Enhancing Utilization of Municipal Solid Waste Bottom Ash by the Stabilization of Heavy Metals

Abstract

Waste-to-energy (WtE) is a key part of modern waste management. In the European Union, approximately 500 WtE plants process more than 100 million tons of waste yearly, while globally, more than 2700 plants handle over 500 million tons. Roughly 20% of the waste processed is bottom ash (BA). However, this ash can contain heavy metals in concentrations that may render it hazardous. This paper presents a study focusing on stabilizing municipal solid waste incineration BA using simple and industrially viable treatments. The Slovenian WtE plant operator wishes to install the stabilization process; thus, the samples obtained from the plant were treated (1) with a CO₂ gas flow, (2) with water spraying, and (3) with a combination of water spraying and a CO₂ gas flow under laboratory conditions. Thermodynamic calculations were applied to define potential reactions during the treatment processes in the temperature range from 0 to 100 °C and to define the equilibrium composition of the treated ash with additions of CO₂ and water. The standard leaching test EN 12457-4 of treated ash shows a reduction of over 40% in barium concentration and over 30% in lead concentration in leachates.

1. Introduction

The generation of municipal solid waste (MSW) is projected to reach 3.4 billion tons annually by 2025 due to the increasing population and urbanization [1]. For approximately 90% of the global population, the primary process of waste management is still landfilling [2]. In the waste management hierarchy, energy recovery through waste-to-energy (WtE) incineration is more favorable than landfilling as it reduces approximately 80 percent of waste by weight and 90 percent by volume [3,4] and utilizes energy.

Global WtE infrastructure has seen significant growth, with over 2700 facilities operating worldwide as of the beginning of 2023, capable of processing approximately 530 million tons of waste annually, and projections indicating a continued expansion to approximately 3000 operational WtE facilities by 2032, anticipating a waste-processing capacity of more than 700 million tons per year [5,6].

However, the incineration of waste also produces air pollution and bottom ash (BA). The constitutes of BA make up 80 wt.% of total MSW incineration residues, with a ratio of mass to fly ash (FA) of 4:1 to 5:1. Both ash types typically contain CaO, SiO₂, compounds of Al, Fe, Mg, Na, K, and Cl, as well as potentially toxic elements like Cr, Ni, Cu, Zn, Cd, Hg, and Pb [7]. As the most significant portion of solid residue, BA has the potential to be repurposed as a secondary building material [8,9], but bearing in mind the fact that BA usually contains various heavy metals, the utilization of these residues can be challenging [8].

If not managed effectively, improper ash disposal, whether through landfilling, co-disposal, or ocean dumping, can lead to environmental contamination [10,11]. The challenge lies in ash’s full integration into circular economy practices, including metal recovery, material valorization, and reducing reliance on landfilling [12,13]. Addressing these challenges through proper treatment, regulation, and innovative utilization methods can unlock the potential benefits of ash from WtE plants. This approach is crucial for minimizing environmental impacts and maximizing its value in various applications.

According to the European Union (EU) Commission’s decision [14], waste materials in the EU are designated with a six-digit waste number. The waste material studied was “19 01 12 bottom ash and slag other than those mentioned in 19 01 11”. The “19 01 11* bottom ash and slag containing hazardous substances” is hazardous waste from the same source, identified as “wastes from incineration or pyrolysis of waste”. Standard procedures according to EU waste [15] and landfill [16] directives are performed to determine the hazardous properties of ash. Also, local legislation sets limits, according to Annex II of the landfill directive [16], like the Slovenian Regulation on landfills, regarding what waste is considered hazardous and non-hazardous.

To promote the reuse of ash as mineral material, it must be utilized as inert as possible. Some physical and/or chemical mechanisms can be implemented to reduce the high concentration of contaminants in leachate. While Van Gerven T. et al. (2005) comprehensively investigated accelerated carbonization to reduce heavy metal leaching from municipal solid waste incineration (MSWI) bottom ash, monitoring the effects of different carbonization parameters on metal leaching behavior [17], Schnabel K. et al. (2021) presented the first successful implementation of accelerated carbonization in a full-scale rotating drum reactor using real exhaust gas, demonstrating effective control of reactor parameters to achieve sufficient carbonization for regulatory compliance and potential reuse of bottom ash [18].

The reaction can be accelerated by adding water, in which case CO₂ dissolves in water, dissociates, and reacts with dissolved and dissociated calcium hydroxide, forming carbonate materials [19]. Tang et al. (2015) found that calcium oxide (CaO) was one of the dominant chemical components in MSW incineration bottom ash, making up 13–14% of the composition [20]. They showed that calcium compounds in the bottom ash contributed to its alkalinity and buffering capacity [20]. Xia et al. studied the leaching efficiency of calcium compounds and found that it varies significantly with pH, time, and temperature [21]. In contrast, Jian-Zhi Wang et al. investigated the impact of recovered calcium from ash on volume reduction, emphasizing its crucial role in enhancing the ash treatment process [22].

Given that the schemes for successful carbonization of ash have already been described and defined, why is this not an established regular procedure? The authors of this paper sought an efficient and straightforward approach to carbonizing and stabilizing BA with maximum use of its physicochemical characteristics, i.e., hydration of metal oxides and carbonization induced by CO₂ absorption [23].

This study aims to stabilize the BA from the incineration of MSW through simple and industrially feasible treatments to reduce its potential environmental hazards. This research is explicitly focused on the WtE plant due to the desire to stabilize BA and find suitable methods and process conditions. The BA under investigation is somehow similar to other MSW BA but not the same since the properties of input waste, WtE technology, and operation parameters play an important role in BA characteristics. The MSW waste prior to input is mechanically and biologically preprocessed; waste combustion takes place on an air-cooled grate, and BA is extracted from the furnace without water quenching. Thus, this research included (1) carbonization of ash by carbon dioxide (CO₂) flow, (2) hydration of ash by spraying with water, and (3) a combination of treatments, i.e., carbonization of ash with CO₂ flow after hydration with water. The effects of the processes of carbonization, hydration, and combined treatment on the stability of BA were investigated using the EN 12457-4 standard leaching test [24]. The mechanism of applied treatments is explained by considering potential reactions and defining the equilibrium composition of the treated ash using the HSC Chemistry program for thermodynamic calculations.

Defining the simplest sustainable practices allows for a more straightforward implementation and a better understanding of the positive environmental impacts. These practices bring significant benefits: CO₂ sequestration reduces greenhouse gas emissions, while accelerated carbonization stabilizes heavy metal migration activity and reduces the risk of environmental contamination.

2. Methods

2.1. Ash Sampling

Ash samples were taken according to EN 14899:2006 [25]. The samples were collected from both the upper openings of the container and all accessible areas within it. The sampling process occurred over one working day, coinciding with ash production and its transport to the container. As the container was filled from the top, newly deposited ash was readily accessible for sampling. Figure 1 shows ash in the original container and as the sample taken.

A representative sample of bottom ash intended for treatment is a cooled sample of fresh ash from the furnace grate. One sample was stored and aged under anaerobic conditions. Extracts of this ash were analyzed on days 1, 15, and 30 from the beginning of storage. Such storage and analysis should show ash stability without ambient influence.

Laboratory experiments simulating the stabilization and aging of ash were conducted using a baseline (so-called “zero”) sample, which was first sieved and cleaned of larger pieces. A fraction of <2 mm was used.

The initial sample for comparison with treated (CO₂, H₂O, and CO₂/H₂O) ash samples and the physical and chemical parameters of analyses of associated ash leachate were compared to those of an untreated aerobically stored ash sample (start sample) and the corresponding leachate.

A representative ash sample is a powder mixture with some larger solid particles. It has a characteristic smell of ash and burnt material, which becomes more intense when water is added. The smell of sulfur compounds is not detected. A large number of metal particles is observed in contact with a magnetic stirrer.

2.2. Ash Treatment Experiments (CO₂, H₂O and CO₂/H₂O Combination)

The design and implementation of experiments for stabilizing or carbonatizing aerobically stored ash aimed at reducing the proportion of leached metals are based on a previously conducted study and research findings [26,27,28,29].

In the laboratory reactor (V = 3 L), three different carbonization processes of a representative ash sample were carried out:

• A—industrial carbon dioxide (CO₂; added 100% ash weight);
• B—water shower (H₂O, addition of 10% ash weight);
• C—a combination of a water shower (10% ash weight) and gas (CO₂).

The ash sample was exposed to CO₂ in a flow-through glass reactor for 2 min at a constant gas flow rate of 3 L/min, following procedures A and C. Prior to gassing, as outlined in procedures B and C, the sample was treated with a water shower, with the volume of water applied being 10% of the weight of the ash. This was followed by mixing on a magnetic stirrer for 5 min (procedures A, B, and C).

Figure 2 shows the preparation of the experiment, adding water before gassing (procedure C) and gassing with carbon dioxide.

The ash samples treated this way were used to prepare leachate at different time intervals (immediately after treatment and after two and five days of aging). Ash was kept in parafilm-covered glass beakers during the aging period at room temperature.

2.3. Thermodynamic Calculations

The HSC Chemistry program package (version 9.9.2.3) was used to determine and analyze the mechanism of applied ash treatments based on thermodynamic parameters.

The thermodynamic calculations are fortified by an extensive integrated database that covers enthalpy, entropy, and heat capacity data for a vast array of over 28,000 chemical compounds, underscoring the robustness of this research.

The Reaction Equations module is used to determine the change in the Gibbs free energy (ΔG) of potential chemical reactions of macroconstituents of ash (CaO, MgO, Fe₂O₃) and impurities (PbO and BaO) with CO₂ and H₂O in the temperature range from 0 to 100 °C. The application of the Equilibrium Compositions module enabled the examination of the influence of the addition of CO₂ from 0 to 1.0 kg and the addition of water from 0 to 0.1 kg, as well as the determination of the changes in the phase composition of the initial material (ash) during the applied treatments with CO₂, H₂O, and H₂O/CO₂. The input data for the CO₂ and H₂O treatment calculation were the phase composition of the ash, presented in

Table 1. The content of macro- and microconstituents in the ash.

Macroconstituents	Content (wt.%)
SiO2	37.98
CaO	33.69
Al2O3	14.39
MgO	3.60
Fe2O3	2.80
Microconstituents	Content (g/kg)
PbO	1.00
BaO	1.00

, and process parameters (amounts of CO₂ and water) used in the ash treatments: 2.3:1.0 kg of ash sample, 1.0 kg of CO₂ and 0.1 kg of water. In contrast, the input data for the H₂O/CO₂ treatment were the output data obtained from the H₂O treatment calculation.

2.4. Preparation of Ash Sample Leachates

The leachates were prepared following standard EN 12457-4:2002 [24] immediately after treatment and after 2 and 5 days of aging.

The L/S ratio of 10 L/kg and the leaching time of 24 h were performed with constant stirring. This was followed by centrifugation and filtration of the upper phase through 1.2 μm, 1 μm, and 0.7 μm filters.

2.5. Analyses of Physical–Chemical Parameters of Leachate Quality

Analyses of the physical–chemical parameters were carried out using standardized methods, some using cuvette tests and some using atomic absorption spectroscopy (AAS), in three parallels. The set of analyzed parameters and the analytical methods used are given in

Table 2. Analyzed parameters and methods used.

Parameter	Method
Temperature (°C)	DIN 38404-4 [30]
pH value after 24 h/(leachate)	ISO 10523 [31]
Conductivity (mS/cm) after 24 h/(leachate)	EN 27888 [32]
Dry solids share (mg/L)	EN 14346 [33] and EN 15169 [34]
Acid capacity Ks 4.3 (mmol/L)	HL LCK 362 (0.5–8.0 mmol/L)
Barium (mg/L Ba)	HL PP 8014 (2–100 mg/L Ba)
Lead (mg/L Pb)	AAS
Chrom total (mg/L Cr)	AAS
Chloride (mg Cl/L)	HL LCK 311 (70–1000 mg/L Cl)
Sulfate (mg SO42−/L)	MN 086 (10–200 mg/L SO42−)

.

3. Results

3.1. Thermodynamic Calculation Results

3.1.1. Treatment with CO₂

To understand the CO₂ treatment, it is necessary to analyze the reactions of macro- (CaO, MgO, and Fe₂O₃) and microconstituents (lead, barium, and chlorides) in the ash sample (Equations (1)–(6)). Figure 3 shows the temperature dependencies of ΔG for these reactions.

CaO + CO₂(g) = CaCO₃

(1)

MgO + CO₂(g) = MgCO₃

(2)

Fe₂O₃ + 3CO₂(g) = Fe₂(CO₃)₃

(3)

BaO + CO₂(g) = BaCO₃

(4)

PbO + CO₂(g) = PbCO₃

(5)

2MeCl₂ + 2CO₂(g) + O₂(g) = 2MeCO₃ + 2Cl₂(g)

(6)

Note: Me—designation for metal.

Calcium oxide, magnesium oxide, barium oxide, and lead oxide reactions are possible (ΔG—change in Gibbs energy is <0) at a defined temperature range. In contrast, the reactions of iron oxide and chlorides (MeCl₂) are impossible. This means that treatment with CO₂ gas flow will not influence the stabilization of chlorides in the ash sample.

According to the chart in Figure 3, variations in the ΔG values show that the dominant reaction is the formation of barium carbonate from barium oxide (Reaction (4)), with the most negative ΔG value of −221.5 kJ. This is followed by the reactions of calcium, magnesium, and lead oxides with CO₂, with ΔG of −131.224, −49.409, and −43.132 kJ, respectively.

Figure 4 and Figure 5 present changes in the equilibrium composition of BA with CO₂ addition. As seen in Figure 4, introducing gaseous CO₂ up to 0.3 kg into the reactor leads to the formation of Ca-aluminosilico-carbonate. Over this amount, this transitional compound disintegrates. As a result of its decomposition, an evident increase in calcium carbonate, silicates in the form of SiO₂, and aluminum oxide is observed. Magnesium carbonate starts to form when the amount of CO₂ gas in the system is greater than 0.2 kg, while the Fe₂O₃ content in the system is constant during the treatment.

As shown by the thermodynamic modeling of microconstituents in Figure 5, CO₂ gas will first react with barium oxide at the beginning of the experiment. In contrast, lead oxide reacts later after introducing amounts of CO₂ greater than 0.2 kg.

3.1.2. Treatment with H₂O

The following reactions were analyzed to determine the influence of H₂O treatment on the ash sample’s macro- and microconstituents (Equations (7)–(11)). The ΔG temperature dependencies for these reactions are shown in Figure 6.

CaO + H₂O = Ca(OH)₂

(7)

MgO + H₂O = Mg(OH)₂

(8)

Fe₂O₃ + 3H₂O = 2Fe(OH)₃

(9)

BaO + H₂O = Ba(OH)₂

(10)

PbO + H₂O = Pb(OH)₂

(11)

The results of ΔG temperature dependencies for the ash treatment with H₂O are shown within the chart in Figure 6. At a defined temperature range, reactions for Ca, Mg, and Ba oxide are possible (ΔG—change in Gibbs energy is <0), while iron oxide and lead oxide reactions, with negative values of ΔG, are not possible.

Gibbs energy change values of −221.584 kJ show that the dominant reaction is the oxidation of barium (Equations (10)), with the most negative Gibbs energy change value and the highest energy release during exothermic reactions, followed by calcium with ΔG of −57.923 kJ and ΔG of magnesium oxide of −27.335 kJ.

Figure 7 and Figure 8 present changes in the equilibrium composition of BA with water addition. Figure 7 shows that introducing water into the system at a value of 0.02 kg decreases the concentration of silicate, aluminum oxide, and calcium oxide due to the formation of the hydrated calcium–aluminum–silicate.

The amount of MgO and Fa₂O₃ remains unchanged during the treatment. Magnesium oxide does not form hydroxide due to the insufficient amount of water in the system, although this reaction is thermodynamically possible. The hydration reaction of Fa₂O₃ is impossible under these thermodynamic conditions.

Figure 8 shows the equilibrium composition of microconstituents in the ash sample during the treatment with water. Barium oxide in the presence of water immediately forms barium hydroxide, while lead oxide does not interact with water.

With the highest amount of water addition (0.10 kg), the hydrated structure (CaAl₂Si₄O₁₂ · 2H₂O) amounts to 0.31 kg of equilibrium composition, which is about 30% of the sample’s mass. The formation of such a calcium–aluminum–silicate hydrate structure could lead to the encapsulation of the contaminants and their stabilization [35].

3.1.3. Treatment with H₂O and CO₂

During the ash treatment with water, its mineral composition changes. The changed composition was simulated, and the following values were obtained, as shown in

Table 3. Composition of ash after treatment with water.

Compound	Content [kg]
SiO2	0.2078
CaO	0.0754
CaAl2Si4O12 · 2H2O	0.3109
Ca(OH)2	0.2925
Al2O3	0.0709
MgO	0.0360
Fe2O3	0.0280
TiO2	0.0120
NaOH	0.0116
CuO	0.0090
KOH	0.0024
Ba(OH)2	0.0011
PbO	0.0010
ZnO	0.0010
Ca3(Al2Si2O8)3 · CaCO3	0.0004

.

Figure 9 and Figure 10 present changes in the equilibrium composition of BA with CO₂ addition after the hydration process. As shown in Figure 9, adding CO₂ to the system transforms the present calcium oxides and hydroxides into calcium carbonate and increases the amount of hydrated calcium–aluminum–silicate. The formed hydrated structure amounts to about 60% of the total ash mass.

Introducing water into the system, followed by CO₂, leads to the formation of Pb and Ba carbonates as more stable forms of metals from their hydroxides [36], as shown in Figure 10.

The combination of H₂O and CO₂ proved to be the most suitable treatment method, where, in addition to chemical stabilization, metals are incorporated into the hydrated calcium–aluminum–silicate structure or carbonized.

3.2. Experimental Results

Immediately after sampling the ash from the ash container at the WtE site, a certain amount of the sample was stored in a glass container, and kept under anaerobic conditions throughout the experiments. This sample was the reference sample on which analyses of physical–chemical parameters without aging were performed, followed by control analyses after 15 and 30 days of aging (

Table 4. Chemical and physical data and pollutant content in leachates of untreated anaerobically stored ash at different ages—reference sample.

Storage Conditions	Anaerobic
Ash Stabilization	None
Ash Aging	Start Sample	15 Days	30 Days
Date of Analysis	26 November	3 December	18 December
Temperature (°C)	21.3	22.3	21.2
pH value after 24 h/(leachate)	12.22	12.17	12.28
Conductivity (mS/cm) after 24 h/(leachate)	10.29	9.8	10.07
Dry solids share (mg/L)	99.686	99.686	99.612
Acid capacity Ks 4.3 (mmol/L)	35.65	21.7	27.45
Barium (mg/L Ba)	29.00	27.5	24.5
Lead (mg/L Pb)	0.266	0.374	0.309
Chrom total (mg/L Cr)	0.104	0.097	0.118
Chloride (mg Cl/L)	1775.0	1525.0	1227.5
Sulfate (mg SO42−/L)	<10	<10	<10
Barium (mg/kg d.s.)	290.91	275.87	245.95
Lead (mg/kg d.s.)	2.668	3.752	3.102
Chrom total (mg/kg d.s.)	1.043	0.973	1.185
Chloride (mg/kg d.s.)	17,806	15,298	12,323
Sulfate (mg/kg d.s.)	<100	<100	<100

).

The results of the analyses compiled in Table 4 show that, as expected, the properties of ash stored under anaerobic conditions do not change significantly regarding metal and salt leaching. The physical–chemical parameters change slightly (a slight drop in pH and acid capacity due to the release of carbon dioxide and the conversions of carbonate equilibrium components).

Table 5. Chemical and physical data and pollutant content in leachates of untreated aerobically stored ash before and after treatment at different ages.

Storage Conditions	Aerobic
Ash Stabilization	None	CO2			H2O			CO2/H2O
Ash Aging	0 Days	0 Days	2 Days	5 Days	0 Days	2 Days	5 Days	0 Days	2 Days	5 Days
Date of Analysis	25 Nov.	25 Nov.	27 Nov.	30 Nov	25 Nov.	27 Nov.	30 Nov.	25 Nov.	27 Nov.	30 Nov.
Parameter
Temperature (°C)	21.1	22.3	20.4	19.4	22.0	21.7	20.4	22.2	21.0	20.1
pH value after 24 h/(leachate)	11.43	11.17	11.76	11.76	12.22	12.02	12.177	12.02	11.17	11.16
Conductivity (mS/cm) after 24 h/(leachate)	5.81	5.33	5.90	5.37	9.30	7.87	8.64	7.63	5.13	5.03
Dry solids share (mg/L)	99.789	99.605	99.605	99.605	94.012	94.012	94.012	94.128	94.128	94.128
Acid capacity Ks 4.3 (mmol/L)	18.80	19.90	19.00	18.75	27.17	21.35	17.20	13.15	21.65	17.60
Barium (mg/L Ba)	26.00	26.50	25.00	27.500	29.00	20.00	20.60	28.50	14.50	14.750
Lead (mg/L Pb)	0.797	0.795	0.419	0.490	0.553	0.753	0.593	0.452	0.457	0.493
Chrom total (mg/L Cr)	<0.001	<0.001	<0.001	<0.001	<0.001	<0.001	<0.001	0.0016	<0.001	<0.001
Chloride (mg Cl/L)	1270.0	1395.0	1320.0	1215.000	920.0	1262.5	1270.0	1447.5	1421.0	1240.000
Sulfate (mg SO42−/L)	<10	<10	<10	<10	<10	<10	<10	<10	<10	<10

shows that aging for up to 5 days in aerobic conditions and treatment with water and CO₂ have various effects on metal and salt leaching. Also, some physical–chemical parameters did change. The greatest changes can be observed in conductivity and acid capacity. The changes are due to the carbonatation process.

The research results in

Table 6. The content of pollutants in ash before and after treatment at different ages according to the mass of dry ash matter depending on storage.

Storage Conditions		Anaerobic	Aerobic
Ash Stabilization		None	None	CO2			H2O			CO2/H2O
Ash Aging		0 Days	0 Days	0 Days	2 Days	5 Days	0 Days	2 Days	5 Days	0 Days	2 Days	5 Days
Date of analysis		25 Nov.	25 Nov.	25 Nov.	27 Nov.	30 Nov.	25 Nov.	27 Nov.	30 Nov.	25 Nov.	27 Nov.	30 Nov.
Parameter	ELV *	Reference Sample	Start Sample
1 Barium (mg/kg d.s.)	100	210.91	188.90	192.89	181.97	200.17	223.64	154.24	158.86	219.51	111.68	113.61
2 Barium		1.12	1.00	1.02	0.96	1.06	1.18	0.82	0.84	1.16	0.59	0.60
3 Barium		1.00	0.90	0.91	0.86	0.95	1.06	0.73	0.75	1.04	0.53	0.54
1 Lead (mg/kg d.s.)	10	5.75	5.78	3.05	3.57	4.03	5.81	4.57	3.49	3.52	3.80	4.68
2 Lead		0.99	1.00	0.53	0.62	0.70	1.00	0.79	0.60	0.61	0.66	0.81
3 Lead		1.00	1.00	0.53	0.62	0.70	1.01	0.80	0.61	0.61	0.66	0.81
1 Chrome (mg/kg d.s.)	10	0.160	0.007	0.007	0.007	0.007	0.008	0.008	0.008	0.012	0.008	0.008
2 Chrome		22.90	1.00	1.04	1.04	1.04	1.10	1.10	1.10	1.76	1.10	1.10
3 Chrome		1.000	0.045	0.045	0.045	0.045	0.048	0.048	0.048	0.077	0.048	0.048
1 Chloride (mg/kg d.s.)	15,000	12,909	9227	10,154	9608	8844	7095	9736	9794	11,149	10,945	9551
2 Chloride		1.40	1.00	1.10	1.04	0.96	0.77	1.06	1.06	1.21	1.19	1.04
3 Chloride		1.00	0.71	0.79	0.74	0.69	0.55	0.75	0.76	0.86	0.85	0.74
1 Sulfate (mg/kg d.s.)	20,000	72.73	72.65	72.79	72.79	72.79	77.12	77.12	77.12	77.02	77.02	77.02
2 Sulfate		1.001	1.000	1.002	1.002	1.002	1.061	1.061	1.061	1.060	1.060	1.060
3 Sulfate		1.00	1.00	1.00	1.00	1.00	1.06	1.06	1.06	1.06	1.06	1.06

Notes: * Regulation on landfills of Pb and Ba [37]; ¹ including the mass of bigger fractions 27.5%; ² fraction of the start sample; ³ fraction of reference sample.

show that the stabilization method holds great potential in reducing the barium content of ash leachate. By adding 10% water to the mass of ash and gassing with pure carbon dioxide (2 min) after two days of aging, a reduction in the barium content in the leachate by >40% was achieved compared to the initial untreated sample and by >47% relative to the reference sample. The lead content in the leachate was also reduced by >30%. Notably, a minimal influence was observed on sulfate, chloride, and chrome leachate concentrations, further highlighting the promising potential of the stabilization method.

Figure 11 shows the relative results of experimental ash stabilization based on the start sample. Stabilization is successful for barium and lead. Treatment with CO₂ in H₂O does not influence chrome, chloride, or sulfate.

4. Discussion

The thermodynamic calculations conducted were beneficial in determining the reactions and their products based on the fresh ash composition and stabilization method utilized. The calculations predicted reactions that would occur at expected ambient temperatures during treatments with water, carbon dioxide, or both for macro- and microconstituents.

Based on thermodynamic calculation results, Pb, Mg, Ca, and Ba oxide reactions with CO₂ were expected. In the case of addition of only water, the reactions of Mg, Ca, and Ba oxides were expected. When water and CO₂ are applied, Pb and Ba oxides form carbonates. Also, a hydrated calcium–aluminum–silicate structure was formed, representing the most significant fraction of stabilization products, followed by calcium carbonate.

According to experimental data, the highest reduction percentage for barium content in leachate occurs after treatment with water and CO₂ for 2 days. This can be attributed to the formation of stable barium carbonate and hydrated calcium–aluminum–silicates. These compounds effectively bind heavy metals within their structures, limiting their solubility and mobility in aqueous environments. In comparison, individual treatments with CO₂ or H₂O yielded moderate reductions, indicating that the synergistic effect of combining carbonatation and hydration plays a crucial role in enhancing stabilization outcomes.

The stabilization processes also induced notable chemical transformations in the ash. During carbonatation, CO₂ reacted with oxides in the ash, forming carbonates. This reaction reduced the pH of the leachate and neutralized the ash’s alkalinity over time. Similarly, hydration triggered the formation of hydroxides and hydrated the calcium–aluminum–silicate structures, further contributing to the immobilization of metals. The combined treatment amplified these effects by simultaneously promoting carbonatation and hydration reactions, resulting in a more stable mineral matrix.

The experimental results proved the absence of interaction between PbO and water. The highest level of lead leaching was obtained in the presence of water, where it was not bound into lead hydroxide. On the other hand, the experimental results showed a reduced concentration of barium in leachates due to the formation of barium hydroxide and a hydrated structure that could physically encapsulate mobile Ba compounds.

Ash stabilization simulations and experimental results show that using water and CO₂ transforms some metal oxides and calcium into carbonates and forms a hydrated calcium–aluminum–silicate structure. These changes make metals more immobile and bound to solid structures, thus making leachate less polluting. The stabilization process increases the mass and volume of the samples.

Thermodynamic calculations showed that adding 0.3 kg of CO₂ (30% of the amount of ash) is sufficient to form the equilibrium composition of the treated ash in the case of treatment with CO₂ and combined treatment with water and CO₂. In the treatments with water, the entire amount of added water (0.1 kg) was consumed to form the equilibrium composition.

As presented in the experimental results and confirmed by thermodynamic calculations, treating ash with CO₂ gas flow and water spraying did not affect stabilizing chlorides and their leaching from the ash samples.

In addition to the stabilization of heavy metals, the treatments led to a change in the physical and chemical characteristics of the ash itself. Namely, the pH values slightly decreased during the sample’s aging, indicating a decrease in alkalinity due to the formation of carbonate compounds. Besides the change in pH value, the conductivity value also changed, which is one of the indicators of the change in the concentration of dissolved ions, probably sulfates and chlorides. The decrease in ion concentration significantly emphasizes the effectiveness of the combined treatment of water and CO₂ for ash stabilization. The results follow the previous research and literature review and build on them. As pointed out in the paper by Tang et al. [20], calcium oxide plays a key role in the buffering capacity of ash, enabling pH stabilization. In addition, it promotes the formation of carbonates and hydroxides through the process of hydration of the sample, and in that way, significantly contributes to stabilizing the metal in the material. These processes are essential to reduce the mobility of metals and bind them into less reactive forms, thereby reducing the potential for environmental contamination. Interestingly, within the framework of the combined treatment, the results are significantly better than in the current practice. Optimization of parameters such as CO₂ flow rate, water content, and aging duration can significantly improve the stabilization process for carbonation and hydration.

5. Conclusions

The results of this work indicate a potential solution in waste-to-energy (WtE) operations; that is, they provide a potential solution for stabilization and ash management. Effective stabilization of heavy metals contributes to reducing the negative impact on the environment. It creates opportunities to use this waste raw material as a by-product for another branch of industry. Especially of note, the results establish explicit stabilization process parameters for the bottom ash properties at the Slovenian WtE plant.

Upon reviewing the literature, it is evident that the concentrations of macro- and microconstituents in the ash from the grate at the Slovenian WtE plant align with global findings. This underscores the universal nature of our research and its relevance to the global context. It also provides an insight to the somewhat different MSW input, combustion process, and ash extraction.

This paper presents various options for ash stabilization. Three procedures were selected for treating bottom ash from the MSW WtE plant and tested with actual samples under laboratory conditions. The aim was to hydrate and carbonatite ash and increase its carbonate content, as carbonates have a high affinity for divalent metals like barium, which is occasionally detected in ash leachates. The goal was to stabilize ash for possible further use.

Thermodynamic calculations predicted the possible reactions during the treatments and the equilibrium composition of treated ash, whose effects were confirmed by experiments.

The research results indicate that stabilizing freshwater-unquenched ash with hydration and carbonatation effectively reduces barium and lead content in leachate. The treatment involving 10% water addition and carbon dioxide exposure with 5 days of aging led to a reduction of over 40% in barium content and a 30% reduction in lead content compared to the untreated samples.

What is also interesting is the consumption of CO₂ in the treatment process itself. Its usage integrates CO₂ into stable mineral forms, aligning with global efforts to reduce carbon dioxide emissions. Nevertheless, despite these encouraging results, numerous challenges persist. It is necessary to carry out experimental tests in industrial environments and to evaluate the results due to changes in the composition of the ash, and the amount, i.e., the ratio of reactants, as well as the limitations of the technology and the value of operating costs. The industrial stabilization process at the Slovenian WtE plant is intended to utilize flue gases after flue gas treatment. It contains up to 10 volume % of CO₂ and generally from 10 to 20 volume % of water vapor. Also, the sediment obtained after this treatment needs to be subjected to further tests, exposure to atmospheric conditions, and temperature changes in short-term and long-term conditions. These factors are crucial to ensure the reliability and safety of fly ash reuse in real-world applications. Future research should aim to test these methods in different types of ash and waste streams to assess their generalizability and effectiveness on a larger scale. It would also be necessary for future studies to include the impact of stabilization on additional components, other heavy metals, or organic compounds that are considered environmental pollutants.

Reducing costs involves lowering landfill expenses and investing in safer alternatives. Stabilizing ash and immobilizing hazardous components allows for reuse and shifts from primary to secondary metal extraction. Utilizing stabilized ash in briquette production and cement reduces environmental impacts and costs for new raw materials. Technological advances enable the recovery of metals from materials with lower concentrations, making this process economically viable. Despite low metal concentrations in ash, the significant volume in disposed tailings warrants further recovery analysis. The results of this work indicate the possibility of ash stabilization using water and carbon dioxide, thus achieving a reduced negative impact of ash on the environment with reduced CO₂ emissions into the atmosphere. With this treatment, the ash becomes a by-product, which promotes the sustainable use of the material but also supports the concept of reducing the rate of CO₂ impact on climate change. Both economic and environmental benefits are achieved by using one type of waste, such as CO₂, to treat another type of waste, such as ash.