Tailoring the Structural and Morphological Properties of Biogenic Nano-Chlorapatite to Enhance the Capture Efficiency Towards Cr(VI)

Abstract

Nano-chlorapatite (nClAP) has been widely used as an efficient and environment-benign material to remediate heavy metal-contaminated water and soil. However, the adsorption capacities of nClAP to heavy metal oxyanions such as Cr(VI) are limited, which restricts its further application in environmental remediation. Herein, a novel carboxymethyl cellulose (CMC)-modified biogenic nClAP (CMC-nClAP) adsorbent was synthesized by a facile wet chemical method and used for Cr(VI) removal from water. The obtained CMC-nClAP materials were characterized by FTIR, XRD, TEM, and TGA analyses. Then, batch experiments were conducted to explore the effects of various factors such as the ratio of CMC and nClAP, pH, adsorbent dosage, adsorption time, and temperature on the adsorption process. The results revealed that the CMC-nClAP adsorbent displayed markedly improved stability against aggregation as well as Cr(VI) adsorption capacity as compared to that of the pristine nClAP. The Cr(VI) adsorption data obeyed the Langmuir isotherm model and pseudo-second-order kinetic model. Site energy distribution analyses revealed that Cr(VI) first occupied the high-energy sites and then diffused to the low-energy adsorption sites on the CMC-nClAP surface. Our experimental results indicated that the CMC-nClAP could be a promising material for the removal of Cr(VI) from water.

1. Introduction

Chromium (Cr), a highly toxic heavy metal, is primarily introduced into water systems via industrial discharges originating from electroplating, leather tanning, textile dyeing, and battery manufacturing processes [1]. While Cr(III) is an essential trace nutrient, Cr(VI) is a well-known carcinogen, teratogen, and mutagen with strong migratory capacities and solubility. It has detrimental effects on all life systems [2]. To reduce the possible harm caused by Cr to receiving water bodies and human health, many governments or organizations have imposed strict standards on the maximum contamination level (MCL) of total Cr and Cr(VI) [3]. In China, the MCL for the industrial effluents is set at 1.5 and 0.5 mg/L for total Cr and Cr(VI), respectively. Therefore, it is necessary to effectively remove Cr(VI) from water.

Since Cr(VI) cannot be biodegraded, methods such as adsorption [4], ion exchange [5], membrane separation [6], electrochemical precipitation [7], and chemical reduction and extraction [8] have been adopted for the elimination of Cr(VI) from wastewater. Among these technologies, adsorption has gained importance due to its low cost, good adaptability, simple operation, and fine efficiency [4]. Up to now, many kinds of adsorbents have been exploited and discussed for the adsorption of Cr(VI), such as carbonaceous nanomaterials [9], iron-based nanomaterials [10], and layered double hydroxides [11]. Although these novel adsorption materials exhibited strong Cr(VI) capture ability, some disadvantages existed for these materials such as the tedious synthesis process, high cost, and/or the use of some toxic additives [4,7,8]. Hence, it was extremely urgent to seek for low-cost, environmentally friendly, and biocompatible adsorbents with high adsorption capacity of Cr(VI) from aqueous media.

Chlorapatite [ClAP, Ca₅(PO₄)₃Cl] has been widely considered as a kind of non-toxic and eco-friendly material, which can be utilized in bone tissue engineering [12], water and wastewater treatment [13], and soil and sediment remediation [14]. Particularly, the use of nano-chlorapatite (nClAP) in the adsorption and immobilization of heavy metals gained importance recently. It was considered that the synthetic nClAP could convert various Pb(II) species into sparingly soluble pyromorphite [Pb₅(PO₄)₃Cl] through the dissolution–precipitation process [15]. Similarly, the dominant mechanism of Cd(II) and Cu(II) scavenging was due to the formation of Cd₅(PO₄)₃Cl and Cu₃(PO₄)₃ precipitates [16,17]. However, compared with the higher affinity towards metal cations, nClAP showed very limited adsorption capacity towards metal oxyanions such as Cr(VI). There are three main reasons for the poor uptake of Cr(VI) onto nClAP. First, Cr(VI) generally exists in the form of oxyanions such as Cr₂O₇²⁻, HCrO₄⁻, and CrO₄²⁻, which cannot be precipitated directly [18]. Second, Cr(VI) adsorption onto the apatite surface was primarily driven by electrostatic interaction or hydrogen bond formation or other types of weak physical interactions [19]. Third, the high surface area and energy of ClAP nanoparticles make them susceptible to aggregation in solutions, resulting in decreased adsorption capacity and restricting their environmental applications [20].

To prevent nanoparticle agglomeration, modification of nanomaterials with various stabilizers, support materials, or dispersants has been studied intensely. For instance, Hoch et al. [21] adopted a strategy to stabilize iron nanoparticles using cost-effective and eco-friendly carboxymethyl cellulose (CMC) and polyacrylate as stabilizers. The stabilized iron nanoparticles exhibited better reactivity as well as Cr(VI) adsorption capacity compared to non-stabilized counterparts. Similarly, sodium dodecyl sulfate [15], CMC [16], rhamnolipid [22], and sodium lignin sulfonate [23] have been applied as stabilizers to enhance the dispersibility of nClAP, and thus significantly promoted the removal efficiency for Pb²⁺, Cd²⁺, Zn²⁺, and Cu²⁺ ions. However, the possibility of using stabilized nClAP to improve Cr(VI) oxyanion removal has not been explored.

This study aims to prepare and evaluate the feasibility of aqueous Cr(VI) oxyanions removal using the CMC-modified nClAP (CMC-nClAP) adsorbent. CMC is chosen because it is a low-cost and environmentally friendly polyelectrolyte with abundant carboxylate and hydroxyl groups, which can effectively prevent agglomeration of nanoparticles and thereby facilitate preparation of highly stable nClAP particles [24]. The specific objectives are the following: to (1) prepare CMC-nClAP and characterize the adsorbent using X-ray diffraction (XRD), Fourier transform infrared (FTIR) spectroscopy, zeta potential determination, transmission electron microscopy (TEM), and thermogravimetric (TGA) analysis; (2) examine the influence of CMC concentration, pH, adsorbent dosage, adsorption time, and temperature on Cr(VI) removal efficiency; (3) elucidate the adsorption behavior and underlying mechanisms through kinetic, isotherms, and site energy distribution analyses.

2. Materials and Methods

2.1. Preparation of Biogenic nClAP and CMC-nClAP

In this work, CaO derived from waste eggshells is used as a calcium source and H₃PO₄ is used as the phosphorus precursor. According to the previous work [13], waste eggshells were cleaned, dried, crushed, and pulverized by using a porcelain mortar and pestle. The raw powders were then calcined at 1100 °C for 5 h to produce CaO, which further reacted with HCl under ambient conditions to produce a CaCl₂ solution [25]. Subsequently, nClAP was prepared using the CaCl₂ solution and H₃PO₄ (85% purity, Kemiou, Tianjin, China) via a precipitation method. The molar ratio of Ca:P was fixed to 5:3, where the H₃PO₄ solution was added to the calcium precursor dropwise under constant stirring. The pH of the solution was adjusted around 11 by using NH₃·H₂O. After the reaction was complete, the precipitated mixture was poured, filtrated, and thoroughly washed with deionized water and ethanol to remove residual impurities. The filtered product was vacuum-dried and ground into fine powders to obtain biogenic nClAP samples.

For the preparation of CMC-nClAP, different concentrations of CMC-Na aqueous solutions (0.25 wt%, 0.5 wt%, and 0.75 wt%) were mixed with the CaCl₂ solution. The concentration range of CMC was selected to avoid excessively high concentrations, which could lead to the sticky solution and hinder the preparation of composite materials. The subsequent procedures were identical to the preparation of pristine nClAP. The prepared products were noted by the ratio of CMC. For instance, nClAP, 0.25CMC-nClAP, 0.5CMC- nClAP, and 0.75CMC-nClAP referred to the nClAP obtained by adding 0, 0.25%, 0.5%, and 0.75% CMC, respectively.

2.2. Characterization

The crystal structures of nClAP samples were identified by XRD. The functional groups of the samples were determined using FTIR spectroscopy. Nanostructures of the samples were investigated using TEM, while the thermal behavior of the samples was examined by TG analyses. The zeta potential (ZP) of the adsorbent was determined using a Malvern Zetasizer (Nano-ZS90, Malvern Instruments Ltd., Malvern, UK). The detailed description of the characterization methods is provided in the Supplementary Materials.

2.3. Batch Adsorption Experiments

Adsorption experiments were conducted by mixing an appropriate amount of nClAP or CMC-nClAP adsorbent with controlled initial Cr(VI) concentrations and pH values. The volume of the Cr(VI) solution used in the experiments was 50 mL. The reaction flasks were then placed on the thermostatic shaker (Zhicheng ZWY-240, Shanghai, China) at constant temperature and 150 rpm speed. The adsorption studies were further performed under various conditions, such as the CMC concentration (0, 0.25 wt%, 0.5 wt%, and 0.75 wt%), pH (3–8), adsorbent dosage (0.05–1.0 g/L), adsorption time (0–120 min), and temperature (283, 298, and 313 K). Each experiment was performed in triplicate and average values were used. After the adsorption experiments, the mixture was centrifuged for 10 min at 5500 rpm (TDL-500C, Baiou, Jinan, China). Thereafter, supernatant samples were filtered and analyzed for Cr(VI) content by the spectrophotometric diphenylcarbazide method, measuring at 540 nm.

2.4. Adsorption Models

To analyze the adsorption mechanism, Langmuir and Freundlich models were used to fit the adsorption isotherms. Among the most common kinetic models, pseudo-first-order (PFO) and pseudo-second-order (PSO) models were used in this study to describe the adsorption kinetics of Cr(VI) onto the adsorbent surface. The equations are listed in the Supplementary Materials.

2.5. Approximate Site Energy Distribution Analysis

To explore the energetic characteristics of Cr(VI) and adsorbent interactions, the average adsorption energy (E^*) and site energy distribution (F(E^*)) of Cr(VI) on CMC-nClAP are described with the following functions [26].

$${\mathrm{q}}_{\mathrm{e}}\left({\mathrm{C}}_{\mathrm{e}}\right)={{\displaystyle \int }}_0^{+\infty }{\mathrm{q}}_{\mathrm{h}}\left(\mathrm{E},{\mathrm{C}}_{\mathrm{e}}\right)\mathrm{F}\left(\mathrm{E}\right)\mathrm{dE}$$

(1)

where q_h(E, C_e) is the homogeneous isotherm on the local adsorption sites with adsorption energy E, and F(E) is the site energy frequency distribution over a range of sites with homogeneous energies. E is the adsorption energy difference between the solute and solvent at the adsorption site.

The condensation approximation approach proposed by Cerofolini [27] was used to express the relationship between E^* and C_e.

$${ \mathrm{C}}_{\mathrm{e}}={\mathrm{C}}_{\mathrm{s}} \exp \left(-{\displaystyle \frac{\mathrm{E}-{\mathrm{E}}_{\mathrm{s}}}{\mathrm{RT}}}\right)={\mathrm{C}}_{\mathrm{s}} \exp \left(-{\displaystyle \frac{{\mathrm{E}}^{\ast }}{\mathrm{RT}}}\right)$$

(2)

where C_s is the maximum solubility of solute (K₂Cr₂O₇, 12.2 g in 50 mL water, 25 °C); C_e is the equilibrium concentration of Cr(VI) in solution; E (kJ/mol) is the adsorption energy at adsorption equilibrium; E_s is the adsorption energy corresponding to C_e = C_s, which is also regarded as the lowest physically realizable adsorption energy [26]. E^⁎ = E − E_S refers to the difference of adsorption energies between the solute and solvent to the adsorbent surfaces based on the reference point E_s.

According to the relationship between C_e and E^⁎, the adsorption isotherm model can be described as a function of q_e(E^⁎) related to E^⁎ [28]. Then, the approximate site energy distribution function F(E^*) can be obtained by deriving q_e(E^⁎). The formula is as follows:

$$\mathrm{F}\left({\mathrm{E}}^{\ast }\right)=-{\displaystyle \frac{{\mathrm{dq}}_{\mathrm{e}}\left({\mathrm{E}}^{\ast }\right)}{{\mathrm{dE}}^{\ast }}}={\displaystyle \frac{{\mathrm{C}}_{\mathrm{e}}}{\mathrm{RT}}}\times {\displaystyle \frac{{\mathrm{dq}}_{\mathrm{e}}\left({\mathrm{E}}^{\ast }\right)}{{\mathrm{dC}}_{\mathrm{e}}}}$$

(3)

where F(E^*) is the energy distribution function of the adsorption site (mg·mol/(g·kJ)); q(E^*) is the solute adsorption state concentration (mg/L).

Based on the Langmuir and Freundlich isotherm models, the distribution of adsorption sites is calculated using Formulas (4) and (5) [29]:

$$\mathrm{F}({\mathrm{E}}^{\ast })={\displaystyle \frac{{\mathrm{Q}}_{\mathrm{m}}{\mathrm{K}}_{\mathrm{L}}{\mathrm{C}}_{\mathrm{s}}}{\mathrm{RT}}}\exp \left({\displaystyle \frac{-{\mathrm{E}}^{\ast }}{\mathrm{RT}}}\right){\left[1+{\mathrm{K}}_{\mathrm{L}}{\mathrm{C}}_{\mathrm{s}}\exp \left({\displaystyle \frac{-{\mathrm{E}}^{\ast }}{\mathrm{RT}}}\right)\right]}^{-2}$$

(4)

$$\mathrm{F}({\mathrm{E}}^{\ast })=-{\displaystyle \frac{{\mathrm{dq}}_{\mathrm{e}}\left({\mathrm{E}}^{\ast }\right)}{{\mathrm{dE}}^{\ast }}}={\displaystyle \frac{\mathrm{d}\left[{\mathrm{K}}_{\mathrm{F}}{\left({\mathrm{C}}_{\mathrm{s}}{\mathrm{e}}^{-\frac{{\mathrm{E}}^{\ast }}{\mathrm{RT}}}\right)}^{\frac{1}{\mathrm{n}}}\right]}{{\mathrm{dE}}^{\ast }}}={\displaystyle \frac{{\mathrm{K}}_{\mathrm{F}}{{\mathrm{C}}_{\mathrm{s}}}^{\frac{1}{\mathrm{n}}}{\mathrm{e}}^{-\frac{{\mathrm{E}}^{\ast }}{\mathrm{nRT}}}}{\mathrm{nRT}}}$$

(5)

Furthermore, the average site energy μ(E^*) and the standard deviation (σ_e^*) are also calculated using the following equations.

$$\mathsf{\mu }\left({\mathrm{E}}^{\ast }\right)={\displaystyle \frac{{{\displaystyle \int }}_{{\mathrm{E}}_1^{\ast }}^{{\mathrm{E}}_2^{\ast }}{\mathrm{E}}^{\ast }\times \mathrm{F}\left({\mathrm{E}}^{\ast }\right){\mathrm{dE}}^{\ast }}{{{\displaystyle \int }}_{{\mathrm{E}}_1^{\ast }}^{{\mathrm{E}}_2^{\ast }}\mathrm{F}\left({\mathrm{E}}^{\ast }\right){\mathrm{dE}}^{\ast }}}$$

(6)

$$\mathsf{\mu }\left({\mathrm{E}}^{\ast 2}\right)={\displaystyle \frac{\left[{{\displaystyle \int }}_{{\mathrm{E}}_1^{\ast }}^{{\mathrm{E}}_2^{\ast }}{\mathrm{E}}^{\ast 2}\times \mathrm{F}\left({\mathrm{E}}^{\ast }\right){\mathrm{dE}}^{\ast }\right]}{\left[{{\displaystyle \int }}_{{\mathrm{E}}_1^{\ast }}^{{\mathrm{E}}_2^{\ast }}\mathrm{F}\left({\mathrm{E}}^{\ast }\right){\mathrm{dE}}^{\ast }\right]}}$$

(7)

$${\mathsf{\sigma }}_{\mathrm{e}}^{\ast }=\sqrt{\mathsf{\mu }\left({\mathrm{E}}^{\ast 2}\right)-\mathsf{\mu }{({\mathrm{E}}^{\ast })}^2}$$

(8)

where μ(E^*) and μ(E^*2) correspond to the mathematical expectation of E^* and E^*2, respectively. E₁^* and E₂^* are the minimum and maximum values of E^*.

3. Results and Discussion

3.1. Characteristics of Biogenic nClAP and CMC-nClAP

3.1.1. XRD Analysis

Figure 1 shows that the XRD patterns of all the nClAP samples, including those of 0.25CMC-nClAP, 0.5CMC-nClAP, and 0.75CMC-nClAP, are consistent with the patterns of crystalline nClAP reported in the literature [30]. The peak positions (2θ) of the nClAP diffraction peaks appeared at 26.25°, 32.46°, and 39.25°, which were in accordance with the (002), (211), and (130) reflection planes of the standard crystallography of chlorapatite (#27-0074) [13]. This similarity of XRD patterns between nClAP and CMC-nClAP samples suggested that CMC did not affect the basic structure of nClAP. However, a decrease in crystallinity with increasing CMC content was observed in Figure 1, presumably owing to the inhibitive effect on the growth of nClAP crystals [16,31].

3.1.2. FTIR Analysis

Figure 2 shows the FTIR spectra of synthesized nClAP and CMC-nClAP samples. As for the pristine nClAP, the band at 3440 cm⁻¹ was attributed to the stretching vibration of the -OH group, and the band at 1620 cm⁻¹ was due to the adsorbed water. The bands observed at 563 cm⁻¹ were due to ν₄ O-P-O in the phosphate group, and the bands of the ν₄ bending mode of the phosphate group were also observed at 603 cm⁻¹. The characteristics bands for ν₁ P-O in PO₄³⁻ and ν₃ P-O in PO₄³⁻ were observed at 961 and 1031 cm⁻¹, respectively [31]. In contrast, the relative intensities of the PO₄³⁻ peaks in all the CMC-nClAP samples slightly declined with the rising CMC ratio, as well as the existence of the bands at 2920 cm⁻¹ (asymmetric C-H stretch) in CMC-nClAP, suggesting the incorporation of CMC on nClAP [32]. Furthermore, the bands attributed to CO₃²⁻ groups were observed at 872, 1412, and 1462 cm⁻¹ in the nClAP sample [33], while the bands at 1608 and 1428 cm⁻¹ for CMC-CAP samples were ascribed to asymmetric and symmetric vibrations of the COO⁻ group [16].

3.1.3. TG Analysis

TG analysis of the prepared nClAP and CMC-nClAP samples was carried out to determine the amount of CMC associated with nClAP. The TG curve of nClAP (Figure 3a) showed two small weight losses at the beginning of the heating process and above 500 °C. The first weight loss was owing to the evaporation of water molecules, where this water adsorbed on the surface of nClAP and was not bound to the crystal structure. The second one was due to the removal of residual carbonates [34]. No obvious weight loss can be observed from 500 °C to 1000 °C. From the TG curve of CMC-nClAP samples shown in Figure 3a, it is found that the mass loss (about 5%) up to 250 °C may be ascribed to evaporation of bonded water, while fast mass losses of 25.5%, 26.7%, and 31.2% between 250 °C and 470 °C were due to the thermal decomposition of the CMC polymer [35]. The TG results could also be reflected by differential thermal analysis (DTA). The corresponding DTA curves (Figure 3b) exhibited two exothermic peaks with maxima at 350 °C and 685 °C that confirmed the structural transformations in the CMC-nClAP sample.

3.1.4. TEM Analysis

The TEM images of pristine nClAP showed significant agglomeration of rod-like nanoparticles (Figure 4), while the less agglomerated samples were obtained for 0.25CMC-nClAP and 0.5CMC-nClAP. These observations were consistent with previous studies conducted by He and Zhao [36], who indicated that CMC could prevent the nanoparticles from agglomeration through electrostatic stabilization. The same argument appeared to be plausible for the stabilization of nClAP by CMC, since the negatively charged CMC could adsorb to the nClAP surface [37]. However, the 0.75CMC-nClAP sample still showed agglomerated morphology and the particles seemed to be attached by unbound CMC molecules. Additionally, the energy-dispersive X-ray spectrometry (EDS) data of pristine nClAP showed the main characteristic signals of calcium, phosphorus, oxygen, and chlorine. Besides, the EDS data of the CMC-nClAP samples showed an increased signal of oxygen evidencing the modification of nClAP with CMC.

3.2. Affecting Factors for Cr(VI) Adsorption by CMC-nClAP

3.2.1. Effect of CMC Content on the Adsorption Ability of nClAP

The adsorption ability of various synthesized CMC-nClAP materials (nClAP, 0.25CMC-nClAP, 0.5CMC-nClAP, and 0.75CMC-nClAP) was compared and the results are shown in Figure 5. It was found that the CMC incorporation significantly improved the adsorption of nClAP towards Cr(VI) at the given conditions. Adsorption of Cr(VI) increased with increasing CMC additions, and the highest adsorption of Cr(VI) was observed for 0.5CMC-nClAP. The adsorption of Cr(VI) by the nClAP modified with 0.75% CMC, however, was similar to or even slightly lower than that by the pristine nClAP. The 0.75CMC-nClAP sample contained a higher amount of CMC, which exhibited agglomerated morphology as indicated by the TEM image (Figure 4). The excess CMC could inhibit the interaction of nClAP and Cr(VI) due to the wrapping effect of CMC, as reported in the literature [37]. To elucidate the adsorption interaction between the adsorbents and Cr(VI), further testing under varied conditions was conducted using the 0.5CMC-nClAP (noted as CMC-nClAP) as adsorbent.

3.2.2. Effect of pH

The effect of pH on the adsorption of Cr(VI) on CMC-nClAP is depicted in Figure 6a. The adsorption capacities were found to be 26.5 and 26.2 mg/g at pH 3.0 and 4.0, respectively, while it dramatically dropped to 2.1 mg/g at pH 8.0, indicating that the removal of Cr(VI) was significantly influenced by pH and low pH facilitated the elimination of Cr(VI).

The observed effect of solution pH on eliminating Cr(VI) might be explained by both the electrostatic attraction mechanism of Cr(VI) oxyanions with CMC-nClAP and the reduction mechanism of Cr(VI) to Cr(III). According to the speciation diagram of Cr(VI), the dominant form of Cr(VI) is ${\mathrm{HCrO}}_4^{-}$ at 2.0 < pH < 6.0, while the major form was ${\mathrm{Cr}}_2{\mathrm{O}}_7^{2-}$ at pH > 7.0 [38]. At pH < pH_ZPC (around 4.1) of CMC-nClAP (Figure 6b), the adsorbent surfaces carried positive charges, which was favorable for adsorption of anionic ${\mathrm{HCrO}}_4^{-}$ species. Furthermore, it is widely accepted that Cr(VI) has high redox potential under acidic media and can be reduced to Cr(III) (${\mathrm{HCrO}}_4^{-}+7{\mathrm{H}}^{+}+3{\mathrm{e}}^{-}\leftrightarrow {\mathrm{Cr}}^{3+}+4{\mathrm{H}}_2\mathrm{O},{\mathrm{E}}^0=1.36 \mathrm{V}$) [39]. Hence, it was reasonable that Cr(VI) oxyanions, either in aqueous solution or binding on the positively charged CMC-nClAP surfaces, could be reduced to Cr(III). Moreover, this reduction process was especially promoted with the decrease in pH since protons participated in this process [40]. At pH > pH_ZPC, the surface of the adsorbent became negatively charged, leading to electrostatic repulsion between the adsorbent and the anionic Cr(VI) species. This repulsion significantly reduced adsorption capacity. Similar trends have been reported in previous literature for Cr(VI) removal by various materials, such as modified polyethyleneimine [38], nanostructured polyaniline [38], and tin-functionalized hydroxyapatite [41]. Considering the effect of pH on Cr(VI) removal, a pH of 3.0 or 4.0 is favorable for the application of CMC-nClAP; however, the apatite materials are partially soluble at pH 3.0 [42], so to avoid the dissolution loss, pH 4.0 was selected in the following experiments.

3.2.3. Effect of Adsorbent Dosage

Adsorption was also performed by adjusting the adsorbent dosage between 0.1 and 5.0 g/L at fixed Cr(VI) initial concentration (10 mg/L) and pH 4.0. As depicted in Figure 7, Cr(VI) adsorption declined remarkably from 32.2 mg/g to 1.9 mg/g as the adsorbent dosage was raised from 0.05 g/L to 5 g/L. Since the concentration of Cr(VI) was fixed, the ratio of Cr(VI) to the adsorption sites decreased with increasing amount of CMC-nClAP material. Accordingly, the decline in adsorption amount was due to the increasing of vacant adsorption sites. Similarly, according to Zhang et al. [43], raising the amount of adsorbent led to a situation of unsaturation of adsorption sites, resulting in a decline in the adsorption capacity. In addition, the aggregation of adsorbents at higher dosage could also retard the adsorption capacity. Figure 7 also demonstrated that the removal rate for Cr(VI) rose from 16.5% to 95.2% as the dosage of CMC-nClAP was raised from 0.05 g/L to 5 g/L. The enhancement in removal rate can also be ascribed to the increase in adsorption sites with the raised dosage of CMC-nClAP material [44]. However, the higher dosage of adsorbent increased the cost of operation. The present findings also implied that it was achievable to eliminate Cr(VI) completely when there was sufficient adsorbent dosage in solution.

3.2.4. Adsorption Kinetics and Isotherms

The results of the adsorption kinetics of Cr(VI) by CMC-nClAP are presented in Figure 8a. It shows that the Cr(VI) adsorption increased rapidly in the initial phase and gradually slowed down as equilibrium was approached. It was evident that Cr(VI) uptake increased rapidly in the first 20 min, contributing to 73.2%, 84.0%, and 74.9% of the final adsorption amount for 5, 10, and 20 mg/L Cr(VI), respectively, and then exhibited a slower adsorption rate for the next 100 min until equilibrium was reached. The adsorption kinetics data were further evaluated by PFO and PSO models. It was evident that the correlation coefficients (R²) (

Table 1. The kinetic fitting parameters of Cr(VI) adsorption on CMC-nClAP (T = 298 K, adsorbent dosage = 0.1 g/L).

Cr(VI) Concentration	qe,exp (mg/g)	PFO Model			PSO Model
		k1	qe,cal	R2	k2	qe,cal	R2
		(1/h)	(mg/g)		(g/(mg·h))	(mg/g)
5 mg/L	11.8	0.099	10.7	0.919	0.008	12.2	0.967
10 mg/L	26.2	0.082	24.4	0.946	0.003	25.4	0.969
20 mg/L	35.8	0.077	33.3	0.934	0.002	36.5	0.978

) of the PSO model were higher than that of the PFO model. Meanwhile, the experimental adsorption capacities (q_e,exp) for Cr(VI) are comparable to the calculated values (q_e,cal) from the PSO model, further suggesting that the adsorption kinetics of Cr(VI) onto CMC-nClAP were well fitted to the PSO model, which suggested that the rate of chemical interactions (i.e., chemisorption between the Cr(VI) and the active sites of CMC-nClAP) was the step that controlled the adsorption kinetics of Cr(VI) [45].

The adsorption isotherms of Cr(VI) onto CMC-nClAP at three different temperatures were studied at the initial Cr(VI) concentrations of 2, 5, 10, 20, 30, 40, and 50 mg/L. The results are depicted in Figure 8b. Langmuir and Freundlich isotherm models for adsorption of Cr(VI) were utilized to analyze the experimental data. The values of q_max, K_L, and K_F increased with rising temperature, implying that the adsorption is endothermic. As shown in Figure 8b, the Langmuir model could fit the experimental data very well, which was signified by higher correlation coefficients (0.985 < R² < 0.991,

Table 2. The isotherm fitting parameters of Cr(VI) adsorption on CMC-nClAP (adsorbent dosage = 0.1 g/L).

Temperature	Langmuir Model			Freundlich Model
	KL	qmax	R2	KF	1/n	R2
	(L/mg)	(mg/g)		[(mg/g)·(L/mg)1/n]
283 K	0.067	47.4	0.985	5.386	0.513	0.933
298 K	0.119	52.1	0.989	9.950	0.407	0.915
313 K	0.151	61.1	0.991	11.745	0.405	0.932

) than the Freundlich model (0.915 < R² < 0.933). Thus, the adsorption of Cr(VI) was related to similar adsorption sites of the CMC-nClAP adsorbent and the interface between Cr(VI) and adsorbent generated a single layer of Cr(VI) [46]. Both K_L and K_F values increased continuously with the increasing temperature, suggesting the adsorption capacity and intensity improved with increasing temperature. As the temperature rose from 283 K to 313 K, the calculated q_max increased from 47.4 mg/g to 61.1 mg/g. This might be due to the improvement in diffusion efficiency, which increased the impact frequency between the CMC-nClAP and Cr(VI), thus improving Cr(VI) uptake [47]. Similarly, the adsorption of Cr(VI) by lignin-based anionic adsorption resin [48] and porous nanocomposite [49] also agreed well with the Langmuir isotherms and exhibited an endothermic process. Furthermore, the empirical parameter, 1/n, in the Freundlich model ranged between 0.1 and 1, implying favorable adsorption of Cr(VI) [13].

Furthermore, the maximum adsorption capacity of the CMC-nClAP material has also been compared with previously reported adsorbents and is demonstrated in Table S1 in the Supplementary Materials. It was found that CMC-nClAP had higher uptake capacity than most of the previous materials, such as zero-valent iron [40], functionalized hydroxyapatite [41], and commercial activated carbon [50]. Moreover, the present adsorbent was prepared from waste eggshells as the starting materials, which could reduce the cost. Thus, CMC-nClAP demonstrates potential for the elimination of toxic Cr(VI) from water.

3.3. Langmuir-Based Analysis of Adsorption Site Energy Distribution

Site energy distribution analysis was considered to be useful in exploring the adsorption mechanisms in different adsorbate–adsorbent systems. As mentioned above, the Langmuir model had the best fitting for the Cr(VI) adsorption process; hence, the Langmuir constants were adopted in the adsorption site energy distribution analysis.

The site energy E^* of Cr(VI) adsorption on CMC-nClAP as a function of q_e at different temperatures is depicted in Figure 8c. The E^* values declined significantly with the increase of q_e at all the three temperatures, suggesting that Cr(VI) initially occupied high energy adsorption sites at low contents and then gradually diffused to low energy adsorption sites on the CMC-nClAP adsorbent [29]. A similar phenomenon was observed for the adsorption of Cr(VI) on activated carbon or biochar [50], as well as on MCM-41 [51]. Meanwhile, the E^* values were in the sequence of 283 K < 298 K < 313 K at a fixed q_e, showing that higher temperature facilitated Cr(VI) adsorption (consistent with isotherm data).

Figure 8d displays the approximate site energy distribution curves of Cr(VI) adsorption at varying temperatures, where the area covered under the curves could theoretically represent the number of available active sites or the saturation adsorption amount [52,53]. Obviously, the site energy distribution curves displayed similar shapes characterized by a solitary apex and a mostly normal distribution. In addition, the frequency function F(E^*) first rose to the peak value with increasing E^* and then fell towards zero. Generally, the upward tendency implies that a small portion of adsorbates are forced to occupy lower energy sites under high content of Cr(VI). Conversely, Cr(VI) was preferentially bound to the high energy site at lower concentrations [54]. Meanwhile, as shown in Figure 8d, the site energy distribution curve peaks moved towards higher E^* values as the temperature rose from 283 K to 313 K, signifying that there were more sites, including high energy sites and low energy sites for Cr(VI) at 313 K [53]. Accordingly, the values of F(E₂^*) at 288, 298, and 308 K were found to be 4.99, 5.26, and 5.91 mg·mol/kg·kJ, respectively. This could be due to the fact that as the temperature increased, the energy of the binding sites increased, resulting in more binding sites being activated and available for Cr(VI) capture, thereby enhancing the adsorption capacity of Cr(VI) [54,55].

Furthermore, the average site energy (μ(E^*)) and the site energy heterogeneity (σ_e^*) were used to evaluate the adsorption affinity of Cr(VI) to CMC-nClAP and the homogeneity of the CMC-nClAP surface, respectively. It was considered that the higher the μ(E^*), the greater the adsorption affinity [56]. As depicted in Figure 8d, the μ(E^*) at 283 K, 298 K, and 313 K of Cr(VI) adsorption on CMC-nClAP was 27.17, 30.19, and 31.96 kJ/mol, respectively, indicating that high adsorption affinity occurred at 313 K [55].

Moreover, the values of the σ_e^*, corresponding to adsorbent site energy heterogeneity, were calculated and shown in Figure 8d. The results showed that σ_e^* increased from 8.71 to 9.62 and 10.35 kJ/mol when the temperature increased from 283 to 298 and 313 K, respectively. The higher σ_e^* at temperatures of 298 K and 313 K signified higher degrees of heterogeneity [57]. The variation of σ_e^* with temperature was consistent with the obtained heterogeneity factor of n from the Freundlich model, which suggested that the higher the temperature, the greater the heterogeneity [58]. Previous researchers argued that the heterogeneity of adsorption sites partly stemmed from the different chemical nature (i.e., chemical composition heterogeneity induced by the tailored groups, especially O-containing groups) of the adsorption sites [54,59]. In this work, CMC-nClAP was made from nClAP with CMC modification. The O-containing groups, such as ${\mathrm{COO}}^{-}$ groups, were tailored to nClAP (as indicated by the FTIR characterizations in Figure 2). The TEM observations (Figure 4) also revealed that CMC-nClAP had an irregular composite structure. Therefore, the particular structure and tailored functional groups led to the energetic heterogeneity of the CMC-nClAP adsorbent.

The mechanism of Cr(VI) adsorption on CMC-nClAP can be discussed on the basis of adsorption data and the values that the average site energies obtained from the model assume. The introduction of CMC enhanced the heterogeneity of adsorption sites of nClAP, providing more binding sites for Cr(VI). Overall, electrostatic attraction and reduction reaction also contributed to the improved adsorption of Cr(VI) by the CMC-nClAP adsorbent.

3.4. Adsorption Thermodynamics

To further clarify the influence of temperature on Cr(VI) adsorption, the thermodynamic parameters standard entropy change (ΔS⁰, J/(mol·K)), standard enthalpy change (ΔH⁰, kJ/mol), and standard free energy of adsorption (ΔG⁰, kJ/mol) were calculated by the following equations:

$$\Delta G^0=-RTln{K}_L$$

(9)

$$ln{K}_L={\displaystyle \frac{\Delta S^0}{R}}-{\displaystyle \frac{\Delta H^0}{RT}}$$

(10)

where R is the universal gas constant (8.314 J/(mol·K)); T is the temperature in Kelvin; K_L is the Langmuir constant that is related with the energy of adsorption.

The values of ΔH⁰ and ΔS⁰ were determined from the slope and intercept of the plot of lnK_L versus 1/T (Figure 9), and the obtained parameters are presented in

Table 3. Thermodynamic parameters of the Cr(VI) adsorption on CMC-nClAP.

Temperature (K)	R2	ΔG0 (KJ/mol)	ΔH0 (KJ/mol)	ΔS0 (J/(mol·K))
283	0.9557	−9.89	19.89	105.62
298		−11.84
313		−13.04

. The calculated value of ΔG⁰ suggested the spontaneous nature of the Cr(VI) adsorption process [60], and the declined ΔG⁰ with rising temperature indicated an increased spontaneity with higher temperature. In addition, the positive value of ΔH⁰ (19.89 kJ/mol) demonstrated that the adsorption of Cr(VI) on CMC-nClAP was an endothermic reaction. Moreover, the positive value of ΔS⁰ (105.62 J/(mol·K)) indicated increasing randomness during the adsorption process. Spontaneous, endothermic, and entropy-raising adsorption processes were widely recognized for Cr(VI) adsorption by activated eucalyptus biochar [61], as well as the graphene oxide–zinc oxide nanohybrid [62].

3.5. Effects of Coexisting Ions

Generally, various ions also coexisted with the actual Cr(VI)-containing wastewater. To evaluate the practical applicability of the CMC-nClAP adsorbent, the effects of coexisting cations (K⁺, Ca²⁺, and Mg²⁺) and anions (Cl⁻, NO₃⁻, and SO₄²⁻) were investigated. As depicted in Figure 10a, the presence of coexisting cations showed minimal influence on the adsorption amount of Cr(VI), while the q_e of Cr(VI) still maintained 95%. This suggested that the cations did not compete with the adsorption sites of Cr(VI) oxyanions [63]. In addition, the competitive effects of monovalent anions (Cl⁻ and NO₃⁻) on the adsorption of Cr(VI) were also neglectable due to the low affinity of these anions [18]. However, when SO₄²⁻ was present at concentrations of 5 and 10 mM, the q_e of Cr(VI) decreased by 10.5% and 16.4%, respectively. Similar findings were reported in the literature [46], which suggested that the influence of coexisting anions on Cr(VI) adsorption were more significant at higher valence states, causing competition for the adsorption sites. Overall, the prepared CMC-nClAP adsorbent exhibited a high tolerance to interfering ions, underscoring the robustness and potential applicability for Cr(VI) removal in wastewater.

3.6. Reusability

In order to assess the reusability of the CMC-nClAP adsorbent, the Cr(VI)-loaded material was treated with 0.5 M NaOH to desorb Cr(VI) [46]. As shown in Figure 10b, the adsorption amount of Cr(VI) still accounted for 90% of the original adsorption amount at the fifth cycle of the adsorption experiment. These results confirmed that the CMC-nClAP adsorbent maintained high adsorption performance after multiple cycles. Thus, it demonstrated excellent reusability in practical application.

4. Conclusions

In this work, a CMC-nClAP composite was prepared using a simple approach and used in this study for aqueous Cr(VI) elimination. Material characterization techniques revealed structural and functional aspects of the CMC-nClAP adsorbent, which resulted in the efficient adsorption of Cr(VI). Batch adsorption results suggested that the CMC-nClAP composite combined the advantages of nClAP and CMC, showing better Cr(VI)-removal ability and favorable dispersibility, as indicated by the sedimentation kinetics data, than pristine nClAP. Studies of a pH dependence revealed that Cr(VI) adsorption by the CMC-nClAP adsorbent was more feasible at lower pH values. The adsorption kinetics data exhibited a better fit to the PSO model, and the adsorption isotherms fitted well with the Langmuir model with the maximum Cr(VI) adsorption capacity value of 52.1 mg/g at 298 K. The adsorption of Cr(VI) was a spontaneous endothermic process. The site energy distribution combined with the Langmuir isotherm model indicated that CMC-nClAP had a heterogeneous energy distribution on its surface, and the high energy sites on the adsorbents were preferentially occupied by Cr(VI), followed by the low energy sites. Common co-existing ions did not significantly affect the Cr(VI) adsorption potential of CMC-nClAP. The reuse test revealed that the adsorption capacity for Cr(VI) remained above 90% after five cycles. Based on these findings, it can be concluded that the CMC-nClAP composite is a promising alternative for eliminating Cr(VI) from wastewater. However, future research is needed to address the limitations of the CMC-nClAP adsorbent, such as its narrow pH range and susceptibility to interference from SO₄²⁻.