Resourcization of Argillaceous Limestone with Mn₃O₄ Modification for Efficient Adsorption of Lead, Copper, and Nickel

Abstract

Argillaceous limestone (AL) is comprised of carbonate minerals and clay minerals and is widely distributed throughout the Earth’s crust. However, owing to its low surface area and poorly active sites, AL has been largely neglected. Herein, manganic manganous oxide (Mn₃O₄) was used to modify AL by an in-situ deposition strategy through manganese chloride and alkali stepwise treatment to improve the surface area of AL and enable its utilization as an efficient adsorbent for heavy metals removal. The surface area and cation exchange capacity (CEC) were enhanced from 3.49 to 24.5 m²/g and 5.87 to 31.5 cmoL(+)/kg with modification, respectively. The maximum adsorption capacities of lead (Pb²⁺), copper (Cu²⁺), and nickel (Ni²⁺) ions on Mn₃O₄-modified argillaceous limestone (Mn₃O₄–AL) in mono-metal systems were 148.73, 41.30, and 60.87 mg/g, respectively. In addition, the adsorption selectivity in multi-metal systems was Pb²⁺ > Cu²⁺ > Ni²⁺ in order. The adsorption process conforms to the pseudo-second-order model. In the multi-metal system, the adsorption reaches equilibrium at about 360 min. The adsorption mechanisms may involve ion exchange, precipitation, electrostatic interaction, and complexation by hydroxyl groups. These results demonstrate that Mn₃O₄ modification realized argillaceous limestone resourcization as an ideal adsorbent. Mn₃O₄-modified argillaceous limestone was promising for heavy metal-polluted water and soil treatment.

1. Introduction

Over the years, large amounts of wastewater containing lead (Pb²⁺), copper (Cu²⁺), nickel (Ni²⁺), and other heavy metal ions have been discharged into surface and underground water. These heavy metals, originating from metal electroplating, mining, battery manufacturing, chemical industries, and other industries [1], are widely recognized as one of the most severe threats to human health, animal well-being, and plants, due to their carcinogenicity, toxicity, non-degradability, persistence in the environment, and ability to bioaccumulate within food chains [2,3,4]. Consequently, there is an urgent need for effective removal of heavy metal ions from wastewater.

One of the most efficient techniques for remediating water contaminated with heavy metals is adsorption due to its advantages, such as availability, cost-effectiveness, and ease of operation [5,6]. The key aspect of the adsorption technique is to select an eco-friendly, inexpensive, and efficient adsorbent. Natural clays, along with other minerals, are considered green adsorbents due to their affordability and abundance as a natural resource [7]. At present, various studies have focused on utilizing minerals for wastewater treatment. Du et al. (2016) investigated the absorption behavior of Pb²⁺ and Cd²⁺ in water using a bacteria–montmorillonite composite material [8]. Zhao et al. (2022) synthesized a silicate-based composite material and demonstrated its remarkable capacity to adsorb Cd²⁺ (63.80 mg/g) [9].

Argillaceous limestone (AL), comprised of carbonate and clay minerals, is widely distributed throughout the Earth’s crust. Due to the precipitation properties of the carbonate minerals and the high cation exchange capacity of the clay minerals, both are effective in removing heavy metals [7,10,11]. However, research on AL as a heavy metal adsorbent remains relatively limited [12]. This is primarily attributed to its low content of functional components, low surface area, and poor active sites [13]. Therefore, it is necessary to modify or functionalize AL before it can be applied as a heavy metal adsorbent. Additionally, AL commonly coexists with grey limestone as an associated mineral. After extracting grey limestone for building materials purposes, AL is often abandoned [14]. Therefore, AL has the advantage of low cost. The modification of AL as an adsorbent for heavy metal removal holds great significance for resource utilization.

Manganese oxide-modified adsorbents have attracted researchers’ attention due to their high electronegativity, surface area, and other unique physicochemical properties [6,15,16,17,18,19]. Moreover, manganese oxides can be firmly loaded onto silica-based surfaces [18]. Additionally, manganese oxides are abundant in geological reserves and therefore, also serve as a low-cost material [19]. To date, reports have confirmed the potential application of Mn₃O₄ as a good adsorbent to remove metal ions from water. Lingamdinne et al. (2022) found that Mn₃O₄-anchored reduced graphene oxide had significant adsorption efficiency for lead and chromium ions and the maximum sequestration capacities were 130.28 and 138.51 mg/g, respectively [6]. In addition, the research of Zou et al. (2016) showed that the maximum adsorption of antimony ions on Mn₃O₄-modified reduced graphene oxide was 151.84 mg/g [17]. Lee et al. prepared Mn₃O₄-coated activated carbon (Mn₃O₄/AC) for use in the removal of Pb and Cu ions, and showed that the adsorption capacities of Mn₃O₄/AC were enhanced 2.2 and 6.1 times for Pb and Cu ions, respectively, compared with pure active carbon [15]. However, to the best of our knowledge, limited studies have been conducted on the adsorption of heavy metals by the Mn₃O₄-modified AL (Mn₃O₄–AL). Previous research has indicated that environmental factors, such as pH, ionic strength, and organic matter, can influence the adsorption capacity of materials. It has been demonstrated that an increase in pH significantly enhances the adsorption of heavy metals [6,20]. Nonetheless, the effect of coexisting ions on adsorption is inconsistent and may hinder or promote the adsorption process [21]. In addition, it has been reported that humic acid can affect the adsorption behavior of heavy metal ions by mineral or synthetic adsorbents [22]. Therefore, it is crucial to investigate the effects of these environmental factors on adsorption for practical applications.

In this study, we modified AL with Mn₃O₄ using an in-situ deposition strategy through a manganese chloride (MnCl₂) and alkali stepwise treatment to improve its adsorption capacity. Consequently, the purposes of this study were to: (1) synthesize and characterize the Mn₃O₄–AL; (2) explore the capacity of Mn₃O₄–AL to adsorb Pb²⁺, Cu²⁺, and Ni²⁺ in mono- or multi-heavy metal systems; (3) examine the impact of environmental factors (pH, ionic strength, and humic acid) on adsorption; and (4) clarify the adsorption mechanism of the adsorption process. The results obtained from this research are expected to provide valuable insights into understanding the ability of Mn₃O₄–AL to adsorb Pb²⁺, Cu²⁺, and Ni²⁺ and contribute towards resource utilization of AL. This work indicated that the Mn₃O₄–AL was a promising candidate for efficient heavy metal removal.

2. Materials and Methods

2.1. Materials and Chemicals

Argillaceous limestone (AL) was collected from the lower Cambrian Mantou formation in the Dabeiwang area of Jiangsu province, China. Our previous research showed that the raw minerals are mainly composed of carbonate minerals, accounting for 41.8%. In addition, clay minerals account for around 12.2% [23].

Sodium nitrite (NaNO₃), calcium nitrite (Ca(NO₃)₂), nitric acid (HNO₃, 70% (w/w)), sodium hydroxide (NaOH), hydrochloric acid (HCl, 37% (w/w)), manganese chloride (MnCl₂), copper nitrate pentahydrate (Cu(NO₃)₂·5H₂O), zinc nitrate hexahydrate (Zn(NO₃)₂·6H₂O), and cadmium nitrate tetrahydrate (Cd(NO₃)₂·4H₂O), lead nitrate (Pb(NO₃)₂), and nickel nitrate hexahydrate (Ni(NO₃)₂·6H₂O, 99.99%) of analytical grade were purchased from Aladdin Reagent Database Inc. (Shanghai, China). Humic acid (HA) was purchased from Sigma-Aldrich (St Louis, MO, USA).

2.2. Modification of Argillaceous Limestone

The modification process for argillaceous limestone (AL) was carried out following the previous study of Sarı et al. (2012) [19]. The synthesis pathway of modifying argillaceous limestone with Mn₃O₄ is shown in Figure 1a. Detailed synthesis steps are as follows: 20 g AL was activated with 80 mL of 6 mol/L NaOH at 90 °C for 2 h. Next, the obtained mixture was added to a 100 mL solution of 2.5 mol/L MnCl₂ (with the pH adjusted to 2.0 using 1 M HCl), and placed at room temperature (25 °C) for 10 h. The mixture was filtered and the resulting precipitate was mixed with 40 mL of a 6 mol/L NaOH solution at 25 °C for 10 h to obtain a manganese hydroxide (Mn(OH)₂) precipitate. The mixture was exposed to the air to oxidize the manganese hydroxide. Then, the modified argillaceous limestone was washed with deionized water and dried at 105 °C.

2.3. Batch Adsorption Experiments

Stock solutions of 1000 mg/L Pb²⁺, Cu²⁺, and Ni²⁺ were prepared using Pb(NO₃)₂, Cu(NO₃)₂·5H₂O, Ni(NO₃)₂·6H₂O. The desired pH of the suspensions was controlled using 0.1 mol/L NaOH or HNO₃. NaNO₃ and Ca(NO₃)₂ were used as background electrolytes to adjust the ionic strength of the suspensions. Adsorption kinetic experiments were conducted by adding 0.0075 g of the Mn₃O₄–AL into 50 mL solutions containing either 50 mg/L Pb²⁺, Cu²⁺, or Ni²⁺ or all three metals. The mixtures were shaken at 25 °C for 0–1080 min. To investigate the impact of the initial concentration and to compare the mono- and multi-heavy metal adsorption capacity of the Mn₃O₄–AL, adsorption isotherm experiments were carried out in multi-heavy metal and single-heavy metal solutions with initial concentrations in the range of 1–200 mg/L. Various concentrations of environmental factors, including pH (2.0, 3.0, 4.0, 5.0, 6.0), ionic strength (sodium ions(Na⁺), calcium ions (Ca²⁺), chloride ions (Cl⁻), and sulfate ions (SO₄²⁻); 0, 0.01, and 0.1 mol/L), and humic acid (HA) (0, 0.1, 0.5, 1, 3, 5, 8, 10, 15, 20, 30 mg/L), were individually controlled to investigate their influence on heavy metal adsorption. In the isotherm and environmental factor experiments, the mixture was shaken for 12 h, and the other details were the same as in the adsorption kinetic experiments.

In addition, to explore the adsorption capacity of the Mn₃O₄–AL in the natural environment, lake water was used as the background water to simulate natural environment conditions. The lake water was collected from Zhaoyang Lake, at the South China Agricultural University, Guangdong Province, China. The pH of the lake water was 7.28, and the dissolved organic carbon (DOC) was 4.02 mg/L. The initial concentrations of total Cd, Pb, Cu, Ni, and Zn of the lake water were 0.08, 1.36, 1.88, 2.06, and 1.22 μg/L, respectively, which was below the environmental quality standards for surface water in China (GB 3838-2002) [24]. The mixed solution was prepared using Pb(NO₃)₂, Cu(NO₃)₂·5H₂O, Ni(NO₃)₂·6H₂O, Zn(NO₃)₂·6H₂O, and Cd(NO₃)₂·4H₂O. The initial concentrations of Pb²⁺, Cu²⁺, Ni²⁺, Cd²⁺, and Zn²⁺ in the multi-metal solution were each 50 mg/L. The mixture was shaken for 12 h, and other details were the same as in the adsorption kinetic experiments.

After shaking, all samples were immediately filtered with 0.22 µm filters. The concentrations of Pb²⁺, Cu²⁺, and Ni²⁺ were determined using flame atomic absorption spectroscopy (AAS) (Z2300, Hitachi, Tokyo, Japan). To ensure the reproducibility and accuracy of data, all of the experiments in this study were performed in triplicate.

The pseudo-first-order and pseudo-second-order models were applied to fit the adsorption kinetics of Pb²⁺, Cu²⁺, and Ni²⁺ onto the Mn₃O₄–AL (Equations (1) and (2)) [21].

$$\log \left({\mathit{Q}}_{\mathit{e}}{-\mathit{Q}}_{\mathit{t}}\right){=\log \mathit{Q}}_{\mathit{e}}-\frac{{\mathit{k}}_1\mathit{t}}{2.303}$$

(1)

$$\frac{\mathit{t}}{{\mathit{Q}}_{\mathit{t}}}=\frac{1}{{\mathit{k}}_2{\mathit{Q}}_{\mathit{e}}^2}+\frac{\mathit{t}}{{\mathit{Q}}_{\mathit{e}}}$$

(2)

where Q_e and Q_t are the adsorption capacity at equilibrium and time t (mg/g), respectively. k₁ is the rate constant of pseudo-first-order (min⁻¹), and k₂ is the rate constant of pseudo-second-order (g/(mg min)).

In addition, the Langmuir and Freundlich models (Equations (3) and (4)) were used to simulate the adsorption isotherm data [6].

$$\frac{{\mathit{C}}_{\mathit{e}}}{{\mathit{Q}}_{\mathit{e}}}=\frac{1}{{\mathit{K}}_{\mathit{L}}{\mathit{Q}}_{\mathit{m}}}\times \frac{{\mathit{C}}_{\mathit{e}}}{{\mathit{Q}}_{\mathit{m}}}$$

(3)

$${\log \mathit{Q}}_{\mathit{e}}{=\log \mathit{K}}_{\mathit{F}}+\frac{1}{\mathit{n}}{\log \mathit{C}}_{\mathit{e}}$$

(4)

where, C_e is the equilibrium concentration of the heavy metal ions in solution (mg/L). Q_e and Q_m are the equilibrium and maximum adsorption capacity, respectively (mg/g). K_L is the Langmuir constant (L/mg), K_F is the Freundlich constant (mg^1−(1/n) L^1/n/g), and 1/n is the heterogeneity factor.

2.4. Characterization Method

The mineralogical compositions of the Mn₃O₄–AL before and after adsorption were analyzed by X-ray diffraction (XRD) (D8 Advance, Bruker, Germany). The morphologies of the Mn₃O₄–AL before and after adsorption were characterized using scanning electron microscopy (SEM) and X-ray energy dispersive spectrometry (EDS) (Quanta FEG 650, FEI, Houston, TX, USA). The surface area of the Mn₃O₄–AL was measured using a gas sorption analyzer (Gemini VII 2390 V1.03, Micromeritics Instrument Corp, Norcross, GA, USA) by the Brunauer–Emmett–Teller (BET) method. A FT-IR instrument (Nicolet iS10, Thermo Fisher, Waltham, MA, USA) was used to obtain the change of functional groups before and after adsorption. Surface chemical compositions of the Mn₃O₄–AL before and after adsorption were evaluated by X-ray photoelectron spectroscopy (XPS) (ESCALAB 250Xi, Thermo Fisher, USA). The zeta potential of the Mn₃O₄–AL under different background solutions was analyzed by a dynamic light scattering analyzer (Nano-ZS90 Zetasizer, Malvern Instruments, Malvern, UK). The total organic carbon (TOC) of the Mn₃O₄–AL was determined with a TOC analyzer (Vario TOC, Elementar, Langenselbold, Germany). The cation exchange capacity (CEC) of the Mn₃O₄–AL was determined based on He et al. (2018) [23]. The pH and dissolved organic carbon (DOC) were determined using the procedure of Zhang et al. (2017) [21].

2.5. Statistical Analysis

Data are presented as the mean ± standard deviation (SD). All data were statistically assessed by one-way ANOVA with the minimum level of significance set at p < 0.05. In addition, all of the functional graphs were produced using OriginPro 2021 software.

3. Results and Discussion

3.1. Characterization of Mn₃O₄ Modified Argillaceous Limestone

To obtain a satisfactory and economical adsorbent, raw argillaceous limestone was modified by in-situ deposition of Mn₃O₄ from a MnCl₂ precursor solution with alkalized treatment. The morphology and chemical composition of the Mn₃O₄–AL were analyzed using SEM and EDS (Figure 1b,c). Previous research indicated that raw argillaceous limestone exhibited dense massive structures [23]. However, after Mn₃O₄ modification, the surface of the argillaceous limestone became fluffy and featured the distribution of numerous small granular particles. This transformation can be attributed to the acidic atmosphere created by the strong acid and weak alkali salt MnCl₂, which dissolved carbonates. Consequently, argillaceous limestone was broken down into smaller particles, leading to an increase in its specific surface area. This observation was further supported by a significant rise in the specific surface area from 3.49 to 24.5 m²/g after modification (Table S1). EDS spectra results revealed that the Mn₃O₄–AL mainly consisted of manganese (Mn), silicon (Si), oxygen (O), calcium (Ca), magnesium (Mg), aluminum (Al), potassium (K), and carbon (C) (Figure 1c). Notably different from our previous research on AL [23], there was a distinct peak corresponding to Mn present in the EDS of the Mn₃O₄–AL. In contrast, there is no Mn peak in AL. These findings indicated the successful loading of Mn onto raw argillaceous limestone.

The FT-IR spectra of AL and the Mn₃O₄–AL are depicted in Figure 2a. For the AL, the broad band at 3437 cm⁻¹ can be attributed to the stretching vibration of –OH groups [25,26]. The peaks at 1396 cm⁻¹ and 1615 cm⁻¹ were the characteristic peaks for bCO₃²⁻ and mCO₃²⁻, respectively [21,27]. Negatively charged functional groups (–OH and COO–) existed on the surface of AL, which facilitates the adsorption of heavy metal ions through electrostatic adsorption and complexation reactions [21]. The bands at 1008 cm⁻¹ and 521 cm⁻¹ refer to Si–O symmetric stretching [28,29]. After modification, the signals of –OH and COO– were still maintained in the Mn₃O₄–AL curve. The intensity of CO₃²⁻ peak at 1473 cm⁻¹ became weaker, indicating partial dissolution of the carbonate mineral [30]. This is consistent with the result of the increasing specific surface area of the Mn₃O₄–AL. Additionally, a new peak appeared at 635 cm⁻¹, which is consistent with the Mn–O–Mn stretching vibration [31,32,33]. It demonstrated that Mn was loaded onto the argillaceous limestone in the form of manganese oxide.

XPS was applied to analyze the chemical states of the Mn₃O₄–AL (Figure 2). The C 1s spectrum exhibited three characteristic peaks at 284.8, 286.3, and 288.8 eV (Figure 2c), corresponding to C–C, C–O–H/C–O–C, and O–C=O, respectively. Combined with the FT-IR result mentioned earlier, it can be concluded that the main form of C was carbonate. Three peaks were observed at 529.9, 531.2, and 532.3 eV in terms of O 1s high resolution (Figure 2d), corresponding to lattice oxygen (Mn–O–Mn), –OH, and C=O [34,35], respectively. The peaks at 641.7 eV and 652.7 eV in the detailed spectra of Mn 2p (Figure 2e) were attributed to 2p_3/2 and 2p_1/2 Mn (III), and the peaks at 644.5 eV and 653.6 eV were attributed to Mn(II) [17,36,37]. Combined with the results of FT-IR, we can speculate that Mn exists in the form of Mn₃O₄ in the Mn₃O₄–AL [35].

The mineralogical compositions of AL before and after modification with Mn₃O₄ were determined by XRD (Figure 2f). The XRD curve depicted that AL was mainly comprised of dolomite (CaMg(CO₃)₂) (2θ = 31.20°) (PDF No. 03-0864) and quartz (PDF No. 82-0513). The peaks at 2θ = 28.98° and 35.98° were regarded as akermanite (Ca₂MgSi₂O₇) (PDF No. 35-0592). With the treatment of NaOH and MnCl₂ sequentially, the location of the dominating diffraction peaks in the Mn₃O₄–AL had not obviously changed, but the relative intensity of those peaks weakened, illustrating that part of the SiO₂ and carbonate minerals had reacted. The diffraction peaks at 2θ = 45.4° and 57.6° were related to hausmannite (Mn₃O₄) (PDF No. 24-0734), which confirmed that Mn₃O₄ had been loaded onto the argillaceous limestone [15]. These results were consistent with those of XPS and FT-IR.

The other physiochemical characterizations of Mn₃O₄–AL are presented in Table S1. With modification, the pH and point of zero charge (PZC) were changed from 8.32 (AL) to 7.92 (Mn₃O₄–AL) and from 1.95 (AL) to 3.20 (Mn₃O₄–AL), respectively. The specific surface area was increased from 3.49 (AL) to 24.5 m²/g (Mn₃O₄–AL) which was consistent with the morphology of the Mn₃O₄–AL as evidenced by SEM (Figure 1a). Those changes were mainly ascribed to the addition of a large amount of acidic MnCl₂ solution in the preparation process of Mn₃O₄–AL, which etched the surface of AL by the neutralization of some carbonate and hydroxyl functional groups. The increased specific surface area can expose more adsorption sites on the surface, which is conducive to the adsorption of heavy metals. In addition, the initial content of total Pb, Cu, and Ni in the Mn₃O₄–AL were 38.1, 12.1, and 20.9 mg/kg, respectively, which were below the environmental quality standard for soils in China (GB 15618-2018) [38]. Therefore, the environmental risk during application of Mn₃O₄–AL can be negligible. Moreover, the cation exchange concentration (CEC) of the Mn₃O₄–AL and AL were 31.5 and 5.87 cmoL(+)/kg, respectively. The higher CEC of the Mn₃O₄–AL could be ascribed to the increased exposure of ion exchange sites, which were not available before Mn modification. The increased CEC value of the Mn₃O₄–AL can further enhance the adsorption capacity of Pb²⁺, Cu²⁺, and Ni²⁺ [21].

3.2. Adsorption Behavior

3.2.1. Adsorption Kinetics

Adsorption kinetics is a significant indicator for evaluating the potential of adsorbents for practical applications. A rapid adsorption process could reduce the cost of remediation. The kinetics and fitting curves of Pb²⁺, Cu²⁺, and Ni²⁺ were investigated in a multi-metal system. The results are depicted in Figure 3, while the corresponding parameter values and correlation coefficients are presented in

Table 1. Kinetic parameters of pseudo-first-order and pseudo-second-order models for competitive adsorption.

Adsorbates	Experimental Qe (mg/g)	Pseudo-First-Order Model			Pseudo-Second-Order Model
		Qe	k1	R2	Qe	k2	R2
		(mg/g)	(min−1)		(mg/g)	(g/(mg min))
Pb	42.48	40.91	0.090	0.743	42.48	0.00092	0.928
Cu	17.31	17.17	0.043	0.895	17.32	0.00519	0.989
Ni	6.272	5.800	0.023	0.926	6.338	0.00449	0.979

. Consistent with previous research, the adsorption rate of Pb²⁺, Cu²⁺, and Ni²⁺ in a multi-metal system increased rapidly during the first 120 min, followed by a slower adsorption stage before reaching equilibrium [39]. Finally, the adsorption equilibrium was reached after 360 min and the adsorption capacity stabilized at 42.48, 17.31, and 6.272 mg/g for Pb²⁺, Cu²⁺, and Ni²⁺, respectively. This was mainly attributed to the large number of adsorption sites on the Mn₃O₄–AL adsorbent and the higher driving force of concentration gradients in the adsorption primary stage of adsorption. As the free adsorption sites were gradually occupied and the repulsive force on the solid–liquid interface increased, the adsorption rate decreased until equilibrium was reached [21,40]. Compared to pseudo-first-order kinetics, the pseudo-second-order kinetic model fit the adsorption kinetic behavior of Pb²⁺, Cu²⁺, and Ni²⁺ on the Mn₃O₄–AL in a multi-metal system more closely, as indicated by the higher regression coefficient (R²). The theoretical equilibrium adsorption values (Q_e) were closer to experimental values (Table 1). The results indicate that chemical adsorption may be the rate-controlling mechanism for the adsorption of Pb²⁺, Cu²⁺, and Ni²⁺ onto Mn₃O₄–AL in multi-metal systems.

3.2.2. Adsorption Isotherm

To gain information on the mechanisms of the adsorption effect, experiments on adsorption isotherms in mono- and multi-heavy metal systems were conducted. Figure 4 depicts the adsorption isotherms and fitting curves of the Langmuir and Freundlich models for adsorption of mono- and multi-heavy metal ions (Pb²⁺, Cu²⁺, and Ni²⁺) by the Mn₃O₄–AL. The corresponding parameters are shown in Table S2. In mono-metal systems, the R² values obtained using the Langmuir model were higher than that of the Freundlich model, and the Q_m for Pb²⁺, Cu²⁺, and Ni²⁺ were 148.73, 41.30, and 60.87 mg/g (Table S2), respectively. Compared to the Freundlich fitting results, the experimental data could be better fitted by the Langmuir model, suggesting the significant role of the monolayer adsorption process in the Mn₃O₄–AL adsorption of Pb²⁺, Cu²⁺, and Ni²⁺.

In the multi-metal system, the adsorption of Pb²⁺, Cu²⁺, and Ni²⁺ by the Mn₃O₄–AL could also be better described by the Langmuir model due to its higher R² values compared to the Freundlich model. In addition, the Q_m values of Pb²⁺, Cu²⁺, and Ni²⁺ were 56.96, 23.77, and 5.78 mg/g, respectively. These values represent 38%, 58%, and 9% of their values in mono-metal systems (Figure 4d). The adsorption selectivity of the Mn₃O₄–AL in the multi-metal system was in the order of Pb²⁺ > Cu²⁺ >Ni²⁺, indicating that the Mn₃O₄–AL exhibited better adsorption selectivity for Pb²⁺. Studies have shown that metal ions with higher electronegativity or a smaller hydrated radius exhibit a greater affinity for adsorption [41,42]. Metal ions with a smaller hydrated radius may have smaller steric hindrance during transport and adsorption to the inner surface of adsorbents [42]. The hydrated ionic radius and hydration energy of the three heavy metal ions were in the order of Pb²⁺ > Cu²⁺ >Ni²⁺ [41,43]. Evidently, the adsorption selectivity of the Mn₃O₄–AL was in accord with the hydrated ionic radius.

The maximum adsorption capacities of the Mn₃O₄–AL in this study were compared with those of other mineral adsorbents (Table S3) [21,23,29,44,45,46,47,48,49,50,51,52,53,54,55,56,57,58]. Although compared to CAL (chitosan-coated argillaceous limestone) or FAL (Fe³⁺-modified argillaceous limestone) [21,23], the Mn₃O₄–AL had lower adsorption of Cu²⁺ and Pb²⁺, the Mn₃O₄–AL had satisfactory adsorption of Ni²⁺. Furthermore, the adsorption capacities of the Mn₃O₄–AL for the three metals were greater than those of most natural and inorganically-modified mineral adsorbents, such as palygorskite, natural and Na-bentonite, goethite, Na-montmorillonite, and zeolite. Some organic material-modified minerals showed higher adsorption capacity than the Mn₃O₄–AL, such as cationic surfactant-modified bentonite, polyacrylamide/sodium montmorillonite, and sodium polyacrylate-grafted bentonite. However, the inorganic modifier has the advantage of being more cost-effective and naturally abundant compared to organic modifiers [21]. Therefore, the Mn₃O₄–AL is a promising adsorbent for removing heavy metals.

3.2.3. Effects of Environmental Factors

The initial solution pH is a significant factor influencing the adsorption process. Acidity can impact the surface charge of the adsorbent and adsorbate and ultimately affects the ability of the adsorbent to adsorb pollutants [42]. The pH of the solution may affect its surface charge due to the exchange of H⁺ [59]. Therefore, the effect of varying the initial solution pH on the adsorption capacity of the Mn₃O₄–AL was investigated to explore the adsorption mechanisms. Figure 5a shows the adsorption behaviors of mono- and multi-heavy metal ions (Pb²⁺, Cu²⁺, Ni²⁺) onto the Mn₃O₄AL from pH 2.0 to 6.0. The adsorption capacity (Q_e) for Pb²⁺, Cu²⁺, and Ni²⁺ by the Mn₃O₄–AL increased from pH 2.0 to 4.0 and reached a plateau at a pH value of 6.0. At low pH, the adsorption of Pb²⁺, Cu²⁺, and Ni²⁺ was inhibited due to the competitive adsorption between H⁺ and Pb²⁺, Cu²⁺, and Ni²⁺ for adsorption sites [60]. In addition, Zeta potential analysis showed that Mn₃O₄–AL in deionized water carried a positive charge at pH levels below 4 (Figure 5b). Therefore, the electrostatic repulsion between the Mn₃O₄–AL and Pb²⁺, Cu²⁺, Ni²⁺ also resulted in a reduction of adsorption capacity. As the initial pH increased beyond the pH_pzc, the surface charge of the Mn₃O₄–AL became negative [61]. Therefore, a stronger electrostatic attraction increases the adsorption capacity through electrostatic attraction [62]. In addition, the surface functional groups of the adsorbent were protonated at low pH, which weakened the complexation between the adsorbent and the heavy metal ions [61].

The adsorption of Pb²⁺, Cu²⁺, and Ni²⁺ by the Mn₃O₄–AL at different ionic strengths and alkali ions (i.e., Na⁺ and Ca²⁺) is presented in Figure 5c–e. The concentration of alkali ions had little effect on the adsorption of Cu²⁺ by the Mn₃O₄–AL. This is similar to the influence of ionic strength on Cu adsorption by Na-bentonite [63] and carbon nanotube-hydroxyapatite [64]. The adsorption capacity of Pb²⁺ and Ni²⁺ onto the Mn₃O₄–AL decreased with increasing concentrations of Na⁺ and Ca²⁺. A similar trend was reported in a previous study, where montmorillonite was employed for heavy metal adsorption [65]. This phenomenon mainly occurs because the adsorbed Pb²⁺ and Ni²⁺ may form outer-sphere surface complexes with the Mn₃O₄–AL. The electrolytes may compete with Pb²⁺ and Ni²⁺ for active sites on the Mn₃O₄–AL. In addition, the electrolytes may compress the electrical double layer thickness of the Mn₃O₄–AL, resulting in a decrease in the electrostatic attraction force between the Mn₃O₄–AL and heavy metal ions. This speculation could be supported by the higher inhibitory effect of Ca²⁺ as compared with Na⁺. Divalent ions can compress the thickness of the electric double layer more effectively than monovalent ions, preventing Pb²⁺ and Ni²⁺ from reaching the adsorbent surface [21]. Moreover, increasing the ionic strength may improve the aggregation of the Mn₃O₄–AL [66,67], which may further lead to a decrease in available adsorption sites of the Mn₃O₄–AL. In addition, the effects of different anionic concentrations were investigated and are shown in Figure S1. The concentration of Cl⁻ had little effect on the adsorption of Pb²⁺, Cu²⁺, and Ni²⁺ onto the Mn₃O₄–AL. The adsorption capacity of Pb²⁺, Cu²⁺, and Ni²⁺ on the Mn₃O₄–AL increased with increasing concentrations of SO₄²⁻, primarily due to the formation of sulfate [23].

Humic acid (HA), a component of humic substances, is widely distributed in surface and groundwater. The ability of HA to complex with heavy metals can enhance the migration of heavy metal ions in solution and impact the removal of heavy metal ions [68]. Figure 5f illustrates the impact of HA on the heavy metal adsorption capacity of the Mn₃O₄–AL. The adsorption of Pb²⁺ and Cu²⁺ by the Mn₃O₄–AL increased as the concentration of HA increased, and reached a plateau at 15 mg/L. However, the adsorption of Ni²⁺ onto the Mn₃O₄–AL decreased as the concentration of HA increased. This distinct difference in the effect of HA can be explained by the following reasons. On the one hand, HA contains numerous functional groups, including phenols (–OH) and carboxylic groups (–COOH), which can offer extra adsorption sites and improve the adsorption of heavy metals [69]. Heavy metals can react with HA in a solution to create stable complexes [70,71]. Chaturvedi et al. (2007) also reported that humic acid can enhance the adsorption capacity of heavy metals on mineral surfaces through the formation of ternary mineral surface–metal–organic ligand complexes [72]. Therefore, the formation of Mn₃O₄–AL–HA–metal ion ternary complexes may enhance the adsorption of heavy metals. Research has shown that the affinities of these ions followed the order of Pb²⁺ > Cu²⁺ > Ni²⁺ [71]. The promotional effect of HA on the adsorption of Pb²⁺ and Cu²⁺ is greater than that of Ni²⁺. On the other hand, studies have shown that HA with a large size and low solubility, can accumulate at the solid/water interface to form a coating on the solid phase. HA attached to the solid phase surface will occupy adsorption sites and reduce the adsorption of heavy metals [73]. Therefore, we hypothesize that the HA adsorbed onto the Mn₃O₄–AL tends to combine with Pb²⁺ and Cu²⁺ rather than Ni²⁺ to form Mn₃O₄–AL–HA–Pb and Mn₃O₄–AL–HA–Cu complexes. This results in an increase in the adsorption capacity of Pb²⁺ and Cu²⁺ on the Mn₃O₄–AL with higher concentrations of HA, while the adsorption capacity of Ni²⁺ decreases.

To simulate natural environmental conditions, lake water was used as the solvent to investigate the adsorption capacity of the Mn₃O₄–AL for Pb²⁺, Cu²⁺, Ni²⁺, Cd²⁺, and Zn²⁺ in a multi-metal system (Figure 6). Results showed that the adsorption capacity of Pb²⁺, Cu²⁺, and Ni²⁺ decreased to 36.72, 13.25, and 4.35 mg/g, respectively. The adsorption capacity of these heavy metals followed the order of Pb²⁺ > Cu²⁺ > Ni²⁺ > Cd²⁺ > Zn²⁺, which is consistent with previous research using montmorillonite and Sphagnum peat [41,65]. These results indicated that the Mn₃O₄–AL can effectively remove various heavy metal ions from lake water-based simulated wastewater.

3.2.4. Adsorption Mechanisms

To gain further insight into the adsorption mechanisms of the Mn₃O₄–AL for Pb²⁺, Cu²⁺, and Ni²⁺, the chemical composition of the Mn₃O₄–AL after the adsorption process of multi-heavy metal ions (Pb²⁺, Cu²⁺, Ni²⁺) was determined using SEM-EDS, XRD, and XPS (Figure 7). Peaks of Pb, Cu, and Ni were obvious in the SEM-EDS curves after adsorption, indicating successful fixation of Pb, Cu, and Ni onto the Mn₃O₄–AL. The mineralogical composition of the Mn₃O₄–AL after competitive adsorption was characterized by powder XRD (Figure 7b). New diffraction peaks of hydrocerussite (2θ = 24.74°, 27.16°, and 51.0°) (PDF No. 28–0529) and cerussite (2θ = 44.1°, 48.94°, and 46.9°) (PDF No. 85–1088) appeared after adsorption, indicating the formation of Pb₃(CO₃)₂(OH)₂ and PbCO₃ [21]. The results indicated that the precipitation of lead carbonate plays a crucial role in the process of the Mn₃O₄–AL adsorbing Pb²⁺.

The metallic state of elements in the Mn₃O₄–AL after competitive adsorptions was analyzed using XPS analysis (Figure 7c–i). The presence of Na, Mn, O, Ca, C, and Si elements in the Mn₃O₄–AL before competitive adsorption can be observed (Figure 2b). After competitive adsorptions, three new peaks Cu 2p, Ni 2p, and Pb 4f appeared in the Mn₃O₄–AL (Figure 7c), indicating the successful adsorption of Pb²⁺, Cu²⁺, and Ni²⁺ [19]. Meanwhile, the disappearance of Na 1s after competitive adsorption demonstrated that ion exchange was an important mechanism of the ability of the Mn₃O₄–AL to adsorb Pb²⁺, Cu²⁺, and Ni²⁺ [20,21]. The high CEC of the Mn₃O₄–AL (31.5 ± 1.82 cmoL(+)/kg) can be attributed in part to the presence of clay mineral components, which also confirmed that the Mn₃O₄–AL has a high ion exchange potential [21]. In addition, we proved that alkali ions can significantly influence the adsorption of Ni and Pb in the study mentioned above (Figure 5). This further confirmed that ion exchange was the primary adsorption process of competitive adsorptions. After adsorption, the C 1s peak was observed at 284.8, 286.3, and 288.8 eV (Figure 7d), which was consistent with the peaks observed before adsorption, indicating that the existing form of carbonate had negligible change after adsorption. Whereas, the O 1s peak changed from 529.9, 531.2, and 532.3 eV to 530.3, 531.9, and 533.4 eV. The O 1s peak shifted to a higher binding energy after adsorption. This shift in binding energy position may be due to electron transfer between the heavy metal ions (M²⁺) and the Mn₃O₄–AL, indicating the formation of –Mn–O–M²⁺, –Si–O–M²⁺ and –COO–M²⁺ compounds [21,74,75]. Furthermore, detailed spectra of the peaks of Pb, Cu, and Ni are shown in Figure 7. The main peaks of Pb 4f_7/2 and Pb 4f_5/2 were observed at 138.4 eV and 143.3 eV, indicating the formation of lead complex compounds during adsorption [76]. The Pb 4f_7/2 peak occurred between the lead hydroxide (137.3 eV) and lead carbonate (138.7 eV) binding energy centers (Figure 7g), suggesting the formation of two lead species as a result of the adsorption by the Mn₃O₄–AL [76,77]. These findings were consistent with the XRD analysis results, which confirmed the formation of hydrocerussite (Pb₂(OH)₂(CO₃)₂) and cerussite (PbCO₃). These results demonstrate that the precipitation of lead carbonate played a crucial role in the adsorption process of Pb²⁺ by the Mn₃O₄–AL. In the XPS spectra of Cu (Figure 7h), peaks at 941.6 and 944.9 eV were attributed to shake-up satellite peaks of Cu 2p_3/2, which were caused by the charge transfer between Cu²⁺ and the ligand [77]. The peak at 962.3 eV confirmed the existing form of Cu(OH)₂, while peaks at 934.5 and 954.1 eV were attributed to the formation of copper complex compounds, Mn₃O₄–AL–O–Cu [77]. Therefore, complexation and precipitation were the main mechanisms of Cu²⁺ adsorption [78]. As for Ni, the peaks at 856.1 and 873.7 eV corresponded to the peaks of Ni 2p_3/2 and Ni 2p_1/2, which were attributed to Ni(OH)₂ [67,79]. The peaks that appeared at 860.2 eV belonged to NiO(OH) and 879.1 eV peak was attributed to the corresponding satellite peak of Ni 2p [67,80], indicating that precipitation was an important mechanism of Ni adsorption on Mn₃O₄–AL.

Furthermore, the introduction of negatively charged Mn₃O₄ makes AL more electronegative, leading to an electrostatic attraction between the Mn₃O₄–AL and heavy metal ions [81]. In addition, the increased specific surface area after Mn₃O₄ modification led to an increase in the number of contact sites on the Mn₃O₄–AL, which is favorable for the adsorption of Pb²⁺, Cu²⁺, and Ni²⁺.

In summary, the adsorption mechanisms of the Mn₃O₄–AL for Pb²⁺, Cu²⁺, and Ni²⁺ are as follows: (1) ion exchange between alkali ions and heavy metal ions; (2) precipitation to form PbCO₃, Pb₂(OH)₂(CO₃)₂, Cu(OH)₂, Ni(OH)₂, and NiO(OH) due to the large proportion of carbonate and other basic functional groups in the Mn₃O₄–AL; (3) electrostatic interaction between the Mn₃O₄–AL and heavy metal ions due to the high negative charge of Mn₃O₄; and 4) formation of heavy metal complexes Mn–O–M²⁺, –Si–O–M²⁺ and –COO–M²⁺ by hydroxyl groups on the clay mineral components of the Mn₃O₄–AL. The Mn₃O₄–AL exhibited excellent performance in removing Pb²⁺, Cu²⁺, and Ni²⁺, which could significantly contribute to the remediation of heavy metal-contaminated water and soil.

4. Conclusions

In this study, we provided a strategy for the resourcization of argillaceous limestone modified with Mn₃O₄. The CEC and specific surface area increased significantly after modification. Mn₃O₄-modified argillaceous limestone has a high adsorption capacity for Pb²⁺, Cu²⁺, and Ni²⁺ in both mono-metal and multi-metal systems. At pH 2–4, the adsorption capacity increased with increasing pH values. At pH 4–6, the adsorption capacity was maximized and remained stable, indicating that the Mn₃O₄–AL is suitable for removing heavy metals in weakly acidic soil or neutral wastewater. As the concentration of HA increased, the adsorption of Pb²⁺ and Cu²⁺ increased, but the adsorption of Ni²⁺ decreased. Higher ionic strength was not conducive to the removal of Pb²⁺, and Ni²⁺, but it had no significant effect on the removal of Cu²⁺. Ion exchange, electrostatic interaction, precipitation, and complexation were the primary adsorption mechanisms. Of course, due to the limitations of laboratory-scale experiments and the intricacy of natural soil, further research is needed to assess the potential of the Mn₃O₄–AL in remediating heavy metal-contaminated soil.