Sorption of Pb(II) from Aqueous Solutions by Acid-Modiﬁed Clinoptilolite-Rich Tu ﬀ s with Di ﬀ erent Si / Al Ratios

: The removal of Pb(II) from aqueous solutions by acid-modiﬁed clinoptilolite-rich tu ﬀ was investigated in this work. Clinoptilolite-rich tu ﬀ samples were treated using H 2 SO 4 at di ﬀ erent concentrations. Prior to and following acid treatment, the samples were characterized using X-ray di ﬀ raction (XRD), scanning electron microscopy (SEM), energy dispersive X-ray spectroscopy (EDS), and Fourier-transform infrared spectroscopy (FTIR). The pH of the point of zero charge (pH PZC ) was also determined as part of this characterization. Batch studies were studied to investigate Pb(II) removal as a function of contact time, initial Pb(II) concentration, adsorbent dosage, and solution pH. The results of the XRD and SEM techniques showed that clinoptilolite is the main mineral of the non- and acid-treated natural zeolite samples. However, EDS analysis indicated that the Si / Al ratio increases as the exchangeable ions decrease with increasing acid concentrations. The optimum conditions for Pb(II) removal for samples with 4.37 ≤ Si / Al ≤ 7.9 were found to be as follows: 6–8, and adsorbent dose of 6 mg g − 1 ; whereas for acid-modiﬁed clinoptilolite-rich tu ﬀ s with 9.01 ≤ Si / Al ≤ 9.52, these conditions were as follows: Contact time of 1440 min, pH: 8–10, and adsorbent dose of 10 mg g − 1 . The experimental data were analyzed by kinetic and isotherms models. The results showed that the sorption of Pb(II) on samples with Si / Al ratios of 4.37, 5.31, and 7.91 were in agreement with the pseudo-second order and Langmuir isotherm with q m = 48.54, 37.04, and 14.99 mg g − 1 , respectively, while the kinetic data and isotherm for samples with 9.01 ≤ Si / Al ≤ 9.52 were found to ﬁt the pseudo-ﬁrst order and Freundlich model. and HCLI-0.2 M ( R 2 ≥ 0.99). Comparison of the Langmuir constants ( q m ) between non- and acid-modiﬁed clinoptilolite-rich tu ﬀ s, shows that Pb(II) removal decreases as Si / Al ratio increases ( q m = 48.54, 37.037, and 14.992 mg g − 1 for Si / Al = 4.37, 5.30, and 7.91, respectively). This result can be explained by the decrease in exchangeable ions from the zeolite network with the increase in sulfuric acid concentration due to the dealuminization of the zeolitic material or the greater competition of Pb(II) with the H + from the ion exchange sites in the zeolite network.


Introduction
Zeolites are hydrous and crystalline alumino-silicates minerals with a regular and microporous structure (pore size from 3.0 to 10 Å) [1]. Their frameworks are comprised of tetrahedral building units TO 4 (T = Si or Al), connected together by sharing one oxygen atom at the corners to form complex three-dimensional open networks of channels or cavities and cages (0.2-1.2 nm) occupied by H 2 O molecules and exchangeable cations [2]. The partial substitution of Si 4+ by Al 3+ in the zeolite framework produces a negative charge that is often neutralized by the exchangeable alkali and/or alkaline earth cations (Na + , K + , Ca 2+ , Sr 2+ , or Ba 2+ ) [3] or by H + [4] for acidic zeolite in order to maintain electrical neutrality throughout the structure. Natural zeolites exhibit interesting structural properties such as a large specific surface, flexibility, and well-defined micropores in molecular dimensions, as well as high chemical and physical stability (e.g., thermal and mechanical) [5], and sorption characteristics obtained by combined ion-exchange and molecular-sieve properties [6]. These outstanding properties allow these materials to offer promising applications in many sustainable processes such as wastewater treatment [7] and petroleum refining [8] as well in agriculture and construction [9]. Over the last decades, many researches have focused on clinoptilolite due to its attractive selectivity for toxic metal cations and organic substances [6,[10][11][12][13][14]. It was reported that the ion-exchange process in natural zeolites depends on factors related to adsorbate and adsorbent characteristics such as pH and temperature of the solution, ion concentration, crystal structure, and chemical composition of the zeolitic material [10,15]. The cation exchange capacity (CEC) of a zeolite is directly dependent on the Si/Al ratio [16]. This important property of the natural zeolites can be used to remove heavy metals form polluted water. For instance, the lead (Pb 2+ ) is a toxic metallic specie which is necessary to eliminate from the environment [17].
It was also reported that a variation in the SiO 2 /Al 2 O 3 ratio of the natural zeolites affects the hydrophilicity/hydrophobicity of this material, which is correlated with the adsorption, catalysis, and ion exchange properties [18,19]. Zeolite dealumination using acid as a chemical agent allows zeolites with a higher SiO 2 /Al 2 O 3 ratio to be obtained, which are usually used for the removal of organic hydrophobic contaminants in water [20] and volatile organic compounds in air [21]; whereas alkali treated zeolite using organic hydroxides [22] or sodium hydrate [16,23] can achieve materials with a lower SiO 2 /Al 2 O 3 ratio for the removal of ammonium [24] and heavy metals [6,25].
In previous studies [16,18,19,22,23,[25][26][27][28][29][30], the researches focused mainly on the effect of the SiO 2 /Al 2 O 3 ratio on the structural properties of zeolites. Moreover, these studies used synthesized zeolites [19,22,[26][27][28]30], whereas the only raw zeolites used being mordenite and clinoptilolite. Kussainova et al. [31] reported the Pb(II) removal from an acid media. However, there was no information in the literature about the adsorption behavior of this heavy metal using acid modified natural zeolites. Therefore, the main objective of this research was to know the correlation between the Si/Al ratio and the Pb(II) removal capacity from aqueous solutions using non-and acid-modified clinoptilolite-rich tuffs. In addition, the structural and morphological properties of the materials were studied by XRD, SEM, and IR techniques. This study was aimed to evaluate the influence of contact time, initial pH, adsorbent dosage, and initial concentration on the removal of Pb(II) from aqueous solutions.

Clinoptilolite-Rich Tuff and Conditioning
A clinoptilolite-rich tuff from Villa de los Reyes (San Luis Potosí State, Mexico) was sieved for a particle size of 40 mesh and washed at 25 • C several times with deionized water (18.2 MΩ cm −1 , resistivity) to remove any water-soluble impurities. After vacuum filtration, the clinoptilolite-rich tuff was dried at 80 • C overnight. This sample was named Nat-CLI. Acid treatment was carried out by treating the Nat-CLI with 0.1 M, 0.2 M, 0.5 M, and 1 M H 2 SO 4 (95-98%, J. T. Baker) with a liquid/solid ratio of 1:20. The suspensions were stirred at 90 • C for 4 h and then the solids were vacuum filtered and Appl. Sci. 2019, 9,2415 3 of 17 washed with excess deionized water until the pH of the rinse water was reduced to 6 approximately. Finally, the samples were dried at 80 • C for 12 h and denoted as HCLI-0.1M, HCLI-0.2 M, HCLI-0.5 M, and HCLI-1.0 M, respectively.
Deionized water was used to prepare synthetic stock solutions (500 mg L −1 ) of Pb(II) using PbCl 2 (98%, Sigma-Aldrich Co, St. Louis, MO, USA). The desired concentrations were prepared from the dilution of the stock solution using deionized water. The pH of each was adjusted with 0.1 M HCl or 0.1 M NaOH.

Samples Characterization
X-ray diffraction (XRD) analysis was carried out using an APD 2000 PRO X-Ray diffractometer with CuK α radiation. The patterns of the samples were recorded between 6 • and 60 • 2θ with a scanning speed of 0.025 degrees s −1 and step time of 10 s. Scanning electron micrographs were obtained with a HITACHI S-3400N scanning electron microscope (SEM) and the chemical composition was determined by energy-dispersive X-ray spectroscopy at different points on the surface of the samples. The working conditions of the SEM-EDS were 10 kV for the energy dispersive analysis, with a current intensity of 30 pA, and work distance of 10.6 mm. Fourier transformed infrared (FTIR) spectroscopy was conducted using a Nicolet Nexus 670 FT-IR infrared spectrometer from 4000 to 400 cm −1 with a resolution of 4 cm −1 in a KBr (spectroGrade, Merk KGaA, Darmstadt, Germany) wafer.
The pH PZC was obtained using batch equilibration technique [12]. The samples were introduced into Erlenmeyer flasks containing 0.01 M NaCl solution with a solid/solution ratio of 1:500. The initial pH was adjusted from 2 to 12 by adding a few drops of 0.5 M HCl or 0.1 M NaOH. The mixtures were shaken for 24 h at 24 • C and the final pH values of the filtered suspensions were measured using Thermo Scientific pH-meter (ORNION 3star pH Benchtop). Plots of pH initial vs. pH final were constructed and the pH PZC was determined from the intersection of these curves.

Kinetic
The kinetic behavior was determined using a batch technique. 10 mL of Pb(II) aqueous solution with a known initial concentration (100 mg L −1 ) were added to 0.10 g of non-and acid-modified clinoptilolite-rich tuffs in a 12 mL conical flask. The mixtures were agitated at 150 rpm in an orbital shaker at 25 • C for different time intervals (15-1440 min). At the end of the each contact time, the tubes were centrifuged at 4500 rpm for 2 min, and the adsorbent was removed by filtration. The concentration of remaining Pb(II) in the supernatants was analyzed using atomic absorption spectroscopy (Thermo Scientific iCE 3000 Series).
The sorbed amount of Pb(II) (q, mg g −1 ) was calculated by a mass balance relationship equation as follows: where V is the solution volume (mL), W is the amount of sorbent (g), C o and C t are the initial and final metal concentrations (mg L −1 ), at time t (min). In order to describe the kinetics of the adsorption process of Pb(II) by the zeolite materials, pseudo-first-order and pseudo-second-order models [32] were applied in this work. Lineal forms for these models are represented in Equations (2) and (3), respectively. ln(q e − q t ) = ln q e − k 1 t (2) where q t (mg g −1 ) and q e (mg g −1 ) are the amounts of metal ions adsorbed at time t and equilibrium, respectively. k 1 (min −1 ) and k 2 (g mg −1 min −1 ) are the pseudo-first order and pseudo-second order reaction rate constants, respectively.

Isotherm
The sorption isotherm experiments were carried out in 12 mL flasks at 25 • C by placing 0.10 g of non-and acid-modified clinoptilolite-rich tuffs with 10 mL of Pb(II) aqueous solution at different initial concentrations ranging from 10 to 500 mg L −1 . The mixtures were agitated in an orbital shaker at 150 rpm for 24 h and after this time, the samples were centrifuged at 4500 rpm for 2 min and the metal concentrations of the supernatants were analyzed by atomic absorption spectrometry (AAS).
In this work, Langmuir and Freundlich models were applied to describe the equilibrium metal adsorption. The lineal forms for these models are expressed by Equations (4) and (5), respectively [33].
ln q e = lnK F + 1 n lnC e where q m (mg g −1 ) is the monolayer adsorption capacity of the adsorbent, q e is the amount of solute adsorbed at equilibrium (mg g −1 ), C e (mg L −1 ) is the equilibrium solute concentration, K L (L mg −1 ) is the adsorption constant related to the energy of sorption. K F (mg g −1 )(L mg −1 ) 1/n and 1/n are empirical constants of the Freundlich isotherm. In order to compare the results obtained for the Pb(II) sorption by the different acid-modified zeolitic materials considering each initial concentration from the isotherm, the partition coefficients (PC) were calculated as follows: PC = q e mg g −1 /C e (µM)

Influence of Adsorbent Dosage and Effect of pH
The influence of adsorbent dosage on Pb(II) sorption was investigated by varying the concentration of the zeolitic materials from 1.0 to 10 mg L −1 , while keeping the initial concentration of the Pb(II) solution (500 mg L −1 ) and agitation time (24 h) constant. The Pb(II) concentration in the remaining solution was analyzed as explained above.
The effect of pH on Pb(II) sorption by non-and acid-modified clinoptilolite-rich tuffs was studied in a pH range of 2-10. pH was adjusted with the dropwise addition of 0.1 M HCl or 0.01 M NaOH. Sorption experiments were conducted using 100 mg L −1 of Pb(II) with 0.10 g of sorbent and a contact time of 24 h. Figure 1 shows the X-ray diffraction (XRD) patterns of non-and acid-modified clinoptilolite-rich tuffs. It can be observed that all samples contained clinoptilolite as the principal mineral which is defined by its main characteristic peaks at 2θ = 9.85 • , 11 [34]. Furthermore, comparative analysis of the XRD patterns did not show significant changes in the positions of most of the reflexions before and after the acid treatments. However, a decrease can be observed in the relative intensity of the reflexions of HCLI with the increase in sulfuric acid concentration from 0.1 to 1 M. These results suggest that the treatment of natural zeolite with H 2 SO 4 caused some structural changes and loss of crystallinity in the clinoptilolite phase. This can be attributed to dealumination of the clinoptilolite as verified by the EDS analyses [29,35]. Wang et al. [36] present the reaction mechanism which take place when the clinoptilolite-rich tuffs are in contact with a strong acid media (HNO 3 ). Basically they mention the break of the primary structure of the zeolite network and the formation of Al(NO 3 ) 3 because the dealumination of the zeolitic material. In the present work could be a similar mechanism with the formation of Al 2 (SO 4 ) 3 . The scanning electron microscopy (SEM) images of non-and acid-modified clinoptilolite-rich tuffs are shown in Figure 2. It can be observed that all of the samples presented similar morphologies and also showed the tabular and coffin shapes typical of heulandite/clinoptilolite crystals [13]. In other investigations, the morphology of synthetic zeolite as chitosan had changed [37], whereas it remained unchanged for mordenite after chemical treatment [38]. This indicates that the change in morphology of the natural zeolites with acid treatment depends on the nature as well as the crystalline structure of these zeolitic materials. Table 1 shows the average chemical composition expressed in terms of % weight as obtained by EDS at different points on the surface of non-and acid-modified clinoptilolite-rich tuffs. As expected, the chemical composition changed drastically following the acid treatment of natural zeolite. A total elimination of Na + and Mg 2+ can be observed from the extra-framework of natural zeolite when the concentrations of H2SO4 are 0.1M and 0.2M, respectively; whereas the amount of Ca 2+ and K + ions decrease with increasing sulfuric acid concentration as result of H + exchange or by the loss of Al 3+ from the zeolite network showing that its cation exchange capacity was lost. It is clear that the H + is preferentially exchanged by Na + and Mg 2+ from the zeolitic network rather than Ca 2+ and K + and this depends of the exchangeable sites that they occupy into the zeolite network [39]. It is important to mention that the Fe could be as an associated mineral (Fe2O3) of the zeolitic rock.

Characterization
Moreover, a significant increase in the Si/Al ratio with increasing acid concentration can be observed, due to the Al leaching from the clinoptilolite framework and the consequent loss of its crystallinity. This result is consistent with those observed by XRD analysis. The scanning electron microscopy (SEM) images of non-and acid-modified clinoptilolite-rich tuffs are shown in Figure 2. It can be observed that all of the samples presented similar morphologies and also showed the tabular and coffin shapes typical of heulandite/clinoptilolite crystals [13]. In other investigations, the morphology of synthetic zeolite as chitosan had changed [37], whereas it remained unchanged for mordenite after chemical treatment [38]. This indicates that the change in morphology of the natural zeolites with acid treatment depends on the nature as well as the crystalline structure of these zeolitic materials. Table 1 shows the average chemical composition expressed in terms of % weight as obtained by EDS at different points on the surface of non-and acid-modified clinoptilolite-rich tuffs. As expected, the chemical composition changed drastically following the acid treatment of natural zeolite. A total elimination of Na + and Mg 2+ can be observed from the extra-framework of natural zeolite when the concentrations of H 2 SO 4 are 0.1M and 0.2M, respectively; whereas the amount of Ca 2+ and K + ions decrease with increasing sulfuric acid concentration as result of H + exchange or by the loss of Al 3+ from the zeolite network showing that its cation exchange capacity was lost. It is clear that the H + is preferentially exchanged by Na + and Mg 2+ from the zeolitic network rather than Ca 2+ and K + and this depends of the exchangeable sites that they occupy into the zeolite network [39]. It is important to mention that the Fe could be as an associated mineral (Fe 2 O 3 ) of the zeolitic rock.
Moreover, a significant increase in the Si/Al ratio with increasing acid concentration can be observed, due to the Al leaching from the clinoptilolite framework and the consequent loss of its crystallinity. This result is consistent with those observed by XRD analysis. Figure 3 shows the IR spectra of the non-and acid-modified clinoptilolite-rich tuffs. The spectra have been shifted by 5% from each other starting from that corresponding to Nat-CLI in order to better appreciate the differences in the absorption bands of all the materials. A broad band appears at 3446 cm −1 for Nat-CLI, corresponding to the symmetric and asymmetric stretching vibration of the O-H groups, while the bending of the H-O-H vibration occurs at 1636 cm −1 . The strong and broad band at 1051 cm −1 was assigned to the asymmetric stretching of Si-O; (from SiO 4 ). The weak bands of the symmetric stretching vibration observed for the O-Si-O bonds occur at 792 cm −1 . A weak signal can be observed at 608 cm −1 corresponding to bending vibrations between tetrahedrons, particularly to double ring vibrations. Finally, the band at 457 cm −1 corresponds to bending vibration T-O, T=Al, Si, and is related to the pore opening. These results agree well with those reported by other authors [40].   Figure 3 shows the IR spectra of the non-and acid-modified clinoptilolite-rich tuffs. The spectra have been shifted by 5% from each other starting from that corresponding to Nat-CLI in order to better appreciate the differences in the absorption bands of all the materials. A broad band appears at 3446 cm −1 for Nat-CLI, corresponding to the symmetric and asymmetric stretching vibration of the O-H groups, while the bending of the H-О-H vibration occurs at 1636 cm −1 . The strong and broad band at 1051 cm −1 was assigned to the asymmetric stretching of Si-О (from SiO4). The weak bands of the symmetric stretching vibration observed for the О-Si-О bonds occur at 792 cm −1 . A weak signal   clinoptilolite was treated with the H2SO4 solutions indicating the break of the bond Al-O-from the zeolite network [41]. As the concentration of the acid increases (0.1 M to 1.0 M) the band due to vibration of the O-H groups, is widened and is shifted at 3436 cm −1 . Additionally, the band at 1051 cm −1 is the most affected, this also appears widened and shifted to 1079 cm -1 , this behavior has been attributed to a decrease in the relative content of Si/Al ratio in zeolite framework with the increase in the length of the bonds Si-O-Si(Al), and with the formation of structured effects. The other band affected is the corresponding to bending vibration T-O, T=Al, Si, this is shifted to 462 cm −1 , which indicates a decrease of pore opening. The observed behavior suggests that the dealumination of Nat-CLI take place after its acid treatment.

Point of Zero Charge pH (pHPZC)
As shown in Figure 4, the pHPZC values before and after the acid-modified clinoptilolite-rich tuffs were 6.00 ± 0.01, and 3.00 ± 0.01, respectively. The decrease in pHPZC is the result of the chemical modification of Nat-CLI by the acid treatment. Therefore, the acid-modified clinoptilolite-rich tuffs present a large pH range where the surface charge of the materials is negative (3-12), as compared to Nat-CLI (6)(7)(8)(9)(10)(11)(12). On the other hand, the chemical treatment promoted a decrease in the negative charge, and consequently, a decrease in the number of cations and the generation of very strong Lewis acid sites on the surface of the H-CLI samples. In this sense, Rožić et al. [39] reported that the following reaction took place >Al-OH + H + → >AlOH2 + after the contact of the clinoptilolite-rich tuff with strong acid solutions. As the concentration of the acid increases (0.1 M to 1.0 M) the band due to vibration of the O-H groups, is widened and is shifted at 3436 cm −1 . Additionally, the band at 1051 cm −1 is the most affected, this also appears widened and shifted to 1079 cm -1 , this behavior has been attributed to a decrease in the relative content of Si/Al ratio in zeolite framework with the increase in the length of the bonds Si-O-Si(Al), and with the formation of structured effects. The other band affected is the corresponding to bending vibration T-O, T=Al, Si, this is shifted to 462 cm −1 , which indicates a decrease of pore opening. The observed behavior suggests that the dealumination of Nat-CLI take place after its acid treatment.

Point of Zero Charge pH (pHPZC)
As shown in Figure 4, the pH PZC values before and after the acid-modified clinoptilolite-rich tuffs were 6.00 ± 0.01, and 3.00 ± 0.01, respectively. The decrease in pH PZC is the result of the chemical modification of Nat-CLI by the acid treatment. Therefore, the acid-modified clinoptilolite-rich tuffs present a large pH range where the surface charge of the materials is negative (3-12), as compared to Nat-CLI (6-12). On the other hand, the chemical treatment promoted a decrease in the negative charge, and consequently, a decrease in the number of cations and the generation of very strong Lewis acid sites on the surface of the H-CLI samples. In this sense, Rožić et al. [39] reported that the following reaction took place >Al-OH + H + → >AlOH 2 + after the contact of the clinoptilolite-rich tuff with strong acid solutions.

Kinetic
Figure 5a presents the effect of contact time on the amount of Pb(II) sorbed by non-and acidmodified clinoptilolite-rich tuffs. It can be observed that the time to reach the adsorption equilibrium increases with increasing Si/Al ratios. This can be explained by the presence of a low number of vacant active binding sites as a consequence of the Si/Al ratio increment which limited the exchange of Pb(II) by the exchangeable ions from the zeolite network. Moreover, Figure 5a shows that the kinetic processes for the Nat-CLI, HCLI-0.1 M and HCLI-0.2 M sorbents consist of two main reaction stages: An initial fast adsorption process followed by a slow continuous sorption reaction; whereas a continuous uptake of Pb(II) from aqueous solution is observed over time for HCLI-0.5 M and HCLI-1.0 M.
The results presented in Figure 5b show are similar in maximum Pb(II) uptake when the Si/Al ratio ranges from 4.3 to 7.91 shows. However, the sorption of Pb(II) decreased drastically as Si/Al ratio increased. This can be attributed to the decrease in exchangeable ions during the acid-treatment (Table 1) and higher competition between H + from the acid-modified clinoptilolite-rich-tuffs and Pb(II) in the solutions.

Kinetic
Figure 5a presents the effect of contact time on the amount of Pb(II) sorbed by non-and acidmodified clinoptilolite-rich tuffs. It can be observed that the time to reach the adsorption equilibrium increases with increasing Si/Al ratios. This can be explained by the presence of a low number of vacant active binding sites as a consequence of the Si/Al ratio increment which limited the exchange of Pb(II) by the exchangeable ions from the zeolite network. Moreover, Figure 5a shows that the kinetic processes for the Nat-CLI, HCLI-0.1 M and HCLI-0.2 M sorbents consist of two main reaction stages: An initial fast adsorption process followed by a slow continuous sorption reaction; whereas a continuous uptake of Pb(II) from aqueous solution is observed over time for HCLI-0.5 M and HCLI-1.0 M.
The results presented in Figure 5b show are similar in maximum Pb(II) uptake when the Si/Al ratio ranges from 4.3 to 7.91 shows. However, the sorption of Pb(II) decreased drastically as Si/Al ratio increased. This can be attributed to the decrease in exchangeable ions during the acid-treatment (Table 1) and higher competition between H + from the acid-modified clinoptilolite-rich-tuffs and  The results presented in Figure 5b show are similar in maximum Pb(II) uptake when the Si/Al ratio ranges from 4.3 to 7.91 shows. However, the sorption of Pb(II) decreased drastically as Si/Al ratio increased. This can be attributed to the decrease in exchangeable ions during the acid-treatment (Table 1) and higher competition between H + from the acid-modified clinoptilolite-rich-tuffs and Pb(II) in the solutions.
The data for Pb(II) sorption by non-and acid-modified clinoptilolite-rich tuffs were fitted to pseudo-first order, pseudo-second order, and intra-particle diffusion models. As shown in Table 2, the kinetic parameters corresponding to the sorption of Pb(II) by Nat-CLI, HCLI-0.1M, and HCLI-0.2 M fit best to the pseudo-second order model (R 2 ≥ 0.99 and q cal ≈ q exp ), while for the HCLI-0.5 M and HCLI-1.0 M sorbents, the results show that the pseudo-first order model was more suitable for describing the kinetic process. Table 2. Kinetic parameters for Pb(II) sorption by non-and acid-modified clinoptilolite-rich tuffs obtained by applying the pseudo-second order and pseudo-first order models.
The k 2 diminished when the Si/Al ratio increased. However, the k 1 slightly increased when the Si/Al increased as well. These results shows that the dealumination of the acid-modified clinoptilolite-rich tuff has repercussion on the kinetic of the Pb(II) sorption and the Si/Al ratio will determine the order of the kinetic reaction (pseudo-second order or pseudo-first order). Figure 6 shows Pb(II) sorption by non-and acid-modified clinoptilolite-rich tuffs. It can be observed that the isotherms for the HCLI-0.1 M and HCLI-0.2 M were similar to those for Nat-CLI; whereas the isotherms for the Pb(II) sorption for HCLI-0.5 M and HCLI-1.0 M presented a different behavior. These differences can be attributed to the dealumination of the clinoptilolite-rich tuff after the treatment with 0.5 and 1.0 M H 2 SO 4 solutions. The amount of aluminum in HCLI-0.5 M and HCLI-1.0 M were 1.7 and 2 times lower than Nat-CLI, respectively. In these cases, it is probably that the Ca 2+ and K + from the zeolitic materials had played a role on the Pb(II) sorption.

Isotherm
The experimental equilibrium data for Pb(II) sorption by non-and acid-modified clinoptilolite-rich tuffs were fitted to the Langmuir and Freundlich models (Equations (4) and (5)) with the purpose of describing the sorption mechanism [6]. The values of the isotherm constants, including the determination coefficients (R 2 ), are summarized in Table 3. It was found that the Langmuir model better described Pb(II) uptake by Nat-CLI, HCLI-0.1 M, and HCLI-0.2 M (R 2 ≥ 0.99). Comparison of the Langmuir constants (q m ) between non-and acid-modified clinoptilolite-rich tuffs, shows that Pb(II) removal decreases as Si/Al ratio increases (q m = 48.54, 37.037, and 14.992 mg g −1 for Si/Al = 4.37, 5.30, and 7.91, respectively). This result can be explained by the decrease in exchangeable ions from the zeolite network with the increase in sulfuric acid concentration due to the dealuminization of the zeolitic material or the greater competition of Pb(II) with the H + from the ion exchange sites in the zeolite network. Appl. Sci. 2019, 9, x FOR PEER REVIEW 10 of 18 The experimental equilibrium data for Pb(II) sorption by non-and acid-modified clinoptiloliterich tuffs were fitted to the Langmuir and Freundlich models (Equations 4 and 5) with the purpose of describing the sorption mechanism [6]. The values of the isotherm constants, including the determination coefficients (R 2 ), are summarized in Table 3. It was found that the Langmuir model better described Pb(II) uptake by Nat-CLI, HCLI-0.1 M, and HCLI-0.2 M (R 2 ≥ 0.99). Comparison of the Langmuir constants (qm) between non-and acid-modified clinoptilolite-rich tuffs, shows that Pb(II) removal decreases as Si/Al ratio increases (qm = 48.54, 37.037, and 14.992 mg g −1 for Si/Al = 4.37, 5.30, and 7.91, respectively). This result can be explained by the decrease in exchangeable ions from the zeolite network with the increase in sulfuric acid concentration due to the dealuminization of the zeolitic material or the greater competition of Pb(II) with the H + from the ion exchange sites in the zeolite network.
For sorbents with Si/Al ratios of 9.05 and 9.52, the Freundlich isotherm adjusted better to the Pb(II) uptake (0.99 ≥ R 2 ≥ 0.98) than the Langmuir isotherm. This suggests that sorption did not occur on a uniform site; instead a multilayer sorption occurred in this system. Moreover, n values were between 1 and 10, indicating that the physical sorption of Pb(II) ions by acid-modified clinoptiloliterich tuffs with Si/Al ratios of 9.04 and 9.5 is favorable [14]. It is important to mention that a unique mechanism is not involved on the sorption processes. For instance, the speciation of the heavy metal in solution, the surface charge of the sorbent as well as the nature and structure of the zeolitic materials could be some of the factors which are involved on the ion exchange or adsorption mechanisms.  For sorbents with Si/Al ratios of 9.05 and 9.52, the Freundlich isotherm adjusted better to the Pb(II) uptake (0.99 ≥ R 2 ≥ 0.98) than the Langmuir isotherm. This suggests that sorption did not occur on a uniform site; instead a multilayer sorption occurred in this system. Moreover, n values were between 1 and 10, indicating that the physical sorption of Pb(II) ions by acid-modified clinoptilolite-rich tuffs with Si/Al ratios of 9.04 and 9.5 is favorable [14]. It is important to mention that a unique mechanism is not involved on the sorption processes. For instance, the speciation of the heavy metal in solution, the surface charge of the sorbent as well as the nature and structure of the zeolitic materials could be some of the factors which are involved on the ion exchange or adsorption mechanisms. Table 4 shows the maximum sorption capacity (q m ) of different materials for Pb(II) obtained from the Langmuir model at different initial concentrations, as was obtained in the present work. It can be observed that carbon foam has 10 times more sorption capacity than clinoptilolite-rich tuff for this heavy metal. However, the zeolitic material has 1.9 times more capacity to remove Pb(II) than pectin-containing substances. In the case of the acid-modified clinoptilolite-rich tuff the maximum sorption capacity is 5.1 times lower than that of the unmodified material. Other carbonaceous materials among them biochar, multi-walled carbon nanotubes functionalized by selenophosphoryl groups, and sultone-modified magnetic activated carbon, present similar maximum sorption capacities for Pb(II).

Partition Coefficients
It was found that the partition coefficients have an inverse behavior with respect the Pb(II) initial concentration in aqueous solution, this means that the PC is high when the initial concentration is lower (Figure 7), because there are less Pb(II) in the solution to be sorbed in the preferential sites of the zeolitic networks before the saturation. It is important to mention that the determination coefficient (R 2 ) was 0.1468. In Figure 8 it is observed that the Si/Al ratios of the clinoptilolite-rich tuffs notably affect the partition coefficients as a function of the initial concentration of the chemical specie of the heavy metal in solution. For this reason, it is important to consider the ion exchange sites and the adsorption sites in each zeolite network before and after the treatment with 1.0 M of H2SO4 solutions for the removal of different chemical species from polluted water. In Figure 8 it is observed that the Si/Al ratios of the clinoptilolite-rich tuffs notably affect the partition coefficients as a function of the initial concentration of the chemical specie of the heavy metal in solution. For this reason, it is important to consider the ion exchange sites and the adsorption sites in each zeolite network before and after the treatment with 1.0 M of H 2 SO 4 solutions for the removal of different chemical species from polluted water.

Influence of Adsorbent Dosage
The percentage of Pb(II) uptake with respect to the non-and acid-modified clinoptilolite-rich tuffs dosages (mg L −1 ) is shown in Figure 9. It can observed that the dosage of all of the sorbents used is a very important parameter for Pb(II) uptake. The increment in sorption capacity was attributed to a higher number of available sorption sites [12]. In the case of sorbents with Si/Al ratios of 4.37, 5.31, and 7.91, sorption efficiency was highly dependent on increase in the sorbents dosage of solution up to 6 mg L −1 when the maximum Pb(II) uptake percentage (99.99%) was achieved. For sorbents with a Si/Al ratio of 9.05 and 9.52, Pb(II) removal increases gradually with dosage, reaching 92.36% and 35.99%, respectively. One explanation of this behavior is the few available sorption sites in the zeolite networks of HCLI-0.5 M and HCLI-1.0 M to be occupied by the Pb(II) chemical species. The H-CLI1.0 M has the lowest amount of Al in the zeolitic material diminishing the ion exchange capacity. The amount of Ca 2+ in the zeolitic material also drastically diminished, limiting the ion exchenge: Pb 2+ (s) + Ca 2+ (z) → Pb 2+ (z) + Ca 2+ (s), where s represents the solution and z the zeolitic material. The sites where the K + are located into the zeolite network [47] do not favor the ion exchange with Pb 2+ . Another explanation of this behavior is by the change of the nature of the zeolitic material from hydrophilic to hydrophobic character [36].

Influence of Adsorbent Dosage
The percentage of Pb(II) uptake with respect to the non-and acid-modified clinoptilolite-rich tuffs dosages (mg L −1 ) is shown in Figure 9. It can observed that the dosage of all of the sorbents used is a very important parameter for Pb(II) uptake. The increment in sorption capacity was attributed to a higher number of available sorption sites [12]. In the case of sorbents with Si/Al ratios of 4.37, 5.31, and 7.91, sorption efficiency was highly dependent on increase in the sorbents dosage of solution up to 6 mg L −1 when the maximum Pb(II) uptake percentage (99.99%) was achieved. For sorbents with a Si/Al ratio of 9.05 and 9.52, Pb(II) removal increases gradually with dosage, reaching 92.36% and 35.99%, respectively. One explanation of this behavior is the few available sorption sites in the zeolite networks of HCLI-0.5 M and HCLI-1.0 M to be occupied by the Pb(II) chemical species. The H-CLI1.0 M has the lowest amount of Al in the zeolitic material diminishing the ion exchange capacity. The amount of Ca 2+ in the zeolitic material also drastically diminished, limiting the ion exchenge: Pb 2+ (s) + Ca 2+ (z) → Pb 2+ (z) + Ca 2+ (s) , where s represents the solution and z the zeolitic material. The sites where the K + are located into the zeolite network [47] do not favor the ion exchange with Pb 2+ . Another explanation of this behavior is by the change of the nature of the zeolitic material from hydrophilic to hydrophobic character [36].

Effect of the Solution pH
The effect of initial solution pH on Pb(II) sorption by non-and acid-modified clinoptilolite-rich tuffs is illustrated in Figure 10. It can be observed that the curves presented in Figure 10a, all show a similar shape. For the first group, the removal percentage of Pb(II) increased by only 1.08%, 8.60%, and 9.47 % in the pH range of 2-6, when the Si/Al ratio was 4.37, 5.31, and 7.91, respectively, while the maximum Pb(II) uptake percentage reached 97.92% to 99.58% for these sorbents at a solution pH of about 5. The lower removal capacity of Pb(II) at pH = 2 is due to a lower selectivity of clinoptilolite at a low pH and the strong competition between the Pb 2+ and the H + for the exchangeable sites on the clinoptilolite framework. The decrease in Pb(II) sorption at higher pH values can be explained by the formation of hydrolyzed Pb(II) species (Kps of Pb(OH)2 = 1.43 × 10 −20 ) which could precipitate or by the electrostatic repulsion between a possible negative Pb(II) species (Pb(OH)3 − or Pb(OH)4 2− ) and the negatively charged surface of the HCLI-0.5 M and HCLI-1.0 M zeolitic materials [48].
For the second group (Figure 10b), which corresponds to sorbents with a Si/Al ratio of 9.05 and 9.52, solution pH has a significant influence on the Pb(II) uptake percentage. In the pH range of 2-6, the removal percentage of Pb(II) increases progressively from 22.22% to 82.02% and from 19.48% to 67.48% for Si/Al ratio of 9.05 and 9.53, respectively. At higher pH values, Pb(II) sorption increased continuously up to 93% and 70% for HCLI-0.5 M and HCLI-1.0 M, respectively. This can be attributed to the high acidity of these materials which has a beneficial effect, disfavoring the hydrolysis of the metal ions in the solution [12].

Effect of the Solution pH
The effect of initial solution pH on Pb(II) sorption by non-and acid-modified clinoptilolite-rich tuffs is illustrated in Figure 10. It can be observed that the curves presented in Figure 10a, all show a similar shape. For the first group, the removal percentage of Pb(II) increased by only 1.08%, 8.60%, and 9.47% in the pH range of 2-6, when the Si/Al ratio was 4.37, 5.31, and 7.91, respectively, while the maximum Pb(II) uptake percentage reached 97.92% to 99.58% for these sorbents at a solution pH of about 5. The lower removal capacity of Pb(II) at pH = 2 is due to a lower selectivity of clinoptilolite at a low pH and the strong competition between the Pb 2+ and the H + for the exchangeable sites on the clinoptilolite framework. The decrease in Pb(II) sorption at higher pH values can be explained by the formation of hydrolyzed Pb(II) species (K ps of Pb(OH) 2 = 1.43 × 10 −20 ) which could precipitate or by the electrostatic repulsion between a possible negative Pb(II) species (Pb(OH) 3 − or Pb(OH) 4 2− ) and the negatively charged surface of the HCLI-0.5 M and HCLI-1.0 M zeolitic materials [48].
FT-IR spectra also are in good agreement with the about reduction of crystallinity of HCL due to the dealumination of Nat-CLI that take place after its acid treatment.