Strontium-doped hydroxyapatite as adsorbent effectively to remove lead ions from water

doi:10.21203/rs.3.rs-1434355/v1

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Strontium-doped hydroxyapatite as adsorbent effectively to remove lead ions from water

https://doi.org/10.21203/rs.3.rs-1434355/v1

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In this study, a strontium-doped hydroxyapatite (Sr-HAP) was synthesized by the sol-gel method, which used as adsorbent to remove lead ions (Pb²⁺) from water. The results shown that the adsorption capacities of the Sr-HAP were obviously higher than the HAP, which the adsorption capacities for Pb²⁺ reached 651.175 mg/g. The proper increasement in the dosage of adsorbent was beneficial to the removal of Pb²⁺ by Sr-HAP. Meanwhile Sr-HAP had a wide applicable pH range for Pb²⁺. And the increasement of temperature could increase the adsorption capacity of Sr-HAP for Pb²⁺ to a certain extent. The Langmuir model was used to fit the isotherm adsorption process of Sr-HAP to Pb²⁺ in water. Compared with HAP, the specific surface area of Sr-HAP has increased by 11.1 %, and the pore size distribution of Sr-HAP tended to be smaller and more uniform. Hence, Sr-HAP could be used as an ideal adsorbent to remove Pb²⁺ in wastewater.

Hydroxyapatite

Adsorption

Strontium-doped

Lead ions

Mechanism

Heavy metals pollution was known to cannot undergo biodegradation, resulting to adverse effects on the environment and biological systems (Imron et al., 2019; Purwanti et al., 2020). Lead was one of the heavy metals, which was the second most toxic metal and often found in the environment, then led to no safe level of lead in drinking water (Adeyemi et al., 2021; Kim et al., 2017; Du et al., 2017). It was acutely toxic to human body when consumed in high concentrations inside cells (Nwaehiri et al., 2020). Under environ-mental lead exposure, the elevated lead blood levels could make the Brain system for the permanent and irreversible damage for human, especially the young children (Cecil et al., 2008). At present, the existing removal technologies for lead-contaminated wastewate mainly include chemical precipitation method (Charerntanyarak L., 1999), ion exchange method (Dabrowski et al., 2004), membrane filtration method (Juang et al., 2000), flotation method (Doyle et al., 2003), adsorption method (Crini G., 2005), biosorption method (Rangabhashiyam et al., 2019) and electrochemical method (Pedersen A. J., 2003). Among them, the adsorption method due to its simple operation process, environmental friendliness and high removal efficiency, was an important method to remove lead ions (Pb²⁺) from wastewate.

Hence, much more attention had been paid to the adsorbent selection. Such as activated carbon (AC) (Kongsuwan et al., 2009), zeolite (Chong et al., 2013), chitosan (Paulino et al., 2007), ion exchange resins (Dong et al., 2010), industrial solid waste (Yong C.B., 2018), nanomaterials (Li et al., 2020), hydroxyapatite (HAP) (Wang et al., 2017; Aklil et al., 2004; Lei et al., 2015; Wen et al., 2013; Lin et al., 2009; Sandrine et al., 2007; Fernane et al., 2013), were widely used to the adsorption materials. Among them, the HAP with high adsorption efficiency and environmental friendliness had greatly promote practical applications. Wang et al. (Wang et al., 2017) studied the magnetic hydroxyapatite-mixed oxidation of multi-walled carbon nanotubes (mHAP-oMWCNTs) to remove Pb²⁺ in water, and the results showed that mHAP-oMWCNTs showed better performance than mHAP, mMWCNT and HAP-oMWCNTs. Aklil et al. (Aklil et al., 2004) used calcined hydroxyapatite to adsorb Pb²⁺, Zn²⁺ and Cu²⁺, and found that at pH 5, the adsorption capacity of HAP on Pb²⁺, Zn²⁺ and Cu²⁺ were 85.6, 29.8 and 20.6 mg/L, respectively. Lei et al. (Lei et al., 2015) prepared a hydroxyapatite/chitosan (HAP/CS) porous material with a pore size of 100 ~ 300mm to adsorb Pb²⁺, and its maximum adsorption capacity was 264.42 mg/g. Meanwhile, adding extra elements could improve the performance of hydroxyapatite. Tang et al. (Wen et al., 2013) prepared Si-CHAP by hydrothermally doping silicon carbon into hydroxyapatite. The results showed that Si-CHAP with a large specific surface area of 323.25 m²/g, thereby enhancing the adsorption capacity.

In this work, a strontium-doped hydroxyapatite (Sr-HAP) was synthesized by the sol-gel method. The Pb²⁺ adsorption experiment was carried out using the prepared material. Discuss the adsorption behavior and influence of Sr-HAP on Pb²⁺ under different adsorption conditions. The adsorption kinetic equation and thermodynamic parameters were calculated, and the recycling rate of Sr-HAP was explored. Meanwhile, the Sr-HAP was investigeted by the various analytical techniques, such as x-ray diffractometer, fourier transform infrared, scanning electron microscope, brunauer-emmett-teller, zeta potential analyzer. Fiannly, the adsorption mechanism of Sr-HAP on Pb²⁺ was explained.

2.1 Adsorbent preparation and characterization

2.1.1 Preparation

The Sr-HAP sample with x = Sr/(Ca + Sr) molar ratio of 0.5 was prepared by sol-gel method. 1 mol/L Sr(NO₃)₂, Ca(NO₃)₂·4H₂O and (NH₄)₂HPO₄ solution were diluted to 0.2 mol/L, respectively. The certain volume of Sr(NO₃)₂·6H₂O solution, Ca(NO₃)₂·4H₂O solution, and (NH₄)₂HPO₄ solution were mixed, in which the (Sr + Ca)/P molar ratio was fixed at 1.67. Meanwhile, ammonia water was added quickly into the above mixture to adjust the pH within 10 ~ 11. Next, the mixed solution was stirred quickly for 1 ~ 2 min, then the magnetic stirring was kept for 30 min, and the obtained solution was transferred in the water bath (ageing) at 50℃ for 24 h. After reaching the reaction time, the pH of the supernatant was measured by natural cooling, the sol-gel was centrifuged (4000 r/min, 5 min) and washed with ultrapure water for three times (to neutral) and then washed with absolute ethanol. After drying at 80°C for 24 h, strontium-doped hydroxyapatite (Sr-HAP) was prepared after grinding finely (sieved through 100 mesh), and stored in the dry environment.

2.1.2 Characterization

The characterization were shown in the Supplementary material.

2.2 Adsorption experiments

Pb²⁺ solution was prepared by analytically pure lead nitrate, nitric acid and ultrapure water. The 0.2% nitric acid solution was used to dilute all samples. The batch experiment was carried out in a set of 100 mL polyethylene centrifuge tubes and KOH or HNO₃ solution were used as pH regulator. The certain amount of Pb²⁺ solution was added into the previous solution and then the mixed solution was placed in a constant temperature water bath shaker at 180 r/min. Oscillating at a constant temperature for 24 h and the temperature was set as 25°C. After the sample was filtered with a syringe filter (0.22 µm filter membrane), the residual Pb²⁺ concentration in the solution was measured.

The influence of different initial concentrations of Pb²⁺ solutions (100, 200 and 300 mg/L) and pH 5.0 (± 0.05) solutions, different dosages of Sr-HAP (0.01-0.1g, < 100 mesh) on adsorption was investigated at 25℃. By using nitric acid and sodium hydroxide (0.1mol/L) solution to adjust the initial pH value of Pb²⁺ solution to study the effect of pH on adsorption. 50ml containing Pb²⁺ wastewater with an initial pH of 5.0 were prepared, which had different concentrations (100, 150, 200, 250, 300, 400, 500, 600, 700, 800, 900, and 1000 mg/L). Then, 40 mg Sr-HAP was added to the Pb²⁺ solution to determine the effect of the initial concentration of the solution on the adsorption at the constant temperature of 25°C, 35°C and 45°C. Meanwhile, the isothermal adsorption and thermodynamic characteristics were analyzed. Water samples were collected at different times (5, 10, 15, 30, 60, 120, 180, 240, 300, 360, 420, 480, 540, 720, 1080, and 1440 min) to study the effect of adsorption time on the adsorption. 40 mg of Sr-HAP solid with particle size less than 100 meshes was added to a mixed solution, which consisted of 50 mL of Pb²⁺ solution with an initial concentration of 100 mg/L and different concentrations (0, 10, 20, 40, 60, 80, 100, 150 and 200 mg/L) of Mg²⁺, Ca²⁺ and EDTA aqueous solutions, or Pb²⁺ solution(50 mL, 100 mg/L) and different concentrations (0, 1, 2, 5, 8, 10, 15, 20, 50 and 100 mg/L) of fulvic acid (FA) or black humic acid (HA) solution to study the influence of interfering substances on adsorption. Inductively coupled plasma emission spectrometer (Opetima 7000DV) was used to determine the residual Pb²⁺ concentration in the solution.

2.3 Desorption experiments

Sr-HAP samples containing Pb²⁺ were desorbed with different eluents. The adsorbent saturated under the optimal adsorption conditions was washed with deionized water for 3 ~ 5 times and dried at 80°C. 0.1 mol/L desorption solution (NaOH, HCl, NaCl), H₂O, TCLP (pH = 2.55 and 4.93), Ca(NO₃)₂ (0.001, 0.005 and 0.01M), Sr(NO₃)₂ (0.001, 0.005 and 0.01M) were oscillated under the same conditions, then measured the mass concentration of the filtrate after filtration, and calculated the desorption rate.

3.1 Characterization of the adsorbent

The X-ray diffraction (XRD) patterns of Sr-HAP with different adsorption concentrations of Pb²⁺ was showed in Fig. 1(A). It could be seen that after Pb²⁺ adsorption, the crystal structure of Sr-HAP changed significantly, and the position of the characteristic peaks changed too. The intensity of the main characteristic peaks increased with the increasement of Pb²⁺ concentration, but it still belonged to the hexagonal crystal structure of apatite, and the space group was P₆³/m. Figure 1(B) showed the angular deviation of the characteristic peaks after magnification. It could be seen that the characteristic peak of the (211) crystal plane gradually divided into two different positions (112 crystal plane and 300 crystal plane) as the concentration of adsorbed Pb²⁺ increased. And the peak height (intensity) was consistent with the increasing trend of the initial solution Pb²⁺ concentration. The formation of a new peak at 2θ of 30.1°~30.5° corresponded to the reference substance Pb₁₀(PO₄)₆(OH)₂(HPY) or PbxCa_5 − x(PO₄)₃OH solid solution with lower calcium content. Because the peaks at 30.3 ° and 31.8 ° were determined as the diffraction characteristic signals of HPY and HAP in the solid, the ratio of these two peaks could semi-quantitate the conversion of HAP to HPY. The ionic radius of Pb²⁺ was 1.20 Å, and it was larger than that of Sr²⁺ (1.12 Å). Compared with Ca²⁺ (0.99 Å), it was difficult for Pb²⁺ to enter the Sr-HAP crystal to replace Ca²⁺ and Sr²⁺. Entering into the HAP lattice through the substitution of calcium ions was difficult for lead ions. Therefore, the lattice-substituted Pb might come from the subsequent diffusion of surface-adsorbed Pb to the particle core. It indicated that the dissolution/precipitation effect might exist in the removal of Pb²⁺ by Sr-HAP.

The Sr-HAP solid phase adsorbed with different concentrations of Pb²⁺ was analyzed by infrared spectroscopy to compare the changes of functional groups (Fig. 2). It could be seen that the typical characteristic peaks of Sr-HAP with different adsorption concentrations of Pb²⁺, which was almost identical to that of pristine Sr-HAP indicating that its functional group structure did not change significantly. With increasing of the Pb²⁺ adsorption concentration, the stretching vibration peak intensity of PO₄³⁻ at the wavelengths of 1033.31 and 561.08 cm^− 1 increased, indicating that PO₄³⁻ participated in the chemical reaction in the adsorption process. After absorbing Pb²⁺, the frequency of most infrared peaks changed very little, which indicated that the main infrared signal was still contributed by a large amount of apatite. Therefore, it was hardly be used to infer the absorption mechanism of Pb²⁺. Meanwhile, the morphology of Sr-HAP was analyzed by scanning electron microscope (SEM). It could be seen from Fig. 3 that the morphology of the material presented scattered rod-like or needle-like structured crystals, with the uneven particle morphology and the large pores between the crystal grains.

Figure 4(A) showed the typical N₂ adsorption-desorption isothermal curves of the synthesized HAP and Sr-HAP materials, respectively. As showed in Fig. 4(A), Sr-HAP and HAP exhibited type IV isotherms with H3 hysteresis loop according to IUPAC's empirical classification the material was composed of mesoporous structure. It is generally believed that the shape of the hysteresis loop is related to the texture of the porous material (for example, pore size distribution, connectivity, and geometry), and hysteresis loop material has wedge-shaped pores with open ends (H3 type isotherm did not show any limited adsorption under high P/Po, which could be observed in non-rigid aggregates of plate-like particles) (Alothman Z., 2012), which was consistent with the images of scanning electron microscope. Compared with HAP (102.314 m²/g), the specific surface area of Sr-HAP (115.121 m²/g) increased by 11.1%. Figure 4(B) showed the pore size distribution of HAP and Sr-HAP. It could be seen that the pore size of the material was mainly distributed in the range of 2–50 nm, which belonged to the of mesoporous materials. The pore size distribution of Sr-HAP was smaller than that of HAP and the pore size was more uniform. It showed that the incorporation of strontium caused the material lattice defects and the reduction of the particle size, and the pores between the particles were also reduced.

From Fig. S1, it could be seen that the zero point electricity of Sr-HAP was pHPZC = 8.40, and the zero electricity point of HAP was pHPZC = 6.92. Due to the special structure of apatite (M(PO₄)₃OH), the material had a certain buffering capacity, which could maintain its own acidity and alkalinity in the acid or alkali environment, and released OH⁻ or PO₄³⁻. When the pH value of the solution was lower than the zero electric point, Me²⁺ would electrostatically repulse the positive charge on the surface of the material. When it was higher than its zero electric point, Me²⁺ would electrostatically attract the positive charge on the surface of the material. For Me²⁺, when the pH value of the solution was higher than the zero electric point of Sr-HAP, Me²⁺ was easy to hydrolyze. At this time, the removal of Me²⁺ was caused by its own characteristics. Heavy metal wastewater is generally weak acid or neutral, so it is more reasonable for removal heavy metal ions to choose the material which the pH value was lower than the zero electric point.

3.2 The influence of adsorption conditions

3.2.1 Influence of adsorbent dosage

For Pb²⁺-containing solutions of different concentrations, the dosage of adsorbent has a certain effect on the removal of Pb²⁺. It could be seen from Fig. 5 that the removal rate of Sr-HAP adsorbent for solutions with different initial Pb²⁺ concentrations (100, 200 and 300 mg/L) increased with the increasement of the dosage. When the dosage of Sr-HAP was 40 mg / 50 mL, the removal rate curves of different initial Pb²⁺ concentrations gradually tended to be flat. The unit adsorption capacity of Sr-HAP on Pb²⁺ gradually decreased with the increasing of the adsorbent dosage. The increasement of available active sites on the adsorbent surface might be attributed to the increasing in the dosage of Sr-HAP, while the amount of Pb²⁺ was limited, resulting in the unit adsorption capacity of Sr-HAP to Pb²⁺ gradually decreased. When the initial concentration of Pb²⁺ was 100 mg/L, the adsorption capacity of Pb²⁺ decreased from 403.481 mg/g to 42.377 mg/g with the adsorbent dose increasing from 10 mg to 100 mg in 50 mL solution. At the same time, the corresponding Pb²⁺ removal rate increased from 99.88–99.91%. For the initial Pb²⁺ concentration of 200 mg/L in a 50 mL solution, the adsorption capacity of Pb²⁺ increased from 330.686 mg/g to 466.733 mg/g with the adsorbent dosage increased from 10 mg to 20 mg, and then the adsorption capacity of Pb²⁺ decreased from 466.733 mg/g to 94.953 mg/g with the adsorbent dosage increased from 0.02 g increased to 0.10 g. The corresponding Pb²⁺ removal rate had risen from 35.51–99.95%. When the initial concentration of Pb²⁺ was 300 mg/L and the solution volume was 50 mL, the dosage of adsorbent increased from 10 mg to 40 mg, and the adsorption capacity of Pb²⁺ increased from 298.559 mg/g to 350.052 mg/g. The dosage increased from 40 mg to 100 mg and the adsorption capacity of Pb²⁺ decreased from 350.052 mg/g to 142.641 mg/g. Meanwhile, the removal rate of Pb²⁺ by Sr-HAP gradually increased from 22.63–100.00%. When the dosage of Sr-HAP was 40 mg/50 mL, the removal rate of Pb²⁺ was above 98.00% at different concentrations (100, 200, and 300 mg/L), and the adsorption capacity was relatively large, so 40 mg/50 mL was used as the optimal adsorbent dosage for further experiments.

3.2.2 Influence of the initial pH value of the solution

The pH value of the solution was one of the important parameters to control the metal ion removal mechanism (Hao et al., 2017). Figure 6(A) showed that when the pH value of the solution from 1.0 to 2.0, the removal rate increased as the pH value increased, and the adsorption capacity also increased. when the pH value from 2.0 to 5.0, the removal rate remained basically unchanged, the adsorption capacity also tended to be stable. However, when the pH value form 5.0 to 8.0, the removal rate did not change much, but the adsorption capacity gradually decreased. The pH value of the solution system not only affected the existence of Pb²⁺ in the water, but also directly affected the surface charge point of the adsorbent, the pore structure and chemical properties of the material surface (Yang et al., 2013). Pb²⁺ in the aqueous solution would exist in different forms under different pH conditions. When the pH exceeded 8.0, the lead hydroxide (Pb(OH)₂) K_sp[Pb(OH)₂=1.2×10^− 15] gradually formed. A part of Pb²⁺ in the aqueous solution could be removed by precipitation, which added some interference to the removal of Pb²⁺. Visual MINTEQ v3.0 software was used to simulate the existence of Pb²⁺ with pH changed at 25°C and the concentration of 300 mg/L. As shown in Fig. 6(B), the formation of Pb²⁺ under different pH conditions included Pb²⁺, Pb(OH)_2(aq), Pb(OH)⁺, Pb₂(OH)³⁺, Pb₃(OH)₄²⁺, Pb₄(OH)₄⁴⁺, Pb(OH)³⁻, and so on. When the pH was less than 6.0, the lead element in the solution mainly existed in the form of Pb²⁺ ions. In the pH range of 6.0 to 9.0, the concentration of Pb²⁺ decreased rapidly, and the positively charged Pb(OH)⁺, Pb₂(OH)³⁺, Pb₃(OH)₄²⁺, and Pb₄(OH)₄⁴⁺ increased gradually. Hydrolysis products gradually appeared, while the neutral Pb(OH)₂ began to precipitate at pH 8.0 and became one of the main existing forms at pH 11. Therefore, the pH value of solution should be controlled within the range of 1.0 to 8.0 to eliminate the influence of Pb²⁺ precipitation.

In the lower initial pH range (lower than pHPZC), protons were consumed by the protonation of the ≡ P-O⁻, ≡Ca-OH⁰ and ≡ Sr-OH⁰ groups on the surface, resulting in the increasement in the final pH of the solution. In the acidic solution, the surface of Sr-HAP were dominated by positively charged ≡ Ca-OH²⁺, ≡Sr-OH²⁺ and neutral ≡ P-OH⁰ sites, so that the surface charges of Sr-HAP was positive in this pH range. On the other hand, the final pH decrease occurred in the range of higher initial pH (higher than pHPZC), due to the consumption of ≡ Ca-OH²⁺, ≡Sr-OH²⁺ and ≡ P-OH⁰ positions by OH⁻ through surface deprotonation point. Therefore, the neutral ≡ Ca-OH⁰, ≡Sr-OH⁰ and the negatively charged ≡ P-O⁻ groups were dominant in the alkaline solution, so that the surface of Sr-HAP was negatively charged in the alkaline solution and could adsorb positively charged hydrolyzed substances (such as Pb(OH)⁺, Pb₃(OH)₄²⁺, Pb₄(OH)₄⁴⁺, etc.), which was beneficial to the overall removal process. In the removal process of Pb²⁺ by Sr-HAP, for the Pb²⁺-containing wastewater with the concentration of 100, 200, and 300 mg/L, when the pH value from 2.0 to 6.0, the removal rate was above 97.51% and the adsorption capacity was 100.177, 209.881 and 301.267 mg/g, respectively. It indicated that Sr-HAP was suitable for removing Pb²⁺ in neutral or weakly acidic aqueous solutions, which was beneficial to the application of Sr-HAP to the actual heavy metal wastewater treatment. In this experiment, pH 5.0 was selected for the next experiment, which was consistent with the pH value in the literature (Yan et al., 2014; Zhu et al., 2018).

3.2.3 Influence of the initial concentration of the solution

It could be seen from Fig. 7 that the initial concentration was the main factor affecting the removal effect of Pb²⁺. The overall trend of Pb²⁺ removal at different temperatures was similar. After the initial concentration of Pb²⁺ was 300 mg/L, the removal rate decreased with the increasement of the initial concentration of Pb²⁺, and the adsorption capacity increased with the increasement of the initial concentration of Pb²⁺. When the initial concentration of Pb²⁺ increased from 100 mg/L to 800 mg/L, the removal amount of Pb²⁺ increased from 128.779, 128.481 and 128.474 mg/g to 651.175, 673.806 and 695.223 mg/g, at 25, 35 and 45°C, respectively. Correspondingly, the removal rate dropped from 99.98%, 100.00%, and 99.99–65.89%, 68.52%, and 71.05%, respectively. The reason why the removal amount of Pb²⁺ by Sr-HAP increased with the increasement of the initial concentration might be that when the dosage of Sr-HAP was determined, the initial concentration of Pb²⁺ gradually increased and the driving force of the mass transfer of the solution increased the passing rate of Pb²⁺ from the bulk solution to the particle surface, more and more Pb²⁺ accumulated on the surface active sites of Sr-HAP. The decrease of removal rate was attributed to the increasement of the initial Pb²⁺ concentration, and the active sites on the surface of the adsorbent were fully occupied. The adsorption was close to saturation, and the removal rate also drops (Chao et al., 2017). It could also be concluded from Fig. 7 that temperature would affect the degree and rate of adsorption. As the temperature increased, Pb²⁺ rapidly diffused and accelerated through the surface of the Sr-HAP particles, and the adsorption capacity and removal rate increased. This indicated that appropriate heating could promote the Pb²⁺ and the removal process might be an endothermic process (Wei et al., 2015).

3.2.4 Influence of adsorption time

It could be seen from Fig. 8 that the amount of Pb²⁺ adsorbed by Sr-HAP gradually increased with the increasement of adsorption time. In the initial stage of adsorption, Pb²⁺ quickly occupied the active sites on the surface of Sr-HAP, and the Pb²⁺ concentration difference at the solid/liquid interface was large, the solution mass transfer driving force was relatively large, and it was easy to overcome the mass transfer resistance of Pb²⁺ at the solid/liquid interface, thereby accelerating the adsorption rate. With the prolonging of adsorption time, the active sites on the surface of Sr-HAP had been fully occupied by Pb²⁺, the Pb²⁺ concentration difference at the solid/liquid interface gradually decreased, the adsorption rate decreased, and the adsorption amount gradually became flat. At about 120 min, the adsorption basically reached equilibrium. At this time, the maximum adsorption capacity at different temperatures were 587.077 mg/g (25°C), 593.863 mg/g (35°C) and 598.138 mg/g (45°C). The increasement of temperature could increase the adsorption capacity of Sr-HAP for Pb²⁺, indicating that the adsorption of Pb²⁺ by Sr-HAP was an endothermic process, but it could not shorten the time to reach the equilibrium of the reaction.

3.2.5 Influence of interfering substances

It could be seen from Fig. S3 that the higher the EDTA concentration, the lower the removal rate of Pb²⁺ by Sr-HAP. The calculation of EDTA by Visual MINTEQ v3.0 software showed that when pH 5.0, EDTA mainly appeared in the form of H₂EDTA²⁻, which could react with Pb²⁺ at a molar ratio of 1:1 to form negatively charged Pb(II)EDTA²⁻. Lead element existed in the solution in the form of Pb²⁺ and Pb(II)EDTA²⁻ complexes. At lower pH, the adsorption of Me(II)EDTA²⁻ complexes could be achieved on the -SOH site of the adsorbent through surface coordination reactions. Formula (1) to explain:

$$\begin{array}{c}\text{S}\text{O}\text{H}\text{+}{\text{H}}^{\text{+}}\text{+}{\text{Pb}\left(\text{Ⅱ}\right)\text{EDTA}}_{\left(\text{aq}\right)}^{\text{2-}}\text{=}{\text{SOH}}_{\text{2}}^{\text{+}}\text{+}{\text{Pb}\left(\text{Ⅱ}\right)\text{EDTA}}^{\text{2-}}\text{ }\text{ }\text{ }\text{(}\text{1}\text{)}\end{array}$$

Lead element was adsorbed on Sr-HAP in the form of Pb²⁺ and Pb(II)EDTA²⁻ through ion exchange and surface coordination mechanism. However, the adsorption affinity of Pb(II)EDTA²⁻ complex to the surface of HAP was low. The presence of EDTA would reduce the efficiency of the adsorption process, thereby increasing the mobility of Pb²⁺ and potential heavy metal environmental pollution.

It could be seen from Fig. S2 that the presence of HA had a certain effect on the removal of Pb²⁺ by Sr-HAP. When the concentration of HA increased from 0 to 20 mg/L, it would reduce the adsorption of Pb²⁺ on Sr-HAP, which might be due to the saturation of active sites for adsorption on SR-HAP. After the concentration was 20 mg/L, a further increase in the concentration of HA would not lead to the significant change in Pb²⁺ adsorption, and the removal rate remained basically constant. The similar phenomenon had been observed in previous studies (Wang et al., 2008; Zhu et al., 2010). Because of the strong affinity between Pb²⁺ and Sr-HAP, Sr-HAP always acted faster on Pb²⁺ than that of HA. Even if its concentration increased the driving force for HA to adsorb to the surface of Sr-HAP was still insufficient to compete with Pb²⁺. The effect of FA on the removal of Pb²⁺ was stronger than that of HA, and the overall removal trend of Pb²⁺ was similar to that of HA.

In the concentration range of 0 ~ 200 mg/L, competitive cations (Ca²⁺ and Mg²⁺) had almost no effect on Pb²⁺ immobilized by Sr-HAP (Fig. 9), and the removal rate almost maintained above 99.95%, the results indicated that SR-HAP was selective to the competitive cations that usually existed in the water environment and could effectively remove the divalent metals. Ca²⁺ was the parent cation of Sr-HAP. Compared with Mg²⁺, Ca²⁺ had the higher affinity for the surface of Sr-HAP. Increasing the initial concentration gradually could be regarded as the increasement in the concentration of adsorbate, but it did not affect the removal rate of Pb²⁺, because the adsorption capacity of SR-HAP for Pb²⁺ had not reached saturation.

3.3 Adsorption-desorption experiment

The comparison of the desorption of Sr-HAP adsorbent after adsorbing Pb²⁺ with different desorption solutions was shown in Fig. 10. It could be seen that except for TCLP1 (pH 4.93), HCl, EDTA and NaOH could desorb Pb²⁺, the desorption effect of the other desorption agents on the equilibrium Pb²⁺ adsorption were almost zero. The above phenomenon showed that Pb²⁺ could be more stably fixed on Sr-HAP, and the desorption rate of Pb²⁺ with stronger acidic desorption solution (such as HCl) was stronger. In the case of high H⁺ concentration, the chemical reaction of Pb²⁺ on the SR-HAP surface, such as surface complexation, would have an inhibitory effect, resulting in the desorption of Pb²⁺. The other part of the reason was that the partial dissolution of apatite group minerals under the condition of pH < 3.0. Due to environmental changes, such as changes in pH, the lead fixed on the surface was easily desorbed into the solution again, while the precipitated and complexed lead was very stable. It could be seen from Table S1 that the desorption rate of Pb²⁺ for most of the desorption liquids after three cycles of adsorption-desorption process was lower than 11.42%, indicating that Sr-HAP could be used as an adsorbent with excellent performance to immobilize heavy metals in wastewater containing Pb²⁺.

3.4 Adsorption kinetics

Among the different kinetic equations, the experimental data of Pb²⁺ adsorption by Sr-HAP was most suitable for the pseudo-second-order kinetic model, which was described as follows:

$$\begin{array}{c}\text{t}/{\text{q}}_{\text{t}}\text{=}\text{q}/\left({\text{K}}_{\text{2}}{\text{q}}_{\text{e}}^{\text{2}}\right)\text{+}\text{t}/{\text{q}}_{\text{e}} \left(2\right)\end{array}$$

In the formula: q_e and q_t were the adsorption amount (mg/g) at the time and equilibrium, respectively; K₂ was the pseudo-second-order kinetic rate constant, g/(mg·min).

The adsorption rate (h) (mg/(g min)) could be expressed as:

$$\begin{array}{c}\text{h}\text{=}{\text{k}}_{\text{2}}{\text{q}}_{\text{e}}^{\text{2}} \left(\text{3}\right)\end{array}$$

The equilibrium adsorption capacity (q_e) and the pseudo-second-order constant (K₂) could be calculated from the slope and intercept of the curve t/q_t vs t (Fig. S3(A)). As shown in Fig. S3(A), the correlation coefficients R² of the pseudo-second-order kinetic equations were all equal to 1.0, which could better describe the process of removing Pb²⁺ by Sr-HAP. Pseudo-second-order kinetics and Elovich kinetic equations were mainly used to describe chemical adsorption, that is, the process of Pb²⁺ and Sr-HAP binding. The reaction was fast in the chemical adsorption stage, and the rate of Sr-HAP adsorption of Pb²⁺ was mainly controlled by membrane diffusion and internal diffusion.

Table 2 showed the kinetic fitting parameters of Sr-HAP on Pb²⁺. The theoretical equilibrium adsorption capacities obtained by the pseudo-second-order kinetic model were 588.235, 588.235 and 588.235 mg/g, respectively, and the experimental results were 587.077, 593.863 and 598.138 mg/g which were relatively close to the theoretical equilibrium adsorption capacities. Besides, the correlation coefficient R² was close to 1.

Table 1

Elements mass after Sr-HAP adsorption of Pb²⁺ (A and B are different scanning positions)
(A)			(B)
The element	Wt (%)	Atomic percentage	The element	Wt (%)	Atomic percentage
O	23.80	69.87	O	26.39	69.05
P	7.10	10.76	P	8.23	11.13
Ca	2.33	2.73	Ca	4.65	4.85
Sr	4.87	2.61	Sr	9.76	4.66
Pb	61.90	14.03	Pb	50.97	10.30
The total	100%	100%	The total	100.00	100.00

Table 2

Kinetic fitting parameters of Sr-HAP to Pb²⁺
Simulation type	T/K	Rate constant/K		q_{e, cal} (mg/g)	q_{e, exp} (mg/g)	R²
Pseudo-second-order dynamics	298		4.9898×10^− 4	588.235	587.077	0.9996
	308	K₂[g/(mg·min)] =	4.9828×10^− 4	588.235	593.863	0.9998
	318		5.1607×10^− 4	588.235	598.138	0.9997

3.5 Adsorption isotherm

For ideal monolayer adsorption, the linear expression of Langmuir isotherm adsorption model were:

$$\begin{array}{c}{\text{C}}_{\text{e}}/{\text{q}}_{\text{e}}\text{=}{\text{C}}_{\text{e}}/{\text{q}}_{\text{m}}\text{+}\text{1}/\left({\text{q}}_{\text{m}}{\text{K}}_{\text{L}}\right) \left(\text{4}\right)\end{array}$$

$$\begin{array}{c}{\text{R}}_{\text{L}}\text{=}\text{1}/\left(\text{1+}{\text{K}}_{\text{L}}{\text{C}}_{\text{0}}\right) \left(\text{5}\right)\end{array}$$

In the formula, q_e was the equilibrium adsorption capacity, mg/g; q_m was the maximum adsorption capacity, mg/g; C_e was the mass concentration of the solution at equilibrium, mg/L; C₀ was the initial concentration of the solution, mg/L; R_L was the Langmuir adsorption Equilibrium constant; K_L was the coefficient of Langmuir adsorption isotherm equation, mg/L.

The correlation coefficients R² of the Langmuir model were 0.997 (25°C), 0.998 (35°C) and 0.999 (45°C), indicating the uniformity of the adsorbent surface (Table 3 and Fig. S3(B)). The maximum adsorption q_m were 588.235 mg/g (25°C), 625.000 mg/g (35°C) and 666.667 mg/g (45°C), respectively. And the experimental results were 575.248 mg/g (25°C), 610.037 mg/g (35°C) and 650.249 mg /g (45°C), respectively, which were similar to q_m. It showed that the Langmuir isotherm adsorption model was very suitable to describe the adsorption behavior of Sr-HAP on Pb(Ⅱ).

Table 3

Fitting parameters of Sr-HAP adsorption Pb²⁺ isothermal model
Temperature/°C	Langmuir model
Temperature/°C	q_m/(mg/g)	K_L/(L/mg)	R_L	R²
25	588.235	0.500	0.002 ~ 0.019	0.997
35	625.000	1.600	0.0006 ~ 0.006	0.998
45	666.667	1.667	0.0006 ~ 0.006	0.999

The adsorption equilibrium constant R_L, its value could indicate the adsorption performance of the material: unfavorable adsorption (R_L>1), favorable adsorption (0 < R_L<1), linear adsorption (R_L=1), irreversible adsorption (R_L=0). In Table 4, R_L was between 0 and 1, indicating that the adsorption process of Sr-HAP on Pb²⁺ was favorable.

3.6 Adsorption mechanism

The downward trend of the final pH change curve in the solution in Fig. 11 was very significant. The rapid removal of Pb²⁺, the high fixed amount of Pb²⁺ and the specific pH dependence strongly indicated that dissolution/precipitation was the main mechanism of Pb²⁺ removal (Chen et al., 1997). Relevant studies had also showed that the process of removing Pb²⁺ by apatite not only involved surface complex adsorption, ion exchange, and dissolution/precipitation, but also depended on the initial lead concentration (Guan et al., 2018). The possible accompanying mechanisms in the process of Sr-HAP removal of Pb²⁺ were: surface complexation, dissolution/precipitation, ion exchange and surface adsorption (Pham et al., 2013). When Pb²⁺ in the solution met Sr-HAP, Pb²⁺ quickly adsorbed to the pores on the surface of the adsorbent, and the functional group (-OH) on the surface of the adsorbent reacted with Pb²⁺ to release H⁺; part of the target adsorbed on the adsorbent Pb²⁺ reacted with calcium ions and strontium ions in Sr-HAP. When the adsorbent released H⁺, the pH of the solution decreased, Sr-HAP would partially dissolve and release PO₄⁻, and Pb²⁺ tended to form Pb₁₀(PO₄)₆(OH)₂ with phosphate. With the progress of the removal process, the available sites on the surface and inside of Sr-HAP were gradually fully occupied, the ionic components of the solution and the surface of the material tended to balance, and the entire removal process was basically completed.

Among them, the surface complexation mechanism could be expressed as:

$$\begin{array}{c}\text{≡}\text{P}\text{O}\text{H}\text{+}{\text{Pb}}_{\left(\text{aq}\right)}^{\text{2+}}\text{↔}\text{≡}\text{P}\text{O}{\text{Pb}}^{\text{+}}\text{+}{\text{H}}_{\left(\text{aq}\right)}^{\text{+}} \left(\text{6}\right)\end{array}$$

$$\begin{array}{c}\text{≡}\text{P}{\text{O}}^{\text{-}}\text{+}{\text{Pb}}_{\left(\text{aq}\right)}^{\text{2+}}\text{↔}\text{≡}\text{P}\text{O}{\text{Pb}}^{\text{+}} \left(\text{7}\right)\end{array}$$

$$\begin{array}{c}\text{≡}\text{C}\text{a}\text{O}\text{H}\text{+}{\text{Pb}}_{\left(\text{aq}\right)}^{\text{2+}}\text{↔}\text{≡}\text{C}\text{a}\text{O}{\text{Pb}}^{\text{+}}\text{+}{\text{H}}_{\left(\text{aq}\right)}^{\text{+}} \left(\text{8}\right)\end{array}$$

$$\begin{array}{c}\text{≡}\text{S}\text{r}\text{O}\text{H}\text{+}{\text{Pb}}_{\left(\text{aq}\right)}^{\text{2+}}\text{↔}\text{≡}\text{S}\text{r}\text{O}{\text{Pb}}^{\text{+}}\text{+}{\text{H}}_{\left(\text{aq}\right)}^{\text{+}} \left(\text{9}\right)\end{array}$$

The dissolution/precipitation mechanism was as follows:

Dissolve:

$$\begin{array}{c}{\text{Sr}}_{\text{5}}{\text{C}}_{\text{a}}{\left(\text{P}{\text{O}}_{\text{4}}\right)}_{\text{6}}{\left(\text{OH}\right)}_{\text{2}}\text{+}\text{14}{\text{H}}^{\text{+}}\text{→}\text{5}{\text{Ca}}^{\text{2+}}\text{+}\text{5}{\text{Sr}}^{\text{2+}}\text{+}\text{6}{\text{H}}_{\text{2}}\text{P}{\text{O}}_{\text{4}}^{\text{-}}\text{+}\text{2}{\text{H}}_{\text{2}}\text{O}\text{ }\text{ }\text{ }\left(\text{10}\right)\end{array}$$

Precipitation:

$$\begin{array}{c}\text{10}{\text{Pb}}^{\text{2+}}\text{+}\text{6}{\text{H}}_{\text{2}}{\text{PO}}_{\text{4}}^{\text{-}}\text{+}\text{2}{\text{H}}_{\text{2}}\text{O}\text{→}\text{14}{\text{H}}^{\text{+}}\text{+}{\text{Pb}}_{\text{10}}{\left({\text{PO}}_{\text{4}}\right)}_{\text{6}}{\left(\text{OH}\right)}_{\text{2}} \left(\text{11}\right)\end{array}$$

The ion exchange was:

$$\begin{array}{c}{\text{Sr}}_{\text{5}}{\text{Ca}}_{\text{5}}{\left({\text{PO}}_{\text{4}}\right)}_{\text{6}}{\left(\text{OH}\right)}_{\text{2}}\text{+}{\left(\text{x+y}\right)\text{Pb}}^{\text{2+}}\text{→}\left(\text{5-x}\right){\text{Ca}}^{\text{2+}}\text{+}{\left(\text{5}\text{-}\text{y}\right)\text{Sr}}^{\text{2+}}\text{+}{{\text{Sr}}_{\text{x}}\text{Ca}}_{\text{y}}{\text{Pb}}_{\left(\text{x+y}\right)}{\left({\text{PO}}_{\text{4}}\right)}_{\text{6}}{\left(\text{OH}\right)}_{\text{2}} \left(\text{12}\right)\end{array}$$

In this work, a strontium-doped hydroxyapatite (Sr-HAP) was synthesized and used as adsorbent to remove lead ions (Pb²⁺) from water. The results shown that strontium-doped hydroxyapatite Sr-HAP showed good adsorption capacity for Pb²⁺. When the dosage of Sr-HAP was 40 mg/50 mL, the removal rate of Pb²⁺ was above 98 % at different concentrations (100, 200, and 300 mg/L). And the initial concentration of Pb²⁺ increased from 100 mg/L to 800 mg/L, the removal amount of Pb²⁺ increased from 128.779, 128.481 and 128.474 mg/g to 651.175, 673.806 and 695.223 mg at 25, 35 and 45°C, respectively. Compared with HAP, the specific surface area of Sr-HAP has increased by 11.1 %, from 102.314 m²/g to 115.121 m²/g. The pore size of Sr-HAP is mostly distributed in 2-50 nm. Meanwhile, the adsorption-desorption experiments show that Pb²⁺ can be relatively stable and fixed in the material, and it is difficult to desorb. Surface complexation, ion exchange and surface adsorption are the main effects of Pb²⁺ adsorption onto Sr-HAP. In addition, co-precipitation and dissolution/precipitation may exist in the adsorption process of Pb²⁺.

Ethical Approval

Yes. Our study does not involve human subjects.

Consent to Participate

Yes.

Consent to Publish

Yes.

Authors Contributions

Zongqiang Zhu: Conceptualization, Methodology, Software, Investigation, Writing-Original Draft, Data Curation.

Huiling Jiang: Validation, Formal analysis, Visualization.

Yinian Zhu: Validation, Formal analysis, Visualization.

Lihao Zhang: Validation, Formal analysis, Visualization.

Shen Tang: Validation, Formal analysis, Visualization.

Xiaobin Zhou: Validation, Formal analysis, Visualization, Resources, Writing-Review & Editing, Supervision, Data Curation.

Yinming Fan: Validation, Formal analysis, Visualization, Resources, Writing-Review & Editing, Supervision, Data Curation.

Funding

This work was financially supported by the National Natural Science Foundation of China (Nos. 51978188, 42063003, 52160017), the Guangxi Natural Science Foundation (2018GXNSFAA050044), the Special Fund for Guangxi Distinguished Experts and Program for High Level Innovation Team and Outstanding Scholar of Universities in Guangxi (GuiCaiJiaoHan [2018] 319).

Competing Interests

We declare that we have no financial and personal relationships with other people or organizations that can inappropriately influence our work, there is no professional or other personal interest of any nature or kind in any product, service and company that could be construed as influencing the position presented in the manuscript.

Availability of data and materials

Not applicable.

Adeyemi NO, Atayese MO, Sakariyawo OS et al (2021) Alleviation of heavy metal stress by arbuscular mycorrhizal symbiosis in Glycine max (L.) grown in copper, lead and zinc contaminated soils. Rhizosph 18. https://doi.org/10.1016/j.rhisph.2021.100325
Aklil A, Mouflih M, Sebti S (2004) Removal of heavy metal ions from water by using calcined phosphate as a new adsorbent. J Hazard Mater 112:183–190
Alothman Z, Review A (2012) :Fundamental Aspects of Silicate Mesoporous Materials. Mater., 5:2874–2902
Cecil KM, Brubaker CJ, Adler CM et al (2008) Decreased brain volume in adults with childhood lead exposure. PLoS Med 5:0741–0749
Charerntanyarak L (1999) Heavy metals removal by chemical coagulation and precipitation. Water Sci Technol 39:135–138
Crini G (2005) Recent developments in polysaccharide-based materials used as adsorbents in wastewater treatment. Prog Poly Sci 30:38–70
Chong ML, Fang YZH M.Y., et al (2013) Effects of modified zeolite and addition of CaCl₂ and MgCl₂ on the adsorption characteristics of heavy metal ions Pb²⁺.Envir. Chem.,803–809
Chao L, Zong QZ, Shuang C et al (2017) Adsorption characteristics of copper in water by the porous biomorph-genetic composite of HAP/C with eucalyptus wood template. Envir. Sci., 2017, 38: 1074–1083
Chen XB, Wright JV, Conca JL et al (1997) Effects of pH on heavy metal sorption on mineral apatite. Envi Sci Tech 31:624–631
Du XD, Wang CC, Zhong J et al (2017) Highly efficient removal of Pb²⁺ by a polyoxomolybdate-based organic-inorganic hybrid material {(4-Hap)₄[Mo₈O₂₆]}. J Environ Chem Eng 5:1866–1873
Dabrowski A, Hubicki Z, Podkoscielny P et al (2004) Selective removal of the heavy metal ions from waters and industrial wastewaters by ion-exchange method. Chemosphere 56:91–106
Doyle FM, Liu Z (2003) The effect of triethylenetetraamine (Trien) on the ion flotation of Cu²⁺ and Ni²⁺. J Collo Inter Sci 258:396–403
Dong L, Zhu Z, Ma H et al (2010) Simultaneous adsorption of lead and cadmium on MnO^2-loaded resin. J Envir Sci 22:225–229
Fernane F, Boudia S, Aiouache F (2013) Removal Cu(II) and Ni(II) by natural and synthetic hydroxyapatites: a comparative study. Desalin Wat Treat 52:2856–2862
Guan DX, Ren C, Wang J et al (2018) Characterization of Lead Uptake by Nano-Sized Hydroxyapatite: A Molecular Scale Perspective. ACS Ear and Spa Chem 2:599–607
Hao L, Lv Y, Song H (2017) The morphological evolution of hydroxyapatite on high-efficiency Pb²⁺ removal and antibacterial activity. Microchem J 135:16–25
Imron MF, Kurniawan SB, Soegiant A (2019) Characterization of mercury-reducing potential bacteria isolated from Keputih non-active sanitary landfill leachate, Surabaya, Indonesia under different saline conditions. J Envir Manage 241:113–122
Juang RS, Shiau RC (2000) Metal removal from aqueous solutions using chitosan-enhanced membrane filtration. J Memb Sci 165:159–167
Kim HT, Lee TG (2017) A simultaneous stabilization and solidification of the top five most toxic heavy metals (Hg, Pb, As, Cr, and Cd). Chemosp 178:479–485
Kongsuwan A, Patnukao P, Pavasant P (2009) Binary component sorption of Cu(II) and Pb(II) with activated carbon from Eucalyptus camaldulensis Dehn bark. J Indus Eng Chem 15:465–470
Li JS, Mao LINK et al (2020) Qin Hengji. Effects of nano-Fe₃O₄ loaded acid-modified biochar on Pb²⁺ and Cd²⁺ adsorption in water. J Agric Res Envir 37:241–251
Lei Y, Guan JJ, Chen W et al (2015) Fabrication of hydroxyapatite/chitosan porous materials for Pb(ii) removal from aqueous solution. Rsc Adv 5:25462–25470
Lin KL, Pan JY, Chen YW et al (2009) Study the adsorption of phenol from aqueous solution on hydroxyapatite nanopowders. J Hazard Mater 161:231–240
Nwaehiri UL, Akwukwaegbu PI, Nwoke BEB (2020) Bacterial remediation of heavy metal polluted soil and effluent from paper mill industry. Environ Heal Toxicol 35:1–10
Purwanti IF, Obenu A, Tangahu BV et al (2020) Bioaugmentation of Vibrio alginolyticus in phytoremediation of aluminium-contaminated soil using Scirpus grossus and Thypa angustifolia. Heliyon 6:05004
Pedersen AJ (2003) Characterization and electrodialytic treatment of wood combustion fly ash for the removal of cadmium. Biom Bioe 25:447–458
Paulino AT, Guilherme MR, Reis AV et al (2007) Capacity of adsorption of Pb²⁺ and Ni²⁺ from aqueous solutions by chitosan produced from silkworm chrysalides in different degrees of deacetylation. J Hazard Mater 147:139–147
Pham MD, Tran ND, Nzihou A et al (2013) Hydroxyapatite gel for the improved removal of Pb²⁺ ions from aqueous solution. Chem Eng J 232:128–138
Rangabhashiyam S, Balasubramanian P (2019) Characteristics, performances, equilibrium and kinetic modeling aspects of heavy metal removal using algae. Biore Tech Rep 5:261–279
Sandrine B, Ange N, Didier BA et al (2007) Removal of aqueous lead ions by hydroxyapatites: equilibria and kinetic processes. J Hazard Mater 139:443–446
Wang Y, Hu L, Zhang G et al (2017) Removal of Pb(II) and methylene blue from aqueous solution by magnetic hydroxyapatite-immobilized oxidized multi-walled carbon nanotubes. J Colloid Interface Sci, 2017, 494: 380–388
Wen QT, Yong LF, Xiao ML (2013) Adsorption characteristics of Pb²⁺ from aqueous solution and synthetic silicon-substituted carbonate hydroxyapatite particles. China Envir Sci 33:1017–1024
Wei L, Qi Y, Bo L et al (2015) Adsorption of Cr(â…¥)on Magnetic Graphene from Aqueous Solution.Envir. Sci.,537–544
Wang S, Terdkiatburana T, Tadé MO (2008) Adsorption of Cu(II), Pb(II) and humic acid on natural zeolite tuff in single and binary systems. Separa Puri Tech 62:64–70
Yong CB (2018) Treatment of wastewater containing heavy metal ions by steel slage power. Envir Eng 36:125–127
Yang S, Zhao Y, Ding D et al (2013) An electrochemically modified novel tablet porous material developed as adsorbent for phosphate removal from aqueous solution. Chem Eng J 220:367–374
Yan Y, Dong X, Sun X et al (2014) Conversion of waste FGD gypsum into hydroxyapatite for removal of Pb²⁺ and Cd²⁺ from wastewater. J. Collo. Inter. Sci., 2014, 429: 68–76
Zhu Y, Jiang Y, Zhu Z et al (2018) Preparation of a porous hydroxyapatite-carbon composite with the bio-template of sugarcane top stems and its use for the Pb(II) removal. J Clean Produ 187:650–661
Zhu Z, Li L, Zhang H et al (2010) Adsorption of Lead and Cadmium on Ca-Deficient Hydroxyapatite. Separ Sci Tech 45:262–268

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Strontium-doped hydroxyapatite as adsorbent effectively to remove lead ions from water

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Abstract

Figures

1. Introduction

2. Materials And Methods

2.1 Adsorbent preparation and characterization

2.1.1 Preparation

2.1.2 Characterization

2.2 Adsorption experiments

2.3 Desorption experiments

3. Results And Discussion

3.1 Characterization of the adsorbent

3.2 The influence of adsorption conditions

3.2.1 Influence of adsorbent dosage

3.2.2 Influence of the initial pH value of the solution

3.2.3 Influence of the initial concentration of the solution

3.2.4 Influence of adsorption time

3.2.5 Influence of interfering substances

3.3 Adsorption-desorption experiment

3.4 Adsorption kinetics

3.5 Adsorption isotherm

3.6 Adsorption mechanism

4. Conclusions

Declarations

References

Table

Supplementary Files

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