smallest grain size and washed with circulated process water, generally shows the highest concentrations and ∆_f of up to 8.5 (Figures 4 and 5a). In contrast, the cleaner process water in Process C results in 50% lower concentrations of availability-controlled Cl in the washed fraction C-rBA2 with respect to the fraction C-rBA1, removed without washing (Figure 5b).

Concerning heavy metals under CO₂-saturated conditions, leaching trends are less distinct with ∆_f predominantly between 0.5 and 1.5 (Figures 4, 5a and b). Especially for Cu and Zn the pattern changes for some fractions. There, removed fine fractions (rBA) show decreased (∆_f < 1) and enhanced treated fractions (eBA) increased leaching (∆_f > 1). Due to high CaO-contents, the fine fractions feature an increased buffer capacity, which results in higher pH-values. Thus, sorption of Cu and Zn to HFO is stronger in these fractions, resulting in lower concentrations and ∆_f < 1. While the importance of pH has been acknowledged already, the buffer capacity is another factor relevant for the leaching behaviour of heavy metals.

Results in relation to waste legislation

Swiss Waste Legislation

For the disposal of bottom ash on Landfill Type D, Swiss Waste Legislation only demands threshold values for TOC and non-ferrous metal content (Swiss Confederation, 2016). In fact, this is for a good reason as shown in Table 1: if other threshold values usually valid for materials disposed on landfill Type D would apply, no fraction would meet these requirements. Mainly this is due to the elevated Sb contents exceeding the Type D threshold value by 1.5−3 times in enhanced treated fractions and by up to 11 times in removed fine fractions. Additionally, other heavy metals exceed Type D threshold values in individual fractions. Consequently, disposal on landfills with lower requirements (Type B) or even recycling of bottom ash fractions as raw material for cement clinker is not possible without applying further treatment steps.

Table 1.

Classification of residual bottom ash fractions according to threshold values of the Swiss Waste Legislation.

Only parameters relevant for bottom ash are considered. Results from XRF measurements, all values in mg kg⁻¹. Green = compliance with threshold values applying to waste materials suitable for use as raw material for cement clinker production. Yellow = compliance with threshold values of Landfill Type B. Red = compliance with threshold values of Landfill Type D. Violet = threshold values of Landfill Type D exceeded.

Dutch Waste Legislation

The categorization along Dutch Waste Legislation is only performed for the selected species discussed in section ‘Influence of leaching tests on the leaching behaviour’. Based on the enrichment observed in Figure 5 the fine fractions (rBA) exceed the threshold values for Cu, Pb and Cl (Table 2). As has been demonstrated, the main factors responsible for this are grain size and treatment with used process water. When these factors are considered, opportunities for recycling exist, as is demonstrated with the coarse, enhanced treated fractions (eBA). Fraction C-eBA2 complies with the considered threshold values for unbound construction material and thus qualifies as secondary raw material according to the Dutch Waste Legislation without further treatment. Fractions C-eBA1 moderately exceeds the threshold values for Cl and Cu, while fraction C-eBA2 clearly exceeds the threshold value for Cl due to the treatment with contaminated process water. However, with adjustments to the treatment, compliance with threshold values in these fractions would be realistic.

Table 2.

Classification of residual bottom ash fractions according to Dutch threshold values based on column tests according to CEN/TS 14405.

Only selected species discussed in section ‘Influence of leaching tests on the leaching behaviour’ are considered. All values in mg kg⁻¹. Green = compliance with threshold values for granular material. Recycling as granular material without restrictions. Red = threshold values for granular material exceeded, thus recycling as raw material for shaped products.

Conclusion

Recently implemented enhanced metal recovery processes in the Swiss MSWI Plants 1 and 2 substantially change the composition of residual bottom ash with respect to traditional conventional metal recovery processes. There, treatment steps involve fractionation in different grain size categories leading to trends of distinct enrichment and depletion of selected elements. Thereby volatile heavy metals tend to be enriched in removed fine fractions, while lithophilic heavy metals are stronger represented in the coarser fractions. Further, treatment with process water fosters the dissolution of salts and the mobilization of availability-controlled constituents (e.g. Cl) and associated heavy metals, in particular Cu and Pb by complexation under alkaline conditions. In contrast, the mobilization of Zn is fostered by CO₂-saturated conditions resulting in slightly acidic conditions. The comparison of different leaching tests highlights the importance of the conditions prevailing in a studied system. Thus, when establishing threshold values based on leachate concentrations, a strong emphasis on the geochemical range covered by the applied leaching test is imperative.

In Switzerland, this is considered by the legal batch leaching test performed for different pH ranges, developed to assess materials for final disposal on special landfills. For the majority of waste materials, however, threshold values in Switzerland are based on total contents and thus independent of geochemical conditions. By this, an explicit framework is granted. However, it implies high barriers for the recycling of bottom ash fractions even with depleted heavy metal concentrations. Due to the non-existent scope concerning the environmental availability of these potential pollutants, bottom ash fractions, even after enhanced metal recovery, fail to comply with Swiss threshold values and thus do not qualify for recycling.

In contrast, Dutch threshold values for bottom ash derive from column tests performed with deionized H₂O, resulting in alkaline conditions. By neglecting changing pH conditions over time, the test is only representative for fresh bottom ash, as is the case for fractions considered for recycling. Thus, this test allows evaluating the recycling potential of the studied fractions based on these alternative legal requirements. Effectively, for selected species, bottom ash fractions after enhanced metal recovery prove suitable for recycling according to the Dutch Waste Legislation.

Thus, the possibility to recycle bottom ash fractions as secondary raw material primarily depends on the approach on which threshold values are based on. In order to recycle these fractions in Switzerland further treatment would be necessary, especially to lower total contents of heavy metals. However, with the strict threshold values valid today, this will prove challenging. In fact, if bottom ash fractions in Switzerland are to be recycled, an approach from two directions should be considered: (1) development of mechanical and (bio)chemical treatments to reduce heavy metal concentrations and (2) adaptations of the waste legislation towards more favourable conditions for the recycling of treated bottom ash fractions. For the latter a political discourse should be initiated to define the extent of circular economy aspired in the future.

Footnotes

Acknowledgements

We thank the MSWI plant operators for providing sample material and information about their metal recovery process. Analytical support by Priska Bähler, Christopher Pichler and Anna Zappatini (University of Bern) is highly acknowledged. Finally, we thank Kaarina Schenk for the valuable inputs enhancing the quality of this manuscript.

Declaration of conflicting interests

The authors declared no potential conflicts of interest with respect to the research, authorship, and/or publication of this article.

Funding

The authors disclosed receipt of the following financial support for the research, authorship, and/or publication of this article: However, the research project in general (dissertation of the corresponding author) was financially supported by the following associations/authorities and companies: (1) Verband der Betreiber Schweizerischer Abfallverwertungsanlagen (VBSA), Switzerland (English: Association of Swiss Waste Recycling Plant Operators), (2) Amt für Wasser und Abfall des Kantons Bern (AWA), Bern, Switzerland (English: Office of Water and Waste of the Canton of Bern), (3) KVA Linth, Niederurnen, Switzerland (English: MSWI Plant Linth), (4) SAIDEF SA, Posieux, Switzerland (English: MSWI Plant SAIDEF), (5) SELFRAG AG, Kerzers, Switzerland.

ORCID iD

Andreas Glauser

References

Allegrini

Maresca

Olsson

, et al. (2014) Quantification of the resource recovery potential of municipal solid waste incineration bottom ashes. Waste Management 34: 1627–1636.

Allegrini

Vadenbo

Boldrin

, et al. (2015) Life cycle assessment of resource recovery from municipal solid waste incineration bottom ash. Journal of Environmental Management 151: 132–143.

Blasenbauer

Huber

Lederer

, et al. (2020) Legal situation and current practice of waste incineration bottom ash utilisation in Europe. Waste Management 102: 868–883.

Blatter

zur Mühlen

Langhein

(2014) Die praktische Umsetzung der Trockenentschlackung (English: The practical implementation of dry bottom ash discharge). In: Thiel

Thomé-Kozmiensky

Pretz

, et al. (eds) Mineralische Nebenprodukte und Abfälle - Flaschen, Schlacken, Stäube und Baurestmassen. Nietwerder: TK Verlag Karl Thomé-Kozmiensky, pp.173–194.

Bunge

(2014) Wieviel Metall steckt im Abfall? (English: How much metal is in the waste?). In: Thomé-Kozmiensky

(ed.) Mineralische Nebenprodukte und Abfälle. Neuruppin: TK Verlag Karl Thomé-Kozmiensky, pp.91–130.

Bunge

(2018) Recovery of metals from waste incineration bottom ash. In: Thomé-Kozmiensky

Holm

(eds) Removal, Treatment and Utilisation of Waste Incineration Bottom Ash. Neuruppin: Thomé-Kozmiensky Verlag GmbH, pp.63–143.

Byrne

Miller

(1984) Medium composition dependence of lead(II) complexation by chloride ion. American Journal of Science 284: 79–94.

CEN/TS 14405 (2017) Characterization of Waste – Leaching Behaviour Tests – Up-Flow Percolation Test (Under Specified Conditions). Brussels, Belgium: European Committee for Standardization.

Chandler

Eighmy

Hartlen

, et al. (1997) Municipal solid waste incineration residues: An international perspective on characterisation and management of residues from municipal solid waste incineration. Studies in Environmental Science 67: 483–505.

10.

Chrysochoou

Dermatas

(2006) Evaluation of ettringite and hydrocalumite formation for heavy metal immobilization: Literature review and experimental study. Journal of Hazardous Materials 136: 20–33.

11.

Cornelis

Van Gerven

Vandecasteele

(2006) antimony leaching from uncarbonated and carbonated MSWI bottom ash. Journal of Hazardous Materials 137: 1284–1292.

12.

Danish

EPA

(2016) Danish Statutory Order no. 1672. Odense, Denmark: Danish Environmental Protection Agency.

13.

Di Gianfilippo

Costa

Verginelli

, et al. (2016) Analysis and interpretation of the leaching behaviour of waste thermal treatment bottom ash by batch and column tests. Waste Management 56: 216–228.

14.

Dijkstra

van der Sloot

Comans

RNJ

(2006) The leaching of major and trace elements from MSWI bottom ash as a function of pH and time. Applied Geochemistry 21: 335–351.

15.

Dzombak

Morel

(1990) Surface Complexation Modeling: Hydrous Ferric Oxide. New York: John Wiley & Sons.

16.

FOEN (2017) Messmethoden im Abfall- und Altlastenbereich: Umweltvollzug nr. 1715 (English: Measurement methods for waste materials and contaminated sites: Environmental enforcement no. 1715). In: Federal Office for the Environment. pp. 1–82.

17.

FOEN (2019) Abfallmengen und Recycling 2018 im Überblick (English: Overview of waste and recycling 2018). In: Federal Office for the Environment (ed).

18.

Glauser

Morf

Weibel

, et al. (2020) Ten-years monitoring of MSWI bottom ashes with focus on TOC development and leaching behaviour. Waste Management 117: 104–113.

19.

Gougar

Scheetz

Roy

(1996) Ettringite and C–S–H Portland cement phases for waste ion immobilization: A review. Waste Management 16: 295–303.

20.

Government of the Netherlands (2012) B-76 Green Deal: More Sustainable Practical Use of EfW Plant Clinker. Den Haag: Ministry of Infrastructure and the Environment & Association of Waste Companies, pp.1–5.

21.

Haupt

Kägi

Hellweg

(2018) Modular life cycle assessment of municipal solid waste management. Waste Management 79: 815–827.

22.

Holleman

Wiberg

(2007) Lehrbuch der Anorganischen Chemie, 102 edn. Berlin; New York: Walter de Gruyter, pp.1042–1081. Chapter XV.3.

23.

Johnson

Kersten

Ziegler

, et al. (1996) Leaching behaviour and solubility—controlling solid phases of heavy metals in municipal solid waste incinerator ash. Waste Management 16: 129–134.

24.

LAGA (1994) Mitteilung 19: Merkblatt über die Entsorgung von Abfällen aus Verbrennungsanlagen für Siedlungsabfälle (Laga M19) (English: Communication 19: Leaflet on the disposal of waste from municipal waste incinerators). In: Länderarbeitsgemeinschaft Abfall (ed).

25.

LAGA (2004) Mitteilung 20: Anforderungen an die stoffliche Verwertung mineralischer Abfälle (Laga M20) (English: Requirements for the recycling of mineral waste). In: Länderarbeitsgemeinschaft Abfall (ed).

26.

Lopez Meza

Garrabrants

van der Sloot

, et al. (2008) Comparison of the release of constituents from granular materials under batch and column testing. Waste Management 28: 1853–1867.

27.

MATTM (2006) Decreto ministeriale 5 aprile 2006, n. 186: “regolamento recante modifiche al decreto ministeriale 5 febbraio 1998 - individuazione dei rifiuti non pericolosi sottoposti alle procedure semplificate di recupero, ai sensi degli articoli 31 e 33 del decreto legislativo 5 febbraio 1997, n. 22” (English: Ministerial decree april 5 2006, No. 186: “regulation amending ministerial decree of 5 February 1998 - identification of non-hazardous waste subject to simplified recovery procedures, pursuant to articles 31 and 33 of legislative decree No 22 of 5 February 1997”. In: Ministero dell’ambiente e della Tutela del territorio (ed).

28.

MEDDTL (2011) Arrêté relatif au recyclage en technique routière des mâchefers d’incinération de déchets non dangereu. (English: Decree on the recycling of non-hazardous waste incineration bottom ash in road construction). In: Ministère de l’écologie, du développement durable, des transports et du logement (ed).

29.

Mehr

Haupt

Skutan

, et al. (2020) The environmental performance of enhanced metal recovery from dry municipal solid waste incineration bottom ash. Waste Management 119: 330–341.

30.

Morf

Brunner

Spaun

(2000) Effect of operating conditions and input variations on the partitioning of metals in a municipal solid waste incinerator. Waste Management & Research 18: 4–15.

31.

NEN 7375 (2004) Leaching characteristics – determination of the leaching of inorganic components from moulded or monolithic materials with a diffusion test – solid earthy and stony materials.

32.

NEN 7383 (2004) Leaching characteristics – determination of the cumulative leaching of inorganic components from granular materials with a simplified procedure of the column test – solid earthy and stony materials.

33.

Powell

Brown

Byrne

, et al. (2007) Chemical speciation of environmentally significant metals with inorganic ligands part 2: the Cu2+-OH-, Cl-, CO32-, SO42-, and PO43- systems (IUPAC Technical Report). Pure and Applied Chemistry 79: 895–950.

34.

Powell

Brown

Byrne

, et al. (2009) Chemical speciation of environmentally significant metals with inorganic ligands. Part 3: the Pb²⁺ –OH^–, Cl^–, CO₃^2–, SO₄^2–, and PO₄³–systems (IUPAC technical report). Pure and Applied Chemistry 81: 2425–2476.

35.

Republic of Austria (2017) Technische Grundlagen für den Einsatz von Abfällen als Ersatzrohstoffe in Anlagen zur Zementerzeugung (English: Technical basics for the use of waste as secondary raw material in cement production). In: Federal Ministry of Agriculture, Environment and Water Management and Austria (ed).

36.

Simon

Schmidt

Carcer

(1995) Alterungsverhalten von MVA-Schlacken (English: Ageing behaviour of MSWI bottom ash). Müll und Abfall 11: 759–764.

37.

Soil Quality Decree (2013) Regeling bodemkwaliteit, VROM, Den Haag: Ruimte en milieu (English: Soil quality regulations, VROM, Den Haag: Space and environment). In: Ministerie van Volkshuisvesting, Ruimtelijke Ordeling en Milieubeheer (ed).

38.

Swiss Confederation (2016) Ordinance on the avoidance and the disposal of waste. 1–46.

39.

van Zomeren

Comans

RNJ

(2004) Contribution of natural organic matter to copper leaching from municipal solid waste incinerator bottom ash. Environmental Science & Technology 38: 3927–3932.

40.

Verbinnen

Billen

Van Caneghem

, et al. (2017) Recycling of MSWI bottom ash: A review of chemical barriers, engineering applications and treatment technologies. Waste and Biomass Valorization 8: 1453–1466.

41.

Wan

Wang

, et al. (2006) A study on the chemical and mineralogical characterization of MSWI fly ash using a sequential extraction procedure. Journal of Hazardous Materials 134: 197–201.

42.

Weibel

Eggenberger

Schlumberger

, et al. (2017) Chemical associations and mobilization of heavy metals in fly ash from municipal solid waste incineration. Waste Management 62: 147–159.