Simultaneous Removal of Chlorides and Calcium from EAF Dust Wastewater

Abstract

This research investigates the sorption efficiencies of various adsorbents—synthesized Hydrotalcite, natural zeolite Clinoptilolite, synthetic zeolite, and waste sludge from aluminum anodic oxidation—for simultaneous removal of Cl⁻ and Ca²⁺ ions from synthetic CaCl₂ solutions and wastewater from EAFD recycling. This study addresses the challenges of wastewater purification options, which were not previously addressed in other studies. The high alkalinity and ionic pollutants in EAFD wastewater make the purification process complex. The fact that adsorbents tested in this study were prepared from metallurgical waste predetermines the process to be more sustainable. Adsorbents were thoroughly characterized before and after calcination and sorption using techniques like AAS, LIBS, XRD, BET, BJH, SEM-EDS, and FTIR spectroscopy. Synthetic zeolite achieved near-complete removal of Ca²⁺ ions, while calcined Hydrotalcite at 500 °C excelled in the simultaneous removal of Cl⁻ and Ca²⁺. Equilibrium sorption capacities of HT were 50.3 mg/g for Cl⁻ and 37 mg/g for Ca²⁺ after 360 min, with efficiencies reaching 85% for Ca²⁺ and 83% for Cl⁻. Additionally, HT effectively removed 82% Pb, 91% Cr, and 40% SO₄²⁻ in 24 h of the sorption process. These findings highlight HT as a promising solution for industrial wastewater treatment, offering sustainable and efficient pollutant removal.

1. Introduction

Globally, more than 1.5 Mt of crude steel are produced yearly [1]. Steel production is closely related to generating of significant amounts of electric arc furnace dust (EAFD). Steel mills are facing the demand of decarbonisation, and together with this activity, there is the necessity to reduce the carbon footprint by decreasing generated amounts of waste or recycling and closing the processing loop appropriately. This should be applied also to EAFD. Pyrometallurgical processes can recycle EAFD to recover metals or use hydrometallurgical processes. Hydrometallurgical methods involve leaching in water, acids, bases, or salts; in this case, it is also eligible to consider the regeneration/recycling of this wastewater. Steel plants or waste recycling companies would require improved, highly effective, and fast water treatment technology. Legacy treatment modes cannot thoroughly remediate water for reclamation, but new techniques show great promise [2]. One major problem that steel and recycling companies face is reducing salts from process water and increasing the reusing of process water, which reduces freshwater and wastewater discharge costs [3]. The problematic issues are organic matter, calcium, chlorides, and sulphates, except for heavy metals. For instance, calcium causes severe scaling in the pipes, chlorides cause metal framework corrosion, and sulphates cause problems with fouling, which increase the operation cost when membrane technologies are applied.

The desalination process is associated with a high demand for energy to crystallize the solid salt from wastewater and brines. Furthermore, salts of poor quality must be landfilled at a very high cost as a by-product of desalination. If the source of water is freshwater, it is becoming more common to introduce “zero discharge” policies for industrial plants. The purpose is to reduce freshwater usage by encouraging water recirculation [4]. Before hydrometallurgical treatment, the EAFD is usually subjected to water washing due to decreased acid/base consumption and contamination by various dissolved ionic substances in the next leaching step. This wastewater often contains chlorides, sulphates, carbonates, heavy metals, calcium, silicon, phosphorus, organic matter, etc.

Multiple methods exist for water desalination, but some are very expensive, and others require high energy. Membrane technologies (reverse osmosis, electrodialysis, microfiltration, nanofiltration, etc.) can be particularly prone to scale formation and fouling, and regular membrane cleaning is required [5,6,7,8,9,10,11].

In addition, water pretreatment, such as coagulation, sedimentation, and sand filtration, is sometimes needed [3,12,13,14,15,16,17].

Another option for water desalination is ion exchange. Various ion exchange resins can remove chlorides from synthetic solutions [18,19,20,21].

Ion exchange is considered one of the best and most effective methods for removing anions from water due to its high efficiency, simplicity, and low cost [22,23].

Also, cation exchange resins were successfully applied to remove calcium and magnesium [24]. The effective removal of chlorides and sulphates was achieved by adsorption with solid waste, namely, iron sand and pulp-and-paper waste from alkaline mining process water. The iron sand removal efficiency for Cl⁻ ions was assigned mainly to the high porosity of sorption material and the small ion diameter of chloride [22,23]. The calcium could be successfully removed from a solution by natural zeolites, widely used as adsorbents in separation and purification processes in the past decades [25,26,27,28].

A promising method for removing pollutants from industrial wastewater is sorption. Several experimental studies have been performed on the simultaneous removal of chlorides and calcium from synthetic solutions by sorption [29,30]. However, industrial wastewater contains various pollutants, such as heavy metals, salts, and organic matter, which could affect the sorption efficiency of sorbents [31]. Some authors published studies where industrial wastewater with different cations and anions was examined for the possibility of removing ions with different sorption materials [32,33].

Different sorption materials were tested for simultaneous removal of cations and anions from different types of synthetic and less from real wastewater. High sorption capacity was recorded by testing Hydrotalcite type. HT is a layered double hydroxide with anion exchange properties. HT transforms into Mg-Al oxide when heated to 500–800 °C. The Mg-Al oxide can intercalate anions in solution, and then the original HT structure can be reconstructed. Additional cations and water molecules occur alternating with the positively charged brucite-like main layers. Another sorption material that could be applied for sorption belongs to the clay mineral group. Clay minerals’ mechanisms for removing ions from wastewater could vary based on chemical composition and microstructure. It could be present ion exchange, physical adsorption, or diffusion into the pores or cavities of the clay minerals framework, where they could be attracted by Van der Waals forces or hydrogen bonds [34]. Natural zeolites are a group of crystalline, hydrated tectoaluminosilicates characterized by varied and extremely valuable physicochemical properties. High cation exchange capacity (CEC) makes these minerals widely used in removing heavy metal cations from aqueous solutions. Also, using zeolites for removing inorganic anions from wastewater is possible by changing the chemical surface properties, e.g., treating with large cationic surfactants, which leads to changes in the external surface charge of the zeolite from negative to positive, or by thermal treatment [35].

Natural zeolites are a group of crystalline, hydrated tectoaluminosilicates characterized by varied and extremely valuable physicochemical properties. One of the most significant is the ability to absorb and separate ions and molecules. High cation exchange capacity (CEC) makes these minerals widely used in removing heavy metal cations from aqueous solutions. Also, using zeolites for removing inorganic anions (for example, chlorides) from wastewater is possible by changing the chemical surface properties, e.g., by treating with large cationic surfactants, which leads to changes in the external surface charge of the zeolite from negative to positive, or by thermal treatment [36,37].

This research took a holistic approach to removing of chloride and calcium from industrial wastewater, specifically from the water washing of EAFD from steel production. EAFD wastewater was previously not studied for purification with any other available technologies. Thus, this study offers new insights into purification and treating wastewater with adsorption using noncommercial materials as sorbents.

This study focuses on the removal of chosen cations and anions from industrial wastewater. The EAFD wastewater in this study has a specific composition and high alkalinity. EAFD wastewater is unique in terms of composition, and there is a gap in the research on contaminant removal from EAFD wastewater by sorptions. Moreover, in this study, we were focused on cleaning water itself and specifically on high concentrations of problematic ions, such as chlorides and calcium, in the presence of concentrations of other ions like sulphates and heavy metals. Therefore, interactions between cations and anions must be considered during industrial wastewater sorption studies. In order to compare the sorption behaviour of four different sorbents, a synthetic solution and real wastewater (EAFD wastewater) were used. The sorbents from different sources (laboratory-synthesized Hydrotalcite, natural zeolite Clinoptilolite, laboratory-synthesized zeolite from waste, and waste alumina sludge) were characterized. Then, the sorption capacity of the sorbents and the efficiency of removing Cl⁻ and Ca²⁺ ions were investigated in the absence and presence of other ions. This study aimed to determine and select a prospective sorbent that is most effective and sustainable in the simultaneous removal of chlorides and calcium from industrial wastewater (alkaline).

2. Experimental Procedures

2.1. Sorption Materials

Four various types of adsorption materials were tested. Two solid materials were prepared for this study in the laboratory: Hydrotalcite (HT) and synthetic zeolite (SZ). The other used sorption materials were the waste product, sludge from anodic oxidation of aluminium (KA), which was provided by a Slovakian aluminium profile producer, and natural zeolite Clinoptilolite (ZC) obtained from Zeocem, Bystré, a.s. (Slovakia).

2.1.1. Preparation of Hydrotalcite

Hydrotalcite (LDH, layered double hydroxide) was synthesized hydrothermally by a co-precipitation method from sulphate salts, as described elsewhere [38].

MgSO₄.7H₂O (208.5 g) and Al₂(SO₄)₃.18 H₂0 (33.3 g) were dissolved together in deionized water (500 mL). The obtained sulphate solution was continually dropped under vigorously stirred conditions into 1M Na₂CO₃ (200 mL) tempered to 40 °C. Simultaneously, 2M NaOH (total volume of 600 mL) was added to the suspension to maintain the pH value at 10. After the synthesis, the temperature of the mixture was increased up to the range of 70–75 °C. Subsequently, the mixture was crystallized for 12 h in a hydrothermal bath under the stirring conditions and then, for the next 12 h, placed in the oven at 70 °C without stirring for further crystals to age. The obtained precipitate was washed ten times with 4000 mL deionized water and dried in the oven at 105 °C for 12 h. The dried precursor (HT-a) sample was calcined for 2 h at 500 °C. The calcined product (HT-b) of Mg-Al oxide was stored in a glass bottle for subsequent use.

2.1.2. Preparation of Natural Zeolite Clinoptilolite

Clinoptilolite is Slovakian natural mineral obtained from the deposit Nižný Hrabovec (Zeocem Bystré, a.s., Slovakia). The obtained sample with granularity from 1 to 3 mm was milled to powder in order to prepare materials of suitable grain size for further experiments. The aim was not to prepare the material with a highly specific surface but to study the influence of surface area or milling time on sorption effectivity. Before usage for sorption tests, the zeolite sample was dried at 105 °C for 24 h and calcined at 500 °C for 2 h. Calcined Clinoptilolite (ZC-b) was stored in a glass bottle.

2.1.3. Preparation of Synthetic Zeolite

Sulphate waste liquor (167 mL) obtained from anodic oxidation of aluminium (hazardous waste) with known Al content (29.835 g Al³⁺/L) was mixed with 0.6 M NaOH (400 mL). The mixture was heated at 80 °C for 4 h under stirring in a beaker where the sodium silica glass with a total volume of 121 mL was discontinuously added. During the synthesis, the product was crystallized, then filtered, washed with deionized water, and dried at 105 °C, for 3 h. The dried product (SZ-a) was stored.

2.1.4. Preparation of Sludge

Alumina sludge (non-hazardous waste) obtained from the aluminium profile surface finishing (after anodic oxidation and waste solutions neutralization) was ground to powder, calcined at 500 °C for 2 h in the electrical resistance oven, and used as an adsorbent (KA-b).

2.2. Preparation of Synthetic and Real Wastewater

The synthetic solution with the content of Cl⁻ and Ca²⁺ ions was prepared by diluting anhydrous calcium chloride (Merck, Kenilworth, NJ, USA) in deionized water. The synthetic solution of CaCl₂ contained 2530 mg/L of Cl and 1307 mg/L of Ca.

Industrial wastewater was obtained by washing EAFD with deionized water for 30 min at room temperature. The ratio of deionized water and EAFD during washing was 10. The partial composition of EAFD wastewater is shown in

Table 1. Partial elemental composition of industrial wastewater from washed EAFD (60 min, L:S ratio = 10).

Elements	Cl−	SO42−	Ca2+	Pb2+	Cr	Na+	K+
Ci [mg/L]	1995.2	1200	1084	6.16	1.69	525	694

.

2.3. Adsorption Test

Batch sorption experiments using the synthetic solution and EAFD wastewater were provided. The sorption capacities of four tested sorption materials for removing Cl⁻ and Ca²⁺ ions were tested for 24 h. In total, 100 mL of CaCl₂ solution and an equal volume of EAFD wastewater were placed into beakers, where 3g of tested sorption materials was added. The beakers were shaken for 24 h to reach sorption equilibrium. After reaching equilibrium in the solution, samples were filtrated. Liquid samples were analysed before and after sorption to obtain concentrations of Cl⁻ and Ca²⁺ ions in every sample. Also, the initial and final pH values of the solutions were recorded. Solid residues were dried at ambient temperature and proceeded for characterization by XRD (X-ray powder diffraction phase analysis) and FTIR (Fourier Transformation Infrared spectroscopy). Selected adsorbents were also subjected to LIBS (Laser-Induced Breakdown Spectroscopy) and SEM-EDS (scanning electron microscopy with Energy Dispersive Spectrometry) analysis. The sorption efficiency, $ɳ$ [%], and sorption capacities, qe [mg/g] of Cl⁻ and Ca²⁺, were calculated by using Equations (1) and (2), respectively:

$$\mathrm{ɳ}[\%]={\displaystyle \frac{({\mathrm{C}}_0-{\mathrm{C}}_{\mathrm{e}})}{{\mathrm{C}}_0}}\times 100$$

(1)

$${\mathrm{q}}_{\mathrm{e}}={\displaystyle \frac{({\mathrm{C}}_0-{\mathrm{C}}_{\mathrm{e}})\mathrm{V}}{\mathrm{m}}}$$

(2)

where C₀ (mg/L) and C_e (mg/L) are initial and equilibrium ion concentrations, respectively; V (L) is the initial volume of solution; and m (g) is the mass of adsorption material.

The sorption rate of all studied sorbents was tested in EAFD wastewater. The sample of sorption material was dispersed in 100 mL of wastewater, and the mixture was stirred for 24 h. Individual samples were collected and filtered at chosen time intervals. The samples were analysed for calcium and chloride contents. Samples collected after 24 h of sorption were further analysed for other ion concentrations presented in tested liquid solutions, such as SO₄²⁻, K⁺, Na⁺, Pb²⁺, and Cr³⁺. The obtained ion concentrations from the initial and final solutions were calculated by using the Equation (3):

$${\mathrm{q}}_{\mathrm{t}}={\displaystyle \frac{({\mathrm{C}}_0-{\mathrm{C}}_{\mathrm{t}})\mathrm{V}}{\mathrm{m}}}$$

(3)

where q_t (mg/L) is the uptake capacity of ions by an adsorbent at time t; C₀ (mg/L) and C_t (mg/L) are initial ion concentration and ion concentration at time t; V (L) is the volume of solution; and m (g) is the mass of adsorption material.

All sorption tests were provided in triplets, and the resulting concentrations of ions were used to calculate the average concentration, which was used to interpret the results further, including statistical representation of results by calculating standard deviations.

2.4. Analytical Methods

The chemical composition of solid and liquid samples was determined by Atomic absorption spectrometry (AAS, Varian Spectrophotometer AA20+, Richmond, VA, USA). The weight 0.2 g sorbent sample was dissolved in 100 mL of concentrated HCl, and the main elements were analysed in each sorbent. AAS analysis was also used to measure calcium, lead, chromium, potassium, and sodium in solutions. Chlorides in liquid samples were analysed using the Argentometric titration method. Each sample was titrated three times, and the average value was used for evaluation. Sulphates were determined by spectrophotometry (HACH DR/2000 Direct Reading). The LIBS analysis using an optical microscope Keyence VHX 970F with a LIBS analyser EA-300 was chosen mainly for monitoring elements within the input sorbent and as a complementary dry analytical method to the standard wet chemical method used in this study. Moreover, it serves also as a comparison of the composition of selected sorbents before and after sorption.

Before and after adsorption, the phases presented in precursors and calcined sorbents were analysed by X-ray diffraction (XRD) using a PANalytical X’Pert PRO MRD powder diffractometer using Co Kα radiation. The XRD data were evaluated using the software X’Pert Pro 4.8. The morphology observation of input sorbents before sorption was performed by using a scanning electron microscope (SEM), Tescan Vega3, and EDS analysis (with detector Bruker, Billerica, MA, USA) for chemical composition analysis of selected sorbents after sorption. The samples were coated with carbon before SEM measurements. Infrared spectroscopy measurements of sorbents were recorded with a Nicolet Avatar 330 spectrometer. Spectra were collected in the middle infrared region (4000–400 cm⁻¹), averaging 32 scans with 4 cm⁻¹ resolution. Samples were obtained using the attenuated total reflection (ATR-FTIR) method. The N₂ adsorption/desorption method (calculated using the BET equation) with a NOVA 2200e gas adsorption system at 77.3 K (Quantachrome Instruments, Boynton Beach, FL, USA) determined the surface areas of adsorbents. The samples were degassed before measurement for 27 h at an outgas temperature of 105 °C. The Barrett, Joyner, and Halenda (BJH) method calculated the pore volume and diameter of pores.

Chemical and phase analyses were provided for each sorbent to understand the sorption behaviour of chlorides and calcium on different sorbent materials. The changes in infrared spectra give us insight into what bonds in the structure of materials could be responsible for the sorption of Cl⁻ and Ca²⁺ ions from solutions.

3. Results and Discussion

3.1. Characterization of Sorption Materials

3.1.1. Chemical Composition, Surface Area, Pore Size Distribution, and Morphology

Table 2. Content of selected elements in sorption materials.

Sample	Metal Content [wt.%]
	-	Mg	Al
HT-b	-	24.42	3.72
	Si	Ca	Al
ZC-b	13.32	3.38	2.01
SZ-a	34.7	0.13	11.5
KA-b	0.27	0.04	30.265

provides an overview of the elemental chemical composition of sorption materials used for removing Ca²⁺ and Cl⁻ ions from CaCl₂ solution and EAFD wastewater.

The specific surface area was determined using the adsorption/desorption method, and the BHJ model was used to calculate pore volume and diameter. The purpose of these analyses was to define sorbent’s porosity. Figure 1 presents the results of analyses of sorption materials on pore size distributions, N₂ adsorption and desorption isotherms, and BHJ models of pore size distributions of sorbents.

Three samples (HT-b, ZC-b, and KA-b) show a hysteresis loop with a specific shape. The results indicate that the N₂ adsorption isotherms belong to Type II and III according to the Brunauer, Deming, Deming, and Teller classification, which indicates multilayer adsorption [39].

According to IUPAC isotherms and hysteresis loop classification, the adsorption/desorption curves of HT-b, ZC-b, and KA-b exhibit an inverse “S” shape, which is similar to those of IV and V and represents type H3 loop [39]. This means that types IV and V belong to mesoporous adsorbents with strong and weak affinities, respectively. Therefore, adsorbents appear to have large sheet grain matrices, and there are three main types of pore geometry: plates, sheet cracks, and mixed pores. On the contrary, SZ-a showed an adsorption/desorption isotherm typical of a uniform, non-porous surface material. Because glassy materials are almost amorphous, this is understandable. According to its chemical analysis, SZ-a contains the highest percentage of Si (34%) in all sorbents.

Based on de Boer’s analysis of the desorption isotherms, Figure 1 illustrates the pore size distribution of samples, showing which pore sizes contribute to the total pore volume. All materials have pore sizes in the mesoporous range (2–50 nm).

Based on the BET analysis, all tested sorption materials’ specific surface areas (SSAs) were determined from N₂ adsorption and desorption data. This test aimed to measure the SSA of used sorbents, pore volume, and diameter. Textural parameters of used sorbents are listed in

Table 3. Textural parameters of sorbents.

Sample	BET Surface Area [m2 g−1]	Single Point Pore Volume [cm3 g−1]	Pore Diameter Dv (d) [nm]
HT-b	48.532	0.209	3
ZC-b	27.060	0.157	4
SZ-a	78.267	0.064	3
KA-b	118.890	0.200	4

.

The sample KA-b exhibits the highest SSA (118.89 m²/g), probably due to mild calcination conditions of hydrated alumina sludge and break-up of the particles and pore evolution while releasing chemically bound water. The sample KA-b has 2.5 times higher SSA than HT-b (48.532 m²/g) but a similar pore volume (0.2 cm³/g). The second-largest surface exhibits a sample of SZ-a, with an area of 78.267 m²/g and the lowest pore volume, which corresponds to BHJ analysis (Figure 1C) and confirms that SZ-a exhibits non-porous particles of small size (~nm). The reason for the low surface area at ZC-b (27.060 m²/g) and HT-b is probably that sorbents have larger particle size (aggregates) compared to KA-b and SZ-a (as shown in the SEM morphology images in Figure 2a and Figure 3a).

SEM analysis of HT-b (Figure 2) showed a broad particle size distribution with particle aggregates mostly of rounded shape with visible layered structure, with a size above ~25 µm. SEM-EDS and LIBS analysis confirm the presence of main elements Mg-O-Al (with higher quantities), but SEM-EDS also pointed out minor impurities based on Na, K, Fe, S, and Si. LIBS analysis also identified some water bound with Mg-O and Mg-O-Al in the HT-b sample, which indicates the quick adsorption of water from humid air even after calcination and the ability of both compounds to adsorb water.

The SEM image of ZC-b (Figure 3) shows mostly irregular prolonged particles with smooth or sharp edges and a size distribution similar to that of the HT-b sample. LIBS analysis confirms the presence of compounds based on Si-Al-O-Na-K, Si-O, and Si-Al-O-K and indicates the occurrence of Fe and Ti.

SZ-a sorbent (Figure 4) exhibits a 3D net morphology structure consisting of individual 2D particles of square/cubic shape (~1 µm), which form more significant clusters (around 40 µm) which are interconnected within the 3D aggregate net (Figure 4a). LIBS confirm that the main elements represent the specific zeolite composition, Al-Si-Na-O. Moreover, some whiskers are visible in the SZ-a sample, which could be related to the crystallization of sodium or sodium silica compounds (from residual solvents), as revealed in SEM-EDS and LIBS results.

SEM image of KA-b sorbent (Figure 5) shows irregular shapes of particles bound in layered aggregates of the size of approximately 10 µm. LIBS analysis confirms that sorbent KA-b predominantly contains Al-O or Al-O-H phases.

3.1.2. Mineralogical Composition of Sorption Materials

XRD patterns of samples of Hydrotalcite HT-a and HT-b are shown in Figure 6. In sample HT-a, phase Mg₆Al₂(OH)₁₆CO₃.4H₂O (JCPDS 00-041-01428) was identified. The additional diffraction lines appear in sample HT-a, which may be attributed to the presence of Nordstrandite Al(OH)₃ (JCPDS 01-085-1049), Brucite Mg(OH)₂ (JCPDS 00-001-1169), and Na₉(Al(OH)₆)₂(OH)₃.(H₂O)₆ (JCPDS 01-086-1306) due to the residual concentration of Al and Mg in the initial Mg-Al solution after HT formation. In the calcined HT-b sample, Periclase MgO (JCPDS 00-001-1235), Spinel MgAl₂O₄ (JCPDS 96-500-0121), Corundum Al₂O₃ (JCPDS 01-089-3072), and sodium sulphate Na₂SO₄ (JCPDS 01-083-1570) were identified.

Natural zeolite Clinoptilolite ZC is composed of a three-dimensional grid that consists of silicate tetrahedrites (SiO₄)⁴⁻ that are connected by oxygen atoms, with a few silicon atoms being substituted for aluminium (AlO₄)⁵⁻ [40]. Clinoptilolite becomes stable over the temperature of 450 °C because all bounded water is removed from Clinoptilolite [41]. Therefore, Clinoptilolite was calcined at 500 °C before being used as an adsorbent. A fine powder with an average particle diameter of 32.8 µm (measured by analyser Mastersizer 3000, Malvern Instruments, Ltd., Malvern, UK) was used in the present study and has a molar ratio of Si/Al 6.62. The main mineralogical phases detected (XRD pattern, Figure 7) in raw zeolite ZC-a were Clinoptilolite-Na (JCPDS file 96-900-1395), Cristobalite SiO₂ (JCPDS file 00-039-1425), Albite Na(AlSi₃0₈) (JCPDS file 00-020-0572), Anorthite CaAl₂Si₂O₈ (JCPDS file 00-041-1486), Muscovite-3T (K, Na)(Al, Mg, Fe)₂ (Si_3.1Al_0.9)O₁₀(OH)₂ (JCPDS file 00-007-0042), Orthoclase KAlSi₃O₈ (JCPDS file 00-031-0966), Quartz SiO₂ (JCPDS file 00-046-1045), and Magadiite NaSi₇O₁₃(OH)₃.3H₂O (JCPDS file 00-020-1156). The ZC-a sample also contains titanium impurity (confirmed by LIBS analysis), which could be in the form of Rutile. The XRD pattern of calcined zeolite ZC-b does not show significant phase compared to ZC-a.

The XRD pattern of synthetic zeolite SZ-a is shown in Figure 8. The main mineralogical phases observed from the XRD pattern of sample SZ-a are zeolite A, sodium (LTA) Na₁₂Al₁₂Si₁₂O₄₈)NaAlO₂(H₂O)₂₉ (JCPDS 01-089-3859), with the possible presence of hydrated sodium carbonate Na₂CO₃.H₂O (JCPDS 01-070-2148 ). This study described this zeolite type’s characterization in detail [42].

From the XRD pattern of sorbent alumina sludge KA-a in Figure 9, it was apparent that KA-a is mainly amorphous, but it may indicate the presence of phases such as Bayerite Al(OH)3 (JCPDS 00-012-0457), Boehmite AlO(OH) (JCPDS 01-088-2112), and Brucite Mg(OH)₂ (JCPDS 00-044-1482). SEM-EDS confirmed Mg presence in KA-a. Except for this, phases based on sodium and aluminium sulphates like Felsobanyaite Al₄(SO₄)(OH)₁₀(H₂O)₄ (JCPDS 01-088-0765) and Mirabilite Na₂SO₄.10H₂O (JCPDS 00-001-0207) could be present, too, as sulphur and sodium are expected impurities from the neutralization step during sludge formation.

The calcination of the sample caused a change in phase composition. The non-stable polymorph of the d-alumina phase (JCPDS 00-047-1770) was formed, and another possible candidate is Quintinite-2H Al₂Mg₄(OH)₁₂(CO₃)(H₂O)₃ (JCPDS 01-087-1138). This phase could form from the Al and Mg phases of the sorbent KA-b during calcination. The phase Quintinite-2H has a similar structure to Hydrotalcite, but the amorphous nature of the sample does not allow for a more precise phase analysis. Mg impurities in the sludge and KA-b sample are present due to the dissolution of Mg as an alloying element during anodic oxidation of Al alloy profiles, followed by precipitation as hydroxide together with aluminium. SEM-EDS analysis confirms the presence of Mg and some other impurities in KA-b, such as Si, S, and Fe, representing standard contaminants in a sludge.

3.2. Results of Chloride and Calcium Adsorption from Synthetic Solution and EAFD Wastewater

The experiments investigate the sorption capacity of Ca²⁺ and Cl⁻ ions of four sorbents from synthetic solutions and EAFD wastewater. The initial concentrations of Cl⁻ and Ca²⁺ ions in synthetic solution of CaCl₂ were 2530 mg/L and 1307 mg/L, respectively. The initial Cl⁻ and Ca²⁺ ions concentrations in EAFD wastewater were 1995 and 1084 mg/L, respectively.

Sorption tests were set up for 24 h in order to ensure that equilibrium in sorption systems will be achieved. Figure 10 compares sorption capacities in 24 h for chlorides and calcium removal from synthetic solutions and EAFD wastewater using a volume of 100 mL of solution and a concentration of adsorbent (Cs) 30g/L. Every sorption test was performed three times, and average value of adsorption capacities were calculated with its standard deviations.

Comparing the sorption capacities of chlorides achieved by four tested sorption materials, depicted in Figure 10A, implies that the most suitable sorbent for removal of Cl⁻ is Hydrotalcite (HT-b). The adsorption capacities of HT-b in CaCl₂ solution and EAFD wastewater were 41.6 and 38.6 mg Cl⁻/g, respectively. A different behaviour was noticed when using other tested sorbents where the sorption of chlorides was less effective from EAFD wastewater. In equilibrium, the sorption capacities were 28.2 and 29.2 mg Cl⁻/g using ZC-b and KA-b, respectively. On the contrary, SZ-a reached an adsorption capacity of only 7.4 mg Cl⁻/g. With sorbents, ZC-b, SZ-a, and KA-b have achieved 0 mg/g Cl⁻ ions removal from the synthetic solution of CaCl₂, which is represented in Figure 10A.

The efficiency of calcium removal (Figure 10B) from both solutions using SZ-a adsorbent was around 100%, representing adsorption capacities from synthetic solution 43.55 mg Ca²⁺/g and 36 mg Ca²⁺/g from wastewater. Sorption capacity for HT-b was 21.5 and 15.6 mg Ca²⁺/g to remove Ca²⁺ ions from synthetic solution and EAFD wastewater, respectively. ZC-b and KA-b reached poor adsorption capacity in the synthetic solution. Whilst there are no differences in reaching equilibrium sorption capacity between KA-b (around 7 mg Ca²⁺/g) in both solutions, the qe reached in synthetic solution (5 mg Ca²⁺/g) by using ZC-b was much lower than that in EAFD wastewater (26.2 mg Ca²⁺/g).

The behaviour of pH values of solutions during sorption varies.

Table 4. The comparison of changes in pH values in the synthetic solution and EAFD wastewater (experimental conditions: V = 100 mL, Cs = 30 g/L, t = 24 h).

	Adsorption Material	Synthetic	EAFD Wastewater
Initial pH value (pHi)	-	6.87	12.78
Final pH value (pHf)	HT-b	11.77	12.54
	ZC-b	8.53	12.24
	SZ-a	12.34	12.87
	KA-b	4.41	7.78

shows the solution’s final pH after synthetic and EAFD wastewater sorption. In the beginning, the pH value of the synthetic solution was equal to 6.9. The rapid pH decrease to 4.41 was determined after 24 h of sorption with KA-b. We can attribute this to sulphates released into the solution from KA-b during sorption. The previous experiment in another study of authors showed that 1.7% sulphates and 1% aluminium ions are released into the solution of synthetic CaCl₂ after 24 h of contact with sorbent KA-b.

3.3. Evaluation of Sorption Results

After sorption, the resulting sorption materials were subjected to further analyses to obtain FTIR records and X-ray diffraction patterns of sorbents. The changes in FTIR and XRD records before and after sorption in synthetic solution and EAFD wastewater were analysed to understand the sorption behaviour of different sorbents better.

Characterization of Adsorbents After Sorption

After the sorption with Hydrotalcite from the synthetic CaCl₂ solution, it can be seen from XRD patterns of HT-c (Figure 11) together with HT-a and HT-b (Figure 6) that the calcined sorbent HT-b was rehydrated in the synthetic solution and in EAFD wastewater.

In the Figure 6 Al₂(OH)₁₆CO₃.4H₂O (JCPDS 00-041-01428) was identified. However, in XRD patterns, some shift in individual peak positions (Bragg angle °2 Theta) of the Hydrotalcite phase in the sample HT-c is visible compared to HT-a. By XRD pattern analysis using software (HighScore X-Pert 4.8), the Hydrotalcite phase in HT-c sorbent exhibited an increased d₀₀₆ basal spacing of the diffraction plane (006) from 7.86 to 8.15 Å. The increase in “d₀₀₆” indicates accommodation of chlorides or other ions (Ca²⁺) in interlayers of the HT-c structure, and not only adsorption onto the external surface of Hydrotalcite or other mineral phases contained in the sample HT-c. In addition, the XRD pattern of HT-c revealed some new phases containing chlorides such as Mg₂(OH)₃Cl(H₂O)₄ (JCPDS 01-073-2119), Mg₃Cl(OH)₅.4H₂O (JCPDS 00-007-0416) or Mg-Al chloride hydrate Mg_xAl (OH)_2x+2Cl.zH₂O (JCPDS 00-019-0748). This finding suggests the important role of magnesium species and the hydration process of HT sorbent in chloride interaction and final capture into the hydrated sorbent structure from wastewater. Ca²⁺ ions were removed by HT-b with high efficiency, too. This was caused by the reaction of Ca²⁺ with carbonate anions present in the structure of the HT-c and intercalation of formed CaCO₃ into the interlayer space of HT-c, or, eventually, it is also possible precipitation from solution at the given conditions (pH). Residual sodium and sulphates together with aluminium present in HT-b and during sorption could probably also contribute and assist in the calcium removal through precipitation and forming the more thermodynamically stable solid phases like CaSO₄(H₂O)₂ (JCPDS 01-072-0596) and Na_3.2(CaSO₄)_2.6.1.5H₂O (JCPDS 00-024-0927) or Ca₄Al₂O₆(SO₄).14.H₂O, as indicated by XRD analysis. XRD pattern of HT-c further indicated phases Mg(OH)₂ (JCPDS 00-044-1482), Norstrandite Al(OH)₃ (JCPDS 00-024-0006), and MgAl₂(OH)₈ (JCPDS 00-035-1274) which were formed by rehydration of phases MgO, MgAl₂O₄, and Al₂O₃ present in calcined HT-b. Non-reacted MgO could also possibly be present in HT-c, as indicated by the LIBS analysis.

LIBS analysis of HT-c showed the elemental composition based on Mg-O-Al-S-H-Cl elements (Figure 12), showing the ability of HT-b to adsorb chlorides together with the present residual sulphates, which are probably coupled within the rehydrated HT-c phases. Moreover, the group of elements Mg-O-Ca-S-H-C was also observed, as well as Mg-O-Ca-H, which indicates the probable affinity of Ca species towards hydrated Mg oxide phases of HT-c. In contrast to LIBS analysis results for HT-b (before sorption), even Ca-Mg-O elemental composition was found, which could signify the ability of the active MgO phase to adsorb Ca even without water assistance. This deduction was also supported by the other LIBS point analysis, which found composition Mg-O-Ca-C, which could indicate Ca capturing into the Mg structure and carbonate formation, which XRD also confirmed. On the other hand, the compositions of individual particles within the HT-c without Al and with Ca could also mean that Ca is exchanged for Al in the HT-c structure. Another interesting cognition was that chlorides could be adsorbed not only by the HT phase but also by the Mg hydrated oxide phase. In addition, in both samples before and after sorption in the synthetic solution, the hydrated and non-hydrated MgO phases could be present, as indicated by the LIBS analysis (Mg-O and Mg-O-H).

Mineralogical changes were investigated after sorption in EAFD wastewater from the XRD pattern of HT-d (Figure 11). Amongst phases of Mg(OH)₂ and Al(OH)₃ (similar to those observed in HT-c using the synthetic solution), other new phases were also identified, like Stichtite Mg₆Cr₂CO₃(OH)₁₆·4H₂O (JCPDS 00-045-1475), Ca₄Al₂Cr_0,5O₈Cl.12H₂O (JCPDS 00-043-0088), Mg₃Cl(OH)₅.4H₂O (JCPDS 00-007-0420), Na(Cr(SO₄)₂).12H₂O (JCPDS 00-008-0039), and CaAl₂((OH)₈(H₂O)₂)(H₂O)_1.84 (JCPDS 01-088-1410), which suggests rehydration of calcined HT-b and at the same time readily binding of Cl⁻, Ca²⁺, and Cr ions into the HT-d structure. Other studies confirmed the high affinity of aluminium hydroxide to chromium species [43]. The mechanism of CrO²⁻₄ sorption was interpreted in terms of reactions between chromates and -OH and/or H₂O groups at the hydroxide/liquid interface. It has been shown that chromates are more tightly sorbed on aluminium hydroxide compared to other anions, e.g., chlorides [43]. Nevertheless, when calcium is present in wastewater, chlorides also seem to bind to Al(OH)₃, not only Cr species alone. Calcium also formed a CaCO₃ phase similar to that in the case of synthetic solution sorption, which suggests the presence of residual carbonate groups within the calcined HT-b. LIBS analysis and composition observation using different places of the sample of HT-d sorbent (Figure 13) indicated that chlorides are adsorbed in the composition O-S-Cl-C-Mg-H (for example, as Mg oxychloride–carbonate–sulphate–hydrate). The composition of O-C-S-Mg-H obtained by LIBS measurements of HT-d suggests the good affinity of sulphate to Mg hydrated phases of HT or absorption of sulphates into the layers of rehydrated Hydrotalcite phase. No Al and Ca were found in those compositions of HT-d, which means that the adsorption process in EAFD wastewater is not homogenous, and those elements are not evenly distributed in the bulk of the sorbent. However, Ca and other elements are present in HT-d and were confirmed by SEM-EDS analysis.

From the SEM morphology image (Figure 14), it is visible that the surface of HT-c particles (evident from bigger aggregates) is covered by adsorbed individual particles/species. In Figure 14, the detail of HT-c particles after sorption from synthetic solution is shown, where a visible hydrated fluffy HT layered structure resembles clouds. This is due to the rehydration of calcined HT-b (absorption of water and incorporation into the HT-c structure) and the expansion of the material volume and interlayer and planar space. This is the difference between calcined HT-b, which is not a hydrated structure, and HT-c, which is after sorption from water solution. Figure 15a shows the SEM morphology of HT-d, and Figure 15b shows the elemental composition map of selected areas of the sample HT-d. The morphology of HT-d (Figure 15a) is slightly different from HT-c. HT-d is characterized by a more chaotic arrangement of particles and rehydrated HT-d layers. Moreover, those layers do not look fluffy like HT-c and have a more distorted structure with a thin and solid appearance. SEM-EDS analysis of HT-d (Figure 15b) together with LIBS confirmed the presence and thus sorption of studied impurities Cr, Pb, Ca, and Cl and, furthermore, Zn from real EAFD wastewater. Results also revealed the presence of other quite common impurities, Si and Fe, in HT-d and residual species Na and K, coming from the preparation process of HT-b. From the SEM-EDS map image, it is also evident that HT-d sorbent particles are practically completely covered by adsorbed elements, and Mg, Al, and O are hidden (the intensity of colour is diminished).

FTIR spectra of sample HT-a synthesized, HT-b calcined, HT-c after sorption from the synthetic solution, and HT-d from EAFD wastewater are shown in Figure 16.

Comparing FTIR samples of Hydrotalcite before and after sorption showed that the sample HT-a reveals a very weak peak at 3694 cm⁻¹, and it is attributed to cation hydroxyl stretching regions. This may be assigned to the cation OH stretching vibration of brucite [44], and it is increasing in intensity probably due to rebuilding the structure after calcination and sorption of Hydrotalcite. In the sample, HT-d did not observe changes in FTIR bands belonging to CO₃²⁻ which possibly affected preferences to adsorb cations in EAFD wastewater, mainly Cr³⁺, which was detected in XRD scan as Mg₆Cr₂CO₃(OH)₁₆.4H₂O phase.

The increased pH value in the CaCl₂ solution from initial 6.87 to 11.77 after sorption is evidence of anion sorption (Table 4). Hydroxide ions in the HT interlayer were released into the bulk solution by the adsorption of other anions [45]. Conversely, the adsorption capacity of Hydrotalcite for Cl⁻ from EAFD wastewater (HT-d) was lower than that of CaCl₂ solution. This was probably due to the different pH of solutions (lower initial and higher final pH of the solution after sorption in HT-d). The high OH- concentration in the solution could prevent Cl⁻ from combining with MgAl₂O₄, thereby hindering Cl- removal and supporting the sorption of Me²⁺ [30].

Diffraction peaks of zeolite after sorption in XRD patterns of the ZC-c samples (Figure 17) did not show substantial changes in phase composition compared to ZC-b, which means the sorbent is relatively stable in the solution under the given conditions. XRD patterns in the main detected phases are similar to ZC-b {main phases: Clinoptilolite-Na (JCPDS 96-900-1395), Cristobalite SiO₂ (JCPDS 00-039-1425), Albite Na(AlSi₃O₈) (JCPDS 00-020-0572), Anorthite CaAl₂Si₂O₈ (JCPDS 00-041-1486), Muscovite-3T (K,Na)(Al,Mg,Fe)₂(Si_3.1Al_0.9)O₁₀(OH)₂ (JCPDS 00-007-0042), Orthoclase KAlSi₃O₈ (JCPDS 00-031-0966), Quartz SiO₂ (JCPDS 00-046-1045), and Magadiite NaSi₇O₁₃(OH)₃.3H₂O (JCPDS 00-020-1156)}, but diffraction peaks are a little bit shifted which could be caused by partial adsorption of Ca and some structural disorder. Except for phases mentioned above, ZC-c also indicated the presence of phases based on Ca after sorption, namely, Ca₂Al₄Si₁₄O₃₆.14H₂O (JCPDS 00-025-0124), Heulandite Ca_3.6K_0.8Al_8.8Si_27.4O₇₂.26.1H₂O (JCPDS 00-053-1176), and Ti in the form of Na₄((TiO₄)(SiO₄)₃).(H₂O)₆ (JCPDS 01-088-0703).

ZC-d has similar main diffraction peaks to ZC-b (associated with abovementioned main phases detected in ZC-c, Figure 17), and the new phases occurred based on hydrated calcium, chromium, lead, zinc, chlorides, and sulphates like KCa₄Si₈O₂₀(OH).8H₂O (JCPDS 00-030-0920), Ti₂₉O₄₂Cl₃₂.110H₂O (JCPDS 00-021-1237), Cr₂(SO₄)₃.17H₂O (JCPDS 00-049-0999), Pb₁₀(SO₄)O₇C_l4(H₂O) (JCPDS 01-0070-3392), and K₂Zn(H₂O)₆(SO4)₂ (JCPDS 01-070-1827). These findings revealed that natural zeolite is suitable to adsorb, except calcium and chlorides also mentioned heavy metals from wastewater.

FTIR spectra of Clinoptilolite ZC–b calcined, ZC-c after sorption from the synthetic solution, and ZC-d from real wastewater are shown in Figure 18.

The FTIR analyses confirmed the Heulandite phase at wavelength 594 cm⁻¹, while the existence of 594 cm⁻¹ indicates the presence of “Heulandite type-II”. In the CaCl₂ solution, the sorption of Ca²⁺ was very low (11%), probably due to the limited ability of present aluminium silicate phases (Ca_3.6K_0.8Al_8.8Si_27.4O₇₂.26.1H₂O) to adsorb or exchange ions at given conditions (low pH around 8 of synthetic solution). The sorption of Cl⁻ was not recorded in the CaCl₂ solution. Surprisingly, the sorption efficiency of Ca²⁺ and Cl⁻ from EAFD wastewater increased to 72% and 43%, respectively. The parameters of the real wastewater solution (presence of more ions and different ionic strength, higher pH above 12) seem to be more convenient for both Ca²⁺ and Cl⁻ ion removal. XRD analysis of ZC-d (Figure 17) indicated the occurrence of new phase KCa₄Si₈O₂₀(OH).8H₂O, which could be associated with better calcium adsorption (chemisorption). Chloride ions from EAFD wastewater are probably adsorbed onto the Clinoptilolite surface via titanium compound Ti₂₉O₄₂Cl₃₂.110H₂O, which was observed by XRD analysis in sample ZC-d (Figure 17). Although, in ZC-c, titanium compound was also detected, Ti was bound probably into a more stable silicate phase (Na₄((TiO₄)(SiO₄)₃).(H₂O)₆) which seems to be less active towards chloride ions.

From FTIR measurements (Figure 18), rehydration was not evident after sorption in the samples ZC-c and ZC-d. The bonds observed at band 1025 cm⁻¹ are more intense, caused by O-Si (Al)-O-bond vibrations. Shifting of the characteristic ring band at 1055 cm⁻¹ at raw Clinoptilolite to the lower wavenumber about 20 cm⁻¹ at sample after calcination results in an increase in the number of ring members. The zeolite can be estimated on the basis of band shape changes in the region of 1055 cm⁻¹. This O-Si/Al-O band is sensitive to the content of the framework silicon and aluminium, so it is also susceptible to the dealumination degree [46]. Due to the low pH value in CaCl₂ solution (Table 4), some portion of Ca (~28%) was dissolved from zeolite ZC-b, and this caused the lower degree of Ca²⁺ (and Cl⁻) adsorption onto the zeolite surface.

By XRD analysis of synthetic zeolite A, sodian (LTA), after sorption from synthetic solution SZ-c and from EAFD wastewater SZ-d (Figure 19), except for the main zeolite phase also the new phase CaCO₃ (JCPDS 01-083-0578) was detected. This was probably caused by Ca ion exchange with sodium from the NaCO₃.H₂O phase, which was identified in the sorbent before the sorption of SZ-a. The concentration of Na in solutions was increased more than five times after sorption, and the pH values were also raised above 12.00, which confirms the release of sodium ions into the solution. Ca sorption was almost 100% from both solutions via forming CaCO₃ (Table 4).

The FTIR spectra of synthetic zeolite before and after sorption from synthetic solution SZ-c and EAFD wastewater SZ-d are illustrated in Figure 20.

The FTIR spectrum of synthetic zeolite A, sodian (LTA), indicates that the sample before sorption SZ-a contains a broadband OH group at 3243 cm⁻¹. The intensity of the bands is attributed to bending vibrations of Al–OH at 1397 cm⁻¹ in the raw SZ-a. The band at 952 cm⁻¹ is associated with asymmetric stretching vibration T-O (T = Al, Si) in aluminosilicates. At lower wavenumbers where positions are presented, vibration is assumed to ν O–Al/Si–O [47]. As evident from the FTIR spectra in Figure 20, the intensity of the peak of OH¯ groups is decreased after sorption in both of the solutions after sorption from synthetic CaCl₂ and after sorption from EAFD wastewater. This could be attributed to the sorption of Ca²⁺ onto the hydroxyl groups presented in synthetic zeolite. The intensity of the bands suggested that the bending vibrations of Al–OH at 1397 cm⁻¹ in the raw SZ-a decreased significantly in SZ-c and SZ-d because of the decomposition of the minerals after sorption [48]. The band at 952 cm⁻¹ in a sample before sorption SZ-a associated with asymmetric stretching vibration T–O (T = Al, Si) in aluminosilicates with zeolite or sodalite structure is shifted upward after sorption from CaCl₂ solution (sample SZ-c). It is connected with decreases in the content of tetrahedrally positioned Al atoms in the system [47,48]. This is in agreement with other literature where shift vibration ν O–Al/Si–O toward higher frequency is connected to the dissolution of Al and Si [49]. In addition, in the sample, SZ-d peaks at 952, 707 and 666 cm⁻¹ were observed to dissolve the zeolite structure due to the high value of the pH solution (12.87).

The sorption with alumina sludge KA-d of Cl⁻ efficiency in EAFD wastewater reached 45% at a final pH of 7.78. During the sorption process, a decrease in pH from 6.87 to 4.41 (Table 4) was recorded in the synthetic solution, and it may be caused by dissolving mainly residual sulphates from calcined KA-b (Figure 21).

The FTIR spectra of KA-b calcined, KA-c after sorption from synthetic solution, and KA-d from EAFD wastewater are illustrated in Figure 22. The OH bands at the wavelength 3408 cm⁻¹ of Mg(OH)₂ and Al(OH)₃ were not reconstructed after sorption in CaCl₂ solution and EAFD wastewater. At the wavelength 1363 cm⁻¹ at the KA-a sample, a very intensive peak was observed, which is assumed to be the CH or CH₂ aliphatic bending group. It could be from degreasing agents in the anodic oxidation process of aluminium profiles. After sorption, FTIR results do not show evidence of this phase anymore in both samples, which is probably the signature of its dissolving into the wastewater. Also, the peak at wavelength 934 cm⁻¹ at sample KA-d diminished after sorption, and it contributed to the Si-OH stretching bend. XRD analysis of KA-c and KA-d after sorption revealed the attained amorphous character and presence of phase d-alumina (JCPDS 00-047-1770) in the sorbents after sorption and disappearance of diffraction peaks associated with compound Quintinite-2H Al₂Mg₄(OH)₁₂(CO₃)(H₂O)₃ (JCPDS 01-087-1138). In KA-c and KA-d, the probable new phases were identified as Na_1.67Mg_0.67Al_10.33O₁₇ (JCPDS 01-084-0214) and Mg_0.4Al_0.6Al_1.8O₄ (JCPDS 01-087-0340). On the opposite, in KA-c appeared another quite well-crystallized phase (diffraction peak visible in XRD diffraction pattern of KA-c sorbent in position 31.08° 2theta) in Figure 21, associated with Mg₃(SO₄)₂(OH)₂ (JCPDS 01-079-1189). In addition, in KA-d, the phase Mg(OH)₂ (JCPDS 01-082-2454) was indicated with some probability in contrast to KA-c. However, it is necessary to mention that the XRD phase analysis of KA sorbents is only indicative due to the amorphous character of samples (a nanometric particle size). That is why, for example, new phases based on Ca or Cl were not detected in KA sorbents after sorption.

3.4. Investigation of Sorption Rate of Individual Sorbents from EAFD Wastewater

Setting an equilibrium time for sorption is an essential parameter for the future economic planning of wastewater treatment systems. The optimum time necessary to reach the equilibrium sorption capacity with HT-b, ZC-b, SZ-a, and KA-b for Ca²⁺ and Cl⁻ ions was investigated by a sorption test using EAFD wastewater up to 360 min for HT-b and 120 min for ZC-b, SZ-a, and KA-b. In each experiment, a control sample was also analysed after 24 h to check the potential changes in the equilibrium system. Each experiment was performed three times, and the obtained values for each time section were calculated as average values. The standard deviations were very low for the whole dataset. The initial concentrations of Cl⁻ and Ca²⁺ ions in wastewater were 1890.4 mg/L and 1266 mg/L, respectively.

Figure 23A represents a comparison of the experimental results showing the progress of the sorption capacity of Ca²⁺ and Cl⁻ ions with HT-b within 360 min. The sorption of Ca²⁺ from wastewater is very fast. Within the first 10 min, the sorption equilibrium is reached, which represents the sorption capacity of 37 mg Ca²⁺/g (the sorption efficiency is 85%). On the contrary, Cl⁻ sorption requires even 360 min to reach a sorption equilibrium where the sorption capacity of HT was 50.3 mg Cl⁻/g (the sorption efficiency is 83%). The results imply that after 24 h of the sorption process with HT-b, the sorption capacity of 42.2 mg Ca²⁺/g and 49.5 mg Cl⁻/g was achieved, with the sorption efficiencies 98.9% and 81% for Ca²⁺ and Cl⁻, respectively.

According to Reaction (4),

Mg_0.80Al_0.20O_1.10 +1,10 H₂O → Mg_0.80Al_0.20OH₂(OH)_0.20

(4)

where the release of OH groups accompanies rehydration and subsequent combination of Mg-Al oxides with anions in solution. In the EAFD wastewater, visible turbidity after sorption was observed. This was probably due to the small Al³⁺ ions from Hydrotalcite, which dissolved and precipitated with OH⁻ during sorption. From the sorption time of 30 min in the EAFD wastewater, a gradual increase in the aluminium concentration from 35 mg/L to 80 mg/L up to 120 min was recorded, forming Al(OH)₄⁻. In 240 min of the process, Al concentration in the solution decreases again, probably due to the precipitation of Al³⁺ ions. OH⁻ ions are released from Hydrotalcite during rehydration and reconstruction by incorporating of anions into its structure according to the general Reaction (5).

$${\mathrm{M}\mathrm{g}}_{1-1}{\mathrm{A}\mathrm{l}}_{\mathrm{x}}{\mathrm{O}}_{1+\mathrm{x}/2}+{\displaystyle \frac{\mathrm{x}}{\mathrm{n}}}{\mathrm{A}}^{\mathrm{n}-}+\left(1+{\displaystyle \frac{\mathrm{x}}{2}}\right){\mathrm{H}}_2\mathrm{O}\to {\mathrm{M}\mathrm{g}}_{1-1}{\mathrm{A}\mathrm{l}}_{\mathrm{x}}{\left(\mathrm{O}\mathrm{H}\right)}_2{\mathrm{A}}_{\mathrm{x}/\mathrm{n}}+\mathrm{x}{\mathrm{O}\mathrm{H}}^{-}$$

(5)

where x—molar ratio Al/(Mg + Al) and Aⁿ n-valent anion. This phenomenon of precipitation of Al with OH⁻ helps to more or less maintain a constant concentration of OH⁻ groups in solution [7].

The sorption study [17] describes the precipitation of Ca²⁺ ions with Hydrotalcite from a solution of CaCl₂ with OH⁻ ions released from Hydrotalcite into Ca(OH)₂. The researchers claim that Cl⁻ ions will be sorbed just when calcium precipitates with released OH⁻ ions in the form of Ca(OH)₂. In addition, chloride sorption is more likely in the sorption system when they are in a solution with over the number of OH⁻ groups. Conversely, in the case of EAFD wastewater, where carbonates are present, Ca(OH)₂ is rather not formed. In part, the OH⁻ groups released from Hydrotalcite appear to inhibit chloride sorption because they have a higher affinity for Hydrotalcite than chlorides. Thus, chloride ions begin to adsorb only when Al³⁺ ions begin to dissolve from Hydrotalcite to form Al(OH)₄⁻ and when no sulphate ion is present in the solution. In addition, divalent anions such as SO₄²⁻ or CO₃²⁻ have higher selectivity than monovalent ions. The selectivity of Mg-Al oxides increases with the increasing electric charge of the anion and decreasing anion size. Thus, Hydrotalcite is not very easy to combine with Cl⁻ ions due to the preference to combine with OH⁻ formed during the Reaction (5) [18].

Sorption of Cl⁻ and Ca²⁺ ions with other tested materials (Figure 23B–D) was relatively fast, and within the first 10–20 min, the equilibrium of the sorption process was reached. Moreover, for KA-b, a sharp decrease in the sorption capacity of Ca²⁺ was observed after around 5 min of the sorption process. We suppose that it is caused by partial desorption of calcium from alumina sludge (KA-b). After 24 h, the sorption capacity did not change. In this specific behaviour, Ca²⁺ ion sorption could also play a role in a pH value since, from the beginning up to 24 h of the sorption process, it decreased from 12.78 to 7.78. In Clinoptilolite ZC-b, the sorption capacity of Ca²⁺ ions decreased after 24 h from 30 mg Ca²⁺/g in 20 min to around 11 mg Ca²⁺/g after 24 h. This phenomenon could be assigned to the possible desorption or dissolution of Ca from ZC-b.

The stability (chemical) of the sorbent is an essential parameter in the sorption process from the point of view of practical applicability of adsorbent in wastewater treatment. The results indicate the possible chemical non-stability of sorbent ZC-b and KA-b under given conditions (pH, present species of EAFD wastewater, sorbent composition, etc.). Therefore, their usage in EAFD wastewater is limited. Another study is required to determine the sorption mechanism of these sorbents under alkaline conditions and specify the proper conditions of the sorption process for mutual Ca and Cl removal.

3.5. The Sorption of Other Species from EAFD Wastewater by Hydrotalcite Sorbent

Based on the achieved results and to discover the behaviour and sorption ability of other present species in EAFD wastewater except Cl and Ca, like heavy metals (Pb, Cr) and sulphates, the 24-h sorption test was performed with the identified best sorbent, HT-b. The experimental conditions for the sorption experiment and the obtained results of sorption capacities, mean, and standard deviation (SD) are shown in Figure 23.

The results in

Table 5. The mean and standard deviations of sorption efficiencies and capacities for studying impurities in EAFD wastewater by sorbent HT-b (24-h test, sorbent concentration 30g/L; V = 300 mL; 20 °C, pH₀ = 13.34), x- were not measured.

Ion	SO42−	SD	K+	SD	Na+	SD	Pb2+	SD	Cr3+	SD
Initial concentration [mg/L]	1910	2.31	8.38	0.1	560	0.7	5.52	0.31	4.52	0.2
Final concentration [mg/L]	1150	1.8	-	0.2	-	0.65	0.97	0.01	0.4	0.22
Sorption capacity [mg/g]	25.33	2	-	-	-	-	0.15	0,02	0.14	0.21
Removal efficiency [%]	39.79	1.8	-	-	-	-	82.43	0.02	91.15	0.13

show a noticeable decrease in heavy metal concentration in EAFD wastewater output despite their low initial concentration in the initial sample. Pb²⁺ ions were removed with an efficiency of 82%, and Cr³⁺ ions were removed with 91% efficiency. The positive finding was that by HT-b, it is possible to achieve some sulphate removal up to around 40%, which has not been studied much yet. The salts, especially sulphate, chloride ions, and calcium in wastewater, represent a significant problem in water purification in different industrial sectors. In particular, the sulphates could inhibit and decrease the efficiency of removing salt and heavy metal impurities when applying membrane technologies. Achieved results confirmed and suggest that HT could serve as a universal and sustainable adsorbent based on magnesium and aluminium oxides with good raw material availability, which can remove a broad spectrum of cationic and anionic character impurities from industrial wastewater. Moreover, it could be helpful as a pre-treatment step in membrane technology applications.

Despite promising results, there are some challenges, and optimizing the HT sorption process for removing selected species from the wastewater is further required. The efficiency of sulphate and other heavy metal ions besides calcium and chlorides could be improved by changing conditions at sorption, at least the sorbent dosage. The process is also influenced by pH value, time, the presence of other species, and other factors. Of course, the important issue in the real industrial applicability of HT for the selected purposes is the effectivity of regeneration of HT (desorption and calcination) in wastewater management of industrial production processes in order to decrease the consumption of fresh water and secure the water recirculation in the process and remain sustainable.

4. Conclusions

The study presented in this paper focused on the purification of wastewater, which was not previously studied elsewhere. This wastewater is generated from hydrometallurgical recycling processes of EAFD by-products for zinc and other metal product recovery.

The EAFD wastewater from washing operation is featured with high alkalinity (~pH=12) and quite a high concentration of calcium and chlorides together with other ions like sulphates, sodium, potassium, heavy metals (Cr, Zn, Pb), and other species in various concentrations. To purify such wastewater by using sorption as a promising method, it is necessary to study and describe it in detail, optimize it, and make it possible to implement it efficiently on an industrial scale. This work was focused on the evaluation of the efficiency of the removal of chlorides and calcium ions by four different sorption materials, namely, synthetized Hydrotalcite, natural zeolite Clinoptilolite, synthetized zeolite A, sodian (LTA), and waste alumina sludge. Experimental results have shown the following important findings:

The possibility of simultaneous removal of chloride anions and calcium cations by sorption from EAFD wastewater (generated during EAFD washing) was approved by laboratory-prepared and calcined hydrotalcite sorbent (500 °C). The sorption equilibrium for HT-b was reached within 10 min for Ca²⁺ ions, representing the sorption capacity of 37 mg Ca²⁺/g. The sorption equilibrium for Cl⁻ was achieved after 360 min of the sorption process with a sorption capacity of 50.3 mg Cl⁻/g. The efficiency of ion removal obtained in the sorption equilibrium for HT sorbent was 85% for Ca²⁺ and 83% for Cl⁻.

The results imply that the sorption process of calcium is faster and preferential to chlorides. This could be explained by different sorption mechanisms of Ca²⁺ and Cl⁻ ions in the Hydrotalcite structure, which occur during the mutual presence of both of ions in wastewater. Calcium is adsorbed onto the positively charged brucite-type outer layers of HT-b. On the other hand, chloride anions are mainly bound into the negatively charged interlamellar space of calcined HT together with water molecules. This process reconstructs the HT structure and is accompanied by an anion exchange process in interlayer space to ensure its electroneutrality. The sorption rate of Ca²⁺ ions is moreover promoted by the presence of aluminium hydroxide in the HT sample, where Ca²⁺ ions are adsorbed, or cation exchange could occur, too.

Other tested sorption materials of ZC-b, KA-b, and SZ-a have shown unsatisfactory results for chloride removal. However, a sample of SZ-a exhibits a very high equilibrium sorption capacity for calcium (43.55 mg Ca²⁺/g from CaCl₂ solution and 36 mg Ca²⁺/g from EAFD wastewater). It could be used in wastewater purification except in very acidic solutions. These results support knowledge and enhance efforts to recycle and reutilize waste materials and to produce high-quality raw materials.

The results also suggest that the mechanism by which all studied adsorbents remove impurities from synthetic and wastewater solutions is quite complex and could integrate individual or even parallel running steps like adsorption (chemisorption), precipitation, intercalation, and ion exchange (cation and anion).

A study focused on removing other ions by HT sorption implies that, besides Ca²⁺ and Cl⁻ ions, calcined HT also quite efficiently removes Pb and Cr (82–91%) and partially removes sulphates (40%) after 24 h, as shown from sorption results for EAFD wastewater.

Moreover, this study brings a new challenge concerning increasing the sorption rate for Cl⁻ and Ca²⁺ ions and increasing the sorption capacity of Hydrotalcite. This could be achieved, for example, by the Mg/Al ratio, the time of crystallisation, the temperature of synthesis and co-precipitation, and the temperature and time of calcination, which are very important factors influencing the final effectivity of Hydrotalcite in the sorption of chlorides and calcium. Knowledge of the desorption process of adsorbed impurities, regeneration possibilities, and further reuse is required for its full-scale application. The sustainability concept would also support the research focused on preparation possibilities of HT from secondary sources like, for example, acid and alkaline waste solutions from surface finishing operations in the aluminium industry, as it was used for zeolite preparation in this study.

As Hydrotalcite seems to be an auspicious material for the sorption of Ca²⁺ and Cl⁻ from wastewater such as EAFD wastewater, further sorption studies would be necessary, including investigation of the effect of HT dosage, temperature, and pH on the sorption process. A detailed study on the removal of other ions presented in EAFD wastewater and sorption thermodynamics and kinetics would also be required.